Cadmium toxicity to embryonic–larval development and survival in red sea bream Pagrus major

ARTICLE IN PRESS Ecotoxicology and Environmental Safety 72 (2009) 1966–1974

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Cadmium toxicity to embryonic–larval development and survival in red sea bream Pagrus major Liang Cao a,b, Wei Huang a,b, Xiujuan Shan a,b, Zhizhong Xiao a, Qiyao Wang a, Shuozeng Dou a, a b

Institute of Oceanology, Chinese Academy of Sciences, Qingdao, 266071, PR China Graduate School, Chinese Academy of Sciences, Beijing, 100049, PR China

a r t i c l e in f o

a b s t r a c t

Article history: Received 27 February 2009 Received in revised form 27 May 2009 Accepted 7 June 2009 Available online 1 July 2009

At 18 1C and 33 psu, 24 and 48 h LC50 values of cadmium (Cd) for red sea bream Pagrus major embryos were 9.8 and 6.6 mg l1, respectively, while 24, 48, 72, and 96 h LC50 values for larvae were 18.9, 16.2, 8.0, and 5.6 mg l1, respectively, indicating that embryos were more sensitive to Cd toxicity than larvae. Cd concentrations at Z0.8 mg l1 led to low hatchability (0–90% in Z0.8 mg l1 solutions vs. 97–100% in lower ones), delay in time to hatch, high mortality (38–100% vs. 1–10%), morphological abnormality (42–100% vs. 1–10%), reduced length (3.55–3.60 vs. 3.71–3.72 mm) in the embryos and larvae. They were Cd concentration dependent and potential biological significant endpoints for assessing the risk of Cd to aquatic organisms. Heart beat and yolk absorption of the larvae were significantly inhibited at some high concentrations but they were not as sensitive as other endpoints to Cd exposure. & 2008 Elsevier Inc. All rights reserved.

Keywords: Cd Toxic bioassay LC50 Embryos and larvae Bioindicator

1. Introduction Cadmium is a widespread aquatic environmental pollutant which is associated with a broad spectrum of human activities and products such as plastics, ceramics, glass, and vehicle tires (Thompson and Bannigan, 2008). It has been well documented that cadmium at excessive amount in aquatic environments may affect the functions in digestive, immune, and reproductive organs of fishes (Leonard, 1979; Miliou et al., 1998; Dutta and Kaviraj, 2001; Thophon et al., 2003; Hallare et al., 2005). Moreover, cadmium is commonly considered as a fish neurotoxic substance and a blood circulating system disruptor due to its competition with intracellular calcium (Karen et al., 2001; Fraysse et al., 2006). Exposure to cadmium could induce toxic effects on various biological processes of fishes (e.g. delay in ontogenetic development, low hatchability, high morphological abnormality), and could even cause individual death directly (Witeska et al., 1995; Williams and Holdway, 2000; Jezierska et al., 2009). Since fish embryos and larvae are particularly sensitive to toxicity of cadmium, which may subsequently affect the recruitment and population wellness of the next cohorts, they are widely used as toxicity test animals in risk assessment on aquatic environments and fishery resources (Witeska et al., 1995; Chandra and KhudaBukhsh, 2004; Gonza lez-Doncel et al., 2004). For instance, the embryonic–larval development is used as biological significant

 Corresponding author. Fax: +86 532 82898925.

E-mail address: [email protected] (S. Dou). 0147-6513/$ - see front matter & 2008 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2009.06.002

endpoint to evaluate effects of pollutants, to investigate the mechanisms of toxic effects on larval survival and growth, and to predict long-term heavy metal toxicity to fish population in the wild (Miliou et al., 1998; Blechinger et al., 2002; Hallare et al., 2005; Thompson and Bannigan, 2008). Red sea bream Pagrus major used to be a popular commercial species in East Asian countries. However, the wild population and resources have rapidly declined in the past decades. Many estuaries and coastal waters of the Bohai Sea, their main spawning and nursery grounds, are polluted with high concentrations of heavy metals such as cadmium, copper, zinc and lead (Li et al., 1996). The range of cadmium concentration in this region has been recently recorded for many localities including the Laizhou Bay (0.01–2.02 mg l1), the Bohai Bay (0.01–0.59 mg l1), and the Liaodong Bay (0.02–16.1 mg l1) (Zhang, 2001; Peng et al., 2009). Heavy metal toxicity to the embryos and larvae is thus considered as one potential cause for wild-population decline. Although a few studies have dealt with the toxic effects of chemicals such as selenium (Se) and 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) on their development (Takayanagi, 2001; Yamauchi et al., 2006), the mechanism by which heavy metals like cadmium cause biological damages to red sea bream during their early life stage (ELS) is still poorly understood. This study served two main purposes: (1) determining the median lethal concentrations (LC50) of Cd on the embryos and larvae and then investigate their sensitivity to Cd toxicity and (2) studying the subchronic toxic effects of waterborne Cd on embryonic–larval development and survival, and then determining the potential bioindicators for assessing cadmium toxicity to

ARTICLE IN PRESS L. Cao et al. / Ecotoxicology and Environmental Safety 72 (2009) 1966–1974

aquatic organisms. To achieve these goals, we investigated the hatching process, yolk absorption, embryonic and larval mortality, morphological abnormality, heart rate, and growth of red sea bream exposed to different Cd concentrations.

2. Materials and methods 2.1. Egg production and collection Eggs produced by natural fertilization were obtained from a fish laboratory of the Institute of Oceanology, Chinese Academy of Sciences, Qingdao, China. Twenty males and ten females (3–4 kg in body weight, 30–40 cm in body length) were reared in a 12 m3 indoor ponds with flowing seawater and aeration. Water temperature was kept at 1871 1C with thermostats. A photoperiod of 14L:10D was provided during fish rearing. The fish was fed with cooked mussel and oyster in the morning every day. Red sea bream commonly courts and chases in pairing in the upper waters immediately before releasing eggs and sperm. It was thus possible to determine the spawning and fertilization time according to their spawning behavior. Floating eggs were collected from the ponds using a screen net 1 h after spawning and were then kept in 20 L incubators at an estimated density of 800 eggs l1 at 1871 1C for another hour. During this time, the normally developing embryos floated in the surface water, while the unfertilized or dead ones deposited at the bottoms of the incubators. Only floating viable embryos were used for toxicity tests. The tests were carried out in April–June 2008, during which red sea bream spontaneously spawned.

2.2. Toxicity tests 2.2.1. Test protocol All tests were carried out in 1000 ml acid-rinsed glass beakers. The beakers were placed randomly in a laboratory with a constant room temperature of 1871 1C by air-conditioning. A photoperiod of 14L:10D was provided throughout the tests. Each beaker was filled with 1000 ml test solution. Water quality parameters of the test seawater were: hardness, 183.273.6 mg l1 as CaCO3; pH, 8.270.1; dissolved oxygen concentration, 7.570.2 mg l1; 3371 psu (practical salinity unit). Analytical reagent CdCl2  2.5H2O (purity 499.9%; Sinopharm Chemical Reagent Co. Ltd., Beijing, China) was used as test chemical for the test solutions. A stock solution of 1000 mg Cd2+ l1 was prepared with deionized water, which was diluted to produce desired Cd concentrations of test solutions for each treatment. Thereafter, 10 l desired solution was prepared for each concentration and 1000 ml was transferred into each of the four replicate experimental beakers (four replicates). The test solutions were not renewed and no aeration was provided throughout the toxicity tests.

1967

contamination of the test solutions. The test was terminated at 96 h post-hatching (hph) when larvae opened mouth and were ready to initiate feeding. In the first set of test, five biological parameters of the embryos and larvae were investigated. (1) Accumulative hatchability was defined as the percentage of the accumulative number of larvae hatched during the test to the number of initially stocked embryos (100) in each beaker. This was achieved by counting and calculating the number of dead embryos every 4 h after exposed to test solutions until no surviving embryos remained in the beakers. (2) Hatching rate was defined as the percentage of larvae that were hatched at four intervals between four consecutive time points (48, 52, 60, and 96 hpf) to the accumulative number of hatched larvae during the test. It was used to investigate the Cd toxic effects on time to hatch. The number of larvae that hatched every day was determined by counting and calculating the dead and viable embryos or larvae at these time intervals. No larva was hatched in any treatment after 96 hpf, the defined deadline of hatching. (3) Accumulative mortality was defined as the percentage of dead embryos and larvae throughout the test to the number of initially stocked embryos (100) in each beaker. This was done by counting and totaling the dead embryos and larvae every day. (4) Morphological abnormality of the larvae was examined by checking the dead larvae using a stereoscopic microscope (Nikon-SMZ1000, digital video camera, Nikon-DS 5M; Software, ACT-2U; Tokyo, Japan) every day after hatching during the test. Abnormally developed survivors at the termination of each treatment, if any, were also checked and included. It was defined as the percentage of the total number of abnormally developed larvae to the accumulative number of larvae hatched during the test. (5) Total length (LT): at the termination of the test, surviving larvae were sampled and sacrificed by drying them with absorptive paper. They were then freshly measured for LT using a stereoscopic microscope (Nikon-SMZ1000). Coefficient of variation (CV) of LT in each test solution was calculated to investigate the effect of Cd on individual size variation in the larvae: CV ¼ (S.D./mean LT)  100. Preliminary tests showed that heart beat was detected in the 28 hpf embryos. In the second set of the test, four viable embryos (38 hpf) or larvae (48, 72, and 96 hpf, respectively) were sampled randomly from each experimental beaker to investigate Cd toxic effects on heart beat. To examine the response of yolk absorption to Cd, larvae were sampled randomly from each experimental beaker at 52 and 80 hpf. A treatment was terminated wherever no surviving larvae were available for sampling. The sampled embryos and larvae were freshly videotaped and photographed to investigate their heart rate and yolk absorption using a microscope video system (microscope, Nikon-ECLIPSE 50i and Nikon-SMZ1000; digital video camera, Nikon-DS 5M; Software, ACT-2U). (1) Heart rate of each sampled embryo or larva was determined by counting a 1 min slow back-playing video image frame acquired by the digital video camera. (2) Yolk-sac size (V, mm3) was measured according to the following formula: V ¼ p*a*b2/6, where a and b (mm) are the major axis and minor axis of yolk sac, respectively. Yolk absorption rate (R) was defined by: R ¼ (V0Vt)t1, where V0 and Vt are the initial and final sac yolk sizes, respectively, while t is the lapsing time in hour.

2.3. Chemical analysis 2.2.2. Acute toxicity tests on embryos and larvae One test was run to determine the lethal concentration to cause 50% mortality (24, 48 h LC50) in embryos. Embryos were exposed to Cd solutions of seven nominal concentrations (in quadruplicates): 0 (control), 1, 2, 4, 8, 16, 32 mg Cd2+ l1, respectively. Other experimental conditions were identical to the experimental protocol. Cd exposure was initiated within 3 h post-fertilization (hpf). Embryos were considered dead when they sank to the bottoms of the beakers. Dead embryos were counted and removed every 4 h to monitor embryonic mortality. The test lasted 48 h. Another acute toxicity test was carried out to determine the LC50 (24, 48, 72, and 96 h) of Cd for larvae. The experimental designing, including the nominal Cd concentrations and the ambient settings of the test solutions, was identical to that in the acute toxicity test on embryos except that the embryos were replaced with 100 newly hatched larvae in each beaker at the start of the test. Larvae were considered dead when they deposited onto the beaker bottoms and no mobility was observed. Dead larvae were counted and removed every 4 h after being introduced into the beakers to record the larval mortality during the test.

2.2.3. Toxicity test on embryonic–larval development and survival This test was to investigate Cd toxicity to the embryonic–larval development, survival, and growth. Based on LC50 values of Cd for embryos and larvae derived from the acute toxicity tests, nominal concentrations for Cd-exposure solutions in this test were set as follows: 0 (control), 0.2, 0.4, 0.8, 1.2, 1.8, 2.4, and 3.2 mg Cd2+ l1. Other experimental settings were identical to those in the acute toxicity tests. 100 fertilized eggs were counted randomly and transferred to each of the experimental beakers within 3 hpf. One set of test was conducted to determine the Cd toxic effects on the morphological development, survival, and growth of the embryos and larvae. Another set of test was run for sampling to investigate their yolk absorption and heart rate in different test solutions. Dead embryos and larvae were counted and removed every 4 h to record mortality and prevent

For chemical analysis, test solutions in the acute and subchronic toxicity tests were sampled as follows: since the four replicate solutions at each concentration were of the same source (i.e. the 10 l desired solution) and all the experimental beakers were acid rinsed prior to test setting, one sample from the designed solution at each concentration was taken at the beginning of the test to represent the initial concentrations of the four replicates; at the end of the tests, solutions were sampled from each experimental beaker. Cd concentrations of the samples were measured using inductively coupled plasma mass spectrometry (ICP-MS). The error was assessed as the percentage of the absolute difference of the measured and nominal concentrations to the nominal concentration of each test solution.

2.4. Statistical analysis Kolmogorov–Smirnov and Levene tests were used to assess the conditions for analysis of variance (ANOVA) testing (Zar, 1999). The data of heat beat, total length and yolk absorption met the two assumptions (normality and homogeneity of variance) for ANOVA. Percentage data (hatchability, mortality, and abnormality) were normalised by using arcsine square roots transformation to do ANOVA. The Tukey test was used for post hoc multiple comparisons between means. Difference was considered significant at po0.05. LC50 values were determined using a probit analysis method (Finney, 1971). The no-observed-effect concentration (NOEC), lowest-observed-effect concentration (LOEC) and maximum acceptable toxicant concentration (MATC, the geometric mean of the NOEC and LOEC) were determined by hypothesis testing of biological parameters in the subchronic toxicity test. The acute-to-chronic ratios (ACR) were determined according to the following formula: ACR ¼ 96 h LC50/MATC. All statistical analyses were done using SPSS 13.0 for Windows (SPSS Inc., Chicago, IL, USA). The 25% and 50% inhibition concentrations (IC25 and IC50) were determined using linear interpolation technique (Norberg-King, 1993).

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3. Results

1

1

Except for the 2.4 mg l (measured to be 2.56 mg l ) solutions, measured Cd concentrations were close to nominal concentrations of other solutions. Errors are 5.0–8.7% and 0.3–13.3% in the acute and the subchronic toxicity tests, respectively (Table 1). 24 and 48 h LC50 values (95% confidence limits) for embryos were 9.8 (8.5–11.3) and 6.6 (5.9–7.3) mg l1, respectively, while 24, 48, 72 and 96 h LC50 values for larvae were 18.9 (16.9–21.4), 16.2 (14.6–18.1), 8.0 (7.1–9.1) and 5.6 (5.0–6.2) mg l1, respectively. Embryos had lower 24 and 48 h LC50 values for Cd toxicity than larvae.

Accumulative hatchability (%)

120 3.1. LC50 of Cd for embryos and larvae

100

a

Acute toxicity tests

Sub-chronic toxicity test

Nominal

Measured

Error (%)

Nominal

Measured

Error (%)

0 1 2 4 8 16 32

0.00970.002 1.0770.01 2.1270.05 3.8070.25 7.4870.81 14.6171.55 30.4173.39

7.0 5.8 5.0 6.5 8.7 5.0

0 0.2 0.4 0.8 1.2 1.8 2.4 3.2

0.00970.002 0.2070.04 0.3970.03 0.7570.03 1.1870.05 1.7770.28 2.0870.68 3.1370.39

0.3 2.5 5.4 1.3 1.9 13.3 2.3

b c d

60

e

40 20 0

3.2. Cd toxicity to embryonic–larval development and survival

Table 1 . Nominal and measured cadmium concentrations (mg l1) of the test solutions (mean7S.D., n ¼ 5).

a

80

0

0.2

0.4

0.8

1.2

1.8

2.4

3.2

a

a

2.4

3.2

a

a

Accumulative mortality (%)

120 100 b

80 c 60 d 40 20 f

ef

e

0 0

0.2

0.4

0.8

1.2

1.8

120 Morphological abnormality (%)

Increasing Cd concentration had significant inhibitory effects on accumulative hatchability (ANOVA, po0.05; Fig. 1A). 97–100% of the embryos were hatched in the r0.4 mg l1 solutions, in comparison with 90%, 69%, 60%, and 46% in the 0.8, 1.2, 1.8, and 2.4 mg l1 solutions, respectively. No larvae were hatched in the 3.2 mg l1 solutions. Accumulative hatchability of the embryos in Z0.8 mg l1 solutions was all significantly lower than that in the controls (ANOVA, po0.05 for each comparison). MATC for accumulative hatchability was 0.57 mg l1, while the ACR was 9.90. IC25 and IC50 for hatchability were 1.14 and 1.76 mg l1, respectively (Table 2). In the r0.4 mg l1 solutions, most of the larvae were hatched within 48 hpf (98%, 98%, 95% in the controls, 0.2, 0.4 mg l1 solutions), in contrast with 2–4% for 48–52 hpf (Fig. 2). In Z0.8 mg l1 solutions, however, hatching rate decreased significantly as Cd concentration increased (85%, 62%, 31%, and 26% in the 0.8, 1.2, 1.8, and 2.4 mg l1 solutions, respectively). In contrast, the percentages of larvae hatched during the periods 48–52 hpf (9–41%), 52–60 hpf (5–24%), and 60–96 hpf (2–11%) increased noticeably with increasing Cd concentration in Z0.8 mg l1 solutions (Fig. 2). In 1.8 and 2.4 mg l1 solutions, 10–11% of the larvae were not hatched until 60–96 hpf. Accumulative mortality of the embryos and larvae significantly increased as Cd concentration was elevated (ANOVA, po0.05; Fig. 1B). It did not significantly differ between the controls (1%) and 0.2 mg l1 (4%) treatments but was significantly higher in Z0.4 mg l1 solutions (10%, 38%, 59%, 74%, 100%, and 100% in 0.4, 0.8, 1.2, 1.8, 2.4, and 3.2 mg l1 solutions, respectively) than in the controls (ANOVA, po0.05 for each comparison; Fig. 1B). All the embryos died within 68 hpf in the 3.2 mg l1 solutions and no survivors remained in the 2.4 mg l1 solutions at the end of the test. MATC for accumulative mortality was 0.28 mg l1, while ACR for accumulative mortality was 19.80. 144 h IC25 and IC50 for

a

100 c

b

80 60 d 40 20

e f

f

0

0.2

0 0.4 0.8 1.2 1.8 2.4 Nominal Cd concentration (mg l-1)

3.2

Fig. 1. Accumulative hatchability (A), accumulative mortality (B), and accumulative morphological abnormality (C) of red sea bream embryos exposed to different concentrations of cadmium (mean7S.D., n ¼ 4; ANOVA, Tukey test, treatments sharing the same superscript letters were not significantly different at po0.05).

accumulative mortality were 0.59 and 0.93 mg l1, respectively (Table 2). Morphological abnormalities observed during the embryonic stage were manifested as spastic body contractions, trembling and hyperactivity within the chorion, edema, blastodermal lesions and undeveloped tail (Fig. 3b–e). After hatching, cardiac edema, degenerated and hooked tails, fins lesions and spinal curvature were commonly observed in the abnormally developed larvae (Fig. 3g–o). Among them, skeletal deformity was the most pronounced morphological alteration (Fig. 3k–o). Only 1–3% abnormally developed embryos and larvae were observed in the controls and the 0.2 mg l1 solutions-, in contrast with 10%, 42%, 79%, 89% and 100% in the 0.4, 0.8, 1.2, 1.8, and 2.4 mg l1 treatments, respectively (Fig. 1C). Morphological abnormalities in Z0.4 mg l1 solutions were significantly higher than in the

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Table. 2 Summary of subchronic Cd toxicity data for red sea bream embryos and larvae. Endpoints

NOEC (mg l1)

LOEC (mg l1)

MATC (mg l1)

ACR

IC25 (mg l1)

IC50 (mg l1)

Accumulative hatchability Accumulative mortality Morphological abnormality Heart ratea Total length

0.4 0.2 0.2 1.8 0.4

0.8 0.4 0.4 2.4 0.8

0.57 0.28 0.28 2.1 0.57

9.9 19.8 19.8 2.7 9.9

1.14 0.59 0.55

1.76 0.93 0.8

b

b

b

b

a b

Newly hatched larvae (48 hpf). Data not available.

100 48 hpf 52 hpf 60 hpf 96hpf

Hatching rate (%)

80

60

repressed in 1.8 (5.97  103 mm3 h1) and 2.4 (5.78  103 mm3 h1) mg l1 solutions compared to the controls (7.56  103 mm3 h1; ANOVA, po0.05 for each comparison), while no significant differences were found among o1.8 mg l1 solutions (7.41, 7.53, 7.45, 7.09  103 mm3 h1 in 0.2, 0.4, 0.8, 1.2 mg l1 solutions, respectively; ANOVA, p40.05 for each comparison; Fig. 5B).

40 4. Discussion

20

4.1. Acute toxicity of Cd to fish

0 0

0.2 0.4 0.8 1.2 1.8 Nominal Cd concentration (mg l-1)

2.4

Fig. 2. Hatching rate (mean7S.D., n ¼ 4) of red sea bream embryos exposed to different concentrations of cadmium at four intervals between four consective time points 48, 52, 60, and 96 hpf.

controls (ANOVA, po0.05 for each comparison). Particularly, no embryo or larva normally developed in 2.4 mg l1 solutions. MATC for morphological abnormality was 0.28 mg l1 and ACR for morphological abnormality was 19.80. 144 h IC25 and IC50 for morphological abnormality were 0.55 and 0.80 mg l1, respectively (Table 2). At 38 hpf, there were no significant differences in embryonic heart rate (64–70 beats min1; ANOVA, p40.05 for each comparison) among treatments (Fig. 4A). At 48 hpf, heart rate of the newly hatched larvae in the 2.4 mg l1 solutions (56 beats min1) was significantly lower than in others (72–78 beats min1) (Fig. 4B). MATC for heart beat rate of newly hatched larvae was 2.10 mg l1 and ACR was 2.70 (Table 2). At 72 hpf, heart rate of the larvae was significantly repressed in Z0.8 mg l1 solutions (110–113 beats min1) in comparison with that in the controls (122 beats min1; po0.05 for each comparison), but it did not significantly differ among o0.8 mg l1 solutions (119–122 beats min1; p40.05 for each comparison; Fig. 4C). Similarly, at 96 hpf, larvae in Z0.8 mg l1 solutions exhibited significantly repressed heart rate (125–139 beats min1) in comparison with those in the controls (147 beats min1; ANOVA, po0.05 for each comparison; Fig. 4D). Cd exposure significantly affected the LT of the larvae at the end of the test (ANOVA, po0.05; Fig. 5A). LT of the larvae was significantly less in Z0.8 mg l1 solutions (3.60, 3.56, and 3.55 in 0.8, 1.2, and 1.8 mg l1 solutions, respectively) than in the controls (3.72 mm) (ANOVA, po0.05 for each comparison), while no significant differences were found among o0.8 mg l1 ones (3.71 and 3.72 mm in 0.2 and 0.4 mg l1 solutions, respectively; ANOVA, p40.05 for each comparison). CV of LT in each Cd-exposed solutions (0.45, 0.56, 0.41, 0.79, 0.53 in 0.2, 0.4, 0.8, 1.2, 1.8 mg l1 solutions, respectively) was apparently larger than in the controls (0.15). MATC for total length was 0.57 mg l1 and ACR was 9.90 (Table 2). Yolk absorption rate during 52–80 hpf was significantly

Literature of acute toxicity of Cd to fish species as of 1985 had been critically reviewed and summarized in USPEA (1985). Studies indicated that Cd toxicity to fish was species specific and fish exhibited different ability to tolerate Cd (USEPA, 1985; Appendix 1). For instance, 48 h LC50 values for adult fish fluctuated considerably from 12.0 mg l1 in Atlantic silverside Menidia menidia (Middaugh and Dean, 1977) to 1101.5 and 1775.6 mg l1 in yellow perch Perca flavescens and rainbow trout Oncorhynchus mykiss, respectively (Lacroix and Hontela, 2004). Similarly, 96 h LC50 values of Cd for fish juveniles varied from r0.0008 mg l1 for three trout species (rainbow trout, Birceanu et al., 2008; cutthroat trout O. clarki, Harper et al., 2008; steelhead trout Salmo gairdneri, Cusimano et al., 1986) to 20.1 and 21.1 mg l1 in white seabass Lates calcarifer (Thophon et al., 2003) and common carp Cyprinus ¨ e nyi and Szakolczai, 1993), respectively. Generally, L50 carpio (Sov of Cd for fish embryos and larvae was less investigated than for juveniles and adults (USEPA, 1985; Appendix 1). Nonetheless, it was recorded in some fish species such as zebrafish Danio rerio (Cheng et al., 2000; Blechinger et al., 2002; Hallare et al., 2005; Chan et al., 2006), Atlantic silverside (Middaugh and Dean, 1977), common carp (Suresh et al., 1993), rainbow trout (Pascoe et al., 1986), guppy Poecilia reticulate (Miliou et al., 1998), tilapia Oreochromis mossambicus (Hwang et al., 1996), spotted rainbow fish Melanotaenia fluviatilis (Williams and Holdway, 2000), and mummichog Fundulus heteroclitus (Middaugh and Dean, 1977). In the current study, 24 and 48 h LC50 values of Cd for red sea bream embryos were 9.8 and 6.6 mg l1 respectively, which were lower than those reported for zebrafish embryos (24 h: 24.1 mg l1; 48 h: 30.1–46.8 mg l1 at 26–33 1C; Hallare et al., 2005; Chan et al., 2006). 48 h (16.2 mg l1) and 96 h (5.6 mg l1) LC50 values of Cd for red sea bream larvae fell within the reported values for other fish larvae (Appendix 1). The sensitivity of fish to Cd toxicity could change with developmental stage. Fishes generally become more tolerant to Cd toxicity as they develop and grow. This has been demonstrated by the Cd LC50 data for various fishes (Appendix 1). For example, 50% of the common carp larvae and juveniles could be killed within 96 h due to Cd exposure at 4.3 and 17.1–21.1 mg l1, respectively, and in contrast the needed concentration is as high ¨ e nyi and Szakolczai, 1993; as 121.8 mg l1 for the adults (Sov Suresh et al., 1993; Muley et al., 2000). However, the tolerance of

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Fig. 3. Morphological abnormalities of red sea bream exposed to cadmium. (a) normally developed embryo (36 hpf, control); (b) embryo with blastodermal lesions (40 hpf , 2.4 mg Cd2+ l1); (c) spastic body contractions in the chorion (32 hpf, 3.2 mg Cd2+ l1); (d) embryo with cardiac edema (44 hpf, 2.4 mg Cd2+ l1); (e) embryo with undeveloped tail (44 hpf, 1.2 mg Cd2 l1); (f) normal larva (1 dph, control); (g) larva with cardiac edema (68 hpf, 2.4 mg Cd2+ l1); (h) and (i) degenerated tail (64 hpf, 1.8 mg Cd2+ l1 and 64 hpf, 1.2 mg Cd2+ l1, respectively); (j) larva with hooked tail (144 hpf, 1.2 mg Cd2+ l1); (k) fins lesions and C-shaped spine (60 hpf, 2.4 mg Cd2+ l1); (l) L-shaped larva (60 hpf , 1.8 mg Cd2+ l1); (m) V-shaped larva (64 hpf, 1.8 mg Cd2+ l1); (n) S-shaped larva (64 hpf, 1.2 mg Cd2+ l1); (o) larva with helical spine (72 hpf, 2.4 mg Cd2+ l1).

fish larvae to Cd toxicity may not follow well this case because they undergo various drastic physiological changes such as functioning of organs and metamorphosis at this stage. These biological processes could cause changes in Cd2+ uptake efficiency and subsequently bring profound impacts on the sensitivity of fish larvae to Cd toxicity (Hwang et al., 1996; Chang et al., 1997). Take Atlantic silverside and mummichog for examples, under the same rearing conditions at 30 psu, the ability of newly hatched larvae to tolerate Cd toxicity could be 3 times as strong as the 14 dph ones (Middaugh and Dean, 1977). When it comes to comparing the sensitivity of fish embryos and larvae to chemical toxicity, larval stage is generally believed to have greater sensitivity than embryonic stage, presuming that chorions functioned to prevent free passage of pollutants to embryos while newly hatched larvae are directly exposed to potential toxicants (Fent and Meier, 1994; Gaikowski et al., 1996; Williams and Holdway, 2000; Johnson et al., 2007; Jezierska et al., 2009). However, a few studies reported that embryos were more sensitive than the newly

hatched larvae (Watling, 1982; Marty et al., 1990; Hamm and Hinton, 2000). The current study indicated that red sea bream embryos were more sensitive to Cd toxicity than newly hatched larvae, similar to what was obtained at zinc exposure (Huang et al., 2009). The acute toxicity of Cd to fish is largely affected by aquatic environmental factors. Salinity usually plays a critical role in determining the acute toxicity of Cd to marine fishes. For instance, Voyer (1975) found that the 96 h LC50 of Cd to juvenile mummichog at 30 psu was about one-half what it was at 10 and 20 psu. In contrast, decreasing the salinity could increase the sensitivity of newly hatched mummichog and Atlantic silverside larvae to Cd (Middaugh and Dean, 1977). Moreover, it was also related to other water quality parameters such as temperature, hardness and pH. 96 h LC50 values of Cd to steelhead trout juveniles increased by 600% as pH decreased from 7.0 to 4.7 (Cusimano et al., 1986), while 48 h LC50 value of Cd to zebrafish embryos at 33 1C (46.8 mg l1) was 10 times as high as that at

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160 3.8

140 Total length (LT, mm)

120 100 80 60 40

a

ab

ab b

3.7

c

3.6

c

3.5 3.4

20 0 0

0.2

0.4

0.8

1.2

1.8

2.4

3.3

3.2

0

0.2

a

a

0.4

0.8

1.2

1.8

140 120 100 80

a

a

a

a

a

a b

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Fig. 5. Final total length (mean7S.D., n ¼ 16; A) and yolk absorption (mean7S.D., n ¼ 16; B) in red sea bream larvae exposed to different concentrations of cadmium (ANOVA, Tukey test, treatments sharing the same superscript letters were not significantly different at po0.05).

80 60 40 20 0 160

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et al., 1986). Like most of the studies listed in Appendix 1, the acute toxicity of Cd to red sea bream was investigated in an aquatic environment (183 mg l1 as CaCO3 in hardness, pH 8.2, aerated dissolved oxygen; 33 psu) optimal for embryonic–larval survival. In the wild, however, interaction of various environmental cues could doubtlessly complicate the process of Cd toxicity to this fish. Therefore, particular care should be taken on the aquatic ambient elements as well as the biological processes of the tested fishes when we use LC50 values of Cd for evaluating aquatic pollution or risk assessment on environments.

100 4.2. Toxic effects of Cd on embryonic–larval development and survival of fish

80 60 40 20 0 0

0.2 0.4 0.8 1.2 Nominal Cd concentration (mg l-1)

1.8

Fig. 4. Heart rate (mean7S.D., n ¼ 16) of red sea bream embryos at 38 hpf (A) and larvae at 48 (B), 72 (C), and 96 hpf (D) at different concentrations of cadmium (ANOVA, Tukey test, treatments sharing the same superscript letters were not significantly different at po0.05).

21 1C (4.8 mg l1) (Hallare et al., 2005). It was also reported that 96 h LC50 of Cd to juvenile rainbow trout at a hardness of as CaCo3 70 mg l1 was about one-half what it was at 280 mg l1 (Pascoe

Since hatching is particularly sensitive to environmental toxicity, hatchability and time to hatch are often used as biological endpoints to investigate chemical toxicity to the ELS of fish (Fraysse et al., 2006; Jezierska et al., 2009). Early literature demonstrated that Cd exposure could reduce the hatchability or cause delay in the time to hatch in fishes such as zebrafish (Fraysse et al., 2006) and spotted rainbow fish (Williams and Holdway, 2000). In common carp, Cd at a concentration of 0.01 mg l1 could remarkably reduce their hatchability (Witeska et al., 1995), while over 40% Atlantic salmon Salmo salar embryos could not be hatched when they were exposed to 0.27 mg l1 Cd solutions (Rombough and Garside, 1982). In this study, accumulative hatchability of red sea bream embryos was significantly reduced in Z0.8 mg l1 solutions and no larvae were hatched in

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3.2 mg l1 ones. Similarly, Cd exposure at concentrations Z0.8 mg l1 could delay the time to hatch with over 10% larvae hatched during 60–96 hpf in 1.8 and 2.4 mg l1 solutions, in comparison with 95–98% larvae hatched within 48 hpf in r0.4 mg l1 ones. Hatching in fish results from an interaction of biochemical (enzymatic), biophysical (mechanical) and osmotic mechanisms. During this process, hatching glands produce choionase which is necessary for egg shell disintegration, while hatching enzymes digest the chorions. With the help of twisting, embryos tear up the chorions and become free pro-larvae (De Gaspar et al., 1999; Fraysse et al., 2006; Jezierska et al., 2009). Like in other fishes, water borne cadmium over a certain level (i.e. 0.8 mg l1) might inhibit the hatching enzyme activities of red sea bream embryos and thus decrease their ability to break through the chorions, which leads to low hatchability or delay in hatching. The findings indicated that hatchability and time to hatch of red sea bream embryos were sensitive to Cd exposure and could be reliable bioindicators for evaluating cadmium toxicity to aquatic organismss. Accumulative mortality in red sea bream embryos and larvae exhibited a Cd concentration-dependent trend. It was significantly higher in Z0.4 mg l1 solutions than in the controls. At the end of the test, 59–74% of the embryos and larvae died in 1.2 and 1.8 mg l1 solutions and no survivors remained in 2.4 and 3.2 mg l1 ones. During the first 24 h after exposure, embryonic mortality was low and varied in a narrow range among treatments. The large majority of mortality occurred at late embryonic stage and during the period between post-hatching to swim-up of newly hatched larvae (data not shown). Due to the protection of egg shells, the amount of Cd that penetrated into the embryos at early embryonic stage might be limited. This protection became weak in late embryonic stage or totally disappeared as the shells broke at hatching, and this subsequently led to accumulation of Cd in the embryos and larvae and caused the cease of development or high mortality. Rombough and Garside (1982) also reported that mortality peaks of Atlantic salmon embryos exposed to Cd occurred during development of vitelline blood circulation and just before hatching. In contrast, Jezierska et al. (2009) found that mortality in common carp embryos exposed to metals (Cd, Cu, Pb) reached a peak in the early embryonic stages just after fertilization due to their particular sensitivity to metal intoxication. Diverse conclusions have also been reached about the effects of heavy metals on the embryonic–larval mortality in other fishes (Villalobos et al., 2000; Jezierska and Witeska, 2001; Jezierska et al., 2009). These studies suggested that the sensitivity of fish embryos and larvae to metal intoxication, to a large extent, depended on fish species, developmental stages, pollutants or elemental cues. Embryonic–larval mortality of red sea bream was Cd concentration dependent and sensitive and could be effective at assessing Cd toxicity to organisms in the Bohai Sea, where Cd concentration reached as high as 2 mg l1 in some estuaries or coastal waters (Zhang, 2001; Peng et al., 2009). Exposure to Cd could have adverse effects on morphological development of various fish species. For instance, common carp eggs exposed to 0.02 mg Cd2+ l1 could lead to 47% spinal deformity in the embryos (Witeska et al., 1995). So far, a number of morphological abnormalities in fish embryos and larvae due to Cd exposure have been reported: blastodermal lesions, malformed yolk sac, exogastrulation, cardiac edema, spinal deformity, deformed skull and eye, altered axial curvature, lack or hooked tail, helical bodies, blistering of fins, etc. (Cheng et al., 2000; Williams and Holdway, 2000; Wong and Wong, 2000; Hallare et al., 2005; Jezierska et al., 2009). In this study, morphological abnormalities such as pericardial edema, blastodermal lesions, trembling and hyperactivity within the chorions were observed in

the late embryonic stage. At 48 hpf, the accumulative abnormalities were 71% and 99% in the 1.8 and 2.4 mg l1 solutions, respectively, while no abnormally developed embryos and larvae were observed in the controls. At the end of the tests, morphological abnormalities in Z0.4 mg l1 solutions were significantly higher than in others. These included cardiac edema, degenerated and hooked tail, fins lesions, body shrinkage and spinal curvature in the larvae. Of them, spinal curvature was most pronounced with 80% and 99% occurring in 1.8 and 2.4 mg l1 solutions, which suggested that skeleton might be a target for Cd toxicity in red sea bream larvae. Like in other fish larvae such as zebrafish (Cheng et al., 2000), tilapia (Wong and Wong, 2000) and common carp (Jezierska et al., 2009), water borne Cd might cause a reduction in Ca2+-ATPase activity or myosin and myotome formation necessary for the normal development of a healthy musculoskeletal system. This could lead to reduced Ca2+ uptake and induced skeleton curvature in the larvae. Since morphological abnormality was sensitive to Cd exposure, it could also be an effective index to investigate Cd toxicity in aquatic environments. Heart rate could be also used as a biological endpoint to evaluate metal toxicity to the ELS of fishes because disturbances in heart beat can induce negative consequences in various biological processes such as metabolism, growth and survival. In zebrafish embryos, heart rate was remarkably promoted by increasing copper concentrations (Johnson et al., 2007), while it was repressed in walleye Stizostedion vitreum larvae exposed to mercury (Latif et al., 2001). Several studies demonstrated that exposure to Cd affected the cardiovascular function of fish embryos and repressed their hear beat in zebrafish, common carp, garpike Belone belone and grass carp Ctenopharyngodon idella (Westernhagen et al., 1975; Jezierska et al., 2002; Hallare et al., 2005). In this study, Cd exposure did not significantly affect the heart rate of red sea bream embryos (38 hpf). However, heart rate of the 48 hpf larvae was noticeably lower in 2.4 mg l1 solutions than in others, while that of the older larvae (72–96 hpf) was significantly repressed in Z0.8 mg l1 solutions. Huang et al. (2009) found that increasing Zn concentration (0–2.5 mg l1) could promote the heart rate of the 7 dph red sea bream larvae, but did not cause significant heart beat disturbances of the 38 hpf embryos. These findings suggested that both Cd and Zn exposures could more easily induce heart beat disturbances in larvae than embryos. Heart beat disturbances at ELS of fish can be induced by various causes such as underdeveloped heart, neurological disturbances or inhibition of acetylcholinesterase (Hallare et al., 2006). Mechanistically, cardiac contraction is considered to be activated mainly by Ca2+ entry through the cell membrane, while Cd is a Ca2+ antagonist that disturbs Ca2+ uptake (Morad et al., 1981). Cd at concentrations Z0.8 mg l1 could impair the heart development of red sea bream larvae (e.g. pericardial edema and visceral hemorrhage), reduce Ca2+ uptake or even induce disruption of nervous functioning, which leads to repressed heart beat in the larvae. At the end of the test, total length (LT) of the larvae was significantly less in Z0.8 mg l1 solutions (3.55–3.60 mm) than in the controls (3.72 mm), while individual size variation (CV) in Cd-exposed larvae (0.41–0.79) was noticeably larger than in the controls (0.15). Similar results were obtained when red sea bream embryos and larvae were exposed to TCDD (Yamauchi et al., 2006) and zinc (Huang et al., 2009). These agreed with the finding that retarded growth was commonly observed in chemical exposed fishes such as tilapia (Hwang et al., 1995) and zebrafish (Johnson et al., 2007; Lema et al., 2007). LT reduction of red sea bream larvae at concentrations Z0.8 mg l1 might be directly related to their abnormal morphological development during embryonic–larval ontogenesis as revealed in this test. Delay in hatching due to Cd exposure could partly account for the retarded growth and

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increased individual size variation of the larvae in the high concentrations. Moreover, extra energy might be consumed for detoxification mechanisms as well as increasing metabolic oxygen demand due to oxidative stress. Consequently, less energy reserve could be traded off for development, which caused retarded growth in the Cd-exposed larvae. Similar to those exposed to TCDD (Yamauchi et al., 2006), yolk absorption rate of the red sea bream larvae was significantly repressed as Cd-exposure level was elevated. This could be explained by the finding that external Cd2+competivitively inhibits Ca2+ uptake of fish embryos and larvae and drop in body Ca2+ uptake can result in retardation of yolk absorption (Chang et al., 1997). Since red sea bream larvae were at endogenous stage and utilized only the yolk sac as nutrient reserve for development and growth during the test period, slow yolk absorption due to Cd exposure should account for, at least partly, the retarded growth of the larvae.

5. Conclusion This study revealed that, at 18 1C and 33 psu under laboratory conditions, red sea bream was more sensitive to Cd exposure at embryonic stage than at larval stage, which was similar to the findings when they were exposed to zinc, mercury and copper. This finding was contrary to the general belief that fish embryos were less sensitive to heavy metals than larvae due to the protection of chorions, suggesting a different toxicity behavior of heavy metals on the ELS of red sea bream compared to some other fish species. A threshold of 0.8 mg Cd2+ l1 could bring out detrimental impacts on various biological aspects at the ELS of this fish. These included low hatchability, delay in hatching, high mortality and morphological malformation, and reduced growth in the embryos and larvae. These biological endpoints were concentration dependent on Cd and could be used as potential bioindicator for assessing the risk of Cd to aquatic organism. In particular, mortality and morphological abnormality such as skeletal deformity became significantly higher at Z0.4 mg Cd2+ l1 concentrations than the controls. In the three bays of the Bohai Sea, where red sea bream used to spawn and feed, the average Cd concentrations were recently recorded to be 0.14–0.89 mg l1 and are as high as 2 mg l1 in some extreme locations. These suggested that malformation and mortality in red sea bream at embryonic–larval stage were most sensitive to Cd and this fact could be used as biologically or ecologically significant endpoints for evaluating water borne Cd toxicity in aquatic environments. Heart beat and yolk absorption of the larvae were significantly inhibited at some high Cd concentrations but were not as sensitive as other endpoints.

Acknowledgments This study was supported by the National Natural Science Foundation of China (NSFC) under the program Science Fund for Creative Research Groups (No. 40821004), The Ministry of Science and Technology of P.R. China under Grant contract no. 2007CB407305, and NSFC under Grant contract no. 40676086. We thank Dr. Li J for providing the red sea bream eggs for this study.

Appendix 1. Supplementary data Supplementary data associated with this article can be found in the online version at doi:10.1016/j.ecoenv.2009.06.002.

1973

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