Ecotoxicology and Environmental Safety 112 (2015) 137–143

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Toxicity of surface water from Huangpu River to luminous bacteria (Vibrio qinghaiensis SP. Q67) and zebrafish (Danio rerio) embryos Lili Zhang a,b, Qian Li a, Ling Chen a, Ai Zhang a, Jieni He a, Zhihao Wen a, Lingling Wu a,n a Key Laboratory of Yangtze Water environment, Ministry of Education, College of Environmental Science and Engineering, Tongji University, Shanghai 200092, China b Fuzhou Research Academy of Environmental Sciences, Fuzhou 350000, China

art ic l e i nf o

a b s t r a c t

Article history: Received 23 June 2014 Received in revised form 27 October 2014 Accepted 30 October 2014

Degradation of water quality is an emerging problem in many developing countries. Bioassay is an effective approach to monitor quality of water in aquatic environments. Studies have used luminescent bacteria and zebrafish embryos as bioassay tools in monitoring river water quality. In this study, luminous bacteria (Vibrio qinghaiensis sp. Q67) assay and zebrafish (Danio rerio) embryo toxicity test were performed to assess the ecotoxicity of surface water from the Huangpu River, China, collected during 2012–2013. River water samples inhibited the luminescence [inhibition rates 0–34.6% ( 74.82%)] of Q67 and increased the lethal rates and induced morphological abnormalities in zebrafish embryos. The toxicity to luminous bacteria and zebrafish embryos were higher in winter than in summer months. In addition, samples collected in industrial area, urban sampling sites near drainage outlets, and at the intersection of the tributary that flows into the Huangpu River showed higher toxicity. & Elsevier Inc. All rights reserved.

Keywords: Huangpu River Water toxicity Luminous bacteria Zebrafish embryos Temporal variation

1. Introduction Degradation of water quality is an emerging problem in many developing countries. Anthropogenic, industrial, and agricultural activities introduce toxic chemicals [e.g., pharmaceuticals and personal care products (PPCPs)] and increase the inputs of nutrients (cultural eutrophication) into the aquatic environment. These phenomena mainly contribute to eutrophication and other changes in the chemical composition of aquatic habitats. They also affect food security and quality of drinking water. In densely populated and fast-developing countries like China, clean water is a challenging issue in prospective politics and environmental planning (Bergmann et al., 2012). Some studies have chemically analyzed isolated pollutants in drinking water (Zhu et al., 2014; Klein et al., 2013; Post et al., 2013). Organisms (including humans) are exposed not only to isolated pollutants but also to complex chemical mixtures, the individual components of which might be present at concentrations too low to raise concern. However, additive or even synergistic effects can render such mixtures dangerously potent. For example, the mixture of five estrogenic compounds at individual concentrations too low to exert effects can elicit detrimental cumulative effects on fish (Brian et al., 2005). Thus, the integrated n

Corresponding author. E-mail address: [email protected] (L. Wu).

http://dx.doi.org/10.1016/j.ecoenv.2014.10.037 0147-6513/& Elsevier Inc. All rights reserved.

influence and possible toxicity of pollution on organisms and the ecosystem cannot be assessed only through chemical analysis (Fernandez et al., 2005), but also with different toxicity tests (Wolfram et al., 2012; He et al., 2014). On account of the consistency between selected organisms and their corresponding living space, biomonitoring can directly provide data on the potential effects and actual integrated toxicities of pollutants without prior information on the chemical components present in water. Hence, biomonitoring can reflect the corresponding degree of deleterious effects on the environment. The Huangpu River is the largest river in Shanghai, one of the most developed and urbanized cities in China. Approximately 83 km of the river flows through the Shanghai metropolitan area. This river is the most important shipping artery and the major drinking water source of Shanghai, providing approximately 76.3% raw drinking water (6,300,000 t/d). However, large amounts of industrial wastewater and domestic sewage are being discharged into the river annually (Zhang et al., 2014). Such discharges have caused great deterioration of water quality in the Huangpu River (Zhang, 2007; Jiang et al., 2011). Some studies have focused on the distribution and sources of pollutants in the Huangpu River; these pollutants include halogenated disinfection by-products (Wei et al., 2013), polycyclic aromatic hydrocarbons (Zhang, 2007; Liu et al., 2009b), perfluorinated compounds (Bao et al., 2010; Chen et al., 2009), antibiotics (Jiang et al., 2011), alkylphenolic compounds (4-tert-octylphenol and 4-nonylphenol) and plasticizer bisphenol A (Zhang,

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2007; Wu et al., 2013). Many of these contaminants raise considerable toxicological and health concerns, particularly when human health-based guideline values are unavailable. Ecotoxicity evaluation is important in monitoring the quality of water in order to protect aquatic organisms and human health. However, very few studies have investigated the toxicity of the Huangpu River (Song et al., 1997; Bai et al., 2008). In the present study, luminous bacteria (Vibrio qinghaiensis sp. Q67) and zebrafish (Danio rerio) embryos were employed to assess the toxicity of surface water in the Huangpu River. The suitability and reliability of ecotoxicity assays using Q67 have been demonstrated by numerous studies that focused on toxicity caused by heavy metals (Liu et al., 2003), herbicides, and insecticides (Liu et al., 2009a; Zhou et al., 2010) and by direct analysis of environmental samples (Liao et al., 2010; Mantis et al., 2005). The early life stage test using zebrafish embryos is currently a widely used biomarker for investigating the detrimental effects of aquatic pollutants on fish because of its relative simplicity and rapidity (Knobel et al., 2012; Wu et al., 2010). This study was primarily designed to evaluate the ecotoxicity of surface water samples from Huangpu River, determine whether ecotoxicity is related to different sampling sites and different seasons, and analyze the contribution of some pollutants to the ecotoxicity of the samples.

2. Materials and methods 2.1. Study area and sampling collection The Huangpu River is 114 km long, originates from the Dianshan Lake of Shanghai municipality, East China, and flows northeast, past Shanghai into the Yangtze River estuary at the Wusong, with a basin area of 24,000 km2 (Fig. 1). The upper reaches of the Huangpu River flow through the suburbs of Shanghai, characterized by intensive agricultural activities, particularly animal breeding operations. Its lower reaches flow through the urban area with intensive industrial and residential activities. The Huangpu River is influenced by various sources, including point sources (e.g., municipal wastewater treatment plant effluents) and non-point sources (e.g., confined animal feeding operations, poultry litter applied to farm fields, and storm-water runoff). Eleven sampling sites were selected along the Huangpu River (Fig. 1). S1 is the headstream of the river. The river reach between S1 and S6 is located in the agricultural area of Shanghai and used as the raw water supply for drinking. S4 is the river intersection

where the Yuanxiejing River flows into the Huangpu River. It accounts for 30% of the water of the Huangpu River. S5 is the river intersection where the Maogang tributary flows into the Huangpu River, and S6 is a water source in Shanghai. S7 is located in Wujing industrial district. S8 is in the middle reach of the Huangpu River. The region between S9 and S11 is located at the downtown residential area of Shanghai and is the busiest section of the city. S9 and S10 are also river intersections where the Suzhou River and Yunzao Brook enter into the Huangpu River, respectively. S11 is the estuary joint with the Yangtze River. Water samples were collected in 2012–2013. The campaigns occurred in August (Aug), October (Oct), and December (Dec) of 2012 and February (Feb), April (Apr), and June (Jun) of 2013, including 66 samples. All water samples were collected from the center of the stream (5–15 cm deep below the water surface) and then stored in pre-cleaned amber glass bottles. Each sample was filtered with a 0.45 mm mesh glass microfiber membrane and stored immediately in the dark at 4 °C until analyzed (Wu et al., 2014). Ecotoxicity tests were initiated within 36 h after sample collection. 2.2. Luminous bacteria bioassay For the acute toxicity test, the luminescent bacterium Q67 (purchased from Beijing Hammatsu Photon Techniques Inc., Beijing, China) freeze-dried as pellets in glass bottles were removed from
20 °C storage before the test, with recovery liquid (0.8% NaCl) added for rehydration at 20 °C for 15 min. The standard methods for culture medium preparation and Q67 incubation were adopted from a previous article (Wu et al., 2013). Luminescence inhibition assay was performed in test tubes using an RS9901 Luminometer (Shanghai Rong Sheng Biological Electronics Co. Ltd., Shanghai, China). For each test, four test tubes were prepared, three for parallel samples and one for blank control (recovery liquid). Q67 was directly exposed to the samples. Samples or control liquid (2 mL) were added into each tube, and the bacterial suspension (50 μL) was added at a 10 s interval. After 15 min exposure of the bacteria to the sample at 20 °C, the relative light unit (RLU) of the luminescent bacterium Q67 was measured. Toxicity was evaluated by the inhibition ratio, which was calculated as a previous study of Zhou et al. (2010).

⎛ LU ⎞ X (%) = ⎜1 − ⎟ × 100% LU ⎝ 0⎠ where LU0 is the RLU of Q67 exposed to the blank control, and LU is the RLU to the same site samples. Each test was repeated thrice, and the average inhibition ratio was taken as the final result. 2.3. Zebrafish embryos bioassay

Fig. 1. Location of the Huangpu River and sampling sites.

Sexually mature zebrafish were raised in 25 L tanks in the laboratory and maintained according to the method described by Wu et al. (2010). The control conditions were as follows: temperature, 26.0 70.5 °C; hardness, 250 mg/L; pH, 7.570.5; dissolved oxygen, 10.5 7 0.5 mg/L; and 14 h light/10 h dark photoperiod. The fish were fed twice each day with commercially frozen red mosquito larvae sterilized by ultraviolet lamps (Gonglin, China). Before the test, several translucent plastic spawning boxes (12 cm  20 cm  12 cm) with a mesh insert (3–4 mm mesh size) were placed individually in the aquaria. Each spawning box held six males and three females. Spawning and fertilization occurred within 30 min after light had been turned on in the morning. Eggs were collected and rinsed with the reconstituted water (ISO 7346/3: 294 mg/L CaCl2  2H2O, 123 mg/L MgSO4  7H2O, 123 mg/L NaHCO3, 5.5 mg/L KCl) that had been aerated to nearly 100%

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oxygen saturation prior to bioassay. Fertilization and quality were assessed under an inverted microscope (Nikon TE2000-U). For control experiments, a fertilization rate should be higher than 90%. The acute toxicity tests used in this study were a modification of the FET-test (OECD, 2013). Fertilized eggs were exposed in 24well cell culture plates with a Pasteur pipette. One fertilized egg was placed in each well with 2 mL of test solutions, amounting to a total of 20 eggs per test. The remaining four wells of each plate were used as controls with the reconstituted water. A fixed concentration of 4 mg/L 3,4-dichloroaniline was also performed with egg batch used for testing as a positive control. The plates were covered with foil and then incubated at 26.0 7 1.0 °C and 14 h light/10 h dark photoperiod. The development of embryos from blastula to early larval stage was monitored at specific time points (48 and 96 h). The endpoints used for assessing the embryo toxicity included mortality and abnormality. The lethal and sub-lethal endpoints were observed and recorded for both reference and treated groups using an inverse microscope. All exposure experiments in the study were performed in triplicates.

2.4. Chemical analysis 2.4.1. Pharmaceuticals and personal care products Analysis of five acidic pharmaceuticals, including four nonsteroidal anti-inflammatory drugs (ibuprofen, ketoprofen, naproxen, and diclofenac) and one lipid regulator (clofibric acid), were carried out in this study because of their wide consumption in China (Dai et al., 2009) (Table S1). The pharmaceuticals were extracted from aqueous samples and analyzed according to the method described by Kimura et al. (2007). All samples were analyzed by TSQ Quantum high-performance liquid chromatography (HPLC) coupled with mass spectrometry (Thermo Fisher Scientific, San Jose, CA, USA). Quantification of the target compounds was performed using an external standard method. 2.4.2. Endocrine disrupting-compounds Among endocrine disrupting-compounds (EDCs), steroid estrogens estrone (E1), 17β-estradiol (E2), estriol (E3), and 17αethinylestradiol (EE2) and phenolic xenoestrogens technical-nonylphenols (t-NP) and plasticizer bisphenol A (BPA) were analyzed (Table S1). The former possess the highest estrogenicity, and the latter have moderate estrogenic potency and widespread applications (Vethaak et al., 2005). Analytical methods were adopted from the study of Zhang et al. (2014).

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3. Results 3.1. Inhibition rate of luminous bacteria All samples collected in Aug, Oct, and Dec of 2012 and Feb, Apr, and Jun of 2013 were tested in the Q67 acute toxicity test. The inhibition rates of the luminescence ranged from 0% to 34.63% (74.82%) in the six tests (Fig. 2). S1–S3 and S7 samples in Aug and S8–S10 samples in Oct inhibited the luminescence of Q67. All samples inhibited the luminescence in Dec and Feb. Most samples in Apr, except for those from S1–S3, exhibited inhibitory effects, but the inhibition rates were lower than 5%. Only S7 had inhibitory effect in Jun. 3.2. Lethality to zebrafish embryos A total of 55 samples were subjected to the zebrafish embryo toxicity test (except the samples collected in Oct of 2012). All living zebrafish embryos exposed to the samples had hatched out from the membrane after 96 h exposure. The lethality of the zebrafish embryos at 96 h is shown in Fig. 3. The lethality of the embryos was less than 5% in control groups. No significant difference was found in the lethality of the zebrafish embryos at S2–S4 in all five campaigns compared with the control group (P o0.05). The lethality of the zebrafish embryos exposed to the water samples from other sites varied from 5.00% to 41.67% (72.89%), which was significantly higher than the control group (Po 0.05; Fig. 3). The

Fig. 2. Inhibition rate of Q67 exposed to 11 samples from Huangpu River in Aug, Oct, Dec of 2012 and Feb, Apr, Jun of 2013.

2.4.3. Organochlorine pesticides and polychlorinated biphenyls The OCPs were analyzed according to EPA8081A (USA) and EPA8082 (USA). An Agilent 6890 plus gas chromatography with a microcell electron capture detector was used to analyze 19 OCPs and 13 PCBs (Table S1).

2.5. Data analysis During Luminous bacteria data analysis, significant (P o0.05) data were obtained using T-test and expressed as mean 7SD. The effects of the water samples from the Huangpu River on the embryos were tested using one-way ANOVA, followed by Dunnett’s multiple comparison tests (when data were normally distributed according to least significant difference tests). All statistical analyzes were conducted using SPSS 17.0.

Fig. 3. Lethality of zebrafish embryos exposed to samples for 96 h in Aug, Dec of 2012 and Feb, Apr, Jun of 2013. Asterisks indicate significant difference from the control (Po 0.05).

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Fig. 4. Percentage of abnormalities observed among zebrafish embryos exposed to samples. These samples were those from S4, S7 and S10 in Aug; S4, S7, S10 and S11 in Dec; and S9 in Feb. The zebrafish embryos from other exposure groups showed no detectable abnormalities. Asterisks indicate significant difference from the control (P o 0.05).

highest lethality in zebrafish embryos were 41.67% at site S5 in Aug, 15% at site S5 in Dec and 10% at S6 in Dec. 3.3. Embryotoxicity Several developmental abnormalities of the zebrafish embryos were observed after 48 and 96 h exposure to Huangpu River water; these abnormalities include pericardial edema, blood stasis, yolk sac edema, tail deformations, and kyphosis (Fig. S1). The percentage of abnormalities in the samples varied from 0% to 16.7% ( 72.89%); these samples were those from S4, S7, and S10 in Aug; S4, S7, S10, and S11 in Dec; and S9 in Feb (Fig. 4). Except for the samples from S2 and S3, other samples induced developmental abnormalities in the zebrafish embryos toxicity test. The highest abnormality of zebrafish embryos were 16.7% (S7 site August) and 10% (S10 site December). The zebrafish embryos from other exposure groups developed very well and showed no detectable developmental abnormalities during the exposure test. Most zebrafish embryos exposed to the samples can continue developing and hatch out from the membrane. 3.4. Chemical analysis The toxic effects of water samples on Q67 and zebrafish embryos in Dec campaign were higher than other campaigns (Figs. S2 and S3). Therefore, some micropollutants were analyzed in water samples taken during December campaign. The investigation focused on analytical measurements of PPCPs, EDCs, OCPs, and PCBs. The concentrations of these micropollutants detected in water samples of December are shown in Table 1. PPCPs and EDCs were detected at all 11 sampling sites. The pharmaceutical concentrations were lower than predicted no-effect concentrations (PNECs) (Table S2) Table 2.

4. Discussion 4.1. Toxicity of water samples from the Huangpu river In this study, the toxicity of surface water in the Huangpu River was assessed using the luminous bacterial assay and zebrafish embryo toxicity test. To rank the acute toxicity to Q67 of the sampling sites, the inhibition rates in this study were compared

with those from the other locations. The inhibition rates of the samples in this study were considerably lower than that from the Le An River in China (6.39–83.56%), which is seriously polluted by heavy metals (Ma et al., 1999). Compared with the water samples from the Zaohe River in Xi’an, China, which is a stream receiving urban runoff, treated domestic effluent and untreated industrial wastewater, those from Huangpu had lower inhibition rates. Moreover, the surface water from the Huangpu River showed effects similar to those of the surface water from the Weihe River, Xingqinghu Lake, and Nanhu Lake in Xi’an, China (Ma et al., 2012). The results of the zebrafish embryo toxicity test suggested that some water samples increased the lethal rates (Fig. 3) and induced morphological abnormalities (Fig. S1). The toxic effect of the samples on zebrafish embryos in this study was much less than that of the surface water from the urban and rural rivers in Texas basins (VanLandeghem et al., 2012) and Xiamen Harbor, where showed signs of poor water quality having high mean level of embryo deformity (20–30%) (Klumpp et al., 2002). In general, the toxicity effects in the Huangpu River can be categorized as low to moderate. The different responses between the biomarkers used in this study can be attributed to the different biological levels and time scales at which these indicators act as well as to the different sensitivity levels to environmental stress (Handy et al., 2003). Both bioassays demonstrated to be sensitive to the pollution found in S4, S7–S9. Even, zebrafish embryos resulted to be more sensitive than bacteria, since the embryos have showed mortality and abnormalities in other sampling sites, such as S1, S4–S11. However, these data indicated the urgency to control pollution in the Huangpu River. 4.2. Seasonal variation of toxicity of water samples from the Huangpu river The overall toxicity to luminous bacteria and zebrafish embryos at the 11 sampling sites were higher in winter than in summer (Figs. S2 and S3). The toxic effects of the samples on Q67 and zebrafish embryos were complex. These effects were not caused by a single factor but by multiple pollutants. The different toxicities to luminous bacteria and zebrafish embryos in the six sampling events might be explained by the flow conditions of the river. The period from May to September is the typical wet season of the Huangpu River, and from November to Apr is the typical dry season. In Feb, Apr, and Jun in 2012 and Aug, Oct, and Dec in 2012, the rainfalls were 50.6, 71.7, 196.2, 224.6, 58.6, and 48.5 mm, respectively (http://www.shanghaiwater.gov.cn). The high flow conditions in summer might considerably dilute the concentrations of pollutants in the surface water. Moreover, the high microorganism activity and strong sunlight in summer might be the reason for the higher bio- and photo-degradation of pollutants in this season than in winter (Karthikeyan and Meyer, 2006; Liu et al., 2009b). Moreover, coal burning for home heating may produce more pollutants in winter than in summer. 4.3. Spatial distribution of toxicity of the Huangpu river S4, S7–S9, at which sites the inhibition rates of the luminescence exceeded 20%, are in the middle and lower reaches (except for S4). The inhibition rates gradually decreased along with the flow direction (Fig. 2). Among the sites with high inhibition rates, S7 is located at one of largest chemical industry area in Shanghai, whereas S8–S10 are located in commercial and urban areas (Zhang et al., 2014). S8 is downstream of site S7 and located in the confluence of Dianpu River. S9 is located close to drainage outlets that mainly discharge domestic wastewater, indicating possible pointsource pollutions from municipal effluents. These results can explain the high inhibition rates at these sites.

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Table 1 Chemical analysis of the 11 samples from the Huangpu River in Dec (ng/L). S1

S2

S3

S4

S5

S6

S7

S8

S9

S10

S11

PPCPs

KEP NPX CA DFC IBP

81.31 29.51 1.56 8.94 2.87

50.44 13.98 0.42 26.65 5.99

24.28 19.29 0.89 20.12 5.21

57.56 21.84 2.12 25.95 10.21

50.81 34.7 0.74 8.72 12.4

56.76 10.69 0.93 8.59 4.17

85.55 13.23 0.9 6.56 5.9

57.88 32.51 2.97 21.3 18.02

89.28 15.64 2.5 29.27 4.64

19.45 28.06 3.88 9.78 11.08

16.42 22.12 3.55 19.31 15.73

EDCs

E3 E2 EE2 t-NP BPA E1

87.22 16.25 25.17 240.6 58.56 11.88

50.47 30.18 27.21 730.9 92.16 70.04

59.45 32.50 18.94 1388 67.27 83.21

55.24 38.27 22.34 631.1 635.4 27.06

94.14 70.91 29.73 671.5 244.8 49.38

59.83 29.16 17.16 642.6 87.23 56.64

76.17 74.79 20.89 786.1 687.5 68.81

65.36 37.97 15.47 797.3 100.8 49.23

74.91 54.13 16.48 667.4 571.7 47.37

60.56 35.71 20.48 200.0 356.2 74.35

55.64 35.15 12.52 516.9 87.72 89.03

HCHs

α-HCH β-HCH γ-HCH δ-HCH ε-HCH

ND 0.70 0.49 30.27 0.07

ND ND 0.14 30.59 0.03

ND 0.86 0.43 36.97 0.06

ND ND 0.22 14.81 0.04

ND 0.45 0.16 ND 0.04

ND 0.99 0.39 18.67 0.07

0.05 0.59 0.4 24.52 0.09

ND 0.13 0.11 18.4 0.07

ND 1.46 0.28 14.49 0.35

ND 1.09 0.24 0.47 0.11

ND 0.74 0.31 0.08 0.13

DDTs

2,4-DDE 4,4-DDE 2,4-DDD 4,4-DDD 2,4-DDT 4,4-DDT

ND 0.50 2.33 ND 0.18 ND

ND 0.14 0.18 0.02 0.52 ND

ND 1.09 1.75 0.26 0.11 0.07

ND 0.35 0.41 0.03 ND 0.01

ND 0.12 0.04 0.25 ND 0.01

ND 0.09 0.47 0.02 ND 0.21

ND 0.66 0.68 0.08 ND 0.04

ND 0.18 0.59 ND ND 0.13

ND 0.05 1.46 0.04 ND ND

ND ND 1.26 0.11 ND ND

0.2 0.45 0.49 ND ND ND

Other OCPs

Aldrin Heptachlor Heptachlor Heptachlor Heptachlor Endosulfan Dieldrin HCB Endrin Endosulfan PCB28/31 PCB194

ND 0.22 0.00 0.22 0.17 ND 0.05 0.27 ND ND 0.67 ND

ND 0.10 0.00 0.10 0.08 ND 0.56 1.18 ND ND 1.00 ND

ND 0.45 0.00 0.45 0.01 ND ND 0.43 ND ND 1.02 0.03

ND 0.60 0.03 0.63 ND ND ND 0.25 0.97 ND 1.08 ND

ND 0.00 0.13 0.13 ND ND ND 0.09 ND ND 0.40 ND

ND 0.46 0.04 0.50 0.10 ND ND 0.38 ND ND 1.02 0.04

ND 0.65 0.04 0.69 0.04 ND 0.12 0.35 ND ND 6.21 ND

ND 0.23 0.01 0.24 0.07 ND 0.04 0.19 ND ND 0.73 ND

ND 0.39 0.06 0.45 0.08 ND 0.07 0.13 ND ND 0.59 ND

ND 0.35 0.05 0.4 0.14 ND 0.63 0.22 ND ND 0.07 ND

ND 0.00 0.02 0.02 0.02 ND ND 0.13 ND ND 0.2 ND

ND 3.80 6.36 ND ND 0.02 ND ND 0.75 0.03

ND 1.36 17.91 ND ND 0.07 ND ND 2.77 0.01

ND 2.75 27.38 ND ND 0.15 ND ND 2.28 ND

ND 12.11 11.15 ND ND 0.03 ND ND 0.88 ND

ND 21.73 2.64 ND ND 0.02 ND ND ND ND

0.24 20.40 9.82 ND ND 0.27 ND ND 1.13 ND

ND 16.86 14.43 ND ND 0.01 ND ND 2.29 ND

ND 9.62 6.39 ND ND 0.06 ND ND 0.81 ND

ND 8.63 8.91 ND ND 0.03 ND ND 0.52 0.01

ND 9.51 12.42 ND ND 0.01 ND ND 0.23 0.04

ND 8.1 5.08 ND ND 0.03 ND ND 0.21 ND

PCBs

PCB118 PCB44 PCB52 PCB153 PCB101 PCB138 PCB183 PCB180 PCB18 PCB209

epoxide-α epoxide-β epoxide I

II

Mean values are presented (n¼ 2, replicate samples taken at the same time). ND: not detected.

As the lethality of the zebrafish embryos is shown in Fig. 3, no clear trend with the flow direction was observed. The lethality of S1 was nearly 10% in Dec, Apr, and Jun S1 is located at Dianshan Lake, where the water disturbance is low and the pollutants do not easily spread and dilute. The lethality of the zebrafish embryos at S4 was not higher than 5%, and there was nearly no lethality except in Dec. S4 is the river intersection where the Yuanxiejing River flows into the Huangpu River and the water quality is relatively better than other estuaries. The lethality of the zebrafish embryos at S5 was the highest in Aug and Dec. S5 is the river intersection where the Damaogang tributary flows into the Huangpu River. The water quality of Damaogang tributary is assessed to be between Quality Class IV and V according to the environmental quality standard for surface water in China, indicating that water from it can only be utilized in industry or agriculture (Jiang et al., 2013; Che et al., 2004). The low water quality of this tributary might account for the high lethality of the zebrafish embryos at S5. The lethality of the zebrafish embryos of S7–S11 ranged from 5% to 10%. S7–S9 are located in industrial and urban

areas where industrial and domestic pollution sources are abundant. S10 and S11 are also in urban areas. In addition, S1–S6 serve as drinking water sources and supply the raw water for Shanghai. However, the embryotoxicity hazard of water quality at sites S1, S5 and S6 was higher than others. Table 2 Pearson correlations between concentration of EDCs, OCPs, PCBs and toxicological indices. Pearson R (P value) 96 h lethality

Inhibition rate (Q67)

EDCs

BPA E3

0.203 0.775nn

0.732n 0.038

OCPs/PCBs

Heptachlor Epoxide β PCB44

0.776nn 0.811nn

0.157 0.309

Single asterisk denotes significant correlation at 0.05 level; double asterisks denote significant correlation at 0.01 level

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4.4. Chemical parameters The PNECs for the acidic pharmaceuticals were adopted from previous literatures (Zhao et al., 2010; Wang et al., 2010; TauxeWuersch et al., 2005). Compared with data in the literature shown in Table S3, the concentrations and detection rates of E1, E2, EE2, E3, t-NP, and BPA in the present study were higher than those reported in the surface water from European countries but were close to those reported in the surface water from China (Table S3). The concentrations of OCPs and PCBs were compared with the criterion maximum concentration (CMC) and criterion continuous concentration (CCC) from the National Recommended Water Quality Criteria for evaluating contamination in the surface water (USEPA, 2009) (Table S4). The concentrations of pollutants exceeding the CMC indicated that adverse effects on living creatures would occur frequently in a short time. Meanwhile, exceeding the CCC indicated that adverse effects on living creatures might increase in incidence from rare to occasional in a long time. In this study, neither did the concentration of OCPs exceed the CMC, nor did the concentration of PCBs, while the maximum measured environmental concentration of PCBs exceeded CCC. Therefore, EDCs and PCBs were regarded as the possible pollutants which contributed to toxicity of the surface water in the Huangpu River.

5. Conclusions Surface water samples from Huangpu River had negative effects on the luminescence of Q67 and the development of zebrafish embryos. Compared with the control group, the samples from the Huangpu River significantly inhibited the luminescence of Q67, increased the lethality, and induced developmental abnormalities of zebrafish embryos. The samples collected in Dec exhibited strong toxicity, which corresponded to the reverse of rainfall and activity of microorganism and sunlight intensity. High toxicity was observed in industrial areas, urban sampling sites near drainage outlets, and in the intersection that the tributary flows into the Huangpu River. Overall, the surface water of the Huangpu River has potential ecotoxicity to aquatic organisms. Therefore, future studies should further elucidate the adverse effects on organisms and not merely show pollutant concentrations in water. Moreover, more attention should be paid to control these pollutants, especially EDCs, and PCBs in the catchment of the Huangpu River.

Acknowledgments This work was supported by the national key technology R&D program of the ministry of science and technology of China (2012BAJ24B01).

Appendix A. Supplementary material Supplementary data associated with this article can be found in the online version at http://dx.doi.org/10.1016/j.ecoenv.2014.10. 037.

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