Jianchao
Liu
,
Guanghua
Lu
*,
Zhenghua
Zhang
,
Yijun
Bao
,
Fuli
Liu
,
Donghai
Wu
and
Yonghua
Wang
Key Laboratory for Integrated Regulation and Resources Development on Shallow Lakes, China Ministry of Education, College of Environment, Hohai University, 1 Xikang Road, 210098 Nanjing, China. E-mail: ghlu@hhu.edu.cn; Fax: +86-25-83787330; Tel: +86-25-83787894
First published on 22nd October 2014
Pharmaceutically active compounds (PhACs) have been universally detected in rivers, lakes and coastal waters that are affected by effluents from sewage treatment plants (STPs). In this study, the biological effects and bioaccumulation of PhACs were assessed in crucian carp (Carassius auratus) caged in an effluent-receiving river for 21 days. Compared with control fish in the laboratory and at a reference site, a significant reduction in hepatosomatic index (HSI) and increase in the biotransformation enzymes ethoxyresorufin O-deethylase (EROD) and glutathione S-transferase (GST) activities were observed in the fish that was caged downstream from the STP outfall. In general, the total concentrations of PhACs detected in fish tissues were in the order as follows: liver > brain > gill > muscle > bile. The bioaccumulation factors (BAFs) for PhACs were between 195 and 2782 in the major storage tissue liver. The corresponding results for both risk quotient (RQ) and enhanced integrated biomarker response (EIBR) based on laboratory and field studies, respectively, indicated that environmental risk for adverse effects to aquatic organisms were clearly higher at the downstream of the STP outfall than at the upstream.
Environmental impactPharmaceutically active compounds (PhACs) have been universally detected in rivers, lakes and coastal waters affected by effluents from sewage treatment plants. In this work, we (1) investigated the occurrence and bioaccumulation of eight PhACs in fish caged in an effluent-receiving river in Nanjing, China; (2) determined biomarker responses of hepatosomatic index (HSI), biotransformation phase I enzyme ethoxyresorufin O-deethylase (EROD) and phase II enzyme glutathione S-transferase (GST) activities in the liver of fish; and (3) assessed the potential ecotoxicological effects of the contaminants using both risk quotient (RQ) and enhanced integrated biomarker response (EIBR) based on laboratory and field studies, respectively. The article could help to implement management strategies for pollution control of PhACs and be useful for policy-makers for upgrading sewage treatment plants. |
A number of studies have reported the fate and occurrence of PhACs in STP effluents and receiving waters. However, reports regarding the distribution of PhACs in the tissues of fish in the natural environment are lacking. This information is important with respect to toxicokinetics, allocation, and potential contaminant-associated effect as a function of tissue/organ.6 It has been reported that PhACs, such as diclofenac, naproxen, ibuprofen, carbamazepine, erythromycin and 17α-ethinylestradiol can be detected in the bile and muscle of wild fish linked to STPs, with high variance for bioaccumulation factors (BAFs).1,7,8 However, the wild fish are not confined to a fixed local environment, which could not reflect the actual environmental effects of STP effluents. Active biomonitoring (ABM) involves the transplantation of organisms that are collected from an unstressed, unpolluted population to selected polluted sites. One distinct advantage, compared with passive biomonitoring, is that of a well-defined exposure time and location. The chemical and biological consequences of this translocation, which usually involves caging the organisms, can then be monitored in space and time to assess the effects of exposure on selected endpoints.9 Our group had previously utilized ABM to evaluate the complex pollution in the northern section of Taihu Lake, China.10 ABM was also used to detect the pharmaceutical residues in fish downstream of STPs.11,12 ABM has been extensively used in environmental health assessments, and it exhibits better results than passive biomonitoring because organisms already present in situ may have adapted to the pollutants.
The objectives of the present study were to investigate the occurrence and bioaccumulation of eight PhACs in fish caged in an effluent-receiving river in Nanjing, China. The biomarker responses of hepatosomatic index (HSI), biotransformation phase I enzyme ethoxyresorufin O-deethylase (EROD) and phase II enzyme glutathione S-transferase (GST) activities of the fish were determined. In addition, both risk quotient (RQ) and enhanced integrated biomarker response (EIBR) based on laboratory and field studies, respectively, were applied to indirectly/directly estimate the potential ecotoxicological effects of the contaminants. The eight PhACs investigated in this study include antibiotics roxithromycin (ROX) and erythromycin (ERY), antifungal ketoconazole (KCZ), anti-inflammatories ibuprofen (IBU) and diclofenac (DIC), β-blocker propranolol (PRO), anti-epileptic carbamazepine (CBZ) and steroid hormone 17α-ethinylestradiol (EE2).
Crucian carp were transferred from the laboratory and caged in the New Qinhuai river watershed (Nanjing, China) adjacent to the Jiangning STP (32°00′11.04′′N; 118°51′25.14′′E) at one upstream and three downstream sites for 21 days in December, 2013. This STP serves 450000 inhabitants, and the treatment process involves anoxic/anaerobic/aerobic (first-stage project) and oxidation processes (second-stage project). New Qinhuai river sampling sites and STPs location are indicated in Fig. 1. One reference site (S1) is located 0.5 km upstream of the Jiangning STP effluent outfall, but 7 riverine km downstream of the effluent outfall from the Chengnan STP. The three downstream stations (i.e., S2, S3 and S4) were 0.1, 1.5 and 3.1 km below the Jiangning STP effluent outfall, respectively (Fig. 1). Fish controls were cultured in dechlorinated municipal water in a laboratory because various pollutants are present in the natural river and a field negative control site for comparison was not available. At each field site, two non-toxic columnar polyethylene cages (diameter 55 cm and height 55 cm) were deployed, holding 15 fishes each. The cages were mesh-shaped with 1 cm holes. Each cage was fixed in place by three wooden pegs at approximately 20 cm below the water surface. Ten fish (i.e., three fish tested for enzyme assays and seven fish for chemical analysis) were collected and delivered to the laboratory at each station after 7, 14 and 21 days of exposure, respectively. In addition, triplicate water samples were collected at each ABM station, using standard depth- and width-integrated compositing techniques, and used for chemical analysis.
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Fig. 1 New Qinhuai River sampling sites (indicated with S1, S2, S3 and S4) and STP locations in the study areas. |
After transporting to the laboratory, the fish were anaesthetized with MS222 (100 mg L−1) and measured for length and weight. Then, the fish were sacrificed by cervical transection. Liver, brain, gill, muscle and bile tissues were immediately excised and weighed. All the tissues were washed with 0.15 M KCl, blotted with filter paper and immediately stored in liquid nitrogen.
Sites | Time | ROX | ERY | PRO | CBZ | DIC | SUM |
---|---|---|---|---|---|---|---|
S1 | 7 days | 8.23 ± 1.87 | 4.12 ± 0.54 | 0.21 ± 0.13 | 1.03 ± 0.34 | 10.54 ± 3.41 | 24.13 |
14 days | 10.41 ± 1.92 | 8.13 ± 2.10 | 0.69 ± 0.21 | 2.33 ± 1.23 | 30.85 ± 6.64 | 52.41 | |
21 days | 7.38 ± 1.19 | 8.47 ± 1.40 | 0.84 ± 0.54 | 2.65 ± 0.98 | 13.67 ± 4.13 | 33.01 | |
Mean | 8.67 ± 1.56 | 6.91 ± 2.42 | 0.58 ± 0.33 | 2.00 ± 0.86 | 18.35 ± 10.94 | 36.32 | |
S2 | 7 days | 20.45 ± 3.90 | 7.90 ± 1.40 | 1.29 ± 0.65 | 2.11 ± 0.84 | 34.10 ± 8.52 | 65.85 |
14 days | 26.76 ± 5.12 | 8.64 ± 2.09 | 0.63 ± 0.18 | 2.97 ± 0.42 | 49.15 ± 10.42 | 88.15 | |
21 days | 33.61 ± 4.18 | 10.22 ± 3.01 | 0.67 ± 0.20 | 3.83 ± 0.56 | 37.93 ± 7.45 | 86.26 | |
Mean | 26.94 ± 6.58 | 8.92 ± 1.19 | 0.86 ± 0.37 | 2.97 ± 0.86 | 40.39 ± 7.82 | 80.09 | |
S3 | 7 days | 18.69 ± 4.09 | 7.93 ± 1.23 | 0.69 ± 0.43 | 2.11 ± 0.69 | 30.41 ± 4.51 | 59.83 |
14 days | 20.70 ± 2.19 | 9.37 ± 2.13 | 0.61 ± 0.51 | 2.12 ± 0.81 | 48.72 ± 9.02 | 81.52 | |
21 days | 28.91 ± 3.43 | 8.31 ± 5.39 | 0.69 ± 0.36 | 1.86 ± 0.41 | 32.34 ± 6.59 | 72.11 | |
Mean | 22.77 ± 5.41 | 8.54 ± 0.75 | 0.66 ± 0.05 | 2.03 ± 0.15 | 37.15 ± 10.06 | 71.15 | |
S4 | 7 days | 13.55 ± 9.42 | 7.71 ± 2.67 | 0.68 ± 0.38 | 2.06 ± 0.27 | 23.00 ± 7.17 | 47.00 |
14 days | 17.44 ± 2.78 | 7.74 ± 2.21 | 0.73 ± 0.57 | 2.35 ± 0.65 | 42.04 ± 10.48 | 70.30 | |
21 days | 19.57 ± 3.75 | 5.62 ± 3.98 | 0.69 ± 0.25 | 1.46 ± 0.72 | 28.78 ± 8.41 | 56.12 | |
Mean | 16.85 ± 3.05 | 7.02 ± 1.22 | 0.70 ± 0.03 | 1.96 ± 0.45 | 31.27 ± 9.76 | 57.81 |
Clearly, PhAC concentrations in the samples from downstream were higher than those from upstream in the river. For example, the mean concentration of ROX was found to be 8.67 ng L−1 at the S1 site (upstream 0.5 km), which had increased to 26.94 ng L−1 at the S2 site (downstream 0.1 km), suggesting a substantial input of this compound through STP effluent. Subsequently, ROX decreased to 16.85 ng L−1 at the S4 site (downstream 3.1 km), which is a clear sign of reduction. The same trend is observed for the other pharmaceuticals in the present and a previous study,22 which further demonstrates that effluents from STP are the primary contributor of PhAC pollution in urban rivers. This suggested that the levels of PhACs in river water depend not only on the loads from STP but on the river flow, biodegradation, photodegradation and adsorption on sediments, which acted upon the pharmaceuticals after their release.26
Unlike ROX, PRO and CBZ, the highest concentration of KCZ was found in fish muscle, which was in agreement with our previous study, where KCZ was only found in the muscle of wild fish collected from the downstream river of STPs.27 Moreover, the relatively high concentrations of KCZ were also found in the liver and brain, which exhibited an obvious time dependence. These results indicate that KCZ has a high potential for bioaccumulation in crucian carp. The previous studies have shown that KCZ acted as a potent inhibitor of CYP1A and CYP3A enzyme activities in fish.28,29 The inhibition of CYP enzyme activities (i.e., CYP1A and CYP3A) is linked to adverse outcomes in fish, such as the bioaccumulation of contaminants.28 No data were available in the previous literature for the bioaccumulation of KCZ in fish.
Overall, the higher total concentrations (9.10–27.68 ng g−1) of PhACs (ΣPhACs) were found to be in the liver, which is one of major targeted organs for accumulation, biotransformation and excretion of contaminants in fish, followed by the brain, gill, muscle and bile. After biotransformation in the liver, the metabolites and parent compounds are excreted to the small intestine via bile. However, the feeding status of the fish has a pronounced effect on the concentration of chemicals and metabolites in the bile. Following feeding, the fish empties the gall bladder into the intestine, and consequently the bile volume and the content of foreign chemicals will be low with no obvious time dependence.30 For the brain, which is a lipid-rich target tissue similar to the liver, appreciable levels (6.06–19.18 ng g−1) of ΣPhACs were observed. Notably, the accumulation of ΣPhACs in the brain steadily increased until the end of exposure (21 days), and the increasing trend was obvious with the decrease of concentrations of ΣPhACs in the surface water. However, brain/liver concentration ratios of ΣPhACs were 0.62 on average, indicating that PhACs are hindered from crossing the blood–brain barrier. Alternatively, these results indicate that brain tissue mainly consists of relatively polar lipids, which constitute unfavorable sites for PhACs accumulation compared with the triacylglycerol-rich composition of the liver.6 Moreover, gills are the first organs in contact with polluted water and suspended sediment particles; therefore, they can be the significant sites of interaction with contaminants. The concentrations of ΣPhACs in gill and muscle are similar, with the highest concentrations of 16.21 and 16.01 ng g−1, respectively, which were consistent with the concentrations of steroids (i.e., estrone, 17β-estradiol, EE2 and estriol) measured in wild fish from Dianchi Lake, China.31
Bioaccumulation is a process in which chemical substances are absorbed by aquatic organisms from all environmental sources.31 Based on the mean concentrations of PhACs in surface water and in the fish tissues at each sample site, the BAFs of five PhACs (ROX, ERY, PRO, CBZ and DIC) were calculated in different fish tissues, as shown in Table 2. BAFs for these compounds were below 1500 with the exception of PRO, which suggests a low bioaccumulation potential. The chemicals are defined as being “bioaccumulative” if the BAF is greater than 5000 in aquatic organisms, and as being “potentially bioaccumulative” if the BAF is in the range of 2000–5000 in aquatic organisms.32 In the present study, PRO exhibited the highest BAFs (2782), suggesting that the potential for bioaccumulation of PRO exists. The BAFs in the liver for ROX, ERY, PRO and CBZ are in the range of 195 to 2782. It is evident that the liver is a steadily accumulative organ for PhACs and a more suitable matrix for monitoring these compounds in aquatic environments. The previous studies have shown that pharmaceuticals at a certain environmental concentration interfere with the biological functions of organ/tissue in fish.33–35 However, no previous data prove that organ/tissue concentrations are at a level that has an impact on the biological functions of the organ/tissue. The knowledge of organ/tissue concentrations of the compounds in question would be of significant value in biological effect studies.
Tissues | Sites | Time | ROX | ERY | PRO | CBZ | DIC |
---|---|---|---|---|---|---|---|
Liver | S1 | 7 days | 390 | 457 | 1425 | 482 | |
14 days | 368 | 356 | 1796 | 473 | |||
21 days | 440 | 539 | 1159 | 589 | |||
S2 | 7 days | 254 | 556 | 1836 | 661 | ||
14 days | 298 | 1255 | 1575 | 471 | |||
21 days | 233 | 812 | 1370 | 698 | |||
S3 | 7 days | 195 | 641 | 2400 | 757 | ||
14 days | 437 | 1275 | 2782 | 687 | |||
21 days | 201 | 396 | 2190 | 669 | |||
S4 | 7 days | 223 | 1221 | 2017 | 535 | ||
14 days | 291 | 1482 | 2223 | 738 | |||
21 days | 328 | 601 | 2037 | 917 | |||
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Brain | S1 | 7 days | 263 | 281 | 396 | ||
14 days | 233 | 323 | 284 | ||||
21 days | 298 | 333 | 349 | ||||
S2 | 7 days | 94 | 972 | 197 | |||
14 days | 179 | 1125 | 246 | ||||
21 days | 140 | 589 | 225 | ||||
S3 | 7 days | 72 | 700 | 240 | |||
14 days | 78 | 872 | 261 | ||||
21 days | 210 | 581 | 379 | ||||
S4 | 7 days | 82 | 315 | 212 | |||
14 days | 90 | 600 | 260 | ||||
21 days | 199 | 499 | 178 | ||||
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Gill | S1 | 7 days | 279 | 412 | 926 | 320 | 20 |
14 days | 259 | 772 | 151 | 10 | |||
21 days | 116 | 171 | |||||
S2 | 7 days | 72 | 192 | 198 | 13 | ||
14 days | 35 | 1198 | 237 | 89 | 42 | ||
21 days | 43 | 548 | 106 | 2 | |||
S3 | 7 days | 55 | 322 | 165 | 6 | ||
14 days | 51 | 1329 | 281 | 146 | |||
21 days | 34 | 434 | 104 | 4 | |||
S4 | 7 days | 58 | 282 | 139 | 3 | ||
14 days | 46 | 710 | 333 | 121 | |||
21 days | 63 | 405 | 148 | 3 | |||
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Muscle | S1 | 7 days | 129 | 371 | 1292 | 494 | 16 |
14 days | 197 | 342 | 591 | 219 | 19 | ||
21 days | 176 | 164 | 485 | 203 | 34 | ||
S2 | 7 days | 99 | 360 | 556 | 184 | 88 | |
14 days | 76 | 758 | 266 | 124 | 78 | ||
21 days | 65 | 305 | 355 | 399 | 45 | ||
S3 | 7 days | 80 | 456 | 539 | 232 | 20 | |
14 days | 63 | 507 | 366 | 170 | 46 | ||
21 days | 77 | 303 | 536 | 240 | 51 | ||
S4 | 7 days | 102 | 751 | 451 | 201 | 21 | |
14 days | 120 | 725 | 569 | 251 | 54 | ||
21 days | 91 | 505 | 410 | 193 | 60 | ||
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Bile | S1 | 7 days | 135 | 102 | |||
14 days | 206 | 153 | |||||
21 days | 174 | 139 | |||||
S2 | 7 days | 94 | 188 | ||||
14 days | 68 | 116 | |||||
21 days | 140 | 365 | |||||
S3 | 7 days | 104 | 180 | ||||
14 days | 113 | 110 | |||||
21 days | 122 | 147 | |||||
S4 | 7 days | 118 | 172 | ||||
14 days | 120 | 173 | |||||
21 days | 140 | 214 |
Liver EROD and GST activities were obviously induced at all of the monitoring sites during the exposure period (P < 0.05), and the induction rate was higher at day 21 than at days 7 and 14 (Fig. 3B and C). The critical toxicity of any specific xenobiotic that is metabolized through both routes I and II depends on the capacity of the phase II enzymes to conjugate the reactive metabolites produced by phase I biotransformation reactions, in which EROD and GST activities play an important role.14 In the present study, EROD and GST activities exhibited a similar pattern, i.e., a positive correlation (Pearson coefficients >0.87; P < 0.05) was found between the induction of EROD and the increase of GST. Moreover, compared to the reference site S1, the induction rate of EROD was higher than that of GST in effluent-exposed fish caged downstream of the STP. The elevated levels of EROD and GST may result from the bioaccumulation of PhACs in fish from the STP effluents, which is known to occur for ROX, ERY and DIC based on the data from the laboratory studies.33,34,40 Generally, the maximum induction/inhibition rates of the three biomarkers (HSI, EROD and GST) were observed downstream of 0.1 km (S2), which clearly indicates that STP effluents induce biological effects in crucian carp, suggesting that crucian carp are among a group of the sensitive species to PhACs in effluent-exposed studies.
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Fig. 4 The RQ values (A) of single and total PhACs and the correlation (B) between the total RQ and EIBR values at different sampling sites. |
Both RQ and EIBR, based on the laboratory and field studies, respectively, have their place in the environmental risk assessment. EIBR presents an opportunity for a more holistic way of assessing the effects of environmental samples and wastes on ecosystems than RQ, which uses the detection of only a few classes of contaminants. However, the RQ could provide an assessment of the total toxicity contribution of the investigated PhACs that were included in the analytical survey, although it does not reflect the overall toxicity of the river near a STP. To quantitatively analyze the relationships between the two approaches, linear regression analysis was conducted (Fig. 4B). The result showed positive correlations between the ΣRQ of PhACs and EIBR values, with a correlation coefficient R value of 0.84 (P < 0.05). Because the EIBR and RQ are indicators of ecological risk, it appears as though the high risk of contamination in the effluent-receiving river is the result of the high contribution of STP effluents. Consistent with the previous studies, our results revealed that the pharmaceuticals- and metals-based RQ indexes and integrated biomarker response indexes showed clear positive correlations.42,43 This suggests that both the RQ and EIBR approaches are efficient instruments to provide an ecologically relevant assessment of the ecotoxicological risk and can facilitate the decision-making process in controlling the contamination at the downstream of STP.
Footnote |
† Electronic supplementary information (ESI) available. See DOI: 10.1039/c4em00472h |
This journal is © The Royal Society of Chemistry 2015 |