Aquatic Toxicology 150 (2014) 83–92

Contents lists available at ScienceDirect

Aquatic Toxicology journal homepage: www.elsevier.com/locate/aquatox

In situ effects of urban river pollution on the mudsnail Potamopyrgus antipodarum as part of an integrated assessment Radka Zounkova a , Veronika Jalova a , Martina Janisova a , Tomas Ocelka b , Jana Jurcikova b , Jarmila Halirova c , John P. Giesy d,e,f,g,h , Klara Hilscherova a,∗ a

Masaryk University, Faculty of Science, RECETOX, Kamenice 753/5, 62500 Brno, Czech Republic Institute of Public Health, Partyzánské nám. 7, 70200 Ostrava, Czech Republic c Czech Hydrometeorological Institute, Kroftova 2578/43, 61600 Brno, Czech Republic d Department Biomedical Veterinary Sciences and Toxicology Centre, University of Saskatchewan, 52 Campus Drive, Saskatoon, SK S7N 5B4 Saskatchewan, Canada e Department of Zoology, and Center for Integrative Toxicology, Michigan State University, East Lansing, MI, USA f Department of Biology and Chemistry and State Key Laboratory in Marine Pollution, City University of Hong Kong, Kowloon, Hong Kong Special Administrative Region g School of Biological Sciences, University of Hong Kong, Hong Kong Special Administrative Region h State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing, People’s Republic of China b

a r t i c l e

i n f o

Article history: Received 13 November 2013 Received in revised form 7 February 2014 Accepted 27 February 2014 Available online 11 March 2014 Keywords: Mortality Reproduction Passive sampling Gastropoda Sediment, In vitro

a b s t r a c t The freshwater mudsnail (Potamopyrgus antipodarum) is sensitive to toxicity of both sediment and water and also to the endocrine disrupting compounds (EDC) at environmentally relevant concentrations. This study determined effects of in situ exposure of P. antipodarum as a part of a complex assessment of the impact of a city metropolitan area with large waste water treatment plant (WWTP) for 0.5 million population equivalents on two urban rivers. The study combined the in situ biotest with detailed chemical analyses and a battery of in vitro bioassays of both sediment and water. Passive sampling of river water was conducted during the course of exposure of the mudsnail. P. antipodarum was exposed for 8 weeks in cages permeable to sediment and water at localities up- and down-stream of the city of Brno, Czech Republic and downstream of the WWTP in two rivers. Greater mortality and significantly decreased embryo production of P. antipodarum were observed immediately downstream of the city of Brno. P. antipodarum exposed at locations downstream of the metropolitan area and WWTP exhibited greater mortality, while numbers of embryos produced by surviving individuals were comparable or slightly greater than for individuals held at the least polluted location. Comparisons with results of chemical analysis and in vitro assays indicate occurrence of groups of compounds contributing to observed effects. Differences in mortalities of mudsnails among sites corresponded well with in vitro cytotoxicity and concentrations of metals. The results of this study confirm the applicability of this novel field biotest with P. antipodarum for the evaluation of the effects of river pollution on metazoans, especially as suitable in situ part of integrative contamination assessment. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Invertebrates are essential elements of aquatic ecosystems. Among key species in aquatic ecosystems are insects, molluscs and crustaceans. Sub-lethal effects of chemicals on different physiological processes, such as reproduction, growth, ecdysis, behavior, or morphological changes have been observed in all these groups

∗ Corresponding author. Tel.: +420 549 493 256; fax: +420 549 492 840. E-mail address: [email protected] (K. Hilscherova). http://dx.doi.org/10.1016/j.aquatox.2014.02.021 0166-445X/© 2014 Elsevier B.V. All rights reserved.

(Mothershead II and Hale, 1992; Lenihan et al., 1995; Oetken et al., 2004; Péry et al., 2008). Changes in these processes might affect populations, and consequently survival of species and structures or functions of ecosystems (Oehlmann et al., 2007; Oetken et al., 2004). Therefore, surrogate and sentinel species are needed for the assessment of the effect of contaminated water and sediment on these organisms. Some organisms in the class Gastropoda have been found to be sensitive to toxicity associated with sediments and surface waters, and specifically effects of (xeno-)hormones (Schulte-Oehlmann et al., 2000; Tillmann et al., 2001). The freshwater mudsnail

84

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

(Potamopyrgus antipodarum) is a cosmopolitan organism with some advantages for use as an indicator organism, including continuous fertility of parthenogenetic females, few maintenance requirements and sensitivity to environmentally-relevant compounds that might affect reproduction (Duft et al., 2003a, b; Jobling et al., 2004; Mazurová et al., 2008; Stange et al., 2012). Because this mudsnail can live under various environmental conditions, which is documented by its worldwide distribution, it is a suitable model organism for use in biotests with water or sediment of different physical and chemical characteristics. Effects of waste water treatment plant (WWTP) effluents and estrogenic compounds have been investigated using several species of molluscs, including freshwater gastropods Viviparus viviparus or Planorbarius corneus (Ramshorn snail) (Benstead et al., 2011). The ramshorn snail exhibited significant concentration-dependent increase in fecundity and in overall duration of the reproductive cycle in adult snails exposed to WWTP effluent (Clarke et al., 2009). P. antipodarum has been shown to be affected by chemicals with endocrine disruptive potential such as bisphenol A, organotins, estradiols, alkylphenols, UV screens or fadrozole. Greater reproduction of this mudsnail was observed after exposure to artificial sediment spiked with these chemicals (Duft et al., 2003a, b; Gust et al., 2010b; Jobling et al., 2004; Oehlmann et al., 2000; Schmitt et al., 2008). P. antipodarum was also successfully used as a test organism in contact tests with whole sediment (Galluba and Oehlmann, 2012; Mazurová et al., 2008; Schmitt et al., 2010a, 2011), in laboratory tests with waste water (Jobling et al., 2003, 2004) or in on-site flow-through tests with whole WWTP effluent (Magdeburg et al., 2012; Stalter et al., 2010). The 28–56 day sediment contact test with reproduction of P. antipodarum as the measurement endpoint was selected for standardization within the OECD framework (Duft et al., 2007; Oehlmann et al., 2007). While instrumental analyses of water or sediment provide information on presence and concentration of known target substances, there can be also a pool of unknown substances contributing to observed effects in these matrices. In addition, all effects of known substances and especially their interactions are not known. For this reason it is appropriate to use biotests and chemical analyses simultaneously, especially in cases of complex pollution. Both in vitro and in vivo tests can be conducted in the laboratory with water and/or sediment collected from the field. However, some characteristics of the sampled materials representing the environmental matrices can change after their removal from the environment. It is therefore difficult to simulate the real environmental situation and varying conditions in laboratory exposure, especially for dynamic river ecosystems. Thus, the most relevant way of exposure is direct field exposure of model species (Burton and Nordstrom, 2004). There is little information on suitable model species of molluscs to be used for studies directly in field. Only two studies have been published on in situ exposure of P. antipodarum, indicating its potential applicability in assessment of contamination in field (Gust et al., 2010a; Schmitt et al., 2010b). The overall objective of the research, results of which are presented here, was characterization of effects of a city with large municipal WWTP on urban rivers pollution. A major focus of this study was characterization of the influence of in situ exposure to river sediments and water on survival and reproduction of P. antipodarum and examination of the sensitivity of this species and its suitability for direct exposure in urban rivers. Another important aim was to investigate the relationship between results of the in situ contact biotest with data from chemical analyses and in vitro biotests of water and sediment. In addition to collection of grab samples also passive sampling of water was conducted to get more representative estimate of time-weighted concentrations of contaminants. In this study, two types of passive samplers were used. These included Polar Organic Chemical Integrative Samplers

(POCIS), which sequester waterborne hydrophilic contaminants, and semipermeable membrane devices (SPMD) for monitoring waterborne hydrophobic pollutants. Detailed characterization of contamination included in vitro biotests on cytotoxicity, dioxin-like toxicity, (anti)estrogenicity and (anti)androgenicity, and chemical analysis of several classes of pollutants, including hydrophobic organic pollutants, pharmaceuticals, pesticides, perfluorinated organic compounds (PFOCs) and alkylphenols, some of which are known as endocrine disrupting chemicals (EDCs) (Groshart and Okkerman, 2000). 2. Materials and methods 2.1. Localities and sampling design Selection of locations was based on a larger project concerned with a long-term assessment of impact of the metropolitan region of Brno (Czech Republic) on fluvial environment in two urban rivers Svratka and Svitava (Grabic et al., 2010). Brno, with 404,000 inhabitants, is the second-largest city in the Czech Republic with traditional sources of urban water pollution as sewage, industrial wastewater and surface runoff from construction sites and urban roads. A large WWTP with a capacity of 513,000 population equivalent is located downstream of the city and is processing waste waters from Brno and surrounding settlements. Waste water is subjected to primary (mechanical) treatment followed by biological stage of activation with pre-denitrification and anaerobic phosphorus removal (system of circulatory activation with change of anaerobic, anoxic and aerated zones). Excess activated sludge is then anaerobically stabilized (Brnˇenské vodárny a kanalizace, 2010; Ministry of the Environment of the Czech Republic, 2010). Samples of sediments and grab and passive samples of water were taken from six locations. The study locations were chosen to examine the contamination in the rivers, its changes along the flow of the rivers through and downstream of the city of Brno and effects on biota. Two sampling locations (upstream and downstream the city) were chosen at each river to observe the influence of the city, and two sampling locations were located downstream the confluence of both rivers and the WWTP effluent discharge to observe the impact of the WWTP. Thus, this field study included the following locations (Fig. 1): Kníniˇcky (location 1a) – Svratka upstream of the city of Brno, downstream of the dam of Brno reservoir; Pˇrízˇrenice 1 (location 1b) – Svratka downstream of Brno, upstream of the confluence with the Svitava River; Bílovice nad Svitavou (location 2a) – a small town on the Svitava River upstream of Brno; Pˇrízˇrenice 2

Fig. 1. Map of metropolitan region of Brno and sampling locations. 1a – Kníniˇcky – Svratka upstream of Brno; 1b – Pˇrízˇrenice 1 – Svratka downstream of Brno, 2a – Bílovice nad Svitavou – Svitava upstream of Brno; 2b – Pˇrízˇrenice 2 – Svitava downstream of Brno; 3 – Modˇrice – Svratka downstream of confluence with Svitava, downstream of WWTP; 4 – Rajhradice – 3 km downstream of the confluence of the rivers Svratka and Svitava, downstream of the regional WWTP.

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

(location 2b) – Svitava downstream of Brno, upstream of the confluence with Svratka River; Modˇrice (location 3) – Svratka downstream of the confluence with the Svitava River, downstream of the WWTP effluent discharge; Rajhradice (location 4) – a small town 3 km downstream of the confluence of the Svratka and Svitava Rivers, downstream of the WWTP effluent discharge. Only sediment was sampled at location 3. Additionally, samples of effluent water were taken and passive samplers were installed into the effluent of the WWTP. Passive samplers were exposed from the beginning of May 2008 until the beginning of June 2008 (4 weeks). Standard sampling arrangement described in Grabic et al. (2010) and Jalova et al. (2013) with a combination of POCIS (Polar Organic Chemical Integrative Sampler) and several SPMDs (Semipermeable Membrane Device) was used. SPMD and POCIS were obtained from Exposmeter AB, Tavelsjo, Sweden. One POCIS was used for both chemical analysis and bioassay testing. Two SPMDs were used in duplicates for chemical analysis, one SPMD was used for toxicity assessment. SPMDs for chemical analysis contained performance reference compounds (PRC) used as onsite SPMDs calibration. Four deuterated PAHs ([2 H10 ]acenaphthene, [2 H10 ]fluorene, [2 H10 ]phenanthrene, and [2 H12 ]chrysene) and four 13 C -labeled PCBs (PCB 3, 8, 37, and 54) were used as PRCs. Passive 12 samplers were placed in special racks in protective shroud places. Samplers were installed and deployed in 0.5–1 m water depth and exposed for a 4-week period. Temperature was recorded in detail (every hour) during the 4-week deployment of passive samplers by temperature loggers placed on the sampling racks. The temperatures and the overall ranges of their fluctuations (day–night, min–max) were very similar across studied river sites (see Table S1 in Supplementary Materials). Composite bottom sediment samples and grab samples of water were collected on May 12, 2008. Samples of sediment were taken from the upper layer of fresh sediments by use of a method designed by the Czech Hydrometeorological Institute, which was developed in accordance with ISO 5667-12 standard (ISO, 1995). Representative composite sediment samples were prepared by thorough mixing and homogenization of surface sediments collected from several spots (six to eight individual grabs) within each sampling locality (10 m2 area). Samples of water were taken by use of a telescopic sampling device. Passive sampling was conducted according to validated protocols and general rules for passive sampling (Huckins et al., 1993, 1999) and EN ISO/IEC 17025 standard (ISO, 2005). All samples were refrigerated after collection (4 ◦ C) and transported to laboratory, where they were stored at −18 ◦ C until the analysis, which was started within one month of the sampling. 2.2. Identification and quantification of residues Samples of water, sediment, and organic extracts of SPMD and POCIS samplers were analyzed for wide range of organic compounds. Extraction and cleanup of passive samplers, as well as the methodology for chemical analysis of studied compounds, have been described previously (Grabic et al., 2010; Jalova et al., 2013). Briefly, SPMDs were dialyzed with hexane, POCIS eluted with methanol:toluene:dichloromethane (1:1:8, v/v/v; Jalova et al., 2013). Sediment samples were homogenized, lyophilized and sieved prior to analysis. Metals (Al, As, Ba, Cd, Co, Cr, Cu, Mo, Ni, Pb, Se, Ti, Zn) were extracted from sediment (fraction < 2 mm) by nitric acid and measured by ICP-MS method (Elan 6100 with autosampler AS-90; Perkin-Elmer Sciex, USA). Determination of the total mercury content in sediments was performed by means of atomic absorption spectrophotometric (AAS) method using a single-purpose cold vapor Advanced Mercury Analyzer AMA-254 (ALTEC Ltd., Czech Republic). Organic pollutants were extracted from sediments (fraction < 1 mm) by microwave extraction with

85

hexane:acetone mixture and cleaned on a silica gel column. A portion of each organic extract of POCIS, SPMD and sediments was transferred into DMSO for testing in bioassays. After removal of particulate matter and addition of internal standards water samples were directly injected on analytical HPLC column Phenomenex Aqua 5 ␮ C18 125 A (50 mm × 2 mm), where individual analytes were separated and further detected by MS/MS system. All methods were validated in accordance with EN ISO/IEC 17025 standard (ISO, 2005). Extracts of SPMD and sediments were analyzed for polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), organochlorine pesticides, hexachlorobenzene (HCB), ␣-, ␤-, ␥-stereoisomers of hexachlorocyclohexane (HCH), dichlorodiphenyltrichloroethane (DDT) and its degradation products dichlorodiphenyldichloroethylene (DDE) and dichlorodiphenyldichloroethane (DDD), triclosan and its environmental transformation product methyl triclosan (Me-triclosan) and polybrominated diphenyl ethers (PBDEs), expressed as the sum of congeners. Eluates from POCIS were analyzed for polar pesticides, pharmaceuticals and perfluorinated organic compounds (PFOCs). Samples of water were analyzed for polar pesticides, pharmaceuticals and alkylphenols. A complete list of individual pollutants analyzed is attached in footnotes to Table 1b. Concentrations of HCB, HCHs, PCBs, PBDEs, DDT and its degradation products as well as triclosan and Me-triclosan after derivatization were determined by GC/MS–MS using isotope dilution. GC/MS was used for quantification of PAHs. PAHs with more rings were analyzed by HPLC using a FLD detector with deuterated internal standards. Polar pesticides, pharmaceuticals, PFOCs and alkylphenols were measured by standard method direct injection HPLC/MS–MS. Set of carbon 13 C12 -labeled internal standards were included in the analyses as described in Jalova et al. (2013). The native standards were purchased from Dr. Ehrenstorfer, AccuStandards, and Absolute Standards via Labicom. 2.3. In vitro bioassays Four transactivation reporter gene bioassays were used to measure receptor-mediated potencies of organic extracts of sediments and passive samplers by procedures described in Jalova et al. (2013). AhR-mediated (dioxin-like) potency was determined by use of the H4IIE-luc bioassay, a rat hepatoma cell line, which contains a luciferase reporter gene under control of dioxin-responsive enhancers (DRE) (Sanderson et al., 1996; Hilscherova et al., 2001; Villeneuve et al., 2002). Estrogen receptor (ER)-mediated potency was evaluated by use of the MVLN bioassay, a human breast carcinoma cell line which has been transfected with a luciferase gene under control of estrogen receptor activation (Demirpence et al., 1993; Hilscherova et al., 2002; Freyberger and Schmuck, 2005). (Anti)androgenicity was assessed in a bioassay with MDA-kb2 cells, a human breast carcinoma cell line stably transfected with luciferase reporter gene under control of functional endogenous androgen receptor (AR) and glucocorticoid receptor (GR) (Wilson et al., 2002). Cells were cultured in dark in incubator at 37 ◦ C and assays conducted on 96-well microplates. Approximately 24 h after plating, cells were exposed to samples, calibration reference or solvent control. Standard calibration was performed with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD; Ultra Scientific, USA; dilution series 1–500 pM) in case of H4IIE-luc, 17␤estradiol (E2; Sigma–Aldrich, Czech Republic; 1–500 pM) for MVLN and dihydrotestosterone (DHT; Sigma–Aldrich, Czech Republic; 1 pM–10 ␮M) for MDA-kb2. Effects of extracts on MVLN and MDAkb2 cells were assessed either singly or in combination with competing endogenous ligand. Antiestrogenicity was determined by simultaneous exposure of sample extract and 17␤-estradiol

86

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

Table 1a Concentration of pollutants in sediment (per g dry mass of sediment) from the six study sites around Brno. Sediment

1a

2a

1b

2b

3

4

Triclosan Me-triclosan Sum of PBDEs Sum of PCBs Sum of HCHs HCB p,p -DDT Sum of DDT and metabolites

ng/g 0.6 0.1 0.07–1.15 5.9 1.0–1.9 0.85 0.38 1.47–1.86

ng/g 2.4 0.73 0–3 137 0–1.85 1.1 1.3 5.2–6.0

ng/g 13 2.7 1.06–2.84 123 0–0.98 2.1 680 757

ng/g 5.5 1.4 0.48–1.84 20 0–0.96 0.74 5.7 12.0–12.2

ng/g 34 10 1.08–3.45 38.3 0–1.68 2.0 5.4 20.5–20.8

ng/g 57 7.2 1.75–3.67 59.5 0–0.79 1.7 5.4 29.7

Sum of PAHs Al As Ba Cd Co Cr Cu Hg Mo Ni Pb Se Ti Zn

␮g/g 2.3 1630 0.35 43 0.12 4.28 5.72 5.86 0.01 0.015 7.84 16.1 0.37 20.4 44.8

␮g/g 15.2 3910 1.47 100 1.72 5.35 25.7 23.5 0.12 0.037 15.6 36.6 0.49 31.8 178

␮g/g 14.6 6070 1.82 179 2.51 8.03 30.1 79.6 0.89 0.089 24.4 68.9 0.56 50.6 329

␮g/g 19.2 3530 1.44 96 0.46 3.09 16.1 24.5 0.68 0.051 11.7 22.9 0.49 32.3 155

␮g/g 12.3 5030 1.89 124 1.32 5.49 24.1 51.7 0.96 0.087 15.5 50.6 0.54 39.6 214

␮g/g 26 5490 0.91 130 1.12 6.19 25.8 59.2 0.96 0.098 18.3 48.7 0.63 47.7 259

Ranges: the sum of detected compounds – the sum of detected compounds plus limits of detection of non-detected compounds.

Table 1b Concentrations of pollutants in water and passive samplers from study sites and WWTP effluent. 1a

2a

1b

2b

WWTP effluent

4

Water Sum of pesticides Sum of sulfonamides Sum of other antibiotics Sum of other pharmaceuticals Sum of alkylphenols + BPA

ng/l 30–1000 11–58 71–179 38–48 44–63

ng/l 885–1909 24–79 102–192 103–115 n.a.

ng/l 194–1151 0–85 69–181 39–68 5–45

ng/l 1050–1949 65–108 80–181 90–101 23–63

ng/l 961–1889 3290–3330 501–563 2150–2160 278–279

ng/l 534–1365 340–384 82–174 280–292 42–73

POCIS Sum of pesticides Sum of sulfonamides Sum of other antibiotics Sum of other pharmaceuticals Sum of PFOCs

ng/POCIS 523–627 44–61 29–87 115–120 2–7

– – – – –

– – – – –

ng/POCIS 3100–3190 165–173 14–23 397–401 24–27

ng/POCIS 10,370–11,720 3990–4140 534–682 12,550–12,610 140–175

ng/POCIS 1310–1420 287–297 30–36 722–726 14–17

SPMD Triclosan Me-triclosan Sum of PBDEs Sum of PCBs Sum of HCHs HCB p,p -DDT Sum of DDT and metabolites

pg/l 127 210 14.8–28.2 597 186–199 103 27.3 260

– – – – – – – –

pg/l 182 257 18.9–31.4 390 157–164 153 50.9 534

pg/l 659 446 24.9–31.2 2127 145–159 189 208 715

pg/l 34,005 13,991 173–175 2077 646–654 354 62.4 480

pg/l 3374 1779 32.6–39.1 1041 183–193 142 103 587

ng/l 14.4



ng/l 26.4

ng/l 60.0

ng/l 41.9

ng/l 36.2

Sum of PAHs

Ranges: the sum of detected compounds – the sum of detected compounds plus limits of detection of non-detected compounds. Sum of PBDEs: PBDE 28, PBDE 47, PBDE 99, PBDE 100, PBDE 153, PBDE 154, PBDE 183; sum of PCBs: PCB 28 + 31, PCB 52, PCB 101, PCB 118, PCB 138, PCB 153 + 168, PCB 170, PCB 180; sum of HCHs: alfa-HCH, beta-HCH, delta-HCH, gama-HCH; sum of DDT and metabolites: op-DDE, pp-DDE, op-DDD, pp-DDD, op-DDT, pp-DDT; sum of PAHs: Phenantrene, Anthracene, Fluoranthene, Pyrene, Benzo(a)anthracene, Chrysene, Benzo(b)fluoranthene, Benzo(k)fluoranthene, Benzo(a)pyrene, Benzo(g,h,i)perylene, Dibenzo(a,h)anthracene, Indeno(1,2,3-c,d)pyrene; sum of pesticides: 2,4,5-T, 2,4-D, 2,4,-DP (dichlorprop), acetochlor, alachlor, atrazin, azoxystrobin, bentazone, bromacil, bromoxynil, carbofuran, cyanazin, desethylatrazin, desmetryn, diazinon, dichlobenil, dimethoat, diuron, ethofumesat, fenarimol, fenhexamid, fipronil, fluazifop-p-butyl, hexazinon, chlorbromuron, chlorotoluron, imazethapyr, isoproturon, kresoxim-methyl, linuron, MCPA, MCPP (mecoprop), metalaxyl, metamitron, methabenzthiazuron, methamidophos, methidathion, metobromuron, metolachlor, metoxuron, metribuzin, monolinuron, nicosulfuron, phorate, phosalone, phosphamidon, prometryn, propiconazole, propyzamide, pyridate, rimsulfuron, simazin, tebuconazole, terbuthylazine, terbutryn, thifensulfuron-methyl, thiophanate-methyl, tri-allate; +clopyralid in POCIS; +desisopropylatrazin, desethyldesisopropylatrazin, 2-hydroxyatrazin in water; sum of sulfonamides: sulfapyridin, sulfamethazin, sulfamethoxypyridazin, sulfachloropyridazin, sulfamethoxazol; sum of other antibiotics: metronidazol, cefalexin, ofloxacin, norfloxacin, ciprofloxacin, enrofloxacin, erythromycin, trimetoprim; +doxycyclin in water; sum of other pharmaceuticals: diaveridin, carbamazepin, diclofenac; sum of alkylphenols + BPA: t-octylphenol, n-octylphenol, 4-n-nonylphenol, p-nonylphenol, monoNPE, diNPE, bisphenol A; sum of PFOCs: PFHxS, FHUEA, FOSA, N-methyl FOSA, PFOA, PFOS, PFNA. n.a. – value not available. – Sample not available.

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

(33 pM) and antiandrogenicity was tested in combination with dihydrotestosterone (1 nM) – given concentrations are near their EC50 value. Several dilutions of extracts were tested in triplicate to provide a concentration–response curve for each sample. After 24 h of exposure, medium was removed, and cells were washed with phosphate-buffered saline (PBS) and lysed. Intensity of luciferase luminescence corresponding to the respective receptor activation was measured by use of Promega Steady Glo Kit (Promega, USA) in case of assays with H4IIE-luc and MVLN cells and with prepared luciferase reagent (Wilson et al., 2002) in MDA-kb2 assay. Noncytotoxic sample concentrations to be used in bioassays with cell lines were determined by use of the neutral red uptake assay (Freyberger and Schmuck, 2005). At the end of the incubation period, neutral red solution (0.5 mg/ml of media) was added and cells incubated for 1 h at 37 ◦ C. Medium was removed, cells washed with PBS and lysed with 1% acetic acid in 50% ethanol. Absorbance was measured in a microplate spectrophotometer at 570 nm. Yeast strain of recombinant Saccharomyces cerevisiae constitutively expressing luciferase was used for detailed cytotoxicity assessment (Leskinen et al., 2005; Michelini et al., 2005). Statistical evaluation of in vitro bioassays was performed by nonlinear logarithmic regression of concentration–response curves (Graph Pad Prism, GraphPad® Software, San Diego, CA, USA). Relative potencies expressed as TCDD equivalents/E2 equivalents/DHT equivalents were calculated by relating the EC50 value of standard calibration with the concentration of the tested sample inducing the same response (Villeneuve et al., 2000). Cytotoxicity, antiestrogenicity and antiandrogenicity corresponded to the decrease in detected luminescence/absorbance signal given by solvent control in case of cytotoxicity and specified amount of competing standard ligand for the other effects. The IC50 values for antiestrogenicity and antiandrogenicity were calculated from concentration–response curves expressed in percentage of signal of competitive concentration of added natural ligand (33 pM E2, 1 nM DHT). For better clarity of the trends the values are expressed as an index of antiestrogenicity (AE) or antiandrogenicity (AA), which corresponds to reciprocal value of IC50 . Similarly, the index of cytotoxicity was derived as the reciprocal value of IC50 (or IC20 values in case the 50% response was not reached) for the cytotoxic response. Concentrations of analyzed compounds as well as the biological potencies determined in bioassays for SPMD extracts were recalculated to the concentrations in water to take into account the differences in sampling rates among SPMDs from different locations. Performance reference compounds (PRC) were used for in situ calibration of sampling rates. Details of the calculation are described in Jalova et al. (2013). Results of POCIS samples were compared on the basis of concentrations and toxic equivalents in sampler extracts (ng/POCIS). No correction of POCIS sampling rates was made because the water flow in sampling localities did not vary significantly. Also, only minor influence of water flow rate on the accumulation of pollutants into POCIS has been demonstrated (Li et al., 2010). 2.4. In vivo biotest P. antipodarum (Gray 1843) (Mollusca, Gastropoda, Caenogastropoda) is a parthenogenetic and ovoviviparous freshwater mudsnail, which is indigenous to New Zealand, but is currently widely distributed in aquatic environments around the world. These snails inhabit upper layers of aquatic sediments and feed on plants and detritus. A population from clean reference unpolluted location in sand-pond Stratov (district Nymburk, Czech Republic) was used in the test. The specimens were collected one month before the test. Adult individuals 3.6–3.9 mm length (Duft et al., 2003a) were used for the in situ test.

87

The field contact biotest with P. antipodarum was conducted at the six locations (see section on locations and sampling design) around the metropolitan area of Brno from April to June 2008. Cages consisted of stainless tube filter ½ (Valvosanitaria Bugatti, Castegnato, Italy), iron cover 5/4 (Pumpa, Brno, Czech Republic) and nylon net. Cages were laid on the river bottom, about one meter from the river bank and fixed ashore. Six cages were placed at every locality one week before start of the test. About 30 g of sediment was added into each cage and they were kept on river bottom to microfilm overgrow. Eight adult mudsnails were placed into each cage one week later. Thus, a total of 48 mudsnails were exposed at each location. Microfilm and sediment particles served as food and exogenous food was not provided. Cages were covered with nylon net and exposed in the river environment for eight weeks. After the exposure, the cages were transported to the laboratory, adult mudsnails were euthanized in 10% MgCl2 solution, dissected and reproduction success and mortality was evaluated. Reproduction was evaluated by counting the number of embryos in the brood pouch of 20 randomly selected maternal snails per site or of all surviving adults in case they were less than 20. Normality was checked by the Kolmogorov–Smirnov test and homogeneity of variance was confirmed by use of Levene’s test. The statistical significance of differences between groups was evaluated using non-parametric Mann–Whitney U test. Calculations were performed using Microsoft Excel® (Microsoft, Redmond, WA, USA) and Statistica® for Windows 6.0 (StatSoft, Tulsa, OK, USA).

3. Results The analyses of all types of samples collected in spring 2008 both by chemical analysis and in vitro bioassays revealed Svratka upstream of the city (1a) as generally the least polluted location (Tables 1a, b and 2). A detailed report of concentrations of individual compounds is included in Supplementary Materials (Tables S2–S5). Unfortunately, POCIS samplers from locations 2a and 1b and SPMD sampler from 2a were damaged or stolen during the exposure period, so it was not possible to conduct all comparisons. Sediments contained greater concentrations of several groups of non-polar pollutants (PCBs, DDTs, Me/triclosan, PAHs) and metals at location 1b compared to 1a, demonstrating impact of the sources in the city on contamination of the River Svratka. The analysis of SPMDs also confirmed greater concentrations of most of these pollutants directly downstream of the city. The Svitava River was more contaminated upstream of the city (location 2a), such that the effect of sources in the city was not obvious. Results from water grab sample and POCIS also document greater concentrations of pesticides and pharmaceuticals in the Svitava than in the Svratka River. Concentrations of triclosan and its metabolite, sulfonamides and some other pharmaceuticals were greater downstream of the WWTP. Also, concentrations of PAHs in sediments were greatest in the Svratka River, 3 km downstream of the WWTP (location 4). Concentration of most metals including the hazardous elements (MoA, 2009) have shown similar spatial trends with greatest concentrations in sediments from site 1b directly downstream of the Brno city and from the two most downstream sites under WWTP (3, 4, Table 1a). The differences were most pronounced for Cu, Pb and Zn. Results of the in vitro bioassays documented the presence of androgenic and estrogenic compounds in the polar fraction of WWTP effluent. Estrogenicity was also detected in POCIS from the Svitava River downstream of Brno (2b) and at the most downstream location (4; Table 2). At the same time, the greatest antiandrogenic potency was observed in sediments from these two locations.

88

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

Table 2 Results of in vitro assessment of extracts of samples from localities where cages with mudsnails were placed. 1a

2a

1b

2b

3

4

Sediment Dioxin-like toxicity [ng TCDD eq./g sed.] Estrogenicity [ng E2 eq./g sed.] Androgenicity [ng DHT eq./g sed.] Index of cytotoxicitya Index of antiestrogenicitya Intex of antiandrogenicitya

9.08 n.d. n.d. 22.4 1301 762

13.8 n.d. n.d. 99.1 2649 655

11.7 n.d. n.d. 166 1473 246

8.15 n.d. n.d. 66.8 325 1332

10.5 n.d. n.d. 160 2270 722

19.4 n.d. n.d. 290 1353 1602

POCIS Dioxin-like toxicity [ng TCDD eq./POCIS] Estrogenicity [ng E2 eq./POCIS] Androgenicity [ng DHT eq./POCIS] Index of cytotoxicitya Index of antiestrogenicitya Index of antiandrogenicitya

n.d. n.d. n.d. 146 220 270

– – – – – –

– – – – – –

n.d. 0.47 n.d. 323 620 324

WWTP effluent 1.77 2.77 27.7 1238 727 n.d.

n.d. 0.56 n.d. 156 423 375

SPMD Dioxin-like toxicity [pg TCDD eq./l] Estrogenicity [ng E2 eq./l] Androgenicity [ng DHT eq./l] Index of cytotoxicitya Index of antiestrogenicitya Index of antiandrogenicitya

n.d. n.d. n.d. 2.58 3.21 n.d.

– – – – – –

23.0 n.d. n.d. 13.1 14.3 8.02

10.6 n.d. n.d. 1.46 0.74 6.51

7.80 n.d. n.d. 11.6 5.83 6.17

7.58 n.d. n.d. 1.31 2.99 3.61

Standard deviations around the determined values were up to 20% for dioxin-like activity, cytotoxicity and anti/androgenicity and up to 15% for anti/estrogenicity. – Sample not available. n.d. Not detected. a Index of cytotoxicity, antiestrogenicity, antiadrogenicity = reciprocal value of IC50 (IC20 in case of cytotoxicity of SPMD extract). Units are [1/(g/ml)] for sediment extracts, [1/(POCIS/ml)] for POCIS extracts or [1/(l/ml)] for SPMD extracts, respectively.

Unfortunately, the passive samples from location 2a in this sampling (spring 2008) were lost, but POCIS exposed at this location in the same period in 2007 elicited estrogenicity (0.18 ng E2/POCIS). In general, extracts of sediments and SPMDs exhibited antiestrogenic and antiandrogenic potencies. Dioxin-like activity was detected in hydrophobic fraction of water (SPMD) downstream of Brno and in all sediments, but only in WWTP effluent in the case of POCIS. Alternatively, in vitro cytotoxicity was detected in all types of samples from all locations. In sediments, it was greatest downstream of the city on the Svratka River and at the localities downstream of the WWTP. Both dioxin-like toxicity and cytotoxicity of sediments and SPMD from the Svratka River were greater directly downstream of the city (location 1b) compared to upstream (1a). The in situ contact test with P. antipodarum after eight weeks exposure to river sediment and water led to various magnitudes of mortality of adult mudsnails and number of embryos at the study locations (Fig. 2). Mortality was proportional to the general magnitudes of pollution at locations. The least mortality was observed at the most upstream location on the River Svratka (1a), where the least contamination was determined by chemical analysis and

toxicity in vitro biotests. Alternatively, there was about three-fold greater mortality at the location on the Svratka River directly downstream of Brno (1b), which exhibited greater contamination of sediments (Tables 1a and b). The greatest mortalities were observed at locations 3 and 4 downstream of the city and of the WWTP, where greater concentrations of triclosan and Me-triclosan, polycyclic aromatic hydrocarbons, sulfonamides and other pharmaceuticals were detected. Reproduction of mudsnails varied among locations (Fig. 3). There were, on average, approximately two-times lesser numbers of embryos in brood pouches of adults exposed at locations directly downstream of Brno (1b, 2b) compared to the locations upstream of Brno on the same rivers (1a, 2a). In the case of the Svitava River, this difference was statistically significant. The number of embryos was greatest in mudsnails held in the Svitava River upstream of Brno, and was almost twice as great as for mudsnails held in the Svratka River upstream of Brno, which is considered to be the least polluted location. Numbers of embryos of surviving mudsnails from localities downstream of WWTP (3, 4), where the greatest mortality was observed, were, on average, slightly greater than those

number of embryos

20

*

*

* * *

15

10

5

0

1a Fig. 2. Mortality of adults of P. antipodarum after 8 weeks in situ exposure. The insert shows the location of sampling sites on the rivers. 48 individuals were exposed at each site (six cages of eight adult mudsnails).

2a

1b

2b

3

4

Fig. 3. Average number of embryos in the brood pouch of adults of P. antipodarum after 8 weeks in situ exposure. Number of examined maternal snails was 20 for 1a, 2a, 1b, 2b; 10 for site 3, and 5 for site 4. *Significant difference.

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

at the least polluted location 1a (by 40 and 20%, respectively). However, this difference was not statistically significant, similarly to the comparison with the most downstream locations on the Svitava River before the confluence (2b). Alternatively, numbers of embryos in mudsnails at locations 3 and 4 were significantly greater (more than twice as great) when compared to mudsnails exposed in the Svratka River, directly downstream of Brno (1b, Fig. 3).

4. Discussion Effects of chemical pollution, especially of compounds classified as endocrine disruptors, were documented in previous studies with P. antipodarum. Also correlations between effects of some known estrogenic chemicals (17␣-ethinylestradiol, EE2, bisphenol-A, and 4-tert octylphenol) and of mixture of pollutants in WWTP effluent water on reproduction of P. antipodarum and estrogenic effect on fish, such as greater production of vitellogenin, have been observed (Jobling et al., 2003). Effects on reproduction of P. antipodarum after laboratory exposure to sediments containing compounds with dioxin-like, estrogenic and anti-androgenic activity were described also in our previous study with sediment from Lake Pilnok, which has been used as a dumping site for powdered waste coal (Mazurová et al., 2008). However, studies conducted in the laboratory cannot accurately simulate natural conditions and their changes to which organisms are exposed in field, as well as the complex mixture of compounds contained in WWTP effluents and in surface waters. Results of the present study revealed effects on reproduction and mortality of mudsnails exposed for 8 weeks at different localities in the Brno metropolitan area. Two studies testing in situ exposure of P. antipodarum have been published previously, but both of them used only 4 week exposure (Gust et al., 2010a; Schmitt et al., 2010b). During this shorter duration of exposure those authors observed as much as 25% mortality, which corresponds to the more upstream locations in the study, the results of which are presented here. The greater mortality at downstream locations observed in this study can be affected not only by the longer duration of exposure, but also by the magnitude and composition of pollution in the studied rivers. Only one of the two previous in situ studies presented data on concentrations of contaminants (Schmitt et al., 2010b), and the comparison to data presented here shows that there were greater concentrations of PAHs in our study, namely at the downstream locations. In this study exposure was characterized by use of both chemical analysis and in vitro bioassays of sediment and water. Since the results from grab water samples represent only the instantaneous conditions, a four week long passive sampling of water for both hydrophobic and hydrophilic pollutants was conducted during the course of exposure of the mudsnails in river to get representative estimate of longer-term exposure. A battery of in vitro tests was employed to provide complementary information to quantification of individual residues, which can only account for the known compounds and do not take into consideration the possible interactions within mixture. Quantification of individual residues showed that concentrations of some pollutants in the Svratka River increased during its course through Brno. The fact that greater concentrations of DDT, PCBs, metals, triclosan, and PAHs, were found in sediments from the Svratka River directly downstream of Brno than in sediment from upstream location, suggests that there are sources of these pollutants in the city such as urban runoff. Concentrations of some pollutants, including triclosan, Metriclosan and some pharmaceuticals increased downstream of the WWTP, which indicates that the WWTP despite its effectiveness and up-to-date methods of treatment could still contribute contaminants to the river. Similar observations were reported

89

during a recent study conducted in the same area in 2007, which also documented efficient treatment of the WWTP for cytotoxic compounds, xenoestrogens and xenoandrogens (Jalova et al., 2013). Results of two previous studies showed different trends in numbers of embryos in P. antipodarum after in situ exposure. Significantly more embryos in the brood pouch at more polluted locations were observed by Schmitt et al. (2010a,b), whereas Gust et al. (2010a) reported fewer embryos downstream of WWTPs. These contradictory results can be explained by different composition and concentrations of mix of compounds contained in water and sediments (e.g. estrogenic compounds may cause inhibition of reproduction at high concentrations (Jobling et al., 2003)). The differences are explained by the results of another study by Schmitt et al. (2011), which tried to identify compounds responsible for these effects. Effect-directed analysis showed that two out of six fractions stimulated reproduction of P. potamopyrgus, while two other fractions inhibited reproduction. Fractions which stimulated reproduction also exhibited greater estrogenic potency in the ER-LUC assay using reporter cell line BG-1. Results of the study demonstrate that some WWTP effluents and thus surface waters can contain compounds both stimulating and inhibiting reproduction. The resulting effect might depend on quantity and ratio of these compounds. These results correspond well with those of this study, where lesser numbers of embryos were observed in snails from both rivers directly downstream of Brno compared to upstream locations. In the case of the River Svratka there was more than two-fold greater mortality and two-fold lesser numbers of embryos at the location directly downstream of Brno (1b), which corresponds with the greatest magnitude of pollution by PCBs, DDTs and metals. This observation also corresponds with the greater in vitro cytotoxic potency of extracts from both sediments and SPMDs from this location (Table 2). In the case of the Svitava River, the situation was different, since the pollution of this river is already greater upstream of Brno. Relatively great pollution with PCBs, PAHs, polar pesticides, some pharmaceuticals and metals was observed upstream of Brno on the Svitava River. That pollution could be linked to recently increasing habitation density due to moving from the center of the city to suburbs upstream of location 2a. There is a small WWTP for this area with insufficient capacity for the new settlements which could contribute to the river pollution. Some of these pollutants could be related to greater numbers of embryos in mudsnails at location 2a. There was no strong influence of the city sources on the pollution in this river. Pollution of the Svitava River was mostly comparable or lower at the location downstream of Brno and mortality and number of embryos of P. antipodarum at this location were comparable to those observed at the least polluted location (1a). Greatest mortalities were observed on locations 3 and 4 downstream of Brno, the confluence of the two rivers and of the spot where WWTP effluent enters the river. Mortality was probably affected by cytotoxic compounds quantified by the in vitro assay and triclosan and Me-triclosan in sediments at both these locations. Moreover, the greatest concentrations of sulfonamides and other pharmaceuticals in water, and of polycyclic aromatic hydrocarbons and dioxin-like potency in sediments were found at the most downstream location 4. Despite greater mortality, number of embryos of surviving mudsnails was greater than in the Svratka River directly downstream of Brno (1b), but not significantly different (20–40% greater) than at the cleanest location. Individuals at these locations (3, 4) are exposed to multiple stressors, including cytotoxic compounds at the same time with hormonally active compounds. This is demonstrated by the estrogenic and androgenic potency in POCIS from WWTP effluent and estrogenic activity also at location 4, but also by ubiquitous

90

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

80

mortality (%)

600

mortality hazardous elements

400

60 40

200

sum of hazardous elements (mg/kg)

A 100

was also observed between mortality of exposed snails and cytotoxicity of organic extracts of sediments detected in vitro (Fig. 4B), which indicates relation to other chemicals than metals. Mudsnails are exposed to the whole mixture of both analyzed and unknown pollutants and it is not surprising that the total cytotoxicity corresponds better to the mortality than any individual group of compounds. These two correlations indicate that both inorganic and organic pollutants affect their survival.

20

5. Conclusions 0

0

1a

2a

1b

2b

3

4

B 100

400

mortality (%)

80

cytotoxicity

300

60 200 40

100

20

0

index of cytotoxicity

mortality

0

1a

2a

1b

2b

3

4

Fig. 4. Mortality of adults of P. antipodarum after 8 weeks in situ exposure and sum of concentrations of metals classified as hazardous elements (As, Cd, Co, Cr, Cu, Hg, Ni, Pb, Zn; MoA, 2009) in sediments (A), or index of cytotoxicity of extracts of sediments from study sites, respectively (B). Number of specimen as in Fig. 2.

antiestrogenic and antiandrogenic potencies. The relatively good reproduction at these locations where greater mortality was observed could be affected by the presence of endocrine disruptive compounds. The hormonal system of molluscs is insufficiently known. The estrogen receptor (ER) has been identified in P. antipodarum (Stange et al., 2012) and some hormones were detected in molluscs (Lafont and Mathieu, 2007), but there is a lack of information about their function. There is little information about effects of non-endocrine disruptive compounds on reproduction of P. antipodarum. It has been shown that nitrates, as well as fluorides and copper (Cu) nanoparticles reduce reproduction of P. antipodarum (Alonso and Camargo, 2011, 2013; Pang et al., 2012). Changes of reproduction and mortality of the freshwater mudsnail P. antipodarum can thus be the consequence of effects on different magnitudes of general stress, endocrine disruption through receptors, changes of metabolism of hormones or enzymes and others. Sediment and water samples from localities downstream of Brno and the regional WWTP contained relatively high concentrations of organic pollutants and metals known for their negative effects on biota. Some of these compounds, such as organochlorine pesticides, polychlorinated biphenyls, polycyclic aromatic hydrocarbons, pharmaceuticals, belong to the group of EDCs (Depledge and Billinghurst, 1999; Groshart and Okkerman, 2000). However, no correlation between concentrations of any single chemical or group of chemicals and numbers of embryos in mudsnails was observed. Alternatively, mortality was generally in a good accord with concentrations of hazardous metals (Fig. 4A). Negative effects of metals on survival of pulmonate snails have been documented in previous studies (Gupta et al., 1981; Laskowski and Hopkin, 1996; Allah et al., 1997). However, greater mortality at localities 3 and 4 might be caused also by contribution of other pollutants, which have not been analyzed in the sediments. Good correspondence

Results of the present study proved the suitability of freshwater mudsnail P. antipodarum as a model organism for in situ assessment of effects of urban rivers contamination on biota. It demonstrated effects of various sources of pollution in the studied area. The in situ assays with P. antipodarum document the presence of toxic compounds in the complex contaminant mixture in sediments as well as effects on reproduction in mudsnails. This is the first study that brings together this in situ test with simultaneous passive sampling for the determination of time-weighted exposure and detailed characterization of exposure through both chemical analysis and in vitro bioassays of both sediment and water. A battery of in vitro tests provided complementary information to chemical analysis taking into account also unanalyzed compounds and interactions within mixture. This approach enabled to indicate groups of compounds contributing to the observed effects. Chemical pollution resulting from runoff and waste waters from the Brno metropolitan area had negative effect on survival of the freshwater mudsnail. Greater mortality was observed to be consistent with concentrations of metals and in vitro cytotoxicity. Number of embryos was also affected by pollution from the Brno metropolitan area as well as by suburb sources with a small WWTP of insufficient capacity. The early development of embryos in the brood pouch reflects effects of those toxicants that immediately affect the general health condition and reproduction. Acknowledgements This research was supported by the Czech Ministry of Education (LO1214) and the European Union Seventh Framework Programme (FP7) under the Project SOLUTIONS with grant agreement No. 603437. Prof. Giesy was supported by the Canada Research Chair program, a Visiting Distinguished Professorship in the Department of Biology and Chemistry and State Key Laboratory in Marine Pollution, City University of Hong Kong, the 2012 “High Level Foreign Experts” (#GDW20123200120) program, funded by the State Administration of Foreign Experts Affairs, the P.R. China to Nanjing University and the Einstein Professor Program of the Chinese Academy of Sciences. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.aquatox. 2014.02.021. References Allah, A.T.A., Wanas, M.Q.S., Thompson, S.N., 1997. Effects of heavy metals on survival and growth of Biomphalaria glabrata Say (Gastropoda: Pulmonata) and interaction with Schistosome infection. J. Mollus. Stud. 63 (1), 79–86. Alonso, A., Camargo, J.A., 2011. Toxic effects of fluoride ion on survival, reproduction and behaviour of the aquatic snail Potamopyrgus antipodarum (Hydrobiidae Mollusca). Water Air Soil Pollut. 219 (1–4), 81–90. Alonso, A., Camargo, J.A., 2013. Nitrate causes deleterious effects on the behaviour and reproduction of the aquatic snail Potamopyrgus antipodarum (Hydrobiidae Mollusca). Environ. Sci. Pollut. R. 20 (8), 5388–5396.

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92 Benstead, R.S., Baynes, A., Casey, D., Routledge, E.J., Jobling, S., 2011. 17betaOestradiol may prolong reproduction in seasonally breeding freshwater gastropod molluscs. Aquat. Toxicol. 101 (2), 326–334. ˇ Brnˇenské vodárny a kanalizace, 2010. Odvádˇení a cˇ iˇstˇení odpadních vod/COV Brno–Modˇrice (Sewage water treatment at WWTP Brno), Available on-line on http://www.bvk.cz/o-spolecnosti/odvadeni-a-cisteniodpadnich-vod/cov-brnomodrice/ (in Czech). Burton, G.A., Nordstrom, J.E., 2004. An in situ toxicity identification evaluation method part II: field validation. Environ. Toxicol. Chem. 23 (12), 2851–2855. Clarke, N., Routledge, E.J., Garner, A., Casey, D., Benstead, R., Walker, D., Watermann, B., Gnass, K., Thomsen, A., Jobling, S., 2009. Exposure to treated sewage effluent disrupts reproduction and development in the seasonally breeding Ramshorn snail (Subclass: Pulmonata, Planorbarius corneus). Environ. Sci. Technol. 43 (6), 2092–2098. Demirpence, E., Duchesne, M.J., Badia, E., Gagne, D., Pons, M., 1993. MVLN cells – a bioluminescent MCF-7-derived cell-line to study the modulation of estrogenic activity. J Steroid Biochem. 46 (3), 355–364. Depledge, M.H., Billinghurst, Z., 1999. Ecological significance of endocrine disruption in marine invertebrates. Mar. Pollut. Bull. 39 (1–12), 32–38. Duft, M., Schulte-Oehlmann, U., Tillmann, M., Markert, B., Oehlmann, J., 2003a. Toxicity of triphenyltin and tributyltin to the freshwater mudsnail Potamopyrgus antipodarum in a new sediment biotest. Environ. Toxicol. Chem. 22, 145–152. Duft, M., Schulte-Oehlmann, U., Weltje, L., Tillmann, M., Oehlmann, J., 2003b. Stimulated embryo production as a parameter of estrogenic exposure via sediments in the freshwater mudsnail Potamopyrgus antipodarum. Aquat. Toxicol. 64 (4), 437–449. Duft, M., Schmitt, C., Bachmann, J., Brandelik, C., Schulte-Oehlmann, U., Oehlmann, J., et al., 2007. Prosobranch snails as test organisms for the assessment of endocrine active chemicals – an overview and a guideline proposal for a reproduction test with the freshwater mudsnail Potamopyrgus antipodarum. Ecotoxicology 16 (1), 169–182. Freyberger, A., Schmuck, G., 2005. Screening for estrogenicity and anti-estrogenicity: a critical evaluation of an MVLN cell-based transactivation assay. Toxicol. Lett. 155 (1), 1–13. Galluba, S., Oehlmann, J., 2012. Widespread endocrine activity in river sediments in Hesse, Germany, assessed by a combination of in vitro and in vivo bioassays. J. Soils Sediments 12 (2), 252–264. Grabic, R., Jurcikova, J., Tomsejova, S., Ocelka, T., Halirova, J., Hypr, D., Kodes, V., 2010. Passive sampling methods for monitoring endocrine disruptors in the Svratka and Svitava Rivers in the Czech Republic. Environ. Toxicol. Chem. 29 (3), 550–555. Groshart, C., Okkerman, P.C., 2000. Towards the Establishment of a Priority List of Substances for Further Evaluation of Their Role in Endocrine Disruption. European commission DG ENV, Delft, Zeist. Gupta, P., Khangarot, B., Durve, V., 1981. The temperature dependence of the acute toxicity of copper to a freshwater pond snail, Viviparus bengalensis L. Hydrobiologia 83 (3), 461–464. Gust, M., Buronfosse, T., Geffard, O., Mons, R., Queau, H., Mouthon, J., Garric, J., 2010a. In situ biomonitoring of freshwater quality using the New Zealand mudsnail Potamopyrgus antipodarum (Gray) exposed to waste water treatment plant (WWTP) effluent discharges. Water Res. 44 (15), 4517–4528. Gust, M., Garric, J., Giamberini, L., Mons, R., Abbaci, K., Garnier, F., Buronfosse, T., 2010b. Sensitivity of New Zealand mudsnail Potamopyrgus antipodarum (Gray) to a specific aromatase inhibitor. Chemosphere 79 (1), 47–53. Hilscherova, K., Kannan, K., Kang, Y.S., Holoubek, I., Machala, M., Masunaga, S., Nakanishi, J., Giesy, J.P., 2001. Characterization of dioxin-like activity of sediments from a Czech river basin. Environ. Toxicol. Chem. 20 (12), 2768–2777. Hilscherova, K., Kannan, K., Holoubek, I., Giesy, J.P., 2002. Characterization of estrogenic activity of riverine sediments from the Czech Republic. Arch. Environ. Contam. Toxicol. 43 (2), 175–185. Huckins, J.N., Manuweera, G.K., Petty, J.D., Mackay, D., Lebo, J.A., 1993. Lipid-containing semipermeable membrane devices for monitoring organic contaminants in water. Environ. Sci. Technol. 27 (12), 2489–2496. Huckins, J.N., Petty, J.D., Orazio, C.E., Lebo, J.A., Clark, R.C., Gibson, V.L., Gala, W.R., Echols, K.R., 1999. Determination of uptake kinetics (sampling rates) by lipid-containing semipermeable membrane devices (SPMDs) for polycyclic aromatic hydrocarbons (PAHs) in water. Environ. Sci. Technol. 33 (21), 3918–3923. ISO, 1995. ISO 5667-12:1995 Water Quality – Sampling – Part 12: Guidance on Sampling of Bottom Sediments. International Standard. International Organization for Standardization, Geneva, Switzerland. ISO, 2005. ISO/IEC 17025:2005 General Requirements for the Competence of Testing and Calibration Laboratories. International Standard. International Organization for Standardization, Geneva, Switzerland. Jalova, V., Jarosova, B., Blaha, L., Giesy, J.P., Ocelka, T., Grabic, R., Jurcikova, J., Vrana, B., Hilscherova, K., 2013. Estrogen-, androgen- and aryl hydrocarbon receptor mediated activities in passive and composite samples from municipal waste and surface waters. Environ. Int. 59, 372–383. Jobling, S., Casey, D., Rodgers-Gray, T., Oehlmann, J., Schulte-Oehlmann, U., Pawlowski, S., Baunbeck, T., Turner, A.P., Tyler, C.R., 2003. Comparative responses of molluscs and fish to environmental estrogens and an estrogenic effluent. Aquat. Toxicol. 65 (2), 205–220. Jobling, S., Casey, D., Rodgers-Gray, T., Oehlmann, J., Schulte-Oehlmann, U., Pawlowski, S., Baunbeck, T., Turner, A.P., Tyler, C.R., 2004. Erratum to:

91

“Comparative responses of molluscs and fish to environmental estrogens and an estrogenic effluent” (vol. 65, pg 205, 2003). Aquat. Toxicol. 66 (2), 207–222. Lafont, R., Mathieu, M., 2007. Steroids in aquatic invertebrates. Ecotoxicology 16 (1), 109–130. Laskowski, R., Hopkin, S.P., 1996. Effect of Zn, Cu, Pb, and Cd on fitness in snails (Helix aspersa). Ecotoxicol. Environ. Saf. 34 (1), 59–69. Lenihan, H.S., Kiest, K.A., Conlan, K.E., Slattery, P.N., Konar, B.H., Oliver, J.S., 1995. Patterns of survival and behavior in Antarctic benthic invertebrates exposed to contaminated sediments: field and laboratory bioassay experiments. J. Exp. Mar. Biol. Ecol. 192 (2), 233–255. Leskinen, P., Michelini, E., Picard, D., Karp, M., Virta, M., 2005. Bioluminescent yeast assays for detecting estrogenic and androgenic activity in different matrices. Chemosphere 61, 259–266. Li, H., Vermeirssen, E.L., Helm, P.A., Metcalfe, C.D., 2010. Controlled field evaluation of water flow rate effects on sampling polar organic compounds using polar organic chemical integrative samplers. Environ. Toxicol. Chem. 29, 2461–2469. Magdeburg, A., Stalter, D., Oehlmann, J., 2012. Whole effluent toxicity assessment at a wastewater treatment plant upgraded with a full-scale post-ozonation using aquatic key species. Chemosphere 88 (8), 1008–1014. Michelini, E., Leskinen, P., Virta, M., Karp, M., Roda, A., 2005. A new recombinant cellbased bioluminescent assay for sensitive androgen-like compound detection. Biosens. Bioelectron. 20, 2261–2267. Mazurová, E., Hilscherová, K., Jálová, V., Köhler, H., Triebskorn, R., Giesy, J.P., Bláha, L., 2008. Endocrine effects of contaminated sediments on the freshwater snail Potamopyrgus antipodarum in vivo and in the cell bioassays in vitro. Aquat. Toxicol. 89 (3), 172–179. Ministry of the Environment of the Czech Republic, Czech Environmental Inspectorate, 2010. Report on the reconstruction and modernization of the WWTP in Brno 2010; Available online on http://www.cizp.cz/(b1obdbr (in 4uyzcvq454q1mucvb)/default.aspx?id=511&ido=362&sh=-711127208. Czech). MoA, 2009. Direction No. 257/2009 Coll. for sediment use on agricultural soils. Ministry of Agriculture and Ministry of Environment of Czech Republic, Prague. (in Czech). Mothershead II, R.F., Hale, R.C., 1992. Influence of Ecdysis on the accumulation of polycyclic aromatic hydrocarbons in field exposed blue crabs (Callinectes sapidus). Mar. Environ. Res. 33 (2), 145–156. Oehlmann, J., Schulte-Oehlmann, U., Tillmann, M., Markert, B., 2000. Effects of endocrine disruptors on prosobranch snails (Mollusca: Gastropoda) in the laboratory. Part I: Bisphenol A and octylphenol as xeno-estrogens. Ecotoxicology 9 (6), 383–397. Oehlmann, J., Di Benedetto, P., Tillmann, M., Duft, M., Oetken, M., Schulte-Oehlmann, U., 2007. Endocrine disruption in prosobranch molluscs: evidence and ecological relevance. Ecotoxicology 16 (1), 29–43. Oetken, M., Bachmann, J., Schulte-Oehlmann, U., Oehlmann, J., 2004. Evidence for endocrine disruption in invertebrates. Int. Rev. Cytol. 236, 1–44. Pang, C.F., Selck, H., Misra, S.K., Berhanu, D., Dybowska, A., Valsami-Jones, E., Forbes, V.E., 2012. Effects of sediment-associated copper to the deposit-feeding snail, Potamopyrgus antipodarum: a comparison of Cu added in aqueous form or as nano- and micro-CuO particles. Aquat. Toxicol. 106, 114–122. Péry, A.R.R., Gust, M., Vollat, B., Mons, R., Ramil, M., Fink, G., Ternes, T., Garric, J., 2008. Fluoxetine effects assessment on the life cycle of aquatic invertebrates. Chemosphere 73 (3), 300–304. Sanderson, J.T., Aarts, J., Brouwer, A., Froese, K.L., Denison, M.S., Giesy, J.P., 1996. Comparison of Ah receptor-mediated luciferase and ethoxyresorufin-O-deethylase induction in H4IIE cells: implications for their use as bioanalytical tools for the detection of polyhalogenated aromatic hydrocarbons. Toxicol. Appl. Pharm. 137 (2), 316–325. Schmitt, C., Oetken, M., Dittberner, O., Wagner, M., Oehlmann, J., 2008. Endocrine modulation and toxic effects of two commonly used UV screens on the aquatic invertebrates Potamopyrgus antipodarum and Lumbriculus variegatus. Environ. Pollut. 152 (2), 322–329. Schmitt, C., Balaam, J., Leonards, P., Brix, R., Streck, G., Tuikka, A., Bervoets, L., Brack, W., van Hattum, B., Meire, P., de Deckere, E., 2010a. Characterizing field sediments from three European river basins with special emphasis on endocrine effects – a recommendation for Potamopyrgus antipodarum as test organism. Chemosphere 80 (1), 13–19. Schmitt, C., Vogt, C., Van Ballaer, B., Brix, R., Suetens, A., Schmitt-Jansen, M., de Deckere, E., 2010b. In situ cage experiments with Potamopyrgus antipodarum – a novel tool for real life exposure assessment in freshwater ecosystems. Ecotoxicol. Environ. Saf. 73 (7), 1574–1579. Schmitt, C., Streck, G., Lamoree, M., Leonards, P., Brack, W., de Deckere, E., 2011. Effect directed analysis of riverine sediments – the usefulness of Potamopyrgus antipodarum for in vivo effect confirmation of endocrine disruption. Aquat. Toxicol. 101 (1), 237–243. Schulte-Oehlmann, U., Tillmann, M., Markert, B., Oehlmann, J., 2000. Effects of endocrine disruptors on prosobranch snails (Mollusca; Gastropoda) in the laboratory. Part II: Triphenyltin as a xeno-androgen. Ecotoxicology 9 (6), 399–412. Stalter, D., Magdeburg, A., Oehlmann, J., 2010. Comparative toxicity assessment of ozone and activated carbon treated sewage effluents using an in vivo test battery. Water Res. 44 (8), 2610–2620. Stange, D., Sieratowicz, A., Horres, R., Oehlmann, J., 2012. Freshwater mudsnail (Potamopyrgus antipodarum) estrogen receptor: identification and expression analysis under exposure to (xeno-)hormones. Ecotoxicol. Environ. Saf. 75, 94–101.

92

R. Zounkova et al. / Aquatic Toxicology 150 (2014) 83–92

Tillmann, M., Schulte-Oehlmann, U., Duft, M., Markert, B., Oehlmann, J., 2001. Effects of endocrine disruptors on prosobranch snails (Mollusca: Gastropoda) in the laboratory. Part III: Cyproterone acetate and vinclozolin as antiandrogens. Ecotoxicology 10 (6), 373–388. Villeneuve, D.L., Blankenship, A.L., Giesy, J.P., 2000. Derivation and application of relative potency estimates based on in vitro bioassay results. Environ. Toxicol. Chem. 19 (11), 2835–2843.

Villeneuve, D.L., Khim, J.S., Kannan, K., Giesy, J.P., 2002. Relative potencies of individual polycyclic aromatic hydrocarbons to induce dioxinlike and estrogenic responses in three cell lines. Environ. Toxicol. 17 (2), 128–137. Wilson, V.S., Bobseine, K., Lambright, C.R., Gray, L.E., 2002. A novel cell line MDA-kb2, that stably expresses an androgen- and glucocorticoid-responsive reporter for the detection of hormone receptor agonists and antagonists. Toxicol. Sci. 66 (1), 69–81.

In situ effects of urban river pollution on the mudsnail Potamopyrgus antipodarum as part of an integrated assessment.

The freshwater mudsnail (Potamopyrgus antipodarum) is sensitive to toxicity of both sediment and water and also to the endocrine disrupting compounds ...
839KB Sizes 0 Downloads 3 Views