Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment

Journal Pre-proof Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment.

Anne-Katrin Müller, Nele Markert, Katharina Leser, David Kämpfer, Sarah E. Crawford, Andreas Schäffer, Helmut Segner, Henner Hollert PII:

S0269-7491(19)34620-2

DOI:

https://doi.org/10.1016/j.envpol.2019.113636

Reference:

ENPO 113636

To appear in:

Environmental Pollution

Received Date:

15 August 2019

Accepted Date:

15 November 2019

Please cite this article as: Anne-Katrin Müller, Nele Markert, Katharina Leser, David Kämpfer, Sarah E. Crawford, Andreas Schäffer, Helmut Segner, Henner Hollert, Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment., Environmental Pollution (2019), https://doi.org/10.1016/j.envpol.2019.113636

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Journal Pre-proof Manuscript Titel: Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment. Authors: Anne-Katrin Müllera*, Nele Markerta, Katharina Lesera, David Kämpfera, Sarah E. Crawfordab, Andreas Schäffera, Helmut Segnerc and Henner Hollertab aRWTH

Aachen University, Institute of Environmental Research, Worringer Weg 1, 52065 Aachen, Germany

bCurrent

affiliation: Goethe University Frankfurt, Department of Evolutionary Ecology and Environmental

Toxicology, Max-von-Laue-Str. 13, 60438 Frankfurt am Main, Germany cCentre

for Fish and Wildlife Health, University Bern, Länggassstr. 122, 3012 Bern, Switzerland

*Corresponding authors: [email protected]

Abstract: Little is known about sediment-bound exposure of fish to endocrine disrupting chemicals (EDC) under field conditions. This study aimed to investigate potential routes of EDC exposure to fish and whether sediment-bound contaminants contribute towards exposure in fish. Tench (Tinca tinca) and roach (Rutilus rutilus) as a benthic and pelagic living fish species, respectively, were sampled at the Luppe River, previously described as a “hotspot” for accumulation of EDC in sediment. A field reference site, the Laucha River, additionally to fish from a commercial fish farm as reference were studied. Blackworms, Lumbriculus variegatus, which are a source of prey for fish, were exposed to sediment of the Luppe River and estrogenic activity of worm tissue was investigated using in vitro bioassays. A 153-fold greater estrogenic activity was measured using in vitro bioassays in sediment of the Luppe River compared the Laucha River. Nonylphenol (NP; 22 mg/kg) was previously identified as one of the main drivers of estrogenic activity in Luppe sediment. Estrogenic activity of Luppe exposed worm tissue (14 ng 17β-estradiol equivalents /mg) indicated that food might act as secondary source to EDCs. While there were no differences in concentrations of NP in plasma of tench from the Luppe and Laucha, vitellogenin, a biomarker for

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Journal Pre-proof exposure to EDCs, was induced in male tench and roach from the Luppe River compared to both the Laucha and cultured fish by a factor of 264 and 90, respectively. However, no histological alterations in testis of these fish were observed. Our findings suggest that sediments substantially contribute to the overall EDC exposure of both benthic and pelagic fish but that the exposure did not impact gonad status of the fish.

Keywords: Endocrine disruption, sediment, Tinca tinca, Rutilis rutilis, Lumbriculus variegatus 1. Introduction: In the last decades, studies worldwide have demonstrated that various environmental pollutants accumulate in river sediments (Brinkmann et al. 2015; Gong et al. 2016; Li et al. 2019; Niehus et al. 2018). Moreover, studies across Europe have investigated endocrine activity in sediments using chemical and bioanalytical tools and found high concentrations of endocrine disrupting chemicals (EDCs). In particular, effect based methods (EBM) such as the yeast estrogen screen (YES), an in vitro bioassay, have demonstrated high concentrations of EDCs measured in terms of 17ß-estradiol equivalents (EEQs) ranging from 0.02 up to 15.6 ng EEQ/g (Peck et al. 2004; Viganò et al. 2008). In combination with chemical analysis, 17ß-estradiol (E2), ethinylestradiol (EE2), estrone (E1) and nonylphenol (NP) have been identified as major active estrogenic compounds in sediments (Hilscherova et al. 2002; Kinani et al. 2010; Li et al. 2019). Buchinger et al. (2013) described two “hot-spots” of EDC contamination in sediments at the Saale River and its tributaries in Germany. Ethinylestradiol equivalents (EEEQs) measured using the YES assay in sediments of the Calbe and Luppe River were approximately 10 and 37 times higher (7 and 55 ng EEEQ/g), respectivly, compared to all other sampling sites in the Saale catchment and, further, exceeded reported literature values by at least a factor of 4 (Grund et al. 2010; Viganò et al. 2008). Furthermore, NP and E1 were detected in high concentrations of 115 mg/kg and 20.4 µg/kg by GC and LC MS/MS (Buchinger et al. 2013) and were found by Müller et al. (2019) to mainly attribute to the endocrine activity of those sediments. Various adverse effects of aqueous EDC exposures have been reported for freshwater fish species in numerous field and laboratory studies (Bergman et al. 2013). However, only a few studies have investigated the bioavailability and impact of sediment-bound EDCs on aquatic organisms, especially fish (Thompson and Iliadou 1990; Verspoor and Hammart 1991). Laboratory exposure to field sediments containing NP, 4-tert-octylphenol, bisphenol A (BPA), E1, E2 and EE2 lead to hepatic vitellogenin (vtg) induction in male Japanese medaka (Oryzias latipes) (Duong et al. 2009). In contrast, Kolok et al. (2007) and Sangster et al. (2014) found no induction at the gene level of vtg (vg1) or the estrogen receptor α (ERα) in male fathead minnow (Pimephales promelas) exposed in either field or laboratory experiments to hormone contaminated sediment-water systems of the Elkhorn River,

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Journal Pre-proof Nebraska, USA. Whereas, hepatic expression of vtg and estrogen receptor α was reduced in female fathead minnow exposed to sediment-water systems of the Elkhorn River indicating an endocrine disruptive effect (defeminisation) in females (Sangster et al. 2016; Sellin et al. 2010; Zhang et al. 2015). Regardless of the different exposure scenarios, all abovementioned studies concluded that sediment-bound EDCs might become bioavailable to fish and serve as a source for exposure. Several direct and indirect routes of exposure must be considered under field conditions to evaluate the bioavailability of sediment associated EDCs. For example, life history traits such as benthic or pelagic habitat preferences have been proposed to influence exposure conditions (Fan et al. 2019; Goncalves et al. 2014; Gu et al. 2016). Direct contact to contaminated sediment due to benthic habitat preferences and living patterns might enhance the bioavailability of sediment-bound EDCs to fish, e.g., feeding on macroinvertebrates in the upper sediment layers or dormancy in sediment during wintertime. Concentrations of alkylphenols in various wild fish species inhabiting the East China Sea have been reported by Gu et al. (2016) to be related to food habits, living patterns and trophic transfer. NP was found to be higher in fish species feeding on benthic organisms (Gu et al. 2016). Benthic macroinvertebrates such as the oligochaete Lumbriculus variegatus and aquatic insect larvae of Chironomus riparius have been shown to accumulate EDCs in laboratory studies with spiked sediment (Liebig et al. 2005; Mäenpää and Kukkonen 2006), thus, it is important to further assess dietary exposure as a secondary route of exposure for sediment-bound EDCs to fish. In this study, we 1) characterized estrogenic activity and EDCs in different environmental compartments, namely water, sediment, macroinvertebrates and fish and 2) evaluate how these relate to biological responses in fish with respect to endocrine disruption. To assess EDC exposure towards fish, water and sediment as well as tench (Tinca tinca) as benthic living fish, roach (Rutilus rutilus) and pike (Esox lucius; caught only at the Luppe River) as pelagic living fish species were sampled at a highly contaminated river (Luppe River), a less contaminated river as field control (Laucha River). In addition to those purchased from a commercial fish farm as reference. EDCs were measured using LC MS/MS or the YES in vitro bioassay in water, sediment and fish blood plasma samples. Additionally, endocrine activity in extracts of blackworm (Lumbriculus variegatus), a macroinvertebrate species serving as prey for tench and roach, previously exposed to Luppe sediment under laboratory conditions was evaluated in the YES assay. Furthermore, evidence of endocrine disruption in fish was investigated by induction of vtg in tench and roach mucus samples in combination with gonad histology

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Journal Pre-proof 2. Material and Methods: 2.1 Study sites: The Luppe River is a tributary of the Saale River located between Leipzig and Merseburg, Germany. It is a slowmoving stream, approximately 50 km long, and is characterized by abundant macrophyte growth and thick layers of sediment (Kammerad et al. 2014) (Table A.1). According to the Water Framework Directive (WFD) evaluation criteria, the Luppe River has been described as a “heavily modified water body” with a “poor ecological status” and a chemical status that is “not good” (Kammerad et al. 2014). The Laucha River is another tributary of the Saale River in close proximity to the Luppe and, therefore, comparable in habitat characteristics but significantly lower in concentrations of EDCs in sediments (Buchinger et al. 2013).

2.2 Sediment, water and fish sampling: Sampling at the Luppe and Laucha Rivers was done in late July/early August 2017 after spawning season of tench and roach (Table A.1). Prior to sediment sampling, physicochemical parameters of the stream were measured (Table A.1). A 2 L water sample was taken by vertical pumping, cooled and stored at 4 °C for a maximum of 2 days before analysis. Sediment was sampled by means of a van Veen stainless-steel grabber and stainless-steel shovels, where possible. Pooled and homogenized sediment samples were stored in 30 L aliquots at 4 °C until use. After lyophilization (Christ Alpha 1-2, Martin Christ GmbH, Osterode am Harz, Germany) for 120 h, sediments were sieved < 1 mm and ground using a pestle and mortar for extraction and further analysis. Sediments and water samples were extracted for analysis of target EDCs (E1, E2, EE2 and NP) by LC MS/MS and through the YES bioassay (ISO/FDIS 19040-1:2018-03 2018). Extraction and bioassay protocols have been previously described and published in further detail by Reifferscheid et al. (2011) and Müller et al. (2019). Non-lethal electro-fishing was used to catch tench, rudd (Scardinius erythrophthalmus) and roach at both rivers. Pike were caught at the Luppe River. Additionally, roach, rudd and tench were purchased from a fish farm where they were cultured under controlled conditions (Fischfarm Schubert, Wildeshausen, Germany) for comparison to the field-collected fish. Mucus samples were taken from each fish (except pike) for vtg analysis (Mucus Collection Set; TECOmedical GmbH, Bünde, Germany). Fish were sacrificed and dissected according to procedures approved by animal welfare (TVT 2010). Routinely, fork length (± 0.1cm), body weight (± 0.1g) and sex were recorded for all fish. Blood samples were taken from the caudal vain or by heart punctuation using heparinized syringes, frozen in liquid nitrogen in 2 mL cryogenic vials and stored at -80 °C until use. Afterwards, gonads were

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Journal Pre-proof collected, weighed (± 0.001 g), placed in histology cassettes (neoLab Migge GmbH, Heidelberg, Germany) and fixated in Davidson’s solution (AppliChem GmbH, Darmstadt, Germany). After 24 h, the cassettes were transferred to 4% formalin (Sigma-Aldrich Chemie GmbH, Steinheim, Germany) and stored at 4 °C. Somatic indices were determined for each fish including Fulton’s condition index (K) as K = 100 x (total weight [g] / fork length3[cm]) and gonadosomatic index (GSI) as GSI = 100 x (organ weight [g]/ total weight [g]). Scales of each fish were carefully scraped off at the side between head and dorsal fin for age determination (further information see appendix).

2.3 EDC analysis in blood: Target EDC substances E1, E2, EE2 and NP were analyzed in fish blood plasma using LC MS/MS. The extraction method was adopted from Anari et al.(2002). Briefly, blood samples were carefully thawed on ice, immediately centrifuged at 10,000 g for 8 min in a temperature-controlled centrifuge (Hettich Retina 420R, Andreas Hettich GmbH & Co.KG, Germany) at 4°C. It was not possible to get enough blood sample during field sampling from every individual, thus, samples less than 50 µL could not be included in further analysis. 50 µL plasma samples were diluted in 0.5 mL milli-Q water spiked with 50 ng deuterated standard (EE2/E2/E1-2,4,16,16-d4 (99%) from CDN Isotopes, Canada, and 4-n-NP-2,3,5,6-d4 (99.2%) from Neochema, Germany) in 4 mL amber glass vials. After mixing, 2 mL ethyl acetate were added and vortexed for 1 min followed by 1 h incubation at room temperature. Organic phase was then transferred into a 1.5 mL amber glass vial and evaporated under a gentle stream of nitrogen. Dried samples were derivatized with 1 mg/mL dansyl chloride (>99 % Sigma-Aldrich, Germany) in acetone as described by Lin et al. (2007). Samples were stored at -20°C until measurement with LC MS/MS (Table A.3). Method blanks (n=6) were produced following extraction protocol without plasma sample. All measured concentrations of target compounds were blank corrected (see appendix for LC MS/MS settings, Table A.3). Limits of detection (LOD) were calculated as the mean blank plus 3 times the standard deviation, and limits of quantification (LOQ) as the mean blank plus 7 times the standard deviation. For NP the mean blank value was 1.6 ng/mL, the LOD 2.7 ng/mL and the LOQ 4 ng/mL (see Table A.3). In the blood samples, some NP values (n=10 out of a total of 53 samples) were very slightly below the LOQ. Since the peaks could be robustly integrated and these values were only slightly below these have been included in the data analysis. Measured concentrations of NP in 5 out of the 7 cultured tench were below the LOD, and these concentrations set at the LOD for the data analysis.

2.4 Sediment contact test with the blackworm and Luppe sediment: 5

Journal Pre-proof Sediment contact test with the blackworm Lumbriculus variegatus and Luppe sediment was conducted based on OECD Guideline 225 (Huff Hartz et al. 2018; OECD 225 2007). Briefly, treatments consisted of 1 L beakers containing 50 g dry weight (d.w.) of (1) artificial sediment, spiked with target compounds E1, E2, EE2 and NP in DMSO (< 0.1%) matching concentrations and general physicochemical properties of Luppe sediment (see Table A.2) as a positive control (Table A.4); (2) the same artificial sediment (OECD sediment) spiked with dimethyl sulfoxide (DMSO; < 0.1%) as a solvent control; (3) the same artificial sediment without spike as a negative control; and (4) native Luppe sediment. Three replicates per treatment were included, each containing 800 ml of overlying reconstituted water and ten individual worms (see appendix for more details). Extraction of whole organisms was done according to Watts et al. (2001) and whole body extracts of pooled test replicates were tested in the YES bioassay (ISO/FDIS 19040-1:2018-03 2018). The YES bioassay protocol has been previously described in further detail in Müller et al. (2019).

2.5 Biomarker analysis: Vtg was measured using the UltraSensitive Salmonid Vitellogenin ELISA kit (TECOmedical GmbH, Bünde, Germany). All steps were performed in accordance with the instruction manual using the provided materials. The absorbance of the colour reaction was measured at 450 nm and 405 nm using a multi-well plate photometer (TECAN infinite M200, Tecan Austria GmbH, Grödig, Austria). A calibration curve was established by plotting the standard range concentrations (0 - 1 ng/mL) against their absorbances and fitted using a four-parameter logistic curve and all calculated sample concentrations were corrected by the total protein concentration in the sample. Thus, vtg concentrations are stated as ng/mL per mgprotein. Total protein concentrations were determined by a colorimetric reaction using the Bicinchoninic Acid Kit (Sigma Aldrich Chemie GmbH, Steinheim, Germany). For measured absorbances which were below the calibration curve, these concentrations were set at the LOD for the data analysis.

2.6 Histopathology of gonads: Gonad samples were dehydrated and embedded in paraffin (Histosec, Merck, Darmstadt, Germany) according to standard protocols. Briefly, 3 μm stepwise paraffin sections (two per fish) were produced using a microtome (Microm HM 340 E, Microm GmbH, Walldorf, Germany), mounted on 3-aminopropyltriethoxysilane coated microscope slides (Sigma-Aldrich, Buchs, Switzerland) and dewaxed. After drying overnight, the slides were strained with Haematoxylin-Eosin (H&E).

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Journal Pre-proof Histopathological evaluation of the H&E stained gonad sections was done according to modified criteria from the “U.S. Biomonitoring of Environmental Status and Trends (BEST) Program” and the OECD Guidance Document No. 123 (Johnson et al. 2010) (see appendix for more detailed information).

2.7 Statistical analysis: Normality assumptions (Kolmogorov-Smirnov test, Shapiro-Wilk test) were tested prior to parametric analysis. To analyze the significance of variations between experiment data one-way analysis of variance (ANOVA) was used in combination with the Tukey post-hoc test. When criteria of Gaussian normal distribution were not met, data were analyzed using the Kruskal-Wallis test in combination with the Dunn’s post-hoc test (Köhler et al. 2007; Rosner 2015). Analysis was done with the open source program R and figures were created with GraphPad Prism 6 software (GraphPad Software, San Diego, USA). Potential outliers were identified with Grubb´s outlier test. The statistical significance was determined with a type I error (α) of 0.05. At the Luppe River only few roach (n=3) but more rudd (n=11) were caught, whereas, at the Laucha River only few rudd (n=3) and more roach (n=8) were present. Due to hybridization between the roach and rudd (Thompson and Iliadou 1990; Verspoor and Hammart 1991), classification was difficult and due to low sampling numbers the roach and rudd were pooled together and are further referred to as roach through this study. Furthermore, there were no differences in concentrations of NP in plasma when examining the influence of fish gender and age for tench (Kruskal-Wallis; p=0.5), therefore, NP values were pooled at each site for statistical analysis.

3 Results and Discussion: 3.1 EDC exposure of fish under field conditions: 3.1.1 EDCs in water: Target EDCs (NP, E1, E2 and EE2) were measured in water samples of both rivers using LC MS/MS and were greater at the Luppe River compared to the Laucha. Concentrations of NP at the Luppe (75.4 ng/L) exceeded that of the Laucha River (42.1 ng/L) by a factor of 2, whereas E1 was only detected in water samples from the Luppe (see Table 1). E2 and EE2 were not found in Luppe or Laucha water samples. The concentrations of the EDCs detected in the present study are within the range of those reported by other studies worldwide for EDC surface water concentrations of maximum 17 ng/L E1 (Kolodziej et al. 2004) and NP ranging from 6 ng/L (Kuch and Ballschmiter 2001) up to 15 µg/L (Petrovic et al. 2004) as reviewed in Campbell et al. (2006). While the measured concentrations of EDCs in the water of both rivers are typical of those from industrial and urban watersheds, they

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Journal Pre-proof were at least five times below the predicted no effect concentration for water (PNECwater) as well as environmental quality standards (EQS) of 330 300 and 100 ng/L for NP and E1, respectively (EU 2008/105/EC; Hillenbrand et al. 2016). As a result, waterborne exposure to EDCs at both rivers would be expected to be minor and subsequently effects on the endocrine system of fish would not be expected.

3.1.2 EDCs in sediment: While there were only minor differences in the relatively low EDC contamination in surface water of both rivers, concentrations of NP and E1 as well as endocrine activity expressed in EEQs detected in sediment samples of the Luppe River exceeded those found in sediment of the Laucha River by a factor of 91 (NP), 33 (E1) and 153 (EEQ), respectively (Table 1). EEQs as well as concentrations of EDCs determined for sediments at both rivers in the present study are consistent with findings of Buchinger et al. (2013) sampled in 2010 (55 and 0.86 µg/kg EEEQs Luppe and Laucha; NP 115 / 0.5 mg/kg; E1 20.4/0.31 µg/kg; E2 1.5/
Sampling site (country)

Matrices/ Reference

Luppe (present study; Germany)

Sediment

Laucha (present study; Germany)

Sediment

Water Water

EEQs [µg/kg] 61.2 ± 18.3 0.38 ± 0.21 Literature

Compound in sediment [µg/kg] or water [ng/L] NP

E1

E2

22342.6

67.3

>LOD
75.4

19.7


245.5

2.0


42.1



8

Journal Pre-proof Arun and Ouse (UK)

Peck et al. 2004

0.021 – 0.03

-

0.024 0.052

0.006 – 0.014

Rivers in Portugal

Céspedes et al. 2004

0.05 – 0.6

up to 1172



Danube (Germany)

Grund et al. 2010

0.03 – 1.3

6.5 - 1364

0.02 – 0.24


Rhine (Germany)

SchulzeSylvester et al. 2016

1.0 – 5.1

-

-

-

0.45 - 1.3

16.6 – 203.8



Yellow River (China) Wang et al. 2012 Lambro (Italy)

Viganò et al. 2008

1.9 – 15.6

3415

12


Pearl River system (China)

Zhao et al. 2011

9.8 - 101

11.4 28839



Limit of quantification: LOQ= meanblank + 7(SDblank); limit of detection: LOD= meanblank + 3(SDblank); see Table A.3 for LOQs and LODs; - not investigated

3.1.3 Estrogenic activity in the macroinvertebrate Lumbriculus variegatus: To further assess whether sediment-bound EDCs might accumulate in macroinvertebrates and, thus, dietary food sources might be a route of exposure to sediment-bound EDC contamination to fish, a sediment contact assay with L. variegatus exposed to Luppe sediment was conducted. EEQs measured in the YES in vitro bioassay of L. variegatus whole body extracts previously exposed to Luppe sediment were 14.4 ± 0.1 ng EEQ per mg worm tissue d.w., whereas extracts from the solvent and negative control treatment were below the LOD (<3.5 ng/L). Moreover, EEQs detected in the positive control, where worms were exposed to artificial sediment spiked with EDCs to match those of the Luppe sediment, were six times higher (81.9 ± 0.4 ng EEQ/mg d.w.). The estrogenic activity of extracts from L. variegatus observed in the YES assay suggests that EDCs were accumulated by L. variegatus, which is in agreement with other studies reporting that EDCs including NP can accumulate in the tissues of L. variegatus (Croce et al. 2005; Liebig et al. 2005; Mäenpää and Kukkonen 2006). For example, concentrations of NP were 123.9 ng/mg fresh weight in field collected oligochaetes from the Lambo River, where concentrations of NP in sediment were three times lower compared to the Luppe River. Moreover, Liebig et al. (2005) demonstrated in an exposure study using an artificial sediment spiked with radiolabelled EE2 that 84 % of the total radioactivity incorporated by the worm consisted of radiolabelled EE2. Similar findings were reported for other benthic macroinvertebrates as larvae of Chironomus riparius (Mäenpää and Kukkonen 2006; Ruhí et al. 2016). Since benthic macroinvertebrates such as worms and larvae of chironomids serve as one of the major food sources for adult fish, including tench and roach (Kammerad et al. 2012), dietary exposure must also be considered as a route of exposure to EDCs for fish (Croce et al. 2005; Liebig et al. 2005; Mäenpää and Kukkonen 2006). This was also reported by Gu et al (2016) who found higher concentrations of linear alkylphenols including NP in fish feeding on benthic organisms compared to other marine species inhabiting the Yangtze delta region.

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Journal Pre-proof 3.1.4 Uptake of EDCs in fish: Target EDCs (NP, E1, EE2 and E2) were analyzed in blood plasma samples of tench and roach to evaluate uptake and exposure conditions to endocrine disruptors under field conditions. However, only NP was detected in the plasma by LC MS/MS measurements. E1, E2 and EE2 were below the limit of detection (LOD; see Table A.3) possibly due to rapid metabolism (Fay et al. 2014). There were no differences in concentrations of NP in plasma when examining the influence of fish gender and age for tench (Kruskal-Wallis; p=0.5), therefore, NP values were pooled at each site for statistical analysis. NP in the plasma of tench ranged from 15 to 27 ng/mL at the Luppe (23 ± 8 ng/mL) and Laucha River (20 ± 8 ng/mL) (Figure 1; Table A.5), and were significantly elevated in fish from both rivers compared to concentrations in cultured fish (6.5 ± 6 ng/mL) (Kruskal-Wallis; p=0.001). However, there was no difference in the concentration of NP in plasma of tench between the two field sites (Kruskal-Wallis; p>0.05). Blood volume drawn from roach caught at the Luppe River was not sufficient for further analysis except for one female with 35 ng/mL NP. Concentrations of NP in plasma from roach caught at the Laucha River were 13 ± 7 ng/mL. Laboratory studies on NP metabolism in various fish species, including roach, demonstrated that concentrations of NP residues after oral administration or waterborne exposure vary among tissues with high concentrations found in liver and bile and lower concentrations in blood (Cravedi and Zalko 2005; Fay et al. 2014). Smith and Hill (2004) reported that both oral administration and waterborne exposure to NP resulted in a rapid uptake and biotransformation in the liver after 24h, followed by a slower depuration phase of the glucuronide conjugate of hydroxylated NP through the bile and feces. Half-life of 3H-NP residues in plasma of juvenile rainbow trout (Oncorhynchus mykiss) after an intravenous injection was 40 h, whereas half-lives for muscle and liver were 99 h (Coldham et al. 1998). The elevated concentrations of NP measured in the plasma of wild roach and tench in the present study suggest an exposure to EDCs under field conditions at the Laucha and Luppe River. While the concentrations of NP in plasma of the fish did not correspond with EDC concentration in the sediment, rapid metabolism of NP through biliary excretion might explain missing accumulation of NP in the blood(Fay et al. 2014). EBMs such as bioassays (e.g. YES) might be helpful for future studies to assess uptake and organ distribution of EDCs in bile as they are integrative and provide lower detection limits (Brack et al. 2018; Brack et al. 2019). However, the fact that NP, a substance banned from European countries, were detected in water and fish in the present study, indicates that dormant contaminants in the sediment still have the potential to impact water quality and are bioavailable for aquatic organisms. Moreover, the fish themselves, the tench and roach, investigated in the present study may also act as a route of dietary exposure for upper trophic level predatory fish such as pike. Concentrations of NP

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Journal Pre-proof measured in two-year-old pike caught at the Luppe River were approximately two-fold greater (49.4 ± 42.0 ng/mL) compared to tench (23 ± 8 ng/mL), although, this was not statistically significant (Kruskal-Wallis test p>0.05) (Table A.5). However, greater NP concentrations in pike might result from interspecific differences in NP metabolism, rather than indicate biomagnification towards a higher trophic level. Whole body burden of NP or estrogenic potential should be assessed in future research to confirm this hypothesis. Nevertheless, a recent study on bioaccumulation of EDCs in fish with different feeding habits showed that NP accumulated (bioaccumulation factor > 5000) in brackish carnivorous, planktivorous, and detritivorous fish, whereby concentrations of NP in carnivores were significantly higher compared to detritivores (Fan et al. 2019). Moreover, concentrations of NP and other EDCs in sediment were positively related to those in detritivores and planktivores (Fan et al. 2019), which supports the findings of the present study that dietary exposure can act as route of exposure to sedimentbound EDCs. The results from the present study highlight the importance of developing environmental quality standards (EQS) for sediment and sediment quality guidelines (SQGs) for EDCs for current legislation such as the European WFD. Several comprehensive approaches have suggested the use of EBMs as an integrative technique that bridge the gap between chemical contamination and ecological effects and EBMs have been recommended for implementation in the WFD (Brack et al. 2019).

3.2 Endocrine disruptive effects in tench and roach: 3.2.1 Biomarker response: Age distribution of the field fish was estimated by scale analysis, whereas age of cultured fish was known. Cultured tench consisted of one- and two-year-old fish. Age distribution of tench caught at both rivers varied from one-year old juveniles up to mature five-year-old individuals. Mean age of tench caught at the Luppe River were about 2 ± 0.9 a (juvenile), 3 ± 0.8 a (male) and 3.3 ± 1.2 a (female) years and were comparable to that of tench from the Laucha (Table A.5 and A.6). Furthermore, cultured roach and roach caught at both rivers were about two years old (Table A.5 and A.6). Somatic indices including GSI and Fulton’s conditions K index as well as histology of the gonads were analyzed to assess effects on the reproductive and overall fitness, while vtg induction was investigated as a biomarker of EDC exposure to tench and roach (Kroon et al. 2017). In this study, fish from both field sites and the cultured fish showed an overall good condition as indicated by Fulton’s condition K index above 1. Both tench and roach caught at the Luppe and Laucha River had a significantly higher condition index than cultured fish (Kruskal-Wallis, p<0.0001), except for tench from the Laucha River (Table A.5). Reduced GSI has been related previously to EDC exposure and occurrence of intersex (Jobling et al. 2002), but no differences in

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Journal Pre-proof GSI of tench and roach between sampling sites were observed in the present study (Kruskal-Wallis, p>0.05) (Table A.5). Measured vtg concentrations in male tench from the Luppe (454.2 ± 858 ng/mL per mgprotein) were significantly greater than concentrations of vtg in cultured male tench (1.5 ± 2.9 ng/mL per mgprotein; Kruskal- Wallis p ≤ 0.0019) and about 100-fold greater than in male tench of the Laucha River (47.7 ± 93 ng/mLper mgprotein) (Figure 1; Table A.5). Similarly, vtg concentrations measured in male roach from the Luppe (162.5 ± 178.4 ng/mL per mgprotein,) were 100-fold greater than the concentration in cultured male roach (1.5 ± 2.2 ng/mL per mgprotein; Kruskal- Wallis p ≤ 0.0007) and ranged from 18 up to 407 ng/mL per mgprotein in male roach from the Luppe (Figure 1). Compared to cultured fish, vtg was induced in male tench and roach from the Luppe River by a factor of 264 and 90, respectively, indicating an exposure to EDCs at the Luppe River. In contrast to the comparably low waterborne exposure to EDCs at the Luppe and Laucha River, vtg induction in males at the Luppe River may reflect the significant higher EDCs concentrations in sediments of the Luppe River. Another interesting finding of the present study was that vtg concentrations in females with similar age, in which vtg is considered to be naturally induced due to an endogenous, age- and development-dependent production during maturation (Allner et al. 2010), appeared to be lower at the Laucha River (5.2 ± 8.7 ng/mL per mgprotein) compared to the Luppe River (10.9 ± 11 µg/mLprotein). One important finding was the vtg induction in male fish at the Luppe River indicating an exposure to EDCs. These findings agree with previous reports on vtg induction in sole (Solea senegalensis) and medaka exposed to sediment-bound EDCs (Duong et al. 2009; Goncalves et al. 2014). Moreover, Hecker et al. (2002) found a weak but significant induction of vtg in plasma samples of male bream (Abramis brama) along the Elbe River with mean vtg concentrations around 260 ng/mL. While measured NP concentrations in water samples and suspended matter (SPM) at the Elbe River were six and forty times less, respectively, compared to concentrations found at the Luppe River in this study, the findings on vtg induction in male fish are consistent with the findings from the present study. Both the Luppe and Laucha Rivers were characterized during sampling in July by similar environmental conditions, with high temperature and low oxygen content (see Table A.1) as well as similarly low estrogenic activities of surface water. In contrast, measured concentrations of EDCs as well as estrogenic activity in the sediment were highly elevated at the Luppe compared to the Laucha River, suggesting that potential endocrine disruptive effects, as indicated by vtg induction, might be attributed to exposure of sediment-bound EDCs, direct or indirect through food. However, cause-effect relationships are difficult to assign under field conditions, where

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Journal Pre-proof multiple environmental factors such as temperature, nutrition, pathogens contribute to physiological responses (Burki et al. 2012) and biomarker induction in fish sampled at the Luppe River cannot be exclusively explained by exposure to sediment-bound EDC.

Figure 1: Mucus vitellogenin (vtg) concentration [ng/mL per mgprotein] (left) and concentrations of nonylphenol in blood plasma (right) of roach (Rutilus rutilus) (A, B) and tench (Tinca tinca) (C, D) from the Luppe, Laucha and of cultured fish. Boxplots show the 25th to 75th percentile with the median as horizontal line. Data points outside of this range indicated as dots are outliers. Kruskal-Wallis test (p ≤ 0.0019) in combination with the Dunn’s multiple comparison post-hoc test was used to analyze significant differences indicated by different letters among juveniles, males and females across sampling sites.

3.2.2 Histopathology of fish gonads: Evaluation of gonad histology revealed no pathological alterations in testis of male roach and tench caught at the Luppe River when compared to cultured and fish caught at the Laucha River. Testes of male tench cultured under controlled conditions at the fish farm ranged from entirely immature (stage 0) to mid spermatogenic testis (stage 2) but was mainly characterized by an early immature maturation phase (stage 1) where spermatogonia and spermatids are dominant but mature spermatozoa may also be present (see appendix, Table A.6, Figure 2 D -F). Field tench from both sampling sites showed testes in similar maturation stages ranging from immature to middevelopment (stage 0 to 2). Testes of cultured male roach, however, were entirely immature (stage 0). On the contrary, male roach from the Luppe River were assigned to a later maturation phase (stage 3), where all stages

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Journal Pre-proof could be observed but mature sperm were predominating (see appendix, Table A.6). Inflammations and perinucleolar oogenic cells in the testes (testis-ova) were observed in testes from cultured and wild fish at minimal to moderate levels. However, testis-ova occurred in 15 % of cultured tench and in only one single tench from the Luppe River (Figure A.2). This is consistent with the spontaneous prevalence of intersex in fish held under control conditions reported by Wolf (2011). Concentrations of NP in plasma in combination with biomarker vtg induction in male fish observed in the present study indicate that tench and roach at the Luppe River were exposed to EDCs. However, despite high, potentially bioavailable EDCs from the sediment, this was not reflected in the gonad histology of male fish and did not result in adverse effects on the organ level. These findings are consistent with other studies investigating endocrine disruption in fish populations inhabiting German rivers as well as the findings of Goncalves et al.(2014) on sole from the Portuguese Sado Estuary (Allner 2003; Allner et al. 2010; Hecker et al. 2002). Despite a significant induction of vtg in male bream (Abramis brama) and roach from the Elbe River or Schwarzbach River, a tributary to the Rhine River, little or no evidence of intersex occurred in these fish (Allner 2003; Hecker et al. 2002). Similarly, while vtg was induced in male fish compared to control animals, no histological alterations such as testis-ova were found in sole from the Sado Estuary in Portugal where EDCs accumulated in sediment (Goncalves et al. 2014). In contrast, multiple field surveys worldwide reported the prevalence of intersex (testis-ova) in wild cyprinid species including bream, common carp (Cyprinus carpio), barbel (Barbus barbus) and roach associated with exposure to anthropogenic pollutants such as WWTP effluents and EDCs (Jobling et al. 1998; Wang et al. 2018).

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Figure 2: Gonadal development stages of female (A-C) and male (D-F) tench (Tinca tinca). Ovaries of females cultured at a fish farm (A) were in stage 0 and 2, whereas female tench from the Luppe River showed different development stages between stages 0-2 (B) and 4 (C). Testis of cultured (D) and field fish (E-F) were similarly developed with stages between 0 to 2. Development stages were determined based on the predominate stage of the sperm cells in the gonads. In female fish oocytes develop from chromatin-nucleolar, perinucleolar (PN) and cortical alveolar follicle (CA) to vitellogenic (VTC) oocytes, undefined cell cluster (CC) were observed. Stages of male germ cells include spermatogonia (SG), spermatocytes (SC), Spermatids (ST) and final spermatozoa (SZ). Sections were stained with H&E.

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Journal Pre-proof Female tench mature with an age of three to four and spawning occurs in central Europe during summer till the end of July (Kammerad et al. 2012; Pinillos et al. 2003). At onset of maturation ovaries develop asynchronously (Breton 1980; Macrì et al. 2011; Pinillos et al. 2003) which was confirmed by the results of the present study where ovaries of female tench with an estimated age of four years from the Luppe River were characterized by oocytes at various development stages ranging from oogonia, early chromatin nucleolar and perinucleolar oocytes to late cortical alveolar or vitellogenic oocytes in the same ovarian section (Figure 2 C). Ovaries of females from the Laucha River as well as half of the female tench from the Luppe River with an estimated mean age of three years were dominated by pre-vitellogenic follicle in the perinucleolar stage (stage 0 to 1) or perinucleolar and cortical alveolar oocytes (stage 1 to 2) (Figure 2 B, Table A.6). Similarly, ovaries from cultured female tench with a mean age of one were predominantly in an early stage of maturation showing primarily cortical alveolar follicles (stage 2; Figure 2 A, Table A.6), which agrees with literature reports on ovarian structure of immature tench (Macrì et al. 2011). Ovaries from cultured female roach as well as roach from the Luppe and Laucha Rivers with a mean age of two varied from an undeveloped to mid-developed stage where most of the developing follicles were earlyto mid-vitellogenic. This was reported before for immature ovarian constitution of two-year-old roach (Allner et al. 2010), and no trend related to sampling sites were observed (see Table A.6). Comparison of maturation between the field sampled tench and cultured fish is difficult, since the cultured tench had a lower age and maturation might be different under optimal culture conditions e.g. due to nutrition or temperature regimes (Horooszewicz 1983). Further research is needed to investigate whether three-year-old female tench from the Luppe and Laucha River which had ovaries in an early stage of maturation might be delayed in their development. Another interesting finding of the present study was that in more than 80 % of the field sampled tench, small, defined clusters of a few or several cells were present, which could not clearly be assigned to frequent cell types in teleost ovaries (see Figure 2 B, Table A.7). Very few studies, however, have mentioned cell clusters in female gonads. While larger clusters of cells at different development stages have been described as female intersex including spermatogenic cysts (Johnson et al. 2010; Thomas and Rahman 2012; Tsai et al. 2011), smaller cell aggregates consisting of mitotically dividing oogonia have been interpreted as oocytes undergoing atypical meiosis (Allner et al. 2010). Since cell clusters in the present study comprised of smaller aggregates and did not resemble spermatogenic cells, these are presumed to consist of abnormal oocytes caused by atypical meiosis as reported by Allner et al. (2009). Meiotic changes as well as disturbed early oogenesis caused by EDC exposure were also reported for other vertebrates, such as mice. Mechanisms might include disturbances of the spindle apparatus and chromosome segregation (Susiarjo et al. 2007). However, further research is needed including

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Journal Pre-proof multiple sampling time points throughout the year and larger numbers of sampled fish to estimate whether this is relevant for the reproductive success of the population. Overall, NP uptake and vtg induction in male tench and roach observed at the Luppe River in the present study suggest that both species were exposed to EDCs. However, patterns of exposure were not clearly influenced by habitat preference, pelagic or benthic living, as reflected by the insignificant alterations of gonad histology in relation to sediment contamination with EDCs. Several unknown factors such as exposure duration of the fish at the sampling sites, migration from less polluted rivers, stocking of juvenile fish from aquaculture, life stage and species sensitivity and mixture effects must be addressed in future research to estimate whether or not sediment contamination has no lasting adverse effects on the endocrine system under field conditions.

4 Conclusion: Our results on vtg biomarker induction in male tench and roach sampled at the Luppe River, a “hotspot” for EDC accumulation in sediment, together with NP plasma concentrations indicate a potential endocrine disruptive effect and suggest that sediments act as a source of EDCs to fish. Besides direct exposure to the sediment, our results on the estrogenic activity of L. variegatus exposed to Luppe suggest that food act as secondary source for exposure to EDCs. However, exposure to EDCs, including sediment-bound exposure (direct or indirect through food), at the Luppe River did not lead to adverse effects on gonad level in the benthic living tench or the pelagic living roach. Endocrine disruptive effects might, thus, not be relevant for the reproductive success of these populations. Another interesting finding, however, that was not related to EDC concentrations in the sediment, was the occurrence of cell clusters in tench ovaries at both field sites. Further research is needed to evaluate the origin of those cell clusters While exposure to sediment-bound EDCs under normal conditions, as demonstrated in the present study, did not adversely affect fish reproduction, bioavailability of sediment-bound EDCs might significantly increase under extreme weather events such as in a flood event (Crawford et al. 2017; Müller et al. 2019; Wölz et al. 2009). Since EDCs have a high tendency to accumulate in sediments worldwide, future work should address how sediment-bound EDCs might affect fish during flood events or other conditions which may result in greater bioavailability of EDCs.

Acknowledgments:

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Journal Pre-proof We thank Gernot Quaschny for conducting the electrofishing, Jörg Ahlheim for his assistance and the local fishing community of Merseburg for their help. We are thankful for the support of the TecoMedical Group for providing help with vtg ELISA training and data evaluation.

Funding: This study is supported by the RWTH Aachen University, as part of the German Excellence Initiative via the German Research Foundation (DFG); and the SOLUTIONS project (Grant No. 603437).

Competing interests: The authors declare that they have no competing interests.

Authors´ contributions: A-KM has been responsible for the concept of the study, sampling campaigns, drafted the manuscript and performed major parts of the practical work. NM contributed to sampling, biomarker and gonad analysis. KL performed the sediment contact assay. DK and AS contributed to LC MS/MS method development and EDC analysis. HS contributed to histopathological evaluation, provided guidance and conceptual input, SC and HH were responsible for the study design, and supervised the study. All authors read, improved and approved the final manuscript.

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24

Appendix Table A.1: Water parameters of the two sampling sites at the Luppe and Laucha River 2017.

Coordinates

Water parameters

Sampling site

N

E

Width [m]

Depth [cm]

Velocity [m/s]

pH

Temper -atur [°C]

Luppe Laucha

51 23 08.0 51 23 47.3

12 00 33.2 11 59 13.4

17.6 1.3

90 40

0.6 0.2

7.8 7.3

22.8 22.6

Conductivity [µS] 1460 1990

O²[mg/L] 3.8 2.7

Table A.2: Physicochemical characteristics and metal concentrations [mg/kg dry weight] of sediment samples from the Luppe and Laucha Rivers 2017. Particle size distribution was done with a laser diffraction system.

Sampling site Luppe Laucha aanalyzed

Sediment characteristicsa sand (2000 - 63 silt (63 - 2 clay (< 2 µm) [%] µm) [%] µm) [%] 8.64 79.31 12.05 25.46 61.55 12.99

using laser diffraction systems

Type of sediment clayey silt clayey silt

Journal Pre-proof Age determination: Scales were chosen for age determination of field sampled fish. Control fish purchased from a fish farm served as references as their ages were known. Cleaned scales were mounted on a glass microscope slides with the convex side up and examined using a light microscope (Nikon Eclipse TS100, Nikon GmbH, Düsseldorf, Germany). Striations representing annular growth zones were counted to estimate the approximate age of the fish. Annular rings were scored as true growth rings if they round the whole anterior edge of the scale. Furthermore, new scales showing a regenerated center without clear rings were not used for age estimation. The examinations were repeated in two readings and results were cross-checked between two readers, independently.

Histopathology of gonad: Histopathological evaluation of the H&E stained gonad sections was done in a two-step process using a light microscope (Axioskop, Zeiss Germany GmbH, Göttingen; Olympus optical co. GmbH, Hamburg, Germany). First, they were analyzed unblinded to enable comparison to control samples and thereby recognition of subtle differences in histology. Examinations were started with control samples to identify the approximate baseline histology of the species. Then secondly, findings were re-evaluated and confirmed in a blinded and coded manner after the first analysis. Gonadal staging for the assessment of endocrine disrupting effects on the reproductive status of the fish was done according to modified criteria from the “U.S. Biomonitoring of Environmental Status and Trends (BEST) Program” and the OECD Guidance Document No. 123 (OECD, 2010; Schmitt and Dethloff, 2000): Female • • • • • • • Male • • • • • •

Juvenile: oogonia exclusively, difficult to confirm the sex. Stage 0 (Undeveloped): entirely immature phases (oogonia to perinucleolar oocytes). Stage 1 (Early development): majority are pre-vitellogenic follicle in the perinucleolar stage. Stage 2 (Early development): majority are follicle in the cortical alveolar stage. Stage 3 (Mid development): majority of developing follicles are early- to mid-vitellogenic. Stage 4 (Late development): majority of developing follicles are late-vitellogenic. Stage 5 (Post-ovulatory): predominately spent and post-ovulatory follicles. Juvenile: spermatogonia exclusively, difficult to confirm the sex. Stage 0 (Undeveloped): entirely immature phases (spermatogonia to spermatids). Stage 1 (Early development): immature phases predominate, but spermatozoa can also be observed Stage 2 (Mid development): spermatocytes, spermatids, and spermatozoa are present in roughly equal proportions, the germinal epithelium is thinner. Stage 3 (Late development): all stages can be observed, mature sperm are predominating, the germinal epithelium is thinner. Stage 4 (Spent): loose connective tissue with some remnant sperm.

Histological pathologies were numerically graded to allow a semi-quantitative estimation of the degree of histological changes. These grading scores were assigned in comparison to the baseline histology of control fish in accordance with the OECD guidance document (OECD, 2010): Semi-quantitative grading system: • • • • •

Not remarkable: no findings Grade 1 (minimal): ≤ 20 % of the tissue were involved (spatial finding) or ≤ 2 findings occurred per section (discrete finding). Grade 2 (mild): 20-50 % were involved or 3-5 findings occurred per section. Grade 3 (moderate): 50- 80% were involved or 6 – 8 findings occurred per section. Grade 4 (severe): ≥ 80 % were involved or ≥ 9 findings occurred per section.

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Chemical analysis LC MS/MS: Target substances in sediment, water and plasma samples were chromatographically separated on a Nucleoshell RP18 column (150 x 3 mm; 2.7 µm particle size, Machery&Nagel, Germany) attached to an Agilent HPLC system (1200 Series). Pure water was used as mobile phase A and acetonitrile as mobile phase B for the separation with a gradient elution program with a flow rate of 0.4 mL/min and a total run time of 20 min. The column was heated to 25°C and the injection volume was 10 µL per sample. Identification and quantification of target substances E1, E2, EE2 and NP was done by mass spectrometry with a LTQ Orbitrap XL equipped with a heated electron spray ionization (HESI) ion source (heater temperature: 400°C, spray voltage: 4000 V, capillary temperature: 250°C, capillary voltage: 24 V, tube lens: 50 V) operated in positive ionization mode. The analyses were done in single reaction monitoring (SRM) mode with a normalized collision energy (NCE) for the fragmentation of 35. For the quantification an external calibration curve ranging from 1 to 1000 ng/mL were measured and at least five points were used for the calibration curve (R2>0.9). Normalization of the response to compensate for fluctuations in ionization and derivatization efficiency were conducted with deuterated standards. Limit of quantification (LOQ) and detection (LOD) were calculated according to Alvarez (2010) as 3- and 7-fold, respectively, the standard deviation (SD) of the mean blank concentration of the compound (Table A.3). Table A.3: Conditions of mass spectrometry for the analyzed steroidal estrogens. Quantifiers are underlined.

Compound E1 E2 de.-E2 EE2 de.-EE2 NP de.-NP 4n-NP

Retention time [min] 10.24-10.49 9.74-9.98 9.76-10.0 9.76-9.99 9.78-10.1 13.6-14.07 14.96-15.1 14.74-15.1

Parent mass, fragment1-3 [m/z] 504, 425.4/440.4/489.4 506, 171/427.4/442.4/491.3 510, 171/431.4/446.4/495.3 530, 171/451.4/466.5/515.2 534, 171/455.4/470.4/519.3 454, 171/375.3/390.4/439.3 458, 171/379.4/394.4/443.3 454, 171/375.3/390.4/439.3

Limit of quantification (LOQ) [ng/mL] 1.7 3.0

0.9 1.6

1.9

1.0

4.0

2.7

4.0

2.7

Limit of detection (LOD) [ng/mL]

E1= estrone; E2= 17β-estradiol; EE2= Ethynylestradiol; NP= nonylphenols; 4n-NP= 4n-nonylphenol; de.= deuterated; LOQ= meanblank + 7(SDblank); LOD= meanblank + 3(SDblank)

Sediment contact test: Sediment contact test with the blackworm Lumbriculus variegatus and Luppe sediment was conducted based on OECD Guideline 225 (Huff Hartz et al. 2018; OECD 225 2007). The test design was slightly modified since native sediment samples were included additional to spiked, artificial sediment. Briefly, treatments consisted of 1 L beakers containing 50 g dry weight (d.w.) of (1) artificial sediment, spiked with target compounds E1, E2, EE2 and NP in DMSO (< 0.1%) matching concentrations and general physicochemical properties of Luppe sediment as a positive control (Table A.2 and Table A.4); (2) the same artificial sediment (OECD sediment) spiked with dimethyl sulfoxide (DMSO; < 0.1%) as a solvent control; (3) the same artificial sediment without spike as a negative control; and (4) native Luppe sediment. Three replicates per treatment were included, each containing 800 ml of overlying

Journal Pre-proof reconstituted water and ten individual worms. The worms were synchronized 14 days prior to start of the test, which was conducted under temperature-controlled conditions (20°C) for 29 days. Afterwards, worms were removed from the sediment and were cleaned in a petri dish with tap water to remove sediment particles from the worm surface. The worms were then euthanized by placing them in liquid nitrogen. After lyophilization for 72 h, dry weights per replicate were determined. Extraction of whole organisms was done according to Watts et al. (2001). In brief, worms from all three test replicates were pooled into one sample. Worm tissue was then extracted with 5 ml of acetone in a sonication bath for 30 minutes. After centrifugation at 4°C for 30 minutes at 1000 g, the supernatant was transferred into 4 ml brown glass vials and dried under a gentle stream of nitrogen. The extracts were re-dissolved in 250 µl DMSO and tested in the YES assay according to ISO guideline (ISO/FDIS 19040-1:2018-03 2018). Table A.4: Nominal concentrations of 4n-nonylphenol (NP), estrone (E1), 17β-estradiol (E2) and ethynylestradiol (EE2) used for the positive control matching conditions of the Luppe sediment in the sediment contact test with Lumbriculus variegatus

Compound NP E1 E2 EE2

Nominal spike concentration [ng/g dry weight] 120,000 20 1.5 0.008

Journal Pre-proof Table A.5: Sex, general health parameters and age [a] determined by analysis of the scales and gonad, vitellogenin (vtg) in ng/mL per mgprotein and concentrations of nonylphenol (NP) in blood plasma (ng/mL) of tench (Tinca tinca), roach (Rutilus rutilus) and pike (Esox lucius) from the river Luppe, Laucha and of cultured fish. Gender was not determined for pike (mixed sex). Vtg was analyzed by ELISA technique and normalized by protein content. Concentrations of NP in the plasma of fish were measured by LC MS/MS. LSI; liversomatic index, GSI; gonadosomatic index, K; Fulton´s condition index.

Sampling site

Luppe

Gender (Nb.)

Mean age ± SD

juvenile (5)

2 ± 0.9

male (6)

3 ± 0.8

female (9)

3 ± 1.2

juvenile (1)

1 2±0

Mean blood plasma NP [ng/mL] ± SD

2.0 ± 0.9

27.0 ± 10.3

Mean weight [g] + SD

LSI

GSI

K

10.1 ± 1.4 14.4 ± 2.1 13.6 ± 2.5

Tench 21.1 ± 2.0 ± 8.7 0.2 61.8 ± 1.8 ± 28.7 0.2 53.9 ± 1.8 ± 26.1 0.5 Roach

0.2 ± 0.05 0.6 ± 0.4 4.2 ± 3.8

2.0 ± 0.3 1.8 ± 0.2 2.0 ± 0.3

0.9

1.6

205.8

1.9 ± 0.4 1.7 ± 0.6

162.5 ± 178.4 165.1 ± 198.0

0.9 ± 0.1

n.a.

10 10.4 ± 1.2 12.6 ± 3.13

15.6

0.9

20.1 ± 1.4 ± 2.8 ± 5.1 0.2 1.4 female 36.2 ± 2.5 ± 2.1 ± 2 ± 0.4 (7) 28.5 2.0 2.0 Pike mixed 2±0 19.7 ± 70.4 ± 1.8 ± 0.1 ± (10) 2.3 21.0 0.5 0.1 Tench juvenile 8.6 ± 12.1 ± 1.7 ± 0.1 ± 1 ± 0.4 (6) 1.1 4.3 0.4 0.08 14.6 ± 57.7 ± 1.4 ± 0.2 ± male (4) 3 ± 0.7 2.4 29.3 0.2 0.08 female 12.5 ± 38.6 ± 1.3 ± 1.2 ± 3 ± 1.3 (11) 3.1 25.4 0.2 0.4 Laucha Roach juvenile 11.2 ± 21.0 ± 1.3 ± 0.6 ± 1 (2) 0.4 1.6 0.3 0.2 male (0) female 11.4 ± 26.7 1.2 ± 1.1 ± 2±0 (8) 2.7 ±15.7 0.5 0.5 Tench male 11.1 ± 20.8 ± 1.5 ± 0.8 ± 1 ± 0.5 (20) 2.2 13.7 0.4 0.4 female 11.4 ± 20.5 ± 1.6 ± 2.3 ± 1 ± 0.5 (10) 2.2 10.5 0.4 0.9 Cultured fish Roach 10.3 ± 11.2 ± 1.9 ± 1.0 ± male (5) 2±0 0.2 1.0 0.8 0.5 female 11.1 ± 16.6 ± 1.2 ± 2.2 ± 2±0 (14) 1.3 6.5 0.4 0.9 - indicates insufficient plasma volume for analysis; n.a.: not analyzed male (5)

Mean vtg ± SD [ng/mL per mgprotein]

Mean length [cm] + SD

1.8 ± 0.3 1.7 ± 0.3 1.6 ± 0.3 1.5 ± 0.04 1.6 ± 0.2 1.4 ± 0.1 1.2 ± 0.1 1.0 ± 0.1 1.1 ± 0.1

454.2 ± 858 10991.1 ± 11490

4.6 ± 4.4 47.7 ± 93 5.2 ± 8.7

21.5 ± 2.8 22.7 ± 7.0 35.5 (n=1) 49.4 ± 42.0 21.9 ± 11.7 15.8 ± 2.1 (n=3) 21.1 ± 6.1(n=10)

5.8 ± 7.7

10.8 ± 3.7

297.4 ± 810.8

13.3 ± 7.2 (n=7)

1.5 ± 2.9

2.7 (n=2)

19.0 ± 38.0

8.1 ±6.7 (n=5)

1.5 ± 2.2

-

11.2 ± 18.4

-

Journal Pre-proof Table A.6: Histopathological staging of gonads from tench (Tinca tinca) and roach (Rutilus rutilus) from the Luppe and Laucha River and cultured fish. Sex was determined by histological analysis and samples were divided into male, female and indifferent (juvenile) fish. Number (Nb.) and mean age [a] of these groups are stated. Gonad stages were divided into juvenile and stage 0 – stage 4

Nb .

Mean age [a]

Juvenile

juvenile male female

6 5 9

2 ± 0.9 3 ± 0.8 3±1

6 (100 %) 0 0

0 2 (22 %)

juvenile male female

1 5 7

1 2±0 2 ± 0.4

0 0

juvenile male female

6 4 11

1 ± 0.4 3 ± 0.7 3 ± 1.3

juvenile male female

2 0 8

juvenile male female juvenile male female

Sampling site

Stage 0

Stage 1

Stage 2

Stage 3

Stage 4

Stage 5

2 (40 %) 2 (22 %) Roach

3 (60 %) 1 (11 %)

0 0

0 4 (44 %)

0 0

0 2 (29 %)

2 (40 %) 2 (29 %) Tench

0 0

3 (60 %) 3 (43 %)

0 0

0 0

6 (100 %) 0 0

3 (75 %) 2 (18 %)

1 (25 %) 9 (82 %) Roach

0 0

0 0

0 0

0 0

1 2±0

2 0

2 (25 %)

1 (12.5%)

0

0

1 (12.5%)

0 20 10

1 ± 0.5 1 ± 0.5

0 0

2 (10 %) 1 (10 %)

3 (15 %) 9 (90 %)

0 0

0 0

0 0

0 5 14

2±0 2±0

0 0

5 (100%) 5 (36 %)

4 (50 %) Tench 15 (75%) 0 Roach 0 0

0 1 (7 %)

0 7 (50 %)

0 1 (7 %)

0 0

Tench

Luppe

Laucha

Culture d fish

Table A.7: Qualitative histopathological findings in tench (Tinca tinca) and roach (Rutilus rutilus) gonads from the Luppe, Laucha River and cultured fish. Numbers of samples showing each histopathological finding and as well as the range of severity are given for all groups.

Species ovary

Tench testis

ovary

Roach

testis

Luppe

Laucha

Range of severity

11

Cultured fish 10

Nb. examinated

9

inflammation

5 (56 %)

9 (82 %)

6 (60 %)

minimal to moderate

atretic follicle

5 (56 %)

8 (73 %)

9 (90 %)

minimal to severe

undefined cell cluster

8 (89 %)

9 (82 %)

2 (20 %)

minimal to moderate

Nb. examinated

5

4

20

inflammation

0

1 (25 %)

0

minimal

intersex

1 (20 %)

0

3 (15 %)

minimal to moderate

Nb. examinated

7

8

14

inflammation

5 (71 %)

2(25 %)

2 (14 %)

minimal to mild

atretic follicle

0

1 (12 %)

2 (14 %)

minimal to moderate

Nb. examinated

5

0

5

inflammation

1 (20 %)

0

0

minimal

Journal Pre-proof

Figure A.2: Testis with testis-ova of male tench (Tinca tinca) caught at the Luppe River. perinucleolar (PN); spermatogonia (SG); spermatocytes (SC); Spermatids (ST); spermatozoa (SZ). Sections were stained with H&E.

Journal Pre-proof

Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests:

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