Bioavailability and impact of sediment-bound endocrine disrupting chemicals on fish in context of flood events
Von der Fakultät für Mathematik, Informatik und Naturwissenschaften der RWTH Aachen University zur Erlangung des akademischen Grades einer Doktorin der Naturwissenschaften genehmigte Dissertation
vorgelegt von
Master of Science Anne-Katrin Müller aus Uedem
Berichter: Univ.-Prof. Dr. rer. nat. Henner Hollert Univ.-Prof. Dr. rer. nat. Andreas Schäffer Univ.-Prof. Dr. rer. nat. Helmut Segner
Tag der mündlichen Prüfung: 27.02.2020
Diese Dissertation ist auf den Internetseiten der Universitätsbibliothek verfügbar.
Eidesstattliche Erklärung
Anne-Katrin Müller erklärt hiermit, dass diese Dissertation und die darin dargelegten Inhalte die eigenen sind und selbstständig, als Ergebnis der eigenen originären Forschung, generiert wurden.
Hiermit erkläre ich an Eides statt
1. Diese Arbeit wurde vollständig oder größtenteils in der Phase als Doktorand dieser Fakultät und
Universität angefertigt;
2. Sofern irgendein Bestandteil dieser Dissertation zuvor für einen akademischen Abschluss oder eine andere Qualifikation an dieser oder einer anderen Institution verwendet wurde, wurde dies klar angezeigt;
3. Wenn immer andere eigene- oder Veröffentlichungen Dritter herangezogen wurden, wurden diese klar benannt;
4. Wenn aus anderen eigenen- oder Veröffentlichungen Dritter zitiert wurde, wurde stets die Quelle hierfür angegeben. Diese Dissertation ist vollständig meine eigene Arbeit, mit der Ausnahme solcher
Zitate;
5. Alle wesentlichen Quellen von Unterstützung wurden benannt;
6. Wenn immer ein Teil dieser Dissertation auf der Zusammenarbeit mit anderen basiert, wurde von mir klar gekennzeichnet, was von anderen und was von mir selbst erarbeitet wurde;
7. Ein Teil oder Teile dieser Arbeit wurden zuvor veröffentlicht und zwar in:
Müller A-K, Markert N, Leser K, Kämpfer D, Crawford S E, Schäffer A, Segner H, Hollert H (2019) Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment. Environmental Pollution. Doi: 10.1016/j.envpol.2019.113636 Müller A-K, Leser K, Kämpfer D, Riegraf C, Crawford SE, Smith K, Vermeirssen E, Buchinger S, Hollert H (2019) Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling. Water Research. doi: 10.1016/j.watres.2019.06.020
Datum Unterschrift
Abstract
Abstract
To date, numerous studies worldwide have demonstrated that sediments function as a sink for a great variety of environmental pollutants, among them are substances interfering with the endocrine system, so called endocrine disrupting chemicals (EDCs). Estrogenic activity evaluated in sediment samples across Europe via in vitro bioassays ranged from 0.02 up to 55 ng 17β-estradiol (E2) equivalents/g sediment. This is of particular concern since it is well documented that waterborne exposure to even low ng/L concentrations of EDCs can impair the reproduction of freshwater fish species. Feminization of male fish is one of the most notable adverse impacts of exposure to EDCs and the production of the female egg yolk protein vitellogenin (vtg) has been observed to coincide with impairment of gonadal development evident as intersex and, ultimately, reproductive failure. In contrast, little is known about the bioavailability and effects of sediment-associated EDCs on fish. Particularly when sediments are perturbed, e.g., during flood events, sediment-bound substances may become bioavailable. During the past decades, several extreme flood events have occurred in central Europe, including Germany. The likelihood and intensity of flood events have been predicted to increase as a result of global climate change. As consequence, the European Parliament established the Directive 2007/60/EC on the assessment and management of flood risk. In order to minimize adverse consequences of flood events to humans and the environment part of such risk assessment is the evaluation of potential sources of environmental pollution as result of flooding.
The main objectives of the present thesis, as part of the interdisciplinary Project House Water – a project supported by the German Excellence initiative - , were to (i) investigate the bioavailability of sediment-bound EDCs under flood-like conditions when the sediment is subject to suspension; (ii) evaluate the uptake of sediment-bound EDCs during such a simulated flood event into fish and whether this leads to endocrine responses in the fish and (iii) assess the impact of sediment-bound EDCs to freshwater fish species inhabiting a “hot-spot” of EDC contamination in sediment under field conditions. In doing so, this thesis aimed to provide implications for risk evaluation of sediment associated contaminants with special emphasis on flood events. In order to assess the bioavailability of sediment- bound EDCs, the Luppe River previously described as a “hotspot” for EDC accumulation in sediment was chosen as a study site. The concentration of target EDCs and estrogenic activity of sediments from the Luppe River were investigated using chemical analysis (LC-MS/MS) in addition to effect-based methods, such as a novel screening tool (planar Yeast Estrogen Screen; p-YES) that utilizes high performance thin-layer chromatography plates in combination with an in vitro bioassay (YES). Estrone (50%, E1) and nonylphenol (35%, NP) accounted for the majority of estrogenic activity reported for sediment with up to 20 ± 2.4 ng 17β-estradiol equivalents (EEQ)/g dry weight in the Luppe sediments. E2 accounted for approximately 14% of the estrogenic effect, whereas the estrogenic effect attributed to 17α-ethynylestradiol (EE2), when present, was negligible (approx. 1%) from sediment across all Luppe sampling sites. Two types of passive samplers (polar organic chemical integrative sampler (POCIS) and Chemcatcher) were used to investigate the bioavailability of EDCs from suspended sediment under laboratory conditions. NP, E1, E2 and EE2 were remobilized from Luppe sediment when subjected to turbulent conditions, such as in a flood event, and were readily bioavailable at ecotoxicologically relevant concentrations (NP 18 µg/L, E1 14 ng/L, E2 0.2 ng/L, EE2 0.5 ng/L). Both types of passive samplers were applicable in a sediment-water suspension system, with the Chemcatcher displaying higher sampling rates compared to the POCIS.
A laboratory exposure study with juvenile rainbow trout (Oncorhynchus mykiss) was conducted to evaluate uptake and ecotoxicological impact of remobilized sediment-bound EDCs from the Luppe River. Therefore, rainbow trout were exposed over 21 days to constantly suspended sediment in the
I Abstract following treatments: i) a contaminated sediment from the Luppe River ii) a control sediment (exhibiting only background contamination), iii) a serial dilution of Luppe sediment with the sediment control (1:8; 1:4; 1:2), and iv) a water-only control. Measured estrogenic activity using in vitro bioassays as well as target analysis of NP and E1 via LC-MS/MS in sediment, water, fish plasma, as well as bile samples, demonstrated that sediment-bound EDCs became bioavailable during the simulated flood event. EDCs were dissolved in the water phase, as indicated by passive samplers, and were readily taken up by the exposed trout. Interestingly, similar patterns of EDCs were observed in the water and fish blood and bile, suggesting that EDCs partitioned from sediment into the water and subsequently absorbed by the fish, indicating that freely dissolved aqueous concentrations of EDCs might be a major route for uptake of EDCs in fish. An estrogenic response of fish to Luppe sediment was indicated by increased abundance of transcripts of typical estrogen responsive genes, i.e. vitelline envelope protein α, in the liver and vitellogenin induction in the skin mucus. Hepatic gene expression profiles by RNA-sequencing were altered in Luppe exposed fish compared to controls, whereas the repression of a great number of genes involved in cell cycle in combination with induction of apoptotic markers suggest a broader response. However, similar downregulation of cell cycle genes was observed with the sediment control. Together with histological alterations, i.e. local areas of cell lysis, infiltration of immune cells and degenerative nuclear alterations, observed in the liver of fish throughout all treatments, indicates that exposure to suspended particles might elicit stress at the cellular level.
Moreover, tench (Tinca tinca) and roach (Rutilus rutilus) as a benthic and pelagic living fish species, respectively, were sampled at the Luppe River. A field reference site, the Laucha River, in addition to fish from a commercial fish farm as a 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. Estrogenic activity of Luppe exposed worm tissue (14 ng EEQ/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, vtg as biomarker for exposure to EDCs was induced in male tench and roach from the Luppe River compared to both the Laucha and commercially cultured fish by a factor of 264 and 90, respectively. However, no histological alterations in testis of these Luppe exposed fish were observed. Our findings suggest that sediments substantially contribute to the overall EDC exposure of both benthic and pelagic fish in the field but that the exposure did not translated to adverse effects on the gonad level and, thus, might not be of relevance for the reproductive success of these populations in the wild.
The present thesis demonstrated that sediments not only function as a sink for EDCs but can turn into a significant source of pollution when sediments are resuspended. The results demonstrated that sediment-bound EDCs were readily bioavailable for fish under conditions similar to those of a flood event. Partitioning of EDCs into the water phase might be a major route for uptake of remobilized sediment-bound EDCs into the fish. Passive sampling was a useful tool to assess the bioavailability of sediment-bound EDCs and could be a good indicator of sediment toxicity in a regulatory context. Overall, the work described in this thesis greatly contribute to the assessment of sediment-bound EDCs in the context of flood risk.
II Zusammenfassung
Zusammenfassung
In den vergangenen Jahren haben eine Vielzahl von Studien gezeigt, dass sich diverse Umweltschadstoffe im Flusssediment anreichern. Darunter befinden sich auch solche Schadstoffe, die das endokrine System stören – sogenannte endokrine Disruptoren (EDs). Europaweit wurde eine Östrogene Aktivität in Flusssedimenten mittels in vitro Biotests von 0.02 bis zu 55 ng 17β-Estradiol (E2) Äquivalenten (EEQs)/g Sediment nachgewiesen. Dies ist von besonderer Bedeutung, da bereits bekannt ist, dass EDs schon in geringster Konzentration im Wasser von einigen Nanogramm pro Liter zur Verminderung der Reproduktionsleistung von Fischen führen können. Vor allem die Feminisierung männlicher Fische wurde als ein adverser Effekt der EDs Exposition mit hoher Umweltrelevanz beobachtet. Kennzeichen für die Feminisierung der männlichen Fische sind die Produktion des weiblichen Eidotterproteins Vitellogenin und die simultane Entwicklung sowohl männlicher als auch weiblicher Geschlechtsmerkmale innerhalb eins Individuums, bekannt als Intersex. Dies hat eine verminderte Reproduktionsleistung zur Folge, was bis hin zum Aussterben von einer ganzen Fischpopulation führte. Hingegen ist bislang wenig erforscht, ob die im Sediment abgelagerten EDs bioverfügbar für Fische sind und das endokrine System negativ beeinträchtigen. Die im Sediment zwischengelagerten Schadstoffe könnten durch Resuspension (und Desorption), hervorgerufen durch z.B. menschliches Einwirken wie Ausbaggern, aber auch natürliche Ereignisse wie ein Hochwasser, demobilisiert werden und so in hohen und potenziell toxischen Konzentrationen für Fische in der Wasserphase verfügbar werden. In den letzten Jahren wurden einige extreme Hochwasser in Mitteleuropa inklusive Deutschland beobachtet. Das Auftreten solcher extremen Ereignisse wird zukünftig im Kontext der globalen Klimaerwärmung mit großer Wahrscheinlichkeit zunehmen. Daher wurde vom Europäischen Parlament die Direktive 2007/60/EC zur Risikobewertung von Hochwasserereignissen erlassen. Bestandteil einer solchen Bewertung ist unteranderem die Evaluierung potenzieller Schadstoffquellen als Folge des Hochwassers.
Die Hauptzielsetzung dieser Arbeit, als Teil des interdisziplinären Projekthaus Wasser – gefördert durch die Deutsche Exzellenz Initiative - , war es daher: (i) die Bioverfügbarkeit der sedimentgebundenen endokrin wirksamen Schadstoffe in Hinblick auf ihre Remobilisation während eines Hochwasserereignisses abzuschätzen; (ii) die damit verbundene Gefahr einer Störung des endokrinen Systems in Fischen während eines solchen Hochwasserereignisses zu evaluieren und (iii) die Auswirkungen der sedimentgebunden EDs auf heimische Fischarten in einem Fluss mit hoher sedimentbürtiger Belastung abzuschätzen. Diese Studie liefert gemäß der Forderung der europäischen Kommission Aufschluss über ein bestehendes Umweltproblem, trägt dazu bei, dieses Risiko einzuschätzen und ermöglicht somit eine regulatorische Handlungsfähigkeit. Um die Bioverfügbarkeit von sedimentgebundenen EDs abzuschätzen, wurde ein Fluss in der Nähe von Leipzig, der zuvor als „Hot-Spot“ für endokrine Wirksamkeit im Sediment identifiziert wurde, die Luppe, als Untersuchungsort gewählt. Die endokrine Aktivität des Sediments in der Luppe sowie die Konzentration einiger EDs wurde mittels chemischer Analytik (LC-MS/MS) untersucht, zusätzlich zu einem neuartigen in vitro Biotest, der in Anschluss an eine chromatische Auftrennung direkt auf der Dünnschichtplatte durchgeführt wird (plate Yeast Estrogen Screen (p-YES)). So konnte gezeigt werden, dass der Großteil der endokrinen Aktivität aus dem Sediment von 20 ± 2.4 ngEEQ/g auf die hohen Konzentrationen von Estron (E1: 50%) und Nonylphenol (NP:35%) zurückzuführen ist. Des Weiteren trug E2 mit ca. 14% maßgeblich zu der endokrinen Aktivität bei, während die Konzentrationen und Wirksamkeit von 17α-Ethinylestradiol (EE2) mit 1% vernachlässigbar waren. Zwei passive Probenahmesysteme (polar organic chemical integrative sampler (POCIS) und Chemcatcher) wurden verwendet, um die Bioverfügbarkeit dieser Stoffe aus dem Sediment der Luppe unter Suspensionsbedingung ähnlich dem eines Hochwassers zu untersuchen. Anhand der passiven Probenahmesysteme konnte gezeigt werde, dass NP, E1, E2, aber auch EE2, unter
III Zusammenfassung
Suspensionsbedingungen in umweltrelevanten Konzentrationen aus dem Sediment in die Wasserphase frei wurden (NP 18 µg/L, E1 14 ng/L, E2 0.2 ng/L, EE2 0.5 ng/L). Beide Probenahmesysteme erwiesen sich dabei als geeignet für die Anwendung in einem Sediment-Wasser-Suspensionssystems.
Darüber hinaus wurde in einer Laborexpositionsstudie mit juvenilen Regenbogenforellen (Oncorhynchus mykiss) die Umweltrelevanz dieser remobilisierten endokrinen Schadstoffe aus dem Sediment abgeschätzt. Hierbei wurde untersucht, ob die Aufnahme der remobilisierten endokrinen Schadstoffe aus dem Sediment der Luppe während eines simulierten Hochwassers zu Veränderungen des endokrinen Systems dieser Fische führte. Dazu wurden die Forellen über 21 Tage gegenüber Sediment unter konstanter Suspension in nachfolgenden Behandlungen exponiert: (i) einem belasteten Sediment aus der Luppe, (ii) einem geringfügig belasteten Kontrollsediment, (iii) einer seriellen Verdünnung aus belastetem Luppe- und Kontrollsediment (1:8; 1:4; 1:2) und (iv) einer Wasserkontrolle. Die mittels in vitro Biotest nachgewiesene Estrogene Aktivität und die detektierten Konzentrationen von E1 und NP mittels LC-MS/MS in den Sedimenten, dem Wasser sowie in Blut- und Gallenproben der exponierten Fische aus den unterschiedlichen Behandlungen bestätigte die Bioverfügbarkeit der sedimentgebundenen EDs unter Hochwasserbedingungen. Interessanterweise wiesen die gemessen Konzentrationen von NP und E1 im Wasser und Fisch, Blut sowie Galle, übereinstimmende Muster auf, was den Schluss nahelegt, dass die Aufnahme der aus dem Sediment remobilisierten Stoffe über die Wasserphase stattfand. Die Induktion von estrogenabhängigen Gene in der Leber der männlichen Fische, die gegenüber dem Sediment aus der Luppe exponiert wurden, zusammen mit erhöhtem Vitellogenin Gehalt gemessen in der Schleimhaut dieser Fische deuten auf eine endokrine Antwort in den männlichen Fischen hin. Darüber hinaus war die Abundanz von Transkripten kodierend für eine Vielzahl von Genen in der Leber, die an der Zellzykluskontrolle beteiligt sind, in den Luppe exponierten Fischen signifikant vermindert. Selbiger Effekt wurde auch in Fischen der Sedimentkontrolle beobachtet. Diese Veränderungen auf Genebene stimmt mit den histologischen Veränderungen in der Leber auf Organebene überein. In den Lebern aller exponierter Fische, ausschließlich der Wasserkontrolle, wurden lokal begrenzte Zelllysen, Veränderungen des Nukleus sowie Infiltration von Immunzellen beobachtet. Dies lässt den Rückschluss zu, dass die Exposition gegenüber den gelösten Sedimentpartikeln einen zellulären Stress ausgelöst hat.
In einer Feldstudie wurde zudem untersucht, ob die an der Luppe heimischen Fische durch den Kontakt zu dem hoch belasteten Sediment endokrine Veränderungen wie Intersex aufweisen. Hierzu wurden Schleien und Rotaugen an der Luppe gefangen sowie an einem Referenzfluss, der Laucha, zusätzlich zu Fischen aus einer kommerziellen Aquakultur, die als Referenz gelten. Benthisch lebende Würmer (Lumbriculus variegatus) wurden dem Luppesediment gegenüber exponiert, um zu evaluieren, ob sedimentgebundene EDs über die Nahrung an die Fische weitergereicht werden können. Die vergleichsweise 153-fach höhere Estrogene Aktivität gemessen mittels in vitro Biotests in den Sedimentproben der Luppe bestätigte die Laucha als Referenzfluss. Des Weiteren wurde eine Estrogene Aktivität von 14 ng EEQ/mg in den Wurmextrakten gemessen, was bestätigt, dass die Nahrung einen Expositionspfad für Fische darstellen kann. Während sich die gemessenen Konzentrationen an NP im Blut der Fische von der Luppe nicht von denen der Laucha unterschieden, wurde eine Induktion des Biomarkers Vitellogenin in den an der Luppe gefangenen männlichen Fischen im Vergleich zur Laucha und den Referenzfischen aus der Aquakultur beobachtet. Untersuchungen der männlichen Gonade erbrachten jedoch keinen Hinweis auf den Einfluss von endokrin wirksamen Schadstoffen. Die Ergebnisse lassen den Schluss zu, dass sedimentgebundene Schadstoffe an der Luppe zu der Gesamtbelastung an EDs in den Fischen beitragen, diese scheint jedoch nicht zu einer Beeinträchtigung der Gonaden Entwicklung zu führen und somit voraussichtlich den Fortbestand der Populationen nicht negativ zu beeinflussen.
IV Zusammenfassung
Diese Arbeit konnte herausstellen, dass Sedimente sowohl eine Senke als auch Quelle für EDs darstellen können. Die sedimentgebundenen EDs wurden während der Suspension des Sedimentes, ähnlich wie in einem Hochwasserereignis, bioverfügbar für die exponierten Fische, wobei eine Aufnahme über die rückgelösten EDs im Wasser den Hauptexpositionsweg darstellen könnte. Die passiven Probenahmetechniken (passive sampling) stellten sich als äußerst hilfreiche Instrumente heraus, um die Bioverfügbarkeit von sedimentgebundenen EDs unter Suspensionsbedingungen zu bewerten. Diese könnte auch in einem regulatorischen Kontext zur Anwendung gebracht werden. Zusammenfassend werden die in dieser Arbeit gewonnenen Ergebnisse zu einer verbesserten Risikobewertung von sedimentgebundenen endokrinen Schadstoffen beitragen.
V
VI Table of contents
Table of contents
Inhalt Abstract ...... I Zusammenfassung ...... III Table of contents ...... VII List of figures ...... XI List of tables ...... XIII 1 Introduction ...... 1 1.1 General introduction ...... 3 1.2 Endocrine disruption in wildlife ...... 3 1.3 The endocrine system of fish ...... 5 1.4 Mechanisms of endocrine disruption - estrogenic activity in fish ...... 10 1.5 Sources, distribution and fate of EDCs ...... 12 1.6 EDCs within European legal frameworks ...... 15 1.7 Risk assessment of EDCs – present challenges ...... 17 1.8 Project House Water – Flood-hydrotox...... 19 1.9 Objectives and hypotheses ...... 20 1.9.1 Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling ...... 20 1.9.2 Impacts of endocrine disruptors from sediment on rainbow trout (Oncorhynchus mykiss) during a simulated flood event ...... 21 1.9.3 Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment ...... 22 2 Material & Methods ...... 23 2.1 Material and Chemicals ...... 25 2.2 Water and sediment sampling ...... 25 2.3 Sediment extraction ...... 26 2.4 Water extraction ...... 26 2.5 Yeast estrogen screen and planar yeast estrogen screen ...... 27 2.6 Derivatisation with dansyl chloride ...... 28 2.7 Quantification by LC-MS/MS ...... 28 2.8 Extraction of EDCs from plasma ...... 29 2.9 Vitellogenin analysis ...... 30 3 Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling ...... 31 3.1 Abstract ...... 34 3.2 Introduction ...... 35
VII Table of contents
3.3 Material and methods ...... 37 3.3.1 Materials and chemicals ...... 37 3.3.2 Water and sediment sampling ...... 37 3.3.3 Water and sediment extraction ...... 38 3.3.4 Yeast estrogen screen and planar yeast estrogen screen with sediment extracts ...... 38 3.3.5 POCIS and Chemcatcher laboratory calibration ...... 38 3.3.6 Suspended sediment experiment ...... 42 3.3.7 Quantification by LC-MS/MS ...... 43 3.4 Results ...... 43 3.4.1 Accumulation of EDCs and related estrogenic activity along the Luppe River ...... 43 3.4.2 Calibration of POCIS and Chemcatcher ...... 45 3.4.3 Bioavailability of ECs from sediment under turbulent conditions ...... 47 3.5 Discussion ...... 49 3.6 Conclusion ...... 53 3.7 Acknowledgments ...... 53 4 Impacts of endocrine disruptors from sediment on rainbow trout (Oncorhynchus mykiss) during a simulated flood event ...... 55 4.1 Abstract ...... 58 4.2 Introduction ...... 59 4.3 Material and Methods ...... 61 4.3.1 Sediment sampling and investigation ...... 61 4.3.2 Experimental fish ...... 62 4.3.3 Experimental design ...... 62 4.3.4 Sediment extraction ...... 63 4.3.5 Passive sampling ...... 63 4.3.6 Fish sampling ...... 64 4.3.7 Extraction of EDCs from plasma ...... 64 4.3.8 Extraction of EDCs from bile ...... 64 4.3.9 Modeling approach ...... 66 4.3.10 Vitellogenin analysis ...... 66 4.3.11 Liver histology ...... 67 4.3.12 RNA extraction, cDNA synthesis and Illumina sequencing ...... 67 4.3.13 Statistical analysis and functional annotation of transcriptome results ...... 68 4.4 Results ...... 69 4.4.1 Exposure conditions and bioavailability of EDCs from sediment ...... 69 4.4.1.1 EDCs in sediment ...... 70 4.4.1.2 EDCs in aqueous phase ...... 70
VIII Table of contents
4.4.1.3 EDC uptake in rainbow trout ...... 71 4.4.2 Impact of remobilized EDCs form the sediment on rainbow trout ...... 75 4.4.2.1 Hepatic differentially expressed gene profiles ...... 75 4.4.2.2 Vtg induction ...... 82 4.4.2.3 Histology ...... 85 4.4.2.4 General fish health indices ...... 86 4.5 Discussion ...... 86 4.5.1 Bioavailability of sediment-bound EDCs ...... 86 4.5.1.1 Estrogenic activity in sediments ...... 86 4.5.1.2 Partitioning of EDCs between sediment-water interphase during flood-like conditions ...... 87 4.5.1.3 Uptake routes of sediment-bound EDCs into rainbow trout ...... 89 4.5.2 Impact of sediment-bound EDCs on rainbow trout during a simulated flood event .... 90 4.5.2.1 EDC responsive hepatic gene expression in rainbow trout ...... 90 4.5.2.2 Altered hepatic gene expression profiles in response to sediment exposure ...... 91 4.5.2.3 Effects on biomarker induction, liver histology and overall fitness ...... 94 4.6 Conclusion ...... 96 4.7 Acknowledgments ...... 96 5 Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment ...... 97 5.1 Abstract ...... 100 5.2 Introduction ...... 101 5.3 Material and Methods ...... 103 5.3.1 Study sites ...... 103 5.3.2 Sediment, water and fish sampling ...... 103 5.3.3 EDC analysis in blood ...... 104 5.3.4 Sediment contact test with the blackworm and Luppe sediment ...... 104 5.3.5 Vitellogenin analysis ...... 105 5.3.6 Age determination ...... 105 5.3.7 Histopathology of gonads ...... 106 5.3.8 Statistical analysis ...... 107 5.4 Results and Discussion ...... 108 5.4.1 EDC exposure of fish under field conditions ...... 108 5.4.1.1 EDCs in water ...... 108 5.4.1.2 EDCs in sediment ...... 108 5.4.1.3 Estrogenic activity in the macroinvertebrate Lumbriculus variegatus:...... 109 5.4.1.4 Uptake of EDCs in fish ...... 110
IX Table of contents
5.4.2 Endocrine disruptive effects in tench and roach ...... 113 5.4.2.1 Biomarker response ...... 113 5.4.2.2 Histopathology of fish gonads ...... 116 5.5 Conclusion ...... 120 5.6 Acknowledgments ...... 121 6 Discussion ...... 123 6.1 General discussion ...... 125 6.1.1 The use of sediment quality standards in sediment risk assessment ...... 125 6.1.1.1 Bioavailability of sediment-bound contaminants ...... 126 6.1.1.2 Routes of exposure to remobilized sediment-bound contaminants ...... 128 6.1.1.3 Mixture toxicity ...... 129 6.1.1.4 Sediment quality assessment as weight of evidence approach ...... 131 6.1.2 Implications for sediment risk assessment under the WFD ...... 132 6.1.3 Bridging data from laboratory to field – effects of sediment-bound EDCs ...... 135 6.1.4 Implications for flood management at the Luppe River ...... 139 6.2 Conclusion and outlook ...... 141 7 References ...... 143 I. Appendix: Supplementary information to chapter 3 to 5 ...... 164 II. Appendix: Supplementary information to chapter 4 ...... 166 III. Appendix: Supplementary information to chapter 5 ...... 168 Contributions to the published articles and chapters ...... 170 Acknowledgement ...... 172 Curriculum vitae ...... 174
X List of figures
List of figures
Figure 1: Schematic overview of hypothalamus-pituitary-gonadal-axis in teleost fish controls the development of reproductive tissues...... 6 Figure 2: Control mechanisms of the hypothalamus-pituitary-gonadal-axis responsible for reproductive cycle in teleost fish...... 7 Figure 3: Biosynthesis of important steroid hormones in teleost fish here exemplary in female fish. .. 8 Figure 4: Estrogenic activity in sediment samples across Europe investigated using the Yeast Estrogen Screen (YES) in vitro bioassay...... 14 Figure 5: Sampling sites along the Luppe River in eastern Germany near Leipzig. Map adapted from ESRI. L1-L5: sediment sampling sites ...... 26 Figure 6: Polar organic chemical integrative sampler (POCIS) and Chemcatcher configuration and experimental setup used in the calibration study...... 39
Figure 7: Stabilization of the average water concentration (Ctwa) and sampling rates (Rs) after 3-6 calculation cycles...... 42 Figure 8: Sediment related estrogenic activities along the Luppe River...... 44 Figure 9: Calibration of Chemcatcher and POCIS for estrone (E1), 17β-estradiol (E2), ethynylestradiol (EE2) and 4n-nonylphenol (4n-NP)...... 46 Figure 10: Depletion test of estrogenic compounds from suspended sediment by passive sampling. . 49 Figure 11: Suspended particles during 21 days of exposure in g/L measured in the different treatments: ...... 70 Figure 12: Exemplary results of water-only control (WC), sediment control (SC), 1:2 dilution of Luppe with the sediment control (1:2 LU:SC) and Luppe River sediment (LU ) analyzed by high performance thin-layer chromatography combined with the planar-Yeast Estrogen Screen (p-YES)...... 72 Figure 13: Linear regression of biliary nonylphenol (NP) concentrations [µg/mL] measured by LC- MS/MS and estrogenic activity by in vitro bioassay expressed as 17β-estradiol equivalents (EEQ) in bile of rainbow trout exposed for 21 days...... 73 Figure 14: Relationship between the measured and predicted concentrations of nonylphenol (NP) in bile of rainbow trout ...... 74 Figure 15: Venn-diagram (Oliveros 2007-2015) of differentially expressed genes (p<0.05) in male rainbow trout (Oncorhynchus mykiss) ...... 76 Figure 16: Pathway enrichment network of significant gene ontology (GO) terms of differentially expressed hepatic genes in male rainbow trout ...... 82
Figure 17: Vitellogenin (vtg) concentration [ng/mL per mgprotein] in mucus samples of mixed-sex rainbow trout ...... 83 Figure 18: Linear regression of mean concentrations of vitellogenin (vtg) in the mucus of 21-day exposed rainbow trout...... 84 Figure 19: Histological alterations in liver of rainbow trout (Oncorhynchus mykiss) ...... 85
XI List of figures
Figure 20: Mucus vitellogenin (vtg) concentration [ng/mL per mgprotein] and concentrations of nonylphenol in blood plasma of roach (Rutilus rutilus) (A, B) and tench (Tinca tinca) (C, D) from the Luppe, Laucha and of cultured fish...... 116 Figure 21: Testis with testis-ova of male tench (Tinca tinca) caught at the Luppe River...... 117 Figure 22: Gonadal development stages of female (A-C) and male (D-F) tench (Tinca tinca)...... 118 Figure 23: Sediment quality triad as described in Chapman et al. 1990, later complemented with additional lines of evidence in 2006 (Chapman and Hollert 2006) and modified according to results on bioavailability of sediment-bound contaminants observed in the present study...... 131 Figure 24: Effect based methods for water monitoring as presented in Brack et al. (2019a) and modified in Altenburger et al. (2019)...... 134 Figure 25: Bridging lab to field: Assessing the bioavailability of endocrine disrupting chemicals during flood events by means of passive sampling, in vitro bioassays such as the planar-Yeast Estrogen Screen (YES) and in vivo exposure studies...... 138
XII List of tables
List of tables
Table 1: Characteristics of selected endocrine disrupting chemicals from the literature...... 13 Table 2: Conditions of mass spectrometry for the analyzed steroidal estrogens...... 29 Table 3: Estrogenic activity (17β-estradiol equivalents; EEQs) and the corresponding concentrations of estrogenic compounds in Luppe sediment investigated by the means of Yeast Estrogen Screen (YES), high performance thin-layer chromatography YES (p-YES), and LC-MS/MS...... 45
Table 4: Sampling rates (Rs; L/day) for Chemcatcher and POCIS as well as average (± SD) water (Cw
[µg/L]) concentrations of estrogenic compounds before sampler deployment (t0) and after 24h (t24) for
Chemcatcher calibration (t24)...... 47 Table 5: Bioavailability of estrogenic compounds from suspended sediment as detected in the passive samplers by means of LC-MS/MS in three suspension tests...... 48 Table 6: EDCs were measured from extracts of the passive samplers (Chemcatcher) for the aqueous phase as well as from extracts of the sediment of all treatments using LC-MS/MS...... 71 Table 7: Concentrations of nonylphenol (NP), 17β-estradiol (E2), 17α-ethynylestradiol (EE2) and estrone (E1) in blood plasma and bile of rainbow trout using LC-MS/MS, p-YES and estrogenic activity measured in the YES assay...... 74 Table 8: Numbers of genes differentially expressed (DEG; p< 0.05) in male rainbow trout (Oncorhynchus mykiss) after exposure to suspended sediment from the Luppe River (LU), a sediment control (SC), a 1:2 dilution of Luppe sediment with the sediment control (1:2 LU:SC) or a water-only control (WC)...... 76
Table 9: Significantly differentially expressed (p<0.05 and log2 fold change of 2 in at least one comparison) genes in male rainbow trout (n=3) as a response to sediment exposures...... 78
Table 10: Significantly differentially expressed (p<0.05 and log2 fold change of 2 in at least one comparison) genes in 21-d exposed male rainbow trout (n=3)...... 80 Table 11: Summary of general health parameters including Fulton’s condition index (K), gonadosomatic index (GSI) and liver somatic index (LSI), concentrations of vitellogenin (vtg) measured in mucus by ELISA of rainbow trout...... 84 Table 12: 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 ...... 105 Table 13: Estrogenic activity in sediment and water samples from the Luppe and Laucha Rivers and literature reports worldwide determined by the Yeast Estrogen Screen (YES) in 17β-estradiol equivalents (EEQs) and concentration of target EDCs nonylphenol (NP), estrone (E1) and 17β-estradiol (E2) by LC-MS/MS (or other analytical methods)...... 109 Table 14: 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
XIII List of tables
(ng/mL) of tench (Tinca tinca), roach (Rutilus rutilus) and pike (Esox lucius) from the river Luppe, Laucha and of cultured fish...... 112
Table A. 1: Habitat characteristics and water parameters of sampling sites along the Luppe, Laucha and Rhine River...... 164 Table A. 2: Physicochemical characteristics of sediment samples from the Luppe, Laucha and Rhine Rivers...... 164 Table A. 3: Concentrations of 17 PCDD/Fs and 12 dl-PCBs in sediments form the Rhine River sampled at Koblenz (Ehrenbreitstein) and the Luppe River...... 166 Table A. 4: Concentrations of indicatior/i-PCBs and metals in sediments from the Rhine and Luppe River...... 167 Table A. 5: Exposure conditions over 21 days exposure...... 167 Table A. 6: Histopathological staging of gonads from tench (Tinca tinca) and roach (Rutilus rutilus) from the Luppe and Laucha River and cultured fish...... 168 Table A. 7: Qualitative histopathological findings in tench (Tinca tinca) and roach (Rutilus rutilus) gonads from the Luppe, Laucha River and cultured fish...... 169
XIV
Chapter 1
1 Introduction
1
2 Chapter 1 - Introduction
1.1 General introduction
Freshwater ecosystems such as rivers, lakes and wetlands are subjected to a great variety of anthropogenic stressors including man-made chemicals (Dodds et al. 2013; Guillén et al. 2012). For instance, more than 22,000 unique chemicals are currently registered in Europe and from these, 2,782 are produced in large quantities above 1,000 tons per year (European Chemicals Agency 2018). Based on their widespread use in industrial processes, agriculture practices, commerce and consumer products, chemicals can reach the aquatic environment (Guillén et al. 2012). Hence, depending on their properties, modes and extent of use, these chemicals might impact aquatic ecosystems with unpredictable effects on the aquatic wildlife especially freshwater fish species in the long term (Posthuma et al. 2016).
Although overall water quality of major rivers might have improved since the 1960s due to the establishment of stricter regulatory legislation in many countries, i.e., the European regulation on
Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) and the Water
Framework Directive (WFD) in EU countries, the Clean Water Act (CWA) in the US, more than 60 % of the global aquatic ecosystems are degraded or unsustainably used (Millennium Ecosystem
Assessment 2005; Mostert 2009). In particular, freshwater ecosystems have a high proportion of threatened and endangered species with declining populations of freshwater fish species (Burkhardt-
Holm et al. 2002; Freyhof and Wright 2011; Keiter et al. 2006; Moyle and Williams 1990).
Over the last two decades, substances interfering with the endocrine system, namely endocrine disrupting chemicals (EDCs), have remained a focal point in terms of scientific and public awareness
(EU COM 734 final 2018). Numerous field and laboratory studies gathered evidence that exposure to
EDCs can negatively impact the reproductive fitness of fish and, ultimately, impair the reproductive success of entire populations (Kidd et al. 2007).
1.2 Endocrine disruption in wildlife
In the early 1960s an environmental movement inspired by Rachel Carson’s book “Silent Spring” first brought attention to adverse effects on reproduction of wildlife caused by exposure to environmental pollutants (Carson 1962). Carson described in her book that the extensively used insecticide dichloro-diphenly-trichloro-ethane (DDT) accumulated in aquatic and terrestrial organisms,
3 Chapter 1 - Introduction biomagnified along the food web, and caused deleterious effects on the reproduction of songbirds
(Carson 1962). Later on, a number of environmental scientists worldwide discovered that exposure to chemicals interfering with the endocrine system was associated with impaired reproduction i.e., in molluscs, reptiles, fish and mammals (Jobling et al. 1998; Oehlmann et al. 1996; Reijnders 1986). For example, the reproductive failure of common seals (Phoca vitulina) inhabiting the Wadden Sea at the coastal area of the Netherland was related to the uptake of polychlorinated biphenyls (PCB) accumulated in fish (Chiu et al. 2000; Reijnders 1986).
Around the same time anglers observed through the UK high incidences (5%) of intersex, the presence of both male and female characteristics within the same individual, in wild population of roach
(Rutilus rutilus) (Jobling et al. 1998). This was followed by both caging experiments of cultured fish in the effluent of wastewater treatment plants (WWTP) and extensive monitoring studies of wild fish populations. Observed reproductive disturbances in caged and wild fish were consistent with the exposure to hormonally active substances, EDCs, associated with discharges from the WWTP (Harries et al. 1997; Jobling et al. 1996; Jobling et al. 1998). Despite the complex chemical mixture of WWTP effluents, only a small number of chemicals were identified as EDCs – namely, estrone (E1), 17β- estradiol (E2) and 17α-ethynylestradiol (EE2) in addition to alkylphenols e.g. nonylphenols (NP)
(Desbrow et al. 1998). Effect directed analysis, a combination of in vitro bioassays, the Yeast Estrogen
Screen (YES), with gas chromatography-mass spectrometry (GC-MS), was used to demonstrate that the steroid estrogens (E1, E2, EE2) as well as the alkylphenols were the major source of estrogenic activity in the WWTP effluents of rivers (Desbrow et al. 1998; Harries et al. 1997; Liney et al. 2005; Rodgers-
Gray et al. 2001). Endocrine disruptive effects observed in wild and caged fish include the production of the female-specific, estrogen-dependent egg yolk protein precursor vitellogenin (vtg) in male and juvenile fish, inhibited ovarian or testicular development, abnormal blood steroid concentrations, intersex and/or masculinization or feminization of the internal or external genitalia, impaired reproductive success i.e., reduced fertility, delayed male and/or female maturation, increased ovarian atresia, reduced spawning success, reduced hatching success and larval survival (Harries et al. 1997;
Harris et al. 2011; Jobling et al. 1996; Jobling et al. 2002a; Tyler et al. 1996). Further, laboratory exposure of i.e., rainbow trout (Oncorhynchus mykiss), fathead minnow (Pimephales promelas),
4 Chapter 1 - Introduction
Japanese medaka (Oryzias latipes) or zebra fish (Danio rerio) to single substances and/or mixtures of steroid estrogens and alkylphenols at environmental relevant concentrations confirmed the findings of the field monitoring studies supporting the argument that exposure to EDCs impact reproductive development with deleterious consequences for the reproductive success (Jensen et al. 2001; Jobling et al. 1996; Länge et al. 2001; Legler et al. 2000; Nash et al. 2004; Thorpe et al. 2000a). Although the prevalence of intersex and induction of vtg in wild roach caught downstream of WWTP in the UK has been demonstrated to coincide with reduced gamete quality and fertilization success (Jobling et al.
2002a; Jobling et al. 2002b), it has been critically discussed whether these changes impact the reproductive success with the potential for population level outcomes (Tyler and Jobling 2008). An exposure study with EE2 conducted in the Experimental Lake Area in Canada provided evidence that exposure to low levels of a potent estrogen has the potential to significantly interfere with the sexual development of fish with an adverse outcome on the population level. Exposure to 5-6 ng/L EE2 induced feminization of male fathead minnow through the production of vtg, impairment of gonadal development evident as intersex and, ultimately, lead to reproductive failure of the entire population of fathead minnow (Kidd et al. 2007). While vtg was highly increased in male fathead minnow as well as pearl dace (Margariscus margarita) and lake trout (Salvelinus namaycush) after EE2 exposure, the effects were less marked in male white sucker (Catostomus commersonii). Moreover, adverse effects on the population level were restricted to fathead minnows, highlighting the importance to account for interspecies differences in endocrine physiology (Kidd et al. 2007; MacLatchy et al. 2009).
1.3 The endocrine system of fish
The endocrine system closely interacts with the nervous systems and regulates several physiological processes including growth, development, reproduction, homeostasis and behaviour of organisms
(Lintelmann et al. 2003). The primary function of the endocrine system is the communication between cells and organs based on specific chemical messengers, the hormones, secreted by endocrine glands directly in the blood (Di Giulio and Hinton 2008). Although hormones are distributed through the bloodstream, their hormonal activity is highly specific due the presence of specified receptor molecules binding the hormones in the respective target tissue (Pandian 2012). Upon receptor binding, hormones activate intracellular second messenger pathways. In addition, the hormone-receptor complex itself
5 Chapter 1 - Introduction might acts as a transcription factor resulting in transcriptional modifications of responsive genes and, thereby, altering cell function (Anstead et al. 1997).
Figure 1: Schematic overview of hypothalamus-pituitary-gonadal-axis in teleost fish controls the development of reproductive tissues. This figure was modified from Sumpter and Jobling (1995). GtH I/II: gonadotropin hormone; FH: follicle-stimulating hormone; luteinizing hormone (LH).
The endocrine system in teleost fish is mainly controlled by hormones secreted by the hypothalamic- pituitary- gonadal (HPG) axis (Figure 1) (Sumpter and Jobling 1995). Basic characteristics of the HPG axis are similar in all fish and comparable to other vertebrates (Janz 2000). Exogenous stimuli, for instance photoperiod, temperature, nutrition and stressors, as well as endogenous factors are integrated in the hypothalamus and translated into a hormonal response (Breton 1980). Various neurotransmitters and neuropeptides are secreted in the hypothalamus influencing the activity of gonadotropin-releasing hormone (GnRH)-producing neurons (Di Giulio and Hinton 2008). In teleost fish, unlike other vertebrates, the hypothalamic GnRH neurons directly innervate the gonadotropin-producing cells, so called gonadotropes (Swanson et al. 2003). Gonadotropes are situated in the adenohypophysis part of the pituitary (reviewed by Di Giulio and Hinton 2008; Parhar et al. 2002). GnRH stimulates via binding to specific receptors on the plasma membrane of the gonadotropes the synthesis and secretion of gonadotropins (Alok et al. 2000) (Figure 2).
6 Chapter 1 - Introduction
Figure 2: Control mechanisms of the hypothalamus-pituitary-gonadal-axis responsible for reproductive cycle in teleost fish. This figure was adapted from Di Giulio and Hinton (2008).CNS: central nervous system; GnRH: Gonadotropin releasing hormone; FSH: follicle stimulating hormone; LH: luteinizing hormone; E2: 17β-estradiol; T: testosterone; 11-KT: 11-ketotetosterone: 17,20 β -P: 17,20 β-dihydroxy-4-pregnen-3-one; 20 β-S: 20β,21- trihydroxy-4-pregnen-3-one.
The reproductive cycle of teleost fish is under the dual-control of two gonadotropins, namely follicle- stimulating hormone (FSH; previously named GtH I) and luteinizing hormone (LH; previously named
GtH II) (Damasceno-Oliveira et al. 2013; Nagahama and Yamashita 2008) (Figure 2). FSH and LH are glycoprotein hormones elucidating individual secretory patterns during the reproductive cycle
(Damasceno-Oliveira et al. 2013; Yaron et al. 2002). Their functions comprise of the synthesis of the egg yolk proteins (vitellogenesis) and early gonadal development (FSH) as well as maturation and release of mature gametes in male and female (LH) (Di Giulio and Hinton 2008). Both bind to specific receptors on the surface of somatic cells in the gonads, whereat LH-specific receptors were identified in granulosa cells in preovulatory ovarian follicles and in Leydig cells in testis (Miwa et al. 1994). Another receptor type, not distinguishing between LH and FSH, is located on both thecal and granulosa cells during vitellogenesis but only present in the thecal cells in the preovulatory follicle. Similarly, this receptor was found in the Sertoli cells in males during all stages of spermatogenesis (Miwa et al. 1994).
The binding of both gonadotropins to their target receptor activates a G-protein, adenylyl cyclase, and calcium-dependent second-messenger signalling pathways initiating the production and secretion of steroid hormones (Nagahama 2000) (Figure 2). Important androgens in teleost fish are testosterone (T) and 11-ketosterone (11-KT), whereas the main estrogen is E2 (reviewed in Pandian 2015).
7 Chapter 1 - Introduction
Figure 3: Biosynthesis of important steroid hormones in teleost fish here exemplary in female fish. Synthetic enzymes include P450 (P540 side chain clevage), 3β-HSD (3β- hydroxysteroid dehydrogenase), 20β-HSD (20β- dehydroxysteroid dehydrogenase), 17α-hydroxylase, 17β-HSD (17β- dehydroxysteroid dehydrogenase) and P450 arom. (aromatase). Testosterone is synthesised in the thecal cells of the ovary and is converted by the P450 arom. to 17β-estradiol in the granulosa cells. This figure was illustrated and modified from Janz (2000).
Steroid hormones consist of four fused hydrocarbon rings with oxygen and carbon substitutions at different positions (Mommsen and Moon 2005). Biosynthesis of the steroid hormones takes place within the somatic cells of the gonads, i.e. thecal and granulosa cells in the ovary and primarily in the Leydig or interstitial cells in the testis (Di Giulio and Hinton 2008). All steroid hormones are synthesised from a common precursor, cholesterol, and involves several biosynthetic steps mainly regulated by the cytochrome P450 enzyme family (Kime 1993) (Figure 3). Briefly, cholesterol (C-21 steroid) is converted to pregnenolone by a P450 enzyme family member. In turn, pregnenolone itself is transformed by steroidogenic enzymes including 3β- hydroxysteroid dehydrogenase (3β-HSD) and other members of the P450 enzyme family into progesterone and subsequently into 17α-hydroxyprogesterone (Idler and
Macnab 1967). The 17α-hydroxylase (P450c17) converts 17α-hydroxyprogesterone to an androgen (C-
19 steroid), androstenedione, which is then converted to T (Zhou et al. 2007a). The P450 aromatase
8 Chapter 1 - Introduction enzyme (P450 arom) subsequently converts T to E2 (C-18 steroids) (reviewed in Guiguen et al. 2010;
Idler 2012; Reinecke et al. 2006) (Figure 3).
The pattern of steroidogenesis changes during the reproductive cycle controlled by pituitary FSH and
LH secretion. Steroid hormones such as T and E2 are transported in the blood bound to specific carrier proteins to target cells in the gonad, liver and brain (Fostier et al. 1983) (Figure 2). A main function of
E2 in female fish is the induction and regulation of hepatic vtg production as well as synthesis of vitelline envelope proteins (Nagahama 2000). Vtg is required during oocyte growth and is later converted into the major egg yolk proteins lipovitellin and phosvitin serving as a nutrition reserve for the embryo
(Hiramatsu et al. 2005). While vtg concentrations during maturation of oocytes naturally increase 102-
106-fold in female fish, vtg concentrations measured in male fish are very low around 1 ng/mL in blood plasma (Jobling and Tyler 2003; Sumpter and Jobling 1995).
Mechanism of hepatic vtg induction by E2 involves binding to the nuclear estrogen receptor (ER) and activation of the ER signaling pathway (Thomas 2000). Binding of E2 to the ER induces a conformational change, dissociation of heat shock proteins (hsp90, hsp70), followed by dimerization and phosphorylation resulting in its activation. The active receptor-hormone-complex is transferred into the nucleus, where it recruits regulatory co-factor proteins to interact with estrogen-responsive elements
(ERE) on the DNA. Binding of the active receptor-hormone-complex to the ERE in the promotor regions of target genes lead to transcriptional activation of the gene downstream of the ERE, i.e. vtg (Le Dréan et al. 1995). In addition to the nuclear ER, another E2 signaling system is via a membrane receptor
(mER) and involves Ca2+/cAMP signalling, kinase cascades and G-protein coupled receptor activation
(Thomas 2000). Two subtypes of the ER have been identified in teleost fish, ER α and β. Moreover, four isoforms of the ER have been found in salmonid species (ERα1, ERα2, ERβ1 and ERβ2) varying in their expression levels among tissues and cell types (Leaños-Castañeda and van der Kraak 2007; Nagler et al. 2007). EDCs that activate the ER pathway have the potential to alter the regulatory control through the HPG axis leading to adverse consequences for the reproductive fitness (Thomas 1999). For instance, the feminization of male fish through the production of hepatic vtg is mediated via ER activation and signaling (Jobling et al. 1996; Segner et al. 2013).
9 Chapter 1 - Introduction
1.4 Mechanisms of endocrine disruption - estrogenic activity in fish
Although EDCs have been shown to target multiple sites by a variety of mechanisms, the HPG axis appeared to be the most sensitive endocrine system affected (Ankley and Giesy 1998; Arukwe 2008).
EDCs have been found to interfere with the hypothalamic neurotransmitter function controlling GnRH secretion or directly modulate GnRH secretion i.e. bisphenol A (BPA) (Qin et al. 2013; Vetillard and
Bailhache 2006) (Figure 2). Altered GnRH secretion in turn would secondarily affect secretion of gonadotropins from the pituitary. In addition, exposure studies have demonstrated that EDCs have the ability to directly affect GnHR signaling pathways in the gonadotrope and gonadotropin secretion
(Harding et al. 2016). Another site of EDC action is the gonad, where EDCs disrupt endocrine functions by altering the gonadotropin second-messenger pathways and synthesis of steroid hormones (Hecker et al. 2007; Hinfray et al. 2006; Thomas and Rahman 2012). In particular, the P450 aromatase activity, the key enzyme converting androgens to estrogens, was affected by EDCs exposure e.g. NP (Hinfray et al.
2006). Alterations of steroid hormone synthesis have direct consequences for the steroid hormone level in the blood and subsequently, action at their target tissues including gonads and liver (Di Giulio and
Hinton 2008). Moreover, increased metabolic clearance of the steroid hormones have been reported due to the induction of hepatic P450 enzymes by PCBs (Sivarajah et al. 1978), whereas NP was found to inhibit the glucuronidation of E2 (Thibaut and Porte 2004).
Another mechanism of endocrine disruption is the interaction of EDCs with both the nuclear and membrane steroid receptor at the target tissue of the steroid hormones. EDCs can act as agonists or antagonists of steroid hormone action (Thorpe et al. 2003). The activation of the nuclear ER i.e. in the liver through binding of EDCs, has been widely studied in fish. Over the past decades numerous laboratory and field studies demonstrated that a great variety of EDCs in the environment elicit estrogenic activity in fish via binding to the nuclear ERs (Ackermann et al. 2002; Thorpe et al. 2001;
Tyler et al. 1996; Wang et al. 2018). As consequence, transcription of estrogen responsive genes is altered, with induction of genes encoding for vtg, vitelline envelope proteins and ERs resulting in disruption of gonadal development as described in the previous section (Arukwe et al. 1997; Schwaiger et al. 2002; Thorpe et al. 2000b). Steroidal estrogens such as E1 and estriol (E3), but also EDCs including o,p-derivates of DDT, NP and other alkylphenols, BPA, hydroxylated metabolites of several PCBs,
10 Chapter 1 - Introduction exhibit relative binding affinities for ERs in teleost fish. Their affinity to bind to the ERs was demonstrated to be 10-3 to 10-5 times lower compared to E2 (Anstead et al. 1997; Bolger et al. 1998).
Hence, their relative potency to induce vtg production is lower, i.e. E1 has been reported to be approximately two up to 10 times less potent than E2 (Thorpe et al. 2003). In contrast, EE2 was consistently the most potent steroid exceeding the relative estrogenic potency of E2 to induce hepatic vtg production by a factor of 10 up to 27 in i.e. rainbow trout (Thorpe et al. 2003). Thus, many of the
EDCs have the potential to cause an estrogenic response at very low concentrations (parts per billion to parts per trillion) (reviewed in Campbell et al. 2006; Kidd et al. 2007).
The induction of vtg in male fish is a widely used biomarker for exposure to EDCs exhibiting estrogenic activity (Hecker et al. 2002; Sumpter and Jobling 1995; Tyler et al. 1996). However, several factors can influence vtg biomarker induction in fish. Those factors include differences in the sensitivity among species as well as developmental stage and might affect both the sensitivity and the quality of the response (Segner 2005). In particular, the magnitude of vtg induction in response to EDC exposure varies between species, wherein, cyprinid species (e.g. zebrafish and roach) have been reported to exhibit lower vtg levels than salmonids (e.g. rainbow trout) when exposed to EDCs (Tyler et al. 2005).
Moreover, early life stages have been shown to be more susceptible towards EDC exposure with adverse effects manifesting later in development during sexual maturation (Liney et al. 2005). Furthermore, steroidal estrogens but also alkylphenols in mixtures interact in an additive manner to e.g. trigger the induction of vtg in i.e. juvenile rainbow trout (Thorpe et al. 2001). Concentrations of estrogens, e.g. E2 and EE2 and NP, below their individual lowest observed effect concentrations (LOEC) were combined more potent and lead to the induction of vtg in male fish (Silva et al. 2002; Thorpe et al. 2006). Predicted- no-effect concentrations (PNECs) for the steroid hormones including E1 (6 ng/L), E2 (2 ng/L), E3 (60 ng/L) and EE2 (0.1 ng/L) have been derived based on species sensitivity distribution or in vivo vtg induction observed in chronic exposure studies for use in aquatic risk assessment of EDCs (Table 1)
(Caldwell et al. 2012).
11 Chapter 1 - Introduction
1.5 Sources, distribution and fate of EDCs
Steroidal estrogens as well as other EDCs including NP and BPA have been commonly detected in global surface waters at environmental relevant concentrations. For instance, literature reviews reported that EE2 was measured in surface waters in Germany, the Netherlands, Switzerland, the UK, Italy and
US at concentrations of 0.5 up to 5 ng/L (reviewed in Campbell et al. 2006; Hillenbrand et al. 2016;
Schneider 2006). Several sources of EDCs have been identified by means of chemical as well as bioanalytical methods, where in WWTP effluents have been implicated as the major source for steroidal estrogens such as E1, E2 and EE2 (Desbrow et al. 1998; Harries et al. 1997; Liney et al. 2005; Rodgers-
Gray et al. 2001). The actual sources of steroidal estrogens and other EDCs, such as NP, are upstream discharges to the WWTP including households and industrial processes (Campbell et al. 2006). For example, NP is widely used as a surfactant in industrial processes, household cleaners and in agrochemical formulations and discharges from industry have been recognized as a main source to the aquatic environment (Lee et al. 2013; Wang et al. 2016). Moreover, discharges of EE2 from its use in birth control pill practises were calculated for the world’s human population of about 7 billion and approximated to be 700 kg/year (Adeel et al. 2017). Kostich et al. (2013) reported that, on average, a pregnant woman excretes between 260-790, and 280-600 µg/day of E1 and E2, respectively. Removal of steroidal estrogens in current WWTP processes have been found to be incomplete, and, thus, WWTP effluents are important sources of steroidal estrogen into the aquatic environment (Fahlenkamp et al.
2004; Hillenbrand et al. 2016). Another source of EDCs, particularly steroidal estrogens, are effluents from livestock feedlots, where a variety of steroidal estrogens are used to increase growth rates and feed efficiency (Campbell et al. 2006). Furthermore, in aquaculture practises steroidal estrogens are commonly and purposefully used for sex reversal and the production of all-female stocks (Hulak et al.
2010). Additional non-point sources of EDC pollution are run-off from agriculture, where plant protection products as well as fertilizers might elicit estrogenic activity (Ali et al. 2018; Sellin et al.
2009; Sellin et al. 2010; Zhang et al. 2015). Other non-point sources of EDC pollution and new chemicals with yet unknown endocrine disrupting properties might exist.
12 Chapter 1 - Introduction
Table 1: Characteristics of selected endocrine disrupting chemicals from the literature.
Relative potency Solubilitya EQSe PNECf,g to E2 in EDC Molecular structure LogK b,c,d [mg/L] oc [µg/L] [ng/L] the YES assayh,i [%] natural steroidal estrogens:
17β-estradiol 13 3.1-4.0 - 2 100 (E2)
estrone (E1) 6-13 2.45-3.34 - 6 40-60
synthetic steroidal estrogen:
17α- 4.8 2.91-3.04 - 0.1 90-178 ethynylestradiol
alkylphenols:
p-nonylphenol 0.01- 4.9-7 3.6-5.4 0.3 33 (NP) 0.05
EQS: environmental quality standards for surface water; PNEC: predicted no effect concentration based on species sensitivity distribution and vitellogenin induction in male fish; YES: yeast estrogen screen; Sources: a reviewed in Campbell et al. 2006, b reviewed in Ma and Yates 2018; c Ferguson et al. 2001; d Li et al. 2011; e EU 2013/39; f Caldwell et al. 2012; g Giesy et al. 2000; h van den Belt et al. 2004; i Thorpe et al. 2006 Once steroidal estrogens as well as other EDCs enter the aquatic environment, they are subjected to biodegradation processes including bio- and photochemical degradation. Estimated half-life for E2 under aerobic conditions in river water ranges from 0.2 up to 9 days (Jürgens et al. 2002). The breakdown product of E2 is E1 for which similar half-life ranges were observed (Jürgens et al. 2002). Photochemical degradation of NP has been observed with 10-15 h half-lives (Ahel et al. 1994). Moreover, due to the hydrophobic (e.g. high log kow) properties of many EDCs, they tend to bind to suspended particular matter (SPM) with high affinity to organic rich matter (dissolved organic matter (DOM)) later depositing in the sediment (Table 1) (Gong et al. 2016; Ma and Yates 2018; Yamamoto et al. 2003; Zhou et al.
2007b).
13 Chapter 1 - Introduction
Figure 4: Estrogenic activity in sediment samples across Europe investigated using the Yeast Estrogen Screen (YES) in vitro bioassay. Colors express measured estrogenic activity in sediments as 17β-estradiol equivalents (EEQs) in ngEEQ/g sediment dry weight. This map is based on the highest EEQs found in literature. References: Buchinger et al. 2013b; Céspedes et al. 2004; Fenet et al. 2003; Hilscherova et al. 2002; Houtman et al. 2006; Legler et al. 2000; Macikova et al. 2014; Peck et al. 2004; Viganò et al. 2008. Base map of Europa was free for usage from d-maps.com.
Studies worldwide investigating the fate and distribution of EDCs in the environment have demonstrated using chemical and bioanalytical methods that EDCs sorb to SPM and, thus, accumulate in river sediments (Hilscherova et al. 2007; Holthaus et al. 2002; Urbatzka et al. 2007; Viganò et al.
2008; Wang et al. 2012; Wölz et al. 2011; Zhao et al. 2011). For instance, estrogenic activity measured in sediment samples via the YES assay ranged from 0.02 up to 55 ngEEQ/g (Figure 4) (Buchinger et al.
2013b; Céspedes et al. 2004; Peck et al. 2004; Schulze-Sylvester et al. 2016; Thomas et al. 2004; Viganò et al. 2008). In combination with chemical analysis, E1, E2, EE2 and alkylphenols such as 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. EE2 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. 2010b; Viganò et al.
2008). Moreover, high levels of NP and E1 were detected by LC and GC MS/MS, and found to
14 Chapter 1 - Introduction contribute most to the observed estrogenic activity from the German river sediments (Buchinger et al.
2013b).
While it has been recognized that EDCs have a high tendency to accumulate in sediment, little is known about their bioavailability and potential for remobilization, transport and redistribution due to extreme weather events such as floods, bioturbation or dredging. However, some studies indicate that estrogenic compounds have the potential to be remobilized from sediments during floods as shown by estrogen-related endpoints (Wölz et al. 2011). The frequency and intensity of flood events is predicted to increase in the context of global climate change (Alfieri et al. 2018; EU 2007/60; Hirabayashi et al.
2013). On account of this, the European parliament initiated the directive 2007/60/EC regarding the assessment and management of flood risks, whereby they established a framework aiming to assess and reduce the adverse consequences of floods for human health and the environment (EU 2007/60). One major aspect of such risk assessment includes the evaluation of sediment contamination and the potential of contaminant remobilization from sediments, whereby the sink might turn into a source of pollution
(Hollert et al. 2007a).
1.6 EDCs within European legal frameworks
EDCs have been recognized to pose a risk to humans and the environment and, thus have been included in several pieces of European Union legislation including the European legislation on chemicals (REACH), biocides, plant protection products (prospective risk assessment) as well as the
WFD (retrospective risk assessment) (EU COM 734 final 2018). EDCs have been defined according to the World Health Organisation as “an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub)populations” (WHO 2002).
To date, over one hundred substances worldwide have been identified as endocrine disruptors including natural (e.g. E2 and E1) as well as synthetic hormones (e.g. EE2), many pesticides, organic pollutants such as PCBs, BPA, phthalates and alkylphenols (e.g. NP) (EU 2013/39). As a consequence, many EDCs have been banned from use in plant protection products and biocides (EU 2017/2100,
2018/605). Furthermore, the Plant Protection Products Regulation (PPPR) and the Biocidal Product
15 Chapter 1 - Introduction
Regulation (BPR) set out specific criteria for the identification of EDCs (EU 2009/1107, 2012/528).
However, the criteria for the evaluation of EDCs in context of environmental risk assessment only refer to adverse effects of EDCs relevant for population outcomes of non-target organisms. Moreover, the
PPPR regulation states that substances identified as EDCs might be used when the risk of exposure for non-target organisms is negligible, without further defining criteria or thresholds of negligible exposure levels (EU 2018/605).
Under the REACH regulation substances identified as EDCs are of similar regulatory concern as substances of very high concern listed in Article 57f which are subject to authorisation (EU 2006/1907).
However, according to the REACH regulation long-term toxicity test evaluating reproductive endpoints in fish should be recommended by the registrant (Annex IX of EU 2006/1907). An evaluation of
REACH dossiers for substances produced in quantities above 1000 tons per year regarding their formal compliance with the requirements set out in the regulation reported that long-term toxicity tests for fish were most often not conducted based on the abovementioned article (Oertel et al. 2018). Nevertheless,
NP and octylphenol (OP) have been placed in the list of substances requiring a specific authorisation.
In addition, 13 substances have been identified as EDCs and are included in the Candidate list of substances for possible inclusion in the authorisation list in the future (EU 2018/605).
In the past years, the scientific community promoted the evaluation of EDCs in context of chemical risk assessment via the concept of Adverse Outcome Pathways (AOPs) (Kramer et al. 2011; Villeneuve et al. 2014). The Principle of the AOP concept is to link a so called molecular initiating event, i.e. ER receptor binding, through several intermediary key events, such as hepatic vtg induction, formation of intersex, impaired gonadal development, to relevant adverse outcomes on population level (Ankley et al. 2010; Villeneuve et al. 2014). AOPs for EDCs, in particular for ER agonists, have been successfully developed and might become a powerful tool for the regulatory decision-making and substance testing in the future (https://aopwiki.org).
In contrast to the prospective risk assessment, the WFD monitors and manage the risk of chemicals that have been released into the aquatic environment. The main goal of the WFD is to achieve a “good ecological and chemical status” in European freshwater systems. The evaluation of the chemical status
16 Chapter 1 - Introduction of lakes and rivers system is based on monitoring of 45 priority substances for which thresholds
“environmental quality standards” (EQS) have been defined. Several EDCs, among them NP, were included in Annex VIII within the group of priority substances of particular concern for which the EQS apply (Table 1) (EU 2013/39). In addition, E1, E2 and EE2 were included in the European watchlist to be monitored, and for which the development of consensus-based risk evaluation strategies for surface waters have been recommended (EU COM 734 final 2018). However, besides the substances listed in the watchlist and Annex VIII of the WFD, no direct screening or monitoring of other EDCs is required under the WFD (EU 2000/60). Although, the priority substances should be monitored within water, sediment and biota, no quality standards exist for sediment (EU 2013/39). Moreover, it is left to the individual member states to decide whether monitoring of sediment contamination might be necessary
(Brils 2008). Based on the uniform high contamination of sediment with EDCs but also other organic pollutants, sediments might be a major source of pollution, especially under conditions where bioavailability of sediment-bound substances is enhanced such as in a flood event. Consequently, sediment quality guidelines should be established considering bioavailability of sediment-bound substances such as EDCs in context of extreme weather events.
1.7 Risk assessment of EDCs – present challenges
Although screening and identification of EDCs in the context of environmental monitoring under the
WFD is restricted to a limited number of substances, the substances listed in Annex III as well as the watchlist should be chemically monitored in the water at the level of their environmental quality standard (EQS) (EU 2000/60). Given the fact that steroid estrogens such as E1, E2 and EE2 elicit estrogenic activity in fish mediated through the ER at very low exposure concentration in the ng/L range, monitoring at biological relevant concentrations can be challenging (Könemann et al. 2018). In particular, acceptable method detection limits at EQS level are 400 pg/L for E1 and E2 and 35 ng/L for
EE2 (EU 2015/495). Analytical methods for estrogen detection include high-performance liquid chromatography (HPLC), GC-MS/MS and liquid chromatography- mass spectrometry (LC-MS/MS)
(Lin et al. 2007). While those analytical methods provide a precise measurement, they are often not sensitive enough to detect EDCs especially estrogens at their environmental relevant concentrations and acceptable detection limits (Lin et al. 2007). Derivatisation of steroid estrogens, i.e. with dansyl chloride,
17 Chapter 1 - Introduction has been shown to improve signal intensity in mass spectrometry and, thus, might, be used to meet the low detection limits (Lin et al. 2007).
A useful tool for monitoring of EDCs at low environmental concentrations is passive sampling
(Alvarez 2010). For instance, the polar organic chemical integrative sampler (POCIS; sorbent OASIS
HLB) as well as the Chemcatcher (SDB-RPS) have been established for the monitoring of estrogens in the aquatic environment (Alvarez et al. 2007; Skodova et al. 2016; Vermeirssen et al. 2005). Monitoring of EDCs via passive sampling might reflect the actual environmental exposure concentrations more realistically compared to traditional grab sampling, since the former technique provides time-integrated information over extended time periods as well as measurements of freely dissolved concentrations
(Alvarez 2010). Concentrations of freely dissolved organic compounds, such as EDCs, are generally assumed to be more readily available for uptake in an organism and thus bioavailable in water (Rand
1995).
Another concern regarding the chemical monitoring of EDCs relates to the fact that aquatic organisms such as fish are exposed to complex mixtures of EDCs. As mentioned earlier concentrations of E2 and EE2 and NP below their individual LOEC combined elicit higher estrogenic potencies in fish resulting in an induction of vtg in male fish (Silva et al. 2002; Thorpe et al. 2006). Consequently, the current monitoring of the limited number of individual EDCs according to WFD is likely to underestimate the risk posed by EDCs for the aquatic environment (Brack et al. 2019a).
With the upcoming revision of the WFD, several recommendations for monitoring, assessing and managing risks of contamination in the aquatic environment have been addressed (Brack et al. 2017;
Brack et al. 2018; Escher et al. 2018). Key aspects were to include passive sampling as an advanced sampling method as well as effect-based methods (EBMs), such as bioassays and biomarkers, as integrative techniques that bridge the gap between chemical contamination and ecological effects (Brack et al. 2019a). EBMs are bioanalytical methods which use the response of organisms (in vivo) or cellular bioassays (in vitro) to measure and quantify effects of chemical mixtures on toxicological endpoint such as the estrogenic activity of complex environmental samples. Based on the research project
SOLUTIONS and the European monitoring network NORMAN a basic bioassay battery has been
18 Chapter 1 - Introduction suggested including in vitro tests for estrogenic activity, i.e. YES assay (Brack et al. 2019a; Escher et al. 2018).
Furthermore, recommendations for improvements of the WFD have included the development of strategies to investigate the bioavailability of sediment-bound contaminants with emphasis on mechanisms that enhance the bioavailable fraction such as resuspension of sediment during flood events
(Brack et al. 2017). Brack et al. (2017) proposed that EQS for sediment should be derived based on the freely dissolved fraction, that can be determined using passive sampling.
1.8 Project House Water – Flood-hydrotox
An interdisciplinary research project named FLOODSEARCH was conducted at the RWTH
Aachen University to assess the risks associated with the remobilization of sediment-bound contaminants, in particular dioxins, PCBs and polyaromatic hydrocarbons (PAHs), to fish under flood- like conditions (Brinkmann et al. 2010; Cofalla et al. 2012; Schüttrumpf et al. 2011; Wölz et al. 2009).
Adverse effects from remobilized dioxins, PCBs, PAHs were identified in in vitro bioassays (Wölz et al. 2010) as well as in laboratory exposure studies with rainbow trout (Brinkmann et al. 2010; Brinkmann et al. 2015). Brinkmann et al. (2015) demonstrated that sediment-bound dioxin-like contaminants and
PAHs were readily bioavailable from sediments during suspension as in a flood event. Uptake of those re-mobilized sediment-bound contaminants by fish lead to changes in hepatic mRNA expression, hepatic P450 enzyme activity, increased micronuclei frequency in blood as well as histopathological alterations in the liver.
The follow-up project, Project House Water, aimed to further investigate the environmental and socioeconomic impacts and risks of sediment-associated contaminants distributed during and after flood events. Several specialized disciplines including ecotoxicologists, hydrologists, water engineers, chemists, geoscientists, water economists, and sociologists worked together to develop new interdisciplinary assessment and management strategies for the distribution of contaminated sediments during flood events (Crawford et al. 2017). Individual “proof-of-concept” studies were conducted within each discipline, among one of them was Flood-hydrotox presented in the current thesis.
19 Chapter 1 - Introduction
The main goals of Flood-hydrotox were to: (i) provide information of sources and sinks of EDCs associated with sediment in context of flood risk assessment; (ii) provide implications and strategies for the current risk assessment of sediment-bound EDCs; (iii) evaluate environmental relevance of sediment-bound EDCs and risk to freshwater fish species with special emphasis on flood events. In particular, this project aimed to investigate whether EDCs accumulated in sediment might turn from sink to source of pollution during extreme weather events such as flood events and, thus, pose a risk to freshwater fish species.
1.9 Objectives and hypotheses
The objectives of the present study aligned with research goals set out for Flood-hydrotox and were investigated in two laboratory experiments and one field study. The main objectives were to a) investigate the accumulation of estrogenic compounds in river sediments; b) the potential of these sediment-bound estrogenic compounds to be remobilized under flood-like conditions in which sediments is resuspended; c) whether sediment-bound EDCs may become bioavailable to fish, i.e., rainbow trout (Oncorhynchus mykiss) and if uptake consequently leads to adverse effects such as alterations of the endocrine system in fish; and d) whether contamination of EDCs in sediment at a field site might relate to biological responses in field fish i.e. tench (Tinca tinca) and roach with respect to endocrine disruption. These objectives were addressed in the three main research Chapters of this thesis described below.
1.9.1 Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling
In a previous study Buchinger et al. (2013b) identified a “hot-spot” for estrogenic activity in sediment at the Luppe River located in Eastern Germany. The Luppe River was chosen as a worst-case scenario for sediment-borne EDC contamination based on the high levels of NP and E1 (Buchinger et al. 2013b).
In order to assess the risks of remobilization of sediment-bound EDCs to aquatic wildlife at the Luppe river during a potential flood event, the bioavailability of sediment-bound EDCs was examined in the present study under turbulent conditions.
Specifically, the hypotheses of the present study were: 1) NP, E1, E2 and EE2 accumulated in sediments along the Luppe River are the main drivers of estrogenic activity, and 2) Bioavailability of
20 Chapter 1 - Introduction
NP, E1, E2 and EE2 in sediments of the Luppe River is enhanced under turbulent conditions such as in a flood event.
The first hypothesis was tested using the recently developed planar YES assay (Buchinger et al.
2013a) which combines high performance thin-layer chromatography (HPTLC) with the traditional in vitro YES assay, in addition to chemical analysis (LC-MS/MS) to investigate concentrations of EDC and associated estrogenic activity in sediment samples along the Luppe River. In order to test the second hypothesis, two types of passive samplers were used to assess the bioavailability of E1, E2, EE2 and
NP, from this sediment under turbulent conditions in laboratory sediment suspension experiments. Polar organic chemical integrative sampler (POCIS) and Chemcatcher were chosen in order to compare two different sampler configurations regarding their application in a sediment-water-suspension system
(Chapter 3).
1.9.2 Impacts of endocrine disruptors from sediment on rainbow trout (Oncorhynchus mykiss) during a simulated flood event
It is well documented that water-borne exposure to low ng/L concentrations of EDCs can impair the reproduction of fish. In contrast, little is known about the bioavailability and effects of sediment- associated EDCs on fish. Particularly when sediments are perturbed, e.g., during flood events, sediment- bound chemicals may become bioavailable.
The hypotheses of the present study were: 1) Sediment-bound EDCs become bioavailable to fish during a simulated flood event, and 2) Remobilization of EDCs from the sediment during a simulated flood event lead to endocrine responses in the fish.
This was tested by exposing juvenile rainbow trout over 21 days to constantly suspended sediment in the following treatments: i) a contaminated sediment from the Luppe River, a “hotspot” for EDC accumulation, ii) a control sediment (exhibiting only background contamination), iii) a serial dilution of
Luppe sediment with the sediment control, and iv) a water-only control. To investigate the bioavailability of sediment-bound EDCs, passive sampling was used to monitor the water concentration of the target EDCs (E2, EE2, E1, NP) during exposure. Target EDCs were characterized in sediment samples, Chemcatcher extracts, as well as in fish bile and plasma samples by LC-MS/MS. To account
21 Chapter 1 - Introduction for mixture effects as well as other substances exhibiting endocrine disruptive characteristics, the YES assay, as an effect-based method (EBM), was used to measure the estrogenic activity of sediment samples and bile samples. Moreover, the planar-YES (p-YES) assay was used to analyze target EDCs in bile samples in relation to their estrogenic activity. For further complementary insight into the potential effects of EDC exposure to fish, several lines of evidence utilized RNA-sequencing in the male liver to analyze changes in the transcriptome, vtg was measured as a biomarker of exposure and, liver histology was evaluated for alterations on the organ level of the fish in the control and exposure treatments (Chapter 4).
1.9.3 Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment
Little is known about sediment-bound exposure of fish to 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.
The hypothesis was 1) adverse effects such as alterations of gonad histology occur in fish sampled at a “hot-spot” for EDC accumulation in sediment.
In order to address this research question, the estrogenic activity and concentrations of EDCs in different environmental compartments, namely water, sediment, macroinvertebrate (food source for fish) and fish were characterized and evaluated how these relate to biological responses in fish with respect to endocrine disruption. Therefore, tench and roach 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, in addition to fish from a commercial fish farm were used as a reference in the presented research. 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. Concentrations of target EDCs (NP, E1, E2 and
EE2) were measured using LC-MS/MS or the YES in vitro bioassay in water, sediment and fish blood plasma samples. Furthermore, evidence of endocrine disruption in fish was investigated by induction of vtg in tench and roach mucus samples in combination with gonad histology (Chapter 5).
22
Chapter 2
2 Material & Methods
23
24 Chapter 2 – Material & Methods
2.1 Material and Chemicals
Acetone (99.9%), methanol (>99.9%), n-heptane (99.9%), acetonitrile (99.9%), dimethyl sulphoxide
(DMSO, 99.9%), ethyl acetate (>99.9%) were purchased from Carl Roth, Germany. Oasis HLB (30 µm particle size) was purchased from Waters. EmporeTM SPE Disks (SDB-RPS, 47 mm), β-glucuronidase from Helix pomatia exhibiting also sulfatase activity (G7017), 17β-estradiol-3-(β-D-glucuronide),
Bicinchoninic Acid Kit, dansyl chloride (>99%), and RNAlater were purchased from Sigma-Aldrich,
Germany. Chromabond® HLB cartridges and weight glass fiber filters (MN-GF 1) were purchased from
Machery & Nagel, Germany. Supor 100 membrane filters (polyethersulfone (PES), 0.1 µm pore size,
47mm) from Pall corporation were purchased from VWR, Germany. As analytical standards, E1, E2, linear alkyl 4-n-nonylphenol (4-n-NP), and NP (technical mixture of ring and chain isomers) were purchased from Sigma-Aldrich, Germany. EE2 was purchased from the European Directorate for
Quality of Medicines and Healthcare, European Pharmacopoeia. Deuterated standards E1-2,4,16,16-d4
(99%), E2-2,4,16,16-d4 (99%), EE2-2,4,16,16-d4 (98.8%) were from CDN Isotopes, Canada, and 4-n-
NP-2,3,5,6-d4 (99.2%) from Neochema, Germany. An UltraSensitive Salmonid Vitellogenin ELISA kit,
Cyprinid Vitellogenin ELISA kit and Mucus collection set were purchased from TECOmedical GmbH,
Bünde, Germany.
2.2 Water and sediment sampling
Sediments were sampled in July/August 2016 and 2017. In 2016 sediments were sampled from five sites along the Luppe River (Figure 5; Table A.1), whereas in 2017 only one site was resampled to obtain additional sediment for testing. Additionally, sediment was sampled from the Rhine River in Koblenz and sampling was done in cooperation with the German Federal Institute of Hydrology (Koblenz,
Germany). Prior to sediment sampling, velocity, depth, and width of the stream were measured, and a 2
L water sample was taken by vertical pumping, cooled and stored at 4 °C for a maximum of 2 days.
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 of approximately 200 g sediment wet weight (Christ Alpha 1-2, Martin Christ GmbH,
25 Chapter 2 – Material & Methods
Osterode am Harz, Germany) for 120 h, sediments were sieved < 1 mm, ground with a pestle and mortar and extracted.
Sediments were analyzed for physicochemical and chemical properties including particle size distribution, concentrations of heavy metals, 17 PCDD/Fs, 12 dl-PCBs, 4 indicator PCBs, water and organic carbon content (Table A.2, A.3 and A.4). Particle size distribution was determined by laser diffraction analysis, and metals were analyzed using X-ray fluorescence spectroscopy by the Institute of
Geology of the RWTH Aachen. Münster Analytical Solutions Gmbh quantified 17 PCDD/Fs, 12 dl-
PCBs, and four indicator PCBs according to the guideline DIN 38414-24 (DIN 2000). Organic carbon content was determined by dry combustion using an PE 2400 Series II CHNS/O Analyzer.
Figure 5: Sampling sites along the Luppe River in eastern Germany near Leipzig. Map adapted from ESRI. L1- L5: sediment sampling sites 2.3 Sediment extraction
Extraction was adopted from Reifferscheid et al. (2011). Briefly, 10 g dry weight (d.w.) of sediment were extracted using pressurized liquid extraction (PLE) in a SpeedExtractor (Büchi Labortechnik
GmbH, Essen, Germany) at 120°C with 24 mL of a solvent mixture of 1:1 acetone : n-heptane at 118 bar. After desulfurization by activated copper for 24 h, the organic phase was evaporated and exchanged with 4 mL n-heptane. Extracts were stored at -20 °C until further used.
2.4 Water extraction
After filtering 1 L water through a 12 µm glass fiber filter (MN-GF 1, Macherey and Nagel, Düren,
Germany), the water was spiked with 50 ng of the internal deuterated standard mixture (E2, EE2, NP).
26 Chapter 2 – Material & Methods
The water was extracted using OASIS hydrophilic-lipophilic balance (HLB) solid phase extraction cartridges conditioned with 5 mL acetone, methanol, and milli-pore water. Substances were then eluted with 50 mL acetone and evaporated to 0.5 mL. Extracts were then transferred to glass vials, dried under a gentle stream of nitrogen and derivatized with dansyl chloride as described below. Finally, exacts were dissolved in 100 µL methanol and target substances were analyzed with LC-MS/MS.
2.5 Yeast estrogen screen and planar yeast estrogen screen
Estrogenic activity of the sediment and bile extracts redissolved in DMSO was investigated using the YES assay (ISO/FDIS 19040-1:2018-03 2018). Prior to testing with the YES, samples were diluted
1:100 with milli-Q water to a final assay concentration of 1% DMSO, which is tolerated by the bioassay.
Thus, final sediment equivalents in the test were 25 mg/mL and bile equivalents were 0.0006. The YES was performed using the Saccharomyces cerevisiae strain BJ3505 (protease deficient; MAT alpha,
PEP4::HIS3, prb-1-delta1.6R, HIS3-delta200, lys2-801, trp1-delta101, ura3-52gal2can1) according to
McDonnell et al. (McDonnell et al. 1991a; McDonnell et al. 1991b).Each sample was measured at seven dilution levels with four technical replicates. Negative control, process control and E2 standard in seven dilutions were included. At least three independent replicates were performed per sample. The estrogenic activity of the samples was calculated using the E2 standard and is expressed in 17β-estradiol equivalents (ng EEQ/L). By multiplying ng EEQ/L by a factor of 40, the results are expressed as EEQ/kg sediment.
The general procedure of the p-YES has already been described in detail in other studies (Buchinger et al. 2013a; Müller et al. 2004; Schoenborn et al. 2017; Schönborn and Grimmer 2013; Spira et al.
2013). Briefly, extracts were diluted 1:10 with methanol and applied on a HPTLC-plate in volumes ranging from 5 to 50 µL by using an Automatic TLC Sampler 4 (ATS 4, CAMAG). Subsequently, the
HPTLC-plates are chromatographically developed in two steps using an Automated Multiple
Development System (AMD 2, CAMAG). Then, the YES is performed directly on the surface of the
HPTLC-plate by spraying a yeast cell suspension (previously adjusted to 2000 ± 50 formazin attenuation units) onto the plate. After 3 h of incubation and subsequent drying of the plate with cold air for 5 min, the enzyme β-galactosidase, which is expressed in response to the recombinant estrogen receptor stimulation, can be detected by the application of the substrate 4-methylumbelliferyl-β-D-
27 Chapter 2 – Material & Methods galactopyranoside (MUG) in lacZ-buffer (Buchinger et al. 2013a) or the substrate chlorophenol red-β-
D-galactopyranoside (CPRG) (Schick and Schwack 2017). MUG is converted to the fluorescent 4- methylumbelliferone and CPRG to chlorophenol red by the enzyme β-galactosidase. The activity of the
β-galactosidase corresponds to the stimulation of the estrogen receptor and thus allows an indirect detection of the estrogenic activity of the separated sample. The fluorescence signals can be detected by means of fluorescence-imaging at an excitation-wavelength of 366 nm and with a TLC Scanner 4
(CAMAG) at λex = 320 nm. Three to five dilutions of a reference substance mixture (NP, E1, E2, EE2 and estriol (E3)) are included on each plate for calibration purposes to express the signals detected in the sample in terms of EEQs. This allows a quantitative assessment of the agonistic potential of the sample in addition to the qualitative assessment. The calculation of the equivalent concentration of the individual signals detected in the samples is based on the inverse five-parameter log-logistic function of the reference compound (Gottschalk and Dunn 2005). Subsequently, values are expressed in ng/kg
(sediment), ng/L (water), ng/mL (bile) of the original sample considering the enrichment factor, application volume and pre-dilution. Only one test was performed.
2.6 Derivatisation with dansyl chloride
Extracts were derivatized using dansyl chloride according to Lin et al.(2007). For this, 20 µL acetone and 100 µL sodium bicarbonate buffer NaHCO3 (100 mmol/L, pH 10.5) were added to dried extracts.
After vortexing for 1 min, dansyl chloride solution (100 µL, 1 mg/mL in acetone) was added and the sample incubated for 3 min at 60°C. The extract was evaporated to dryness under a stream of nitrogen and dissolved in desired volume of methanol. Target substances E1, E2, EE2 and NP were analyzed with LC-MS/MS.
2.7 Quantification by LC-MS/MS
Target substances 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 flow (0-0.8 min, 60% B, 0.8-3 min, 60 to 80% B, 3-5 min, 80% B, 5-9 min, 80 to 100% B, 9-
28 Chapter 2 – Material & Methods
14 min, 100% B, 14-16 min. 100 to 60% B, 16-20 min, 60% B) 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, sheath gas flow: 30 arb, aux gas flow: 5 arb, sweep gas flow: 0 arb, spray voltage: 4000 V, capillary temperature: 250°C, capillary voltage: 24 V, tube lens: 50 V) performed 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 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 2). Blank samples were produced by using solvents and derivatization without passive sampler.
Table 2: Conditions of mass spectrometry for the analyzed steroidal estrogens. Quantifiers are underlined.
Limit of Retention time Limit of detection Compound Parent mass, fragment1-3 [m/z] quantification [min] (LOD) [ng/mL] (LOQ) [ng/mL] E1 10.24-10.49 504, 425.4/440.4/489.4 1.7 0.9 E2 9.74-9.98 506, 171/427.4/442.4/491.3 3.0 1.6 de.-E2 9.76-10.0 510, 171/431.4/446.4/495.3 EE2 9.76-9.99 530, 171/451.4/466.5/515.2 1.9 1.0 de.-EE2 9.78-10.1 534, 171/455.4/470.4/519.3 NP 13.6-14.07 454, 171/375.3/390.4/439.3 4.0 2.7 de.-NP 14.96-15.1 458, 171/379.4/394.4/443.3 4n-NP 14.74-15.1 454, 171/375.3/390.4/439.3 4.0 2.7 E1= estrone; E2= 17β-estradiol; EE2= Ethynylestradiol; NP= nonylphenols; 4n-NP= 4n-nonylphenol; de.= deuterated; LOQ= meanblank + 7(SDblank); LOD= meanblank + 3(SDblank)
2.8 Extraction of EDCs from plasma
Methods for the extraction of EDCs from plasma were adopted from Anari et al. (2002). Plastic materials were avoided, and all glass materials were solvent-cleaned prior to use. Briefly, 50-µL plasma samples, carefully thawed on ice, were diluted in 4-mL glass vials with 0.5 mL ultrapure water spiked with 50 ng deuterated E2, EE2 and 4n-NP standards. After mixing, 2 mL ethyl acetate was added and
29 Chapter 2 – Material & Methods vortexed for 1 min followed by 1 h incubation at room temperature. The organic phase was transferred into a new vial, dried, derivatized with dansyl chloride and dissolved in 100 µL MeOH. Samples were stored at -20°C until measurement with LC-MS/MS. Method blanks (n=4) were produced following the extraction protocol without plasma sample. Mean blank values for NP were 2.9 ± 0.9 ng/mL. Extraction efficiency (n=4) was tested by spiking blood from the water-only control with 10 ng NP, E1, E2, EE2 standard mix. The recovery for NP was 111 %.
2.9 Vitellogenin analysis
Vitellogenin (vtg) was measured in skin mucus of fish using the UltraSensitive Salmonid or Cyprinid
Vitellogenin ELISA kits (TECOmedical GmbH, Bünde, Germany), respectively. All steps were performed in accordance with the instruction manual using the provided materials. The absorbance of the color reaction was measured at 450 nm (reference filter at 650 nm) using a multiwell plate photometer (TECAN infinite M200, Tecan, Austria). A vtg calibration curve (0.012 - 1 ng/mL or 0 – 1 ng/mL) was fitted using a four-parameter logistic curve. To evaluate the assay performance two salmonid/cyprinid vitellogenin controls were included. Vtg concentrations were corrected by the total protein concentration in the sample and are expressed as ng/mL per mgprotein. Total protein concentrations in mucus samples were determined by a colorimetric reaction using the Bicinchoninic Acid Kit. Briefly,
200 µL of the BCA solution (Bicinchoninic Acid and copper (II) sulfate pentahydrate (4 %) in a ratio of 1:50) was added to 25 µL sample. Additionally, a standard curve using the BCA kit and bovine serum albumin (BSA) as external standard (0.0625 – 1 mg/mL) was prepared. Directly after the incubation time of 30 min at 37 °C standard, blank and samples were measured at the extinction wavelength of 562 nm at 25 °C (TECAN infinite M200, Tecan, Austria). Protein content was then calculated from the BSA standard curve.
30
Chapter 3
3 Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling
31
32 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling
Anne-Katrin Müllera · Katharina Lesera · David Kämpfera · Carolin Riegrafab · Sarah E. Crawforda
· Kilian Smitha · Etienne Vermeirssenc · Sebastian Buchingerb · Henner Hollerta
aRWTH Aachen University, ABBt - Aachen Biology and Biotechnology, Institute of Environmental Research,
Department of Ecosystem Analysis, Worringer Weg 1, 52065 Aachen, Germany bFederal Institute of Hydrology, Section G3 – Biochemistry and Ecotoxicology, Am Mainzer Tor 1, 56068 Koblenz,
Germany cSwiss Centre for Applied Ecotoxicology, Überlandstrasse 133, 8600 Dübendorf, Switzerland
This chapter has been developed based on the following peer-reviewed article:
Müller A-K, Leser K, Kämpfer D, Riegraf C, Crawford SE, Smith K, Vermeirssen E, Buchinger S,
Hollert H (2019) Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling. Water Research. doi: 10.1016/j.watres.2019.06.020
33 Chapter 3 – Bioavailability of estrogenic compounds from sediment
3.1 Abstract
Studies worldwide have demonstrated through in vitro bioassays and chemical analysis that endocrine-disrupting chemicals (EDCs) can accumulate in river sediments. However, remobilization of sediment-bound EDCs due to bioturbation or re-suspension during flood events remains poorly understood. The aim of this study was to evaluate the bioavailability of EDCs, more specifically estrogenic compounds (EC), from sediment under turbulent conditions using a passive sampling approach. Sediment was sampled along the Luppe River, Germany, previously described as a “hotspot” for ECs. The concentration of target ECs and estrogenic activity were investigated using chemical analysis (LC-MS/MS) in addition to a novel screening tool (planar Yeast Estrogen Screen; p-YES) that utilizes high performance thin-layer chromatography plates in combination with an in vitro bioassay
(YES). Estrone (50%, E1) and nonylphenol (35%, NP) accounted for the majority of estrogenic activity reported of up to 20 ± 2.4 µg E2 equivalents per kg dry weight in the Luppe sediments. Two types of passive samplers (polar organic chemical integrative sampler (POCIS) and Chemcatcher) were used to investigate the bioavailability of ECs from suspended sediment under laboratory conditions. NP, E1, E2 and ethynylestradiol (EE2) were remobilized from Luppe sediment when subjected to turbulent conditions, such as in a flood event, and were readily bioavailable at ecotoxicologically relevant concentrations (NP 18 µg/L, E1 14 ng/L, E2 0.2 ng/L, EE2 0.5 ng/L).
Keywords: Endocrine disrupting chemicals, sediment remobilization, passive sampling, p-YES
34 Chapter 3 – Bioavailability of estrogenic compounds from sediment
3.2 Introduction
Over the last two decades, substances interfering with the endocrine system, namely endocrine disrupting chemicals (EDCs), have remained a focal point in terms of scientific and public awareness
(EU COM 734 final 2018). Various adverse effects from exposure to EDCs have been documented in laboratory and field studies of fish (Depiereux et al. 2014; Kidd et al. 2007). Exposure to even low concentrations (5-6 ng/L) of potent EDCs such as ethynylestradiol (EE2) were found to significantly interfere with the sexual development of fish leading to the feminization of male fish through the production of the female egg yolk protein vitellogenin (vtg), impairment of gonadal development evident as intersex and, ultimately, reproductive failure of whole populations (Kidd et al. 2007). To date, over one hundred substances worldwide have been identified as endocrine disruptors including natural
(e.g. 17β-estradiol (E2), estrone (E1)) as well as synthetic hormones (e.g. EE2), many pesticides, organic pollutants such as polychlorinated biphenyls (PCBs), bisphenol A (BPA), phthalates and alkylphenols
(e.g. nonylphenol (NP)) (EU 2013/39). As a consequence, several EDCs, among them NP, were included in the Water Framework Directive (WFD, Annex VIII) within the group of priority substances of particular concern (EU 2013/39). In addition, E2 and EE2 were included in the European watchlist to be monitored, and for which the development of consensus-based risk evaluation strategies for surface waters have been recommended (EU COM 734 final 2018). With the upcoming revision of the WFD, several recommendations for monitoring, assessing and managing risks of contamination in the aquatic environment have been addressed (Brack et al. 2017; Brack et al. 2018; Escher et al. 2018). One key aspect was to include effect-based methods (EBMs), such as bioassays and biomarkers, as integrative techniques that bridge the gap between chemical contamination and ecological effects (Brack et al.
2019a). Based on the research project SOLUTIONS and the European monitoring network NORMAN a basic bioassay battery has been suggested including in vitro tests for estrogenic activity (Brack et al.
2019a; Escher et al. 2018)
Moreover, studies across Europe investigating the fate and distribution of EDCs in the environment have demonstrated using in vitro bioassays and chemical analytical methods that EDCs sorb to suspended particles and, thus, accumulate in river sediments (Hilscherova et al. 2007; Holthaus et al.
2002; Urbatzka et al. 2007; Viganò et al. 2008; Wölz et al. 2011). In the UK, Peck et al. 2004)
35 Chapter 3 – Bioavailability of estrogenic compounds from sediment demonstrated through an in vitro bioassay (Yeast Estrogen Screen (YES)) in combination with chemical analysis that estrogenic activity in sediments downstream from a major sewage treatment plant were about 30 ng E2 equivalents (EEQ)/kg and that the major active chemicals were E1 followed by E2. They concluded that sediments act as sinks for EDCs more specifically estrogenic compounds (EC), which is to be expected due to the hydrophobic (e.g. high log kow) properties of many EDCs. Even higher estrogenic activity in terms of EE2 equivalents (EEEQs) ranging from 0.25 up to 55 µg EEEQ /kg were reported from the YES assay by Buchinger et al. (2013b) in a sediment survey along the Saale River and its tributaries in Germany. Moreover, high levels of NP and E1 were detected by LC and GC
MS/MS, and found to contribute most to the observed estrogenic activity from the German river sediments.
While it has been recognized that ECs have a high tendency to accumulate in sediment, little is known about their bioavailability and potential for remobilization, transport and redistribution due to extreme weather events such as floods, bioturbation or dredging. However, some studies indicate that estrogenic compounds have the potential to be remobilied from sediments during floods as shown by estrogen- related endpoints (Oetken et al. 2005). The frequency and intensity of flood events is predicted to increase in the context of global climate change (Petrow and Merz 2009). On account of this, the
European parliament initiated the directive 2007/60/EC regarding the assessment and management of flood risks, whereby they established a framework aiming to assess and reduce the adverse consequences of floods for human health and the environment (EU 2007/60). One major aspect of such risk assessment includes the evaluation of sediment contamination and the potential of contaminant remobilization from sediments, whereby the sink might turn into a source of pollution (Hollert et al. 2007a). It has been recognized that “good chemical status” and “good ecological status” according to the WFD is not always attainable due to sediment contamination, and recommendations for improvements have included the establishment of environmental quality standards (EQS) for sediments that consider the bioavailability of sediment contaminants (Brack et al. 2017). Bioavailability has been described as the portion of the total quantity or concentration of a chemical in the environment that is potentially available for biological action, such as uptake by an organism. Concentration of freely dissolved organic compounds,
36 Chapter 3 – Bioavailability of estrogenic compounds from sediment such as ECs, are generally assumed to be more readily available for uptake in an organism and thus bioavailable in water(Rand 1995).
The bioavailability of sediment-bound ECs was examined in the present study under turbulent conditions in order to estimate the risk of remobilized ECs to fish during flood events. Specifically, the objectives of the present study, as a part of the interdisciplinary Project House Water consortium
(Crawford et al. 2017), were to investigate 1) the accumulation of ECs and their activity in river sediment and 2) the potential of these compounds to be remobilized from the sediment under turbulent conditions such as in a flood event and thus become bioavailable for fish. The Luppe River was chosen as a worst- case scenario for sediment-borne EC contamination based on the high levels of NP and E1 previously reported by Buchinger et al. (2013b). The recently developed planar YES assay (Buchinger et al. 2013a) which combines high performance thin-layer chromatography (HPTLC) with the traditional in vitro
YES assay, was used in addition to chemical analysis (LC-MS/MS) to investigate concentrations of ECs and associated estrogenic activity in sediment samples along the Luppe River. Two types of passive samplers were used to assess the bioavailability of ECs, specifically E1, E2, EE2 and NP, from this sediment under turbulent conditions in sediment suspension experiments. Polar organic chemical integrative sampler (POCIS) and Chemcatcher were chosen in order to compare two different sampler configurations regarding their application in a sediment-water-suspension system.
3.3 Material and methods
3.3.1 Materials and chemicals
See Chapter 2.1.
3.3.2 Water and sediment sampling
Sediments were sampled in July/August 2016 and 2017. In 2016 sediments were sampled from five sites along the Luppe River (see Figure 5), whereas in 2017 only one site was resampled to obtain additional sediment for testing. Sampling procedure is described in detail in Chapter 2.2.
37 Chapter 3 – Bioavailability of estrogenic compounds from sediment
3.3.3 Water and sediment extraction
Extraction protocols are given in Chapter 2.3 and 2.4. For use in bioassays, 1 mL sediment extract was evaporated to dryness under a gentle stream of nitrogen and dissolved in 1 mL DMSO. For LC-
MS/MS measurements, 500 µL of the sediment extract in n-heptane from sampling site 1 was evaporated to dryness under a stream of nitrogen after adding 5 µg deuterated E2, EE2 and NP standard. Extracts were then derivatized using dansyl chloride according to Lin et al.(2007) (Chapter 2.6).
3.3.4 Yeast estrogen screen and planar yeast estrogen screen with sediment extracts
Estrogenic activity of the sediment extracts redissolved in DMSO was investigated using the YES assay (ISO/FDIS 19040-1:2018-03 2018). YES assay details are described Chapter 2.5. The estrogenic activity of the samples was calculated using the E2 standard and is expressed in 17β-estradiol equivalents (ng EEQ/L). By multiplying ng EEQ/L by a factor of 40, the results are expressed as EEQ/kg sediment.
Additionally, sediment extracts from 2016 were tested in a planar YES (p-YES) assay, which combines high performance thin-layer chromatography (HPTLC) and the in vitro YES bioassay. The general procedure of the p-YES has previously been described in detail in Chapter 2.5. The calculation of the equivalent concentration of the individual signals detected in the samples was based on the inverse five-parameter log-logistic function of the reference compound. Subsequently, values are expressed in ng/kg or ng/L of the original sample considering the enrichment factor, application volume and pre- dilution.
3.3.5 POCIS and Chemcatcher laboratory calibration
A laboratory calibration for E1, E2, EE2 and NP was performed with POCIS and Chemcatcher. For
POCIS, polyethersulfone (PES) membranes were washed three times with 20% methanol in water followed by two times 100% methanol each for 24 h. The cleaned membranes were dried at 50 °C, and afterwards stored in solvent-rinsed aluminum foil at -20 °C. OASIS HLB was cleaned by washing three times with methanol. After the final wash, the OASIS HLB was allowed to dry in the fume hood overnight and was afterwards stored at room temperature. The Empore disks were conditioned by placing them first in methanol for 30 min followed by two times 30 min in milli-Q water. They were
38 Chapter 3 – Bioavailability of estrogenic compounds from sediment stored in milli-Q water at 4 °C until use. POCIS consisted of a membrane-sorbent-membrane sandwich
(see Figure 6). To obtain a surface area per mass of sorbent ratio of 180 cm2/g as described in Alvarez et al. 2007), 54.5 ± 0.5 mg Oasis HLB was sandwiched between two PES membranes (9.8 cm2 surface area) held in place by compression between two stainless steel washers. Empore disks were mounted similarly between two stainless steel washers.
Figure 6: Polar organic chemical integrative sampler (POCIS) (left) and Chemcatcher (middle) configuration and experimental setup used in the calibration study (right).
The calibration of both the POCIS and Chemcatcher was performed using a static renewal exposure setup, in which the samplers were exposed to the target compounds under stirred conditions (ca. 300 rev./min) at 20 ± 2°C in darkness. One single Chemcatcher was applied per beaker with 1 L milli-Q water spiked with a nominal concentration of 5 µg/L of E1, E2, EE2 and 4n-NP and stirred by a magnetic stirrer at 300 rev./min at 20°C. The water was renewed every day and the pH measured. Chemcatchers were removed after exposure times of 2, 4, 7 or 10 days. Duplicates were done per time point.
Additionally, a water control without chemicals was included. Calibration of POCIS was done similarly with slight modifications. Four POCIS were applied per beaker with 1 L milli-Q water spiked with 5
µg/L of each substance and exposed for 4, 7, 14 and 21 days. POCIS calibration was performed in triplicate per sampling time point and one control exposed to water only was included per sampling time point.
Exposed POCIS and Chemcatcher were extracted individually according to Alvarez et al. 2007;
Skodova et al. 2016. Before extraction, 5 µg deuterated E2, EE2 and 4n-NP standards were added to each sampler. Briefly, Chemcatchers were extracted with 10 mL acetone followed by 10 mL methanol for 30 min, each under slow shaking (105 rpm) at room temperature. POCIS were carefully dismantled and the OASIS HLB were washed with methanol from the membrane into a glass column plugged with glass wool and a glass fiber filter. To account for losses, the sorbent mass was determined after extraction
39 Chapter 3 – Bioavailability of estrogenic compounds from sediment and the individual mass of sorbent recovered from each sampler was used in subsequent calculations.
Sorbed compounds were extracted from the OASIS HLB using 40 mL acetone. Dried samples were derivatized with dansyl chlorid and dissolved in 1 mL 40% methanol: milli Q-water. Target substances were determined using LC/MS/MS. To verify the nominal water concentration of 5 µg/L, water samples were taken at several time points throughout the Chemcatcher calibration experiment.
Sampling rates were approximated as linear uptake using the slope of the linear regression of the amount of substance found in each sampler plotted against the sampling time points. Uptake of a substance into a passive sampler during the linear uptake phase can be described by equation 1 (Eq.),
Ms = Rs* Ctwa* t [Eq.1]
where Ms = mass in sampler (ng), Rs = sampling rate (L/day), Ctwa = time weighted average concentration in the water (ng/L) and t = time in days. The sampling rate (Rs) was calculated from m/Ctwa, where m is the slope of the linear regression.
Rs for POCIS were calculated using an iterative approach to correct for the changing number of samplers during the exposure:
Variables
S (x) = no. of samplers in the beaker at timepoint T(n)
Ms (n) = mean mass in samplers at timepoint T(n) [ng]
T(n) = sampling timepoint n [days] t(n) = time span between sampling timepoints (T(n+1) – T(n)) [days]
Ctwa (n) = time weighted water concentration at timepoint T(n) [ng/L]
C0 = expected water concentration at the start [ng/L]
Rs (n) = sampling rate at timepoint T(n) [L/day]
Calculations for sampling timepoint T1 (after 4 d)
Step 1 the initial sampling rate was calculated using the water concentration at the start of the exposure period
(C0) based on the determined average concentration of the substance in water (n=6).
푀푆(푛) 푅푆 (푛) = ∗ 퐶0 [Eq.2] 푇(푛)
40 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Step 2
The new average water concentration was calculated using the initial sampling rate and the amount of samplers
퐶 +퐶 ∗(1−푆 ∗ 푅 ) 퐶 = 0 0 (푥) 푠 (푛) [Eq.3] 푇푊퐴 (푛) 2
Step 3
The sampling rate was calculated again, using the newly determined average water concentration
푀푆 (푛) 푅푆 (푛) = ∗ 퐶푇푊퐴 (푛) [Eq.4] 푇(푛)
Step 2 and step 3 of this calculation were repeated, using the newly determined sampling rate and water concentration, until both values became stable. This is usually the case after 4-7 calculation cycles.
Calculations for sampling timepoint T2, T3 and T4 (after 7, 14, 21 d)
Step 1
The same as for T1, the initial sampling rate was calculated using the water concentration at the start of the exposure period (C0) based on the determined average concentration of the substance in water (n=6).
푀푆(n) 푅푆 (푛) = ∗ 퐶0 [Eq.2] 푇(n)
Step 2
The new water concentration was calculated using the initial sampling rate and the amount of samplers in the beaker at that time, as well as the average water concentration determined for the prior sampling time point (Ctwa (n-1)).
퐶0+퐶0∗(1−푆 ∗ 푅 ) (푥) 푠(푛) ∗푡 +퐶 ∗푇 2 (n) 푇푊퐴 (푛−1) (n−1) 퐶푇푊퐴 (푛) = [Eq.5] 푇(n)
Step 3
Like for T1, the sampling rate recalculated, using the newly determined average water concentration
푀푆(n) 푅푆 (푛) = ∗ 퐶푇푊퐴 (푛) [Eq.4] 푇(n)
Same as for T1, step 2 and step 3 were repeated until both factors became stable.
In an ideal model system, the Rs for all 4 sampling timepoints become stable for the same value. The test with the model data showed that the values became stable after only 3-6 calculation cycles for all sampling timepoints (Figure 7).
41 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Figure 7: Stabilization of the average water concentration (Ctwa) (left) and sampling rates (Rs) (right) after 3-6 calculation cycles. 3.3.6 Suspended sediment experiment
To investigate the bioavailability of EDCs (NP, E1, E2, EE2) present in Luppe sediment, 5 g d.w. from sampling site Luppe1 (2017) was taken, sieved to 2 mm and suspended in 1 L of tap water. The sediment was stirred at 300 rpm at 20°C for 24h in the dark, to allow for equilibration before a passive sampler was added to the sediment suspension. Individual POCIS and Chemcatcher samplers were exposed for 10 or 5 days, respectively, to the suspended sediment under constant stirring (suspension test A).
It is possible that the samplers would cause a depletion of the target compounds in the water phase during the time of exposure or that the system would not be able to sustain the equilibrium state over several days. Therefore, a depletion test was conducted, where a pair of each POCIS or Chemcatcher were successively exposed to the same suspended sediment for 10 or 5 days, respectively. As a first step, single Chemcatcher or POCIS were exposed as described above (suspension test B), after their removal, a new set of samplers were individually deployed into the same suspended sediment. This experiment was repeated for the Chemcatcher (suspension test C). All sediment exposure experiments with POCIS and Chemcatcher (suspension test A-C) were performed in duplicates.
Upon removal, all samplers were cleaned carefully with deionized water before extraction as described above. Target analysis was done by LC-MS/MS. The amount of internal standard for LC-
MS/MS was 50 ng deuterated standard mix of E2, EE2 and NP for all tests with sediment. In addition, one Chemcatcher extract was tested in the p-YES as described above.
42 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Ctwa were calculated using the experimentally determined respective sampling rates and Eq.1. The mass distribution of the different EDCs was used as an indicator for bioavailability in this study setup, and was estimated as:
퐶 ∗푉 푀푎푠푠 푑푖푠푡푟푖푏푢푡푖표푛 [%] = 푡푤푎 푤 ∗ 100 [Eq.6] 퐶푠∗푀푠
Where the initial concentration in sediment [ng/g] is expressed as Cs; mass of sediment [g] as Ms and water volume [L] as Vw.
3.3.7 Quantification by LC-MS/MS
See Chapter 2.7.
3.4 Results
3.4.1 Accumulation of EDCs and related estrogenic activity along the Luppe River
Sediment along the Luppe River was characterized by particle size distribution as clayey silt with a high total organic carbon (TOC) content of 9% except for sampling sites 2 and 5, which had 3% TOC and were classified as loamy or silty sand (Table A.2).
Estrogenic activity expressed as EEQ in sediments measured by the YES varied along the Luppe.
EEQs found at sampling sites 1, 3 and 4 (17.5 ± 1.5, 15.3 ± 0.9 and 20.1 ± 2.5 µg EEQ/kg) were relatively high compared to sites 5 (8.1 ± 0.5 µg EEQ/kg) and 2 (3.4 ± 0.2 µg EEQ/kg) which were about 2.5 and
6 times lower, respectively (Figure 8). Measured EEQs were found to be three times higher during the resampling of site 1 a year later (Table 3). Moreover, in all Luppe sediment extracts, three or even four distinct fluorescence signals could be detected after separation by HPTLC followed by subsequent p-
YES. The retention times (Rf) of these signals aligned with the standard compounds (Rf NP: 0.819 ±
0.010; E1: 0.681 ± 0.014; EE2: 0.567 ± 0.010; E2: 0.4233 ± 0.0079). Figure A.1 shows the HPTLC plate after chromatographic and YES development. Considering the corresponding Rf, as well as the reporter assay induction, NP, E1, and E2 were found in all Luppe sediments, whereas EE2 was only present in samples from sites 1 and 2. Included negative and process controls were below the limit of detection for both tests, the p-YES and YES (see Figure A.1; Blank). Quantitative evaluation of the p-
YES revealed high concentrations of NP (8.7 up to 37.6 mg/kg) and E1 (5.7 up to 92.4 µg/kg), whereas
43 Chapter 3 – Bioavailability of estrogenic compounds from sediment
E2 and EE2, when found, ranged from 0.1 up to 0.9 µg/kg or 0.03 to 0.04 µg/kg, respectively.
Fluorescence signals for the compounds NP, E1 and EE2 that were found in the sediment extracts were compared to the corresponding E2 signal to evaluate their relative potency in EEQs. These qualitative results are summarized in Figure 8, where the area of the circles is proportional to the measured YES
EEQs in µg EEQ/kg and the slices are proportional to the corresponding EEQs for the single substances.
Figure 8: Sediment related estrogenic activities along the Luppe River. The measured estrogenic activities as 17β- estradiol equivalents (EEQs in µg/kg d.w.) in the Yeast Estrogen Screen (YES) are proportional to the area of the respective circles. Each slice is proportional to the related EEQs for the single target substances estrone (E1), 17β- estradiol (E2), ethynylestradiol (EE2) and nonylphenols (NP) investigated with high performance thin-layer chromatography YES (p-YES). Mean EEQ values (Ø) and standard deviation (σ) for each sampling site are indicated.
Relative contributions of the four target ECs to the sediment estrogenic activity were uniform along the Luppe River (Figure 8). E1 accounted for approx. 50% of the total estrogenic effects in µg EEQs/kg
(measured via p-YES), followed by NP accounting for approx. 35%. Although only low concentrations of E2 were detected in the Luppe sediments, due to the high potency E2 accounted for approx. 14% of the estrogenic effect. Finally, the estrogenic effect attributed to EE2, when present, was negligible
(approx. 1%) from sediment across all Luppe sampling sites.
Target compound analysis with LC-MS/MS confirmed the presence of NP, E1 and E2 in sediment extracts from sampling site 1, whereas EE2 was not detected above the limit of detection. Concentrations
44 Chapter 3 – Bioavailability of estrogenic compounds from sediment for NP and E1 were comparable with both methods, p-YES as well as LC-MS/MS (see Table 3), whereas
E2 could be detected but not quantified by means of LC-MS/MS. Additionally, NP (75.4 ng/L) and E1
(19.7 ng/L) were both detected by means of LC-MS/MS in water samples from the Luppe River at sampling site 1. Calculated sediment partition coefficients (Kd) between water and sediment
(Kd=Ctwa/Cs) were 0.001 for NP and 0.067 for E1, indicating strong association of these substances with the sediment phase.
Table 3: Estrogenic activity (17β-estradiol equivalents; EEQs) and the corresponding concentrations of estrogenic compounds in Luppe sediment investigated by the means of Yeast Estrogen Screen (YES), high performance thin- layer chromatography YES (p-YES), and LC-MS/MS.
Concentration by p-YES Concentration by LC-MS/MS Sampling site EEQ [µg/kg] [µg/kg] dry weight [µg/kg] dry weight (year of dry weight NP NP sampling) E1 E2 EE2 E1 E2 EE2 [mg/kg] [mg/kg]
Luppe2 (2016) 3.4 σ=0.2 8.7 5.7 0.1 0.04 ND ND ND ND
Luppe3 (2016) 15.3 σ=0.9 26.2 92.4 0.9 - ND ND ND ND
Luppe4 (2016) 20.1 σ=2.4 31.3 26.1 0.2 - ND ND ND ND
Luppe5 (2016) 8.1 σ=0.5 33.3 15.9 0.1 - ND ND ND ND
ND ND ND ND
E1= estrone; E2= 17β-estradiol; EE2= ethynylestradiol; NP= nonylphenols; LOQ= meanblank + 7(SDblank); LOD= meanblank + 3(SDblank), ND = not determined; σ= standard deviation; - = below LOD 3.4.2 Calibration of POCIS and Chemcatcher
Laboratory calibration of POCIS and Chemcatcher was performed to determine sampling rates (Rs) for the four target compounds NP, E1, E2 and EE2. Uptake of these compounds into Chemcatcher and
POCIS over time using a static renewal system with spiked water is shown in Figure 9. Both passive sampling devices exhibited an integrative (linear) uptake for the steroids and NP over time (R2 ranged from 0.98 and 0.99), except for uptake of NP in POCIS in which no linear uptake was observed (Figure
9).
45 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Figure 9: Calibration of Chemcatcher (left) and POCIS (right) for estrone (E1), 17β-estradiol (E2), ethynylestradiol (EE2) and 4n-nonylphenol (4n-NP). Mass per sampler (Ms, µg) is plotted against sampling timepoints (days). Solid line indicating linear regression and dashed line 95 % confidence intervals.
46 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Table 4: Sampling rates (Rs; L/day) for Chemcatcher and POCIS as well as average (± SD) water (Cw [µg/L]) concentrations of estrogenic compounds before sampler deployment (t0) and after 24h (t24) for Chemcatcher calibration (t24)
Rs [L/day] Chemcatcher calibration Mean C t Compound w 0 Mean C t POCIS Chemcatcher [µg/L] ± w 24 [µg/L] ± SD SD E1 0.081 0.65 6.3 ± 1.0 2.1 ± 0.8 E2 0.090 0.56 5.8 ± 0.9 2.2 ± 1.2 EE2 0.093 0.66 5.2 ± 0.8 1.5 ± 0.7 4n-NP - 0.53 4.9 ± 1.2 0.6 ± 0.3 2 *Rs refer to samplers with a surface area of 9.8 cm .
Water concentration, Cw, of all spiked compounds before sampler deployment (t0) differed up to 25% from nominal water concentration of 5 µg/L (Table 4). Cw for all compounds after 24h of sampling by
Chemcatcher were reduced by at least a factor of three. Temperature was constant during the whole experiment at 20°C and pH was 5.6. All negative controls in both Chemcatcher and POCIS calibration had concentrations that were below the LOD. Data obtained from the laboratory calibrations were used to calculate Rs as described above and in the supplementary information. Rs values of Chemcatcher varied between 0.53 and 0.66 L/d, whereas Rs values obtained for POCIS were approx. seven times lower, ranging from 0.081 to 0.093 L/d (Table 4).
3.4.3 Bioavailability of ECs from sediment under turbulent conditions
E1 and NP were found in the water phase in all sediment re-mobilization tests evaluated by passive sampling, with both POCIS and Chemcatcher, and subsequent LC-MS/MS analysis. However, neither
E2 nor EE2 could be directly detected by LC-MS/MS. Ctwa were calculated for NP and E1 based on the experimentally determined Rs values (see Eq.1), except for NP in the POCIS due to the missing calibration discussed above. Calculated Ctwa for E1 were comparable between POCIS and Chemcatcher and varied from 9.3 up to 28.8 ng/L and 7.5 up to 22.3 ng/L between the suspension tests (A-C), respectively (Table 5). The calculated Ctwa for NP ranged from 2.2 to 28.0 µg/L between the suspension tests A-C and were about 1000 times higher compared to E1.
47 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Table 5: Bioavailability of estrogenic compounds from suspended sediment as detected in the passive samplers by means of LC-MS/MS in three suspension tests
POCIS Chemcatcher E1 E1 NP Test setup Mean Ctwa + Mass Mean Ctwa + Mass Mean Ctwa + Mass SD [ng/L] distribution SD [ng/L] distribution SD [µg/L] distribution [%] [%] [%] Suspension test A 9.3 ± 0.7 3.2 7.5 ± 0.03 2.5 2.2 ± 0.2 2.1 Suspension test B 28.8 ± 6.3 9.8 11.5 ± 0.5 3.9 28.0 ± 3.6 26.7
Suspension NC - 22.3 ± 5.2 7.6 25.7 ± 1.3 24.5 test C Mean + SD 19.1 ± 10.7 6.5 13.8 ± 6.9 4.7 18.6 ± 11.9 17.8 E1: estrone; NP: nonylphenol; NC: not conducted
The equilibrium capacity of the sediment was investigated by consecutive deployment of passive samplers within a constant sediment-water system. The POCIS did not deplete E1 from the sediment- water re-suspension system, since the calculated Ctwa did not differ significantly (Mann-Whitney test; p=0.60) between the first (28.8 ± 6.3 ng/L) and second (24.8 ± 5.0 ng/L) sampling period (10 d later).
In contrast, the Ctwa of both, E1 and NP, for the Chemcatcher were about 2 (6.1 ± 0.4 ng/L ) and 8 (3.5
± 2.8 ng/L) times lower, respectively, after the second sampling period (Mann-Whitney test; p=0.028)
(Figure 10). The Chemcatcher, therefore, depleted E1 and NP from the re-suspension system used in this experimental setup. However, the mass distribution, as an indicator for the bioavailability in this study, was comparable between the two passive sampling systems for E1, varying between 2.5 and
9.8%. Greater bioavailability was observed for NP with 17.8% ending up in the dissolved phase. In addition to LC-MS/MS analysis, one single Chemcatcher extract was investigated in the p-YES. In this extract, E1 (4.8 ng/L), E2 (0.2 ng/L), EE2 (0.5 ng/L) and NP (24.5 µg/L) were detected based on comparison to the included standard compounds.
48 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Figure 10: Depletion test of estrogenic compounds from suspended sediment by passive sampling. Time weighted average concentrations (Ctwa) of estrone (E1) and nonylphenol (NP) detected in POCIS (dark grey bars) and Chemcatcher (light grey bars) by means of LC-MS/MS. Plain and striped bars refer to subsequent deployed samplers in the same suspended sediment to evaluate depletion caused by the passive sampler. Asterisk indicates significant differences detected by Mann- Whitney test. Mean values are given for POCIS (n=2) and Chemcatcher (n=4) + SD as error bars. 3.5 Discussion
In the present study, concentrations of ECs accumulated in sediments along the Luppe River, which had earlier been described as a “hotspot” for EC contamination, could be related to the corresponding estrogenic activity. We used a novel screening tool (p-YES), which combines high performance thin- layer chromatography with an in vitro bioassay (YES), to conclude that mostly E1 and NP accounted for the high estrogenic activity found in the Luppe sediments. EEQs as well as concentrations of ECs determined for sediments along the Luppe River in the present study were similar to that previously found at the same sampling site (55 µg EEEQ/kg; NP 115 mg/kg; E1 20.4 µg/kg; E2 1.5 µg/kg
Buchinger et al. 2013b). The fact that our findings agree with the previous study provides further support for the assumption of Buchinger et al.(2013b) that the sediment-associated estrogenic activity of the
Luppe is historical rather than the result of modern sources of input. As previously reported, the natural riverbed of the Luppe is divided into an upper reach (Neue Luppe) and lower reach (Luppe) as a result of flood protection measures (Buchinger et al. 2013b; Kammerad et al. 2014). Consequently, the flow velocity in the Luppe is low (Table A.1), leading to low sediment transport along the Luppe and small amounts of newly deposited sediment. Buchinger et al. 2013b) thus concluded that old sediment layers are located close to the sediment surface, which is further supported by the results of the present study
49 Chapter 3 – Bioavailability of estrogenic compounds from sediment demonstrating a stable EC contamination pattern, as well as uniform estrogenic activity along the lower reach of the Luppe. Despite the overall high estrogenic activity along the river course determined from the current study, measured activities at sampling site 2 and 5 appeared to be lower compared to the other sites. The lower estrogenic activity might be explained by the differences in physicochemical properties of the sediment among the sites, with the sediment type at these sites having fewer fine particles (approx. 20% less silt), 3-fold lower TOC content, and more sandy material compared to the other sites (Table A.2). The findings of our study agree with previous literature reports which showed that EDC concentrations and estrogenic activity are higher in fractions of fine particles, and that organic matter has a significant influence on the sorption of EDCs to sediment (Duong et al. 2009; Yamamoto et al. 2003; Zhou et al. 2007b). The fact that EE2 was only detected in sediment samples from site 1 and
2 indicates that EE2 pollution might result from a different source of this pollutant, possibly from surrounding households or sewage treatment plants (Adeel et al. 2017; Wang et al. 2016). Moreover, the 3-fold higher estrogenic activity in Luppe sediment from 2017 compared to 2016 together with the unchanged NP and E1 concentrations might indicate that additional ECs contribute to the estrogenic activity in Luppe sediment from 2017. In agreement with Buchinger et al. (2013b), the present study demonstrates that the Luppe serves as a “hotspot” for ECs, with contamination in sediments exceeding the mean estrogenic activity found in the nearby catchment area of the Saale River by at least 37 times.
Comparable contamination profiles of EDCs in Luppe sediments have been reported in sediments in rural, urbanized or industrialized areas worldwide, mostly identifying NP E1, E2 and EE2 but also BPA and other alkyphenols as the active compounds and were exceeded by at least a factor of four in the present study (Peck et al. 2004; Schmitt et al. 2012; Viganò et al. 2008). The fact that NP (banned from
EU countries) was detected in water samples of the Luppe by LC-MS/MS in the current study demonstrates how sediment contamination might negatively impact water quality. The results from the present study highlight the importance of integrating sediment risk assessment in current legislation such as the WFD. However, currently there are no sediment quality guidelines (SQGs) and EQSs established for EDCs or other chemical pollutants worldwide (Brack et al. 2017). Several comprehensive methods and approaches for establishing SQGs for EDCs have suggested the combined use of chemical- analytical techniques and EBMs such as bioassays (Kwok et al. 2014; MacDonald et al. 2000).
50 Chapter 3 – Bioavailability of estrogenic compounds from sediment
Furthermore, EBMs have been recommended for implementation in the WFD in order to bridge the gap between chemical contamination and ecological effects (Brack et al. 2019a). The p-YES used in the current study was demonstrated to be a robust and fast screening tool for the detection of ECs that is applicable to sediment extracts and provides information about the quality, in terms of EEQs, and the quantities of target ECs in environmental samples. Moreover, data obtained in the present study by the p-YES were comparable to conventional chemical analysis but had an advantage in that they had lower
LODs and LOQs.
Another question addressed in this study was whether sediment-bound ECs might become bioavailable under turbulent conditions such as a flood event. Passive sampling was used to evaluate the bioavailability of ECs from suspensions of Luppe sediment, since it targets freely dissolved concentrations of ECs. Freely dissolved compounds represent the effective concentration for diffusive transport and partitioning and, thus, is generally assumed to be bioavailable for organisms (Rand 1995).
Two types of passive sampling devices, POCIS and Chemcatcher, were evaluated for their suitability to determine freely dissolved concentrations of ECs in sediment suspension systems. Chemcatcher exhibited, as reported by previous studies, higher sampling rates compared to the POCIS, that favors monitoring of shorter time frames during which flood events might occur under field conditions. The sampler calibration performed in the present study was simplified compared to other calibration studies
(Vermeirssen et al. 2013), since only a few data points (four) were used for linear regression. However, it was not within the scope of the current study to focus on the determination of ECs sampling rates, but rather the aim was to derive Rs values suitable for studying bioavailability of sediment-borne ECs under conditions close to that of the calibration. The sampling rate for the target compounds E1, E2, EE2 and
NP derived from laboratory calibration with POCIS and Chemcatcher were in the range of previously reported values for these sampling devices (Alvarez et al. 2007; Li et al. 2010; Skodova et al. 2016;
Vermeirssen et al. 2013). Although successful sampling of NP with the POCIS has been reported in the past (e.g. linear uptake; Li et al. (2010)), the non-linear uptake kinetics for NP in POCIS observed in the present study is also supported by similar findings described in other publications (Vallejo et al. 2013).
Moreover, E1 and NP were depleted in the sediment suspension (5gdw/L) experiments with the
Chemcatcher (suspension test B-C). Our results suggest that the uptake of the freely dissolved
51 Chapter 3 – Bioavailability of estrogenic compounds from sediment substances in the water into the Chemcatcher in this experimental setup was faster than the dissociation of the substance from the suspended sediment. This effect is important to consider in laboratory studies with a defined amount of sediment, but might be negligible under field conditions where the sediment acts as an infinite source. Regardless, the usefulness of passive sampling devices in the monitoring of environmental concentrations of EDCs in the context of water quality but also sediment impacts, has been demonstrated in a number of studies (Novák et al. 2018; Vermeirssen et al. 2005), and is further supported by the results of the current study with regard to the monitoring ECs in suspended sediment.
The bioavailability of ECs, inferred by the measure of dissolved concentrations of NP, E1, E2 and
EE2, in the sediment suspension experiments ranged from 2 up to 27%. Freely dissolved concentrations of NP were approximately 250-fold greater during suspension of sediment than under static conditions measured in water samples of the Luppe River. The results of the present study highlight the importance of considering remobilisation and bioavailability of sediment-bound ECs in terms of flood events and demonstrate how a sink for pollutants can turn into a source. Moreover, NP and E1 in the test water were present at environmentally relevant concentrations (NP 18.6 µg/L and E1 19.1 ng/L) in which adverse effects on fish health and reproduction have been documented in field and laboratory studies. A three week exposure to concentrations of 30 µg/L NP caused significant elevations of up to 1000-fold vtg induction in male rainbow trout accompanied by inhibited testicular growth and disturbed spermatogenic development in testes (Jobling et al. 1996). Similar findings were reported by Thorpe et al. (2000), who reported induction of vtg in juvenile rainbow trout after 14 days of exposure to 16 µg/L
NP. Comparable exposure studies conducted with rainbow trout exposed to E1, showed vtg induction at 3.3 ng/L after 14 days (Thorpe et al. 2003) or 50 ng/L after 21 days (Routledge et al. 1998). Although many EDCs, including NP and E1, reveal lower estrogenic potencies relative to E2 or EE2 in terms of biomarker induction, additive effects have been observed for exposure to mixtures of EDCs (Thorpe et al. 2003). However, the magnitude of EDCs exposure in fish is strongly dependent on life-stage, duration, concentrations, uptake-routes and species sensitivity. High amounts of suspended particular matter (SPM) were found in rainbow trout intestine after exposure to suspended sediment and can, thus, serve as an additional route of exposure (Hudjetz et al. 2014). Our results indicate that, despite the presence of environmentally relevant concentrations of freely dissolved ECs in the water, major parts
52 Chapter 3 – Bioavailability of estrogenic compounds from sediment of the estrogenic activity during re-suspension of sediment are still attributed to SPM. Thus, it is important to not only consider the estrogenic activity of freely dissolved concentrations of EDCs in the water phase but also of the sediment phase (e.g. SPM) when assessing the risks of remobilized EDCs to fish in the context of flood events.
3.6 Conclusion
Our results show that ECs accumulated in river sediments but can be remobilized and become readily bioavailable for aquatic organisms when sediment is re-suspended. While further research is needed to investigate and quantify possible adverse effects of remobilized sediment-bound ECs on biota, our findings suggest that events like flooding have the potential to remobilize dormant contaminant loads from sediments. Increased bioavailability of sediment-bound ECs during remobilization events may result in concentrations that cause long-term adverse effects on the ecosystem. Future research should address the remobilization potential in the context of sediment composition with special emphasis on particle size and TOC content of sediment.
3.7 Acknowledgments
This study has generously been supported by a Project House of the Exploratory Research Space
(ERS) at RWTH Aachen University, as part of the German Excellence Initiative via the German
Research Foundation (DFG). The study was also supported by the SOLUTIONS project (European
Union's Seventh Framework Programme for research, technological development and demonstration under Grant Agreement No. 603437). The work at the Federal Institute of Hydrology (BfG) was supported by the Federal Ministry of Education and Research (BMBF) and the Ministry of Science,
Technology and Space (MOST), Israel in the framework of the German-Israeli Cooperation in Water
Technology Research (Project: TREES, Grant No. 02WIL1387).
53
54
Chapter 4
4 Impacts of endocrine disruptors from sediment on rainbow trout (Oncorhynchus mykiss) during a simulated flood event
55
56 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout
Impacts of endocrine disruptors from sediment on rainbow trout (Oncorhynchus mykiss) during a simulated flood event
Anne-Katrin Müllera · Nele Markerta · Katharina Lesera · David Kämpfera · Sabrina Schiwyab · Carolin
Riegrafac · Sebastian Buchingerc · Lin Gand · Ali T. Abdallahd · Bernd Denecked · Helmut Segnere ·
Markus Brinkmannf · Sarah E. Crawfordab and Henner Hollertab
aRWTH Aachen University, Institute for Environmental Research, Worringer Weg 1, 52074 Aachen, Germany bCurrent affiliation: Goethe University Frankfurt, Department of Evolutionary Ecology and Environmental
Toxicology, Max-von-Laue-Str. 13, 60438 Frankfurt am Main, Germany cFederal Institute of Hydrology, Am Mainzer Tor 1, 56068 Koblenz, Germany dIZKF Genomics Facility, University Hospital Aachen, Pauwelsstr. 30, 52074 Aachen, Germany eUniversity of Bern, Centre for Fish and Wildlife Health, Länggassstr. 122, 3012 Bern, Switzerland fUniversity of Saskatchewan, School of the Environment and Sustainability & Toxicology Centre, Saskatoon,
Canada
This chapter will be submitted as manuscript to be published in a peer-reviewed journal:
Müller A-K, Markert N, Leser K, Kämpfer D, Schiwy S, Riegraf C, Buchinger S, Gan L, Abdallah AT,
Denecke B, Segner H, Brinkmann M, Crawford SE, Hollert H (2020). Impacts of endocrine disruptors from sediment on rainbow trout (Oncorhynchus mykiss) during a simulated flood event. To be submitted to Aquatic Toxicology.
57 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout
4.1 Abstract
Numerous environmental pollutants have the potential to accumulate in sediments, and among them are endocrine-disrupting chemicals (EDCs). It is well documented that water-borne exposure to low ng/L concentrations of EDCs can impair the reproduction of fish. In contrast, little is known about the bioavailability and effects of sediment-associated EDCs on fish. Particularly when sediments are perturbed, e.g., during flood events, sediment-bound chemicals may become bioavailable. The main objectives of this study were to evaluate a) whether during a simulated flood event sediment-bound
EDCs become bioavailable to fish, and b) whether this leads to endocrine responses in the fish. Juvenile rainbow trout (Oncorhynchus mykiss) were exposed over 21 days to constantly suspended sediment in the following treatments: i) a contaminated sediment from the Luppe River, a “hotspot” for EDC accumulation, ii) a control sediment (exhibiting only background contamination), iii) a serial dilution of
Luppe sediment with the sediment control, and iv) a water-only control. Measured estrogenic activity using in vitro bioassays as well as target analysis of nonylphenol and estrone via LC-MS/MS in sediment, water, fish plasma, as well as bile samples, confirmed that sediment-bound EDCs became bioavailable during the simulated flood event. EDCs were dissolved in the water phase, as indicated by passive samplers, and were readily taken up by the exposed trout. An estrogenic response of fish to
Luppe sediment was indicated by increased abundance of transcripts of typical estrogen responsive genes i.e. vitelline envelope protein α in the liver and vitellogenin induction in the skin mucus.
Keywords: Endocrine disruption, estrogenic potential, sediment remobilization, flood events,
Oncorhynchus mykiss, nonylphenol
58 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout
4.2 Introduction
During the past decades, several severe flood events have occurred in central Europe, including
Germany (Kundzewicz et al. 2005; Schwandt and Hübner 2014). The likelihood and intensity of flood events have been predicted to increase as a result of global climate change (Alfieri et al. 2018). These have raised public and scientific awareness for historical contamination of river sediments that are subject to remobilization, transport, and redistribution during flood events and, thus, may pose a risk to the environment as well as human health (Hollert et al. 2007a; Hollert et al. 2014). To minimize adverse consequences of flood events, the European Parliament initiated the Directive 2007/60/EC on the assessment and management of flood risk (EU 2007/60). One key aspect of such risk assessment is the evaluation of potential sources of environmental pollution as a consequence of floods (EU 2007/60).
Even under normal weather conditions, it has become evident that “good chemical status” and “good ecological status” according to the Water Framework Directive (WFD) is not always attainable due to sediment contamination (Brack et al. 2018; Brack et al. 2019b). As a consequence, the establishment of environmental quality standards (EQS) for sediments that consider the bioavailability of sediment pollution has been suggested for implementation into the WFD during its periodical revisions (Brack et al. 2017; Brack et al. 2019b).
To date, numerous studies worldwide have demonstrated that sediments function as a sink for a great variety of environmental pollutants, including heavy metals as well as many organic pollutants, e.g., polychlorinated biphenyls (PCBs), bisphenol A (BPA), polyaromatic hydrocarbons (PAH), polychlorinated dibenzo-p-dioxins (PCDDs) and furans (PCDFs) (Brinkmann et al. 2015; Niehus et al.
2018). Moreover, substances interfering with the endocrine system, so-called endocrine-disrupting chemicals (EDCs), such as 17β-estradiol (E2), estrone (E1), 17α-ethynylestradiol (EE2) and alkylphenols, including nonylphenol (NP), have been found to accumulate in sediments up to concentrations of 101 µg/kg E2 equivalents (EEQs) using in vitro bioassays (Li et al. 2019; Macikova et al. 2014; Peck et al. 2004; Viganò et al. 2008). This is of particular concern since waterborne exposure to even low concentrations of EDCs in the ng/L range has been shown to severely impair the reproduction of fish (Kidd et al. 2007). Feminization of male fish is one of the most notable adverse impacts of exposure to EDCs and the production of the female egg yolk protein vitellogenin (vtg) has
59 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout been observed to coincide with impairment of gonadal development evident as intersex and, ultimately, reproductive failure (Jobling et al. 1996; Kidd et al. 2007; Thorpe et al. 2003). Estrogenic activity in water has been reported to increase during suspension of sediments in laboratory experiments as well as during flood events due to remobilization of EDCs from the sediment (Ding et al. 2019; Li et al. 2013;
Müller et al. 2019; Weert et al. 2010; Wölz et al. 2011). However, few studies have investigated the potential effects of sediment-bound EDCs on aquatic wildlife, with only a few examining the effects on fish under static conditions (Ali et al. 2018; Kolok et al. 2007; Sangster et al. 2014; Sangster et al. 2016;
Sellin et al. 2010). To the best of our knowledge, no study has evaluated the risk of remobilized sediment-bound EDCs to fish under flood-like conditions, despite the potentially greater bioavailability of EDCs.
The main objective of this study was to evaluate the risks of sediment-bound EDCs during flood events regarding a) whether they might become bioavailable to fish, i.e., rainbow trout (Oncorhynchus mykiss), when sediments are resuspended, and b) whether uptake of EDCs consequently lead to alterations of the endocrine system in fish when exposed to a complex environmental mixture.
Therefore, juvenile rainbow trout were exposed over 21 days to suspensions of sediment from 1) the
Luppe River as sediment highly contaminated with EDCs, 2) the Rhine River as sediment control, 3-5) three dilutions of Luppe sediment with sediment control: 1:2, 1:4, and 1:8, and 6) a water-only control.
Passive sampling (Chemcatcher, SDB-RPS disks; Müller et al. 2019) was used to monitor the water concentration of the target EDCs (E2, EE2, E1, NP) during exposure. Target EDCs were characterized in sediment samples, Chemcatcher extracts, in fish bile and in plasma samples by LC-MS/MS. To account for mixture effects as well as other substances exhibiting endocrine disruptive characteristics, the yeast estrogen screen (YES) assay was used for bioanalytical determination of the overall estrogenic activity of sediment and bile samples (ISO/FDIS 19040-1:2018-03 2018; Vermeirssen et al. 2005).
Moreover, the planar-YES (p-YES) assay was used to analyze target EDCs in bile samples in relation to their estrogenic activity (Buchinger et al. 2013a). For further complementary insight into the potential effects of EDC exposure to fish, we employed RNA-sequencing of the male liver to analyze changes in the transcriptome, presence of vtg in the skin mucus as a biomarker of exposure and liver histology to assess adverse effects at the organ level.
60 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout
4.3 Material and Methods
4.3.1 Sediment sampling and investigation
Sediment was sampled from the Luppe River, a tributary to the Saale River close to Leipzig (N:51
23 08.0; E:12 00 33.2) and the Rhine in Koblenz at the location Ehrenbreitstein (N: 50 21 19.80; E:7 36
33.56). Sampling was done in cooperation with the German Federal Institute of Hydrology (Koblenz,
Germany) in August 2016 and July 2017, respectively (Figure 5). The Luppe River was earlier identified as a “hotspot” for accumulation of EDCs in sediment (Buchinger et al. 2013b; Müller et al. 2019), whereas sediment from the Rhine, at the location Ehrenbreitstein in Koblenz, was classified as mildly contaminated and used as reference sediment (later referred to as sediment control) in previous studies
(Brinkmann et al. 2010; Brinkmann et al. 2015). Fine grain-sized sediment was sampled from sedimentation zones with reduced velocity by use of a 20-L van Veen stainless steel grab sampler or stainless-steel shovels. Pooled and homogenized sediment samples were stored in 30-L aliquots at 4 °C until use.
Both sediments were analyzed for physicochemical and chemical properties, including particle size distribution, organic carbon content, as well as the concentrations of heavy metals, 17 PCDD/Fs, 12 dioxin-like PCBs (dl-PCBs), and four indicator PCBs (Table A.2, A.3, and A.4). Particle size distribution was determined by laser diffraction analysis, and metals were analyzed using X-ray fluorescence spectroscopy by the Institute of Geology of the RWTH Aachen (Müller et al. 2019).
Organic carbon content was determined by dry combustion using an PE 2400 Series II CHNS/O
Analyzer. Münster Analytical Solutions Gmbh quantified 17 PCDD/Fs, 12 dl-PCBs, and four indicator
PCBs according to the guideline DIN 38414-24 (DIN 2000). Prior to exposure, the estrogenic activities in the sediment samples were quantified using the YES assay. The EEQs determined were approx. sixty times greater in the Luppe sediment (32.8 µg EEQ/kg dry weight (d.w.)) compared to the sediment control (0.57 µg EEQ/kg d.w.). Furthermore, target substances including NP (Luppe: 37.3 mg/kg; sediment control: 39.6 µg/kg), E1 (Luppe: 93.9 µg/kg; sediment control: 1.9 µg/kg), E2 (Luppe:
LOQ 61 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.3.2 Experimental fish Mixed-sex juvenile rainbow trout (weight:18.8 ± 5.9 g; standard length:11.8 ± 1.2 cm) were purchased from commercial aquaculture (Mohnen Aquaculture, Stolberg, Germany). Fish were held in groups of 150 individuals in 1000-L glass fiber-reinforced plastic tanks in a recirculating system with a 400-L biofilter and UVC-sterilizer for at least six months to acclimatize to laboratory conditions. Water was exchanged at a rate of approx. three full replacements per day using municipal tap water. Water quality, including pH, temperature, conductivity, nitrate, nitrite, ammonium, and total hardness, was monitored once per week. Light and dark phases were 12 h each, and fish were fed commercial trout pellets (Star Alevin 2.0 mm, Alltech® Coppens, Netherlands) at a rate of 1% body weight per day. All experiments were conducted in accordance with the Animal Welfare Act and with permission of the federal authorities (Ministry for Environment, Agriculture, Conservation and Consumer Protection of the State of North Rhine-Westphalia, Germany, registration number 84-02.04.2017.A193; 2010/63/EU 2010; TvT e.V. 2010). 4.3.3 Experimental design Juvenile rainbow trout were exposed over 21 days in groups of 24 fish in six different treatments: 1) municipal tap water as water-only control; 2) Rhine sediment sampled at the location Ehrenbreitstein as a sediment control; 3) Luppe sediment as highly contaminated sediment with EDCs and 4-6) three dilutions of Luppe sediment with the sediment control: 1:2, 1:4, and 1:8. A concentration of 5 g/L d.w. of sediment was used in all treatments and the sediment control by adding the corresponding amount of wet sediment to the 400 L exposure water tanks. The sediment dilutions were prepared on d.w. basis: for the 1:2 dilution 2.5 g d.w. Luppe sediment were diluted in 2.5 g d.w. Rhine sediment (sediment control). Similarly, the 1:4 and 1:8 dilution were prepared using 1.25/ 3.75 (1:4) or 0.63/ 4.37 (1:8) g d.w. Luppe and Rhine sediment, respectively. Sediment was constantly suspended using submersible pumps installed on the bottom of the 500-L glass fiber-reinforced plastic tanks, separated through a wide-meshed grid from the fish. This design was successfully used for sediment suspension experiments previously described by Brinkmann et al. (2015). Water quality parameters were monitored throughout the experiment. The temperature was held constant at 7.7 °C, and the oxygen concentration was approx. 11 mg/L during the entire exposure period. Sediment suspension samples were collected every second 62 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout day as 0.5-L duplicates. Samples were filtered through pre-weight 0.7-μm glass fiber filters using a Büchner funnel. After drying overnight at 105 °C, samples were weighed and the suspended particulate matter (SPM) concentration in g/L was calculated. A complete water and sediment exchange was done for all treatments after ten days of exposure. Fish were held during this time (approx. 30 min) in separate aerated tanks filled with municipal tap water. Light and dark phases were 12 h each, and fish were fed daily at 1% of their body weight throughout the experiment. Food was withheld one day prior to sampling of fish on day 21. 4.3.4 Sediment extraction Sediment samples from all five treatments (sediment control, Luppe, and the three dilutions) were taken prior to the start of fish exposure. Extraction techniques were adopted from Reifferscheid et al. (2011) and were described in further detail in Chapter 2.3. Target substances including NP, E1, E2 and EE2 were analyzed using LC-MS/MS (see Chapter 2.7 for details). Additionally, sediment extracts were stored for later analysis of the estrogenic activity through the in vitro YES bioassay (Chapter 2.5). 4.3.5 Passive sampling To monitor the dissolved water concentration (Cfree) of EDCs during exposure, passive samplers (i.e., Chemcatcher; Empore disks SDB-RPS) were included in the experimental set-up. Prior to the exposure study, Chemcatcher samplers were laboratory-calibrated for the target substances E1, E2, EE2, and 4n- NP.Sampling rates (Rs) [L/day] were 0.65 (E1), 0.56 (E2), 0.66 (EE2) and 0.53 (4n-NP) as described in the previous Chapter 3. Cfree were calculated using the experimentally determined respective Rs and Eq.1 (details on calibration Chapter 3). Two Chemcatcher samplers per exposure tank were installed, each in a wide-meshed stainless-steel box to avoid damage through the fish and were replaced by a new pair every seven days. Upon removal, Chemcatcher samplers were carefully cleaned with ultrapure water and immediately extracted according to protocols published by Vermeirssen et al. (2013). Details on conditioning and extraction of Chemcatcher are given in Chapter 3. Target substances were determined using LC-MS/MS (Chapter 2.7). 63 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.3.6 Fish sampling Before and after 21 days of exposure, skin mucus samples for vtg analysis were taken from each fish using swabs from the Mucus Collection Set. Routinely, the fork length (± 0.1 cm) and body weight (± 0.1 g) were recorded before and after exposure. Since there is no sexual dimorphism in juvenile rainbow trout, sex could only be determined after exposure by examination of the gonads. Fish were rapidly killed and dissected according to procedures approved by animal welfare (TvT e.V. 2010). A heparinized syringe was used to sample blood through the caudal vein, immediately cooled on ice, centrifuged at 10,000 g for 8 min, and plasma was transferred to glass vials stored at -80°C until use. Gonad, liver, and bile were carefully isolated. Gonad and liver were weighed (± 0.001 g), and sex noted. The liver was split into three pieces of approximately the same size, where two pieces were transferred into 2-mL cryogenic vials for RNA sequencing, and one piece was taken for histology. Liver samples for histology were transferred into histology cassettes (neoLab Migge GmbH, Heidelberg, Germany), fixated according to Johnson et al. 2010) in Davidson’s solution (AppliChem GmbH, Darmstadt, Germany) for 24 h, and afterwards transferred into 10 % formalin (Sigma-Aldrich Chemie GmbH, Steinheim, Germany) and stored at 4°C. Bile for EDC analysis, as well as the liver for RNA sequencing, were snap-frozen with liquid nitrogen in 2-mL cryogenic vials, kept on dry ice, and then stored at -80°C until further analysis. Somatic indices were determined for each fish including Fulton’s condition index (K), the gonadosomatic index (GSI) and the liver somatic index (LSI). 4.3.7 Extraction of EDCs from plasma Extraction method for EDCs form plasma was adopted from Anari et al. (2002) and is described in detail in Chapter 2.8. 4.3.8 Extraction of EDCs from bile Bile extraction was adapted from protocols published by Vermeirssen et al. (2005) and Gibson et al. (2005). Plastic material was avoided, and all glass material was solvent cleaned before use. Briefly, 15 µL bile per individual fish were added to 10 µL β-glucuronidase and 500 µL sodium acetate buffer (50 mM, pH 5.06) in 4-mL glass vials. The efficiency of the deconjugation enzymes and methods blank (n=3) was tested by using 1 µg 17β-estradiol 3-(β-D-glucuronide) following the extraction protocol, and 64 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 50 % of conjugated E2 was recovered as free E2. Since some bile samples were <15 µL, these samples were excluded from the analysis. After incubation at 37°C for 18 h, the samples were diluted with 2 mL ultrapure water and passed through HLB cartridges previously conditioned with 5 mL each of acetone, methanol, and ultrapure water. After washing with 5 mL ultrapure water and drying under vacuum, samples were eluted with 5 mL methanol. The eluates were then evaporated under a gentle stream of nitrogen to a total volume of 500 µL (water and sediment control) or 1,000 µL (Luppe and dilution of Luppe with sediment control). In addition to targeted analysis with LC-MS/MS, total estrogenic activity was also investigated in the YES assay (Chapter 2.5). For LC-MS/MS measurement, 50 ng internal deuterated standard mix was added to 250 µL of the samples (water and sediment control), or 10 µL (Luppe and dilution of Luppe with sediment control) before derivatization with dansyl chloride (Chapter 2.6). Methanol extracts were solvent exchanged to DMSO for testing with the YES bioassay. Prior to testing with the YES according to ISO/FDIS 19040-1:2018-03 2018, samples were diluted 1:100 with ultrapure water to a final assay concentration of 1% DMSO, which is tolerated by the bioassay. A final enrichment factor of 0.0006 of bile was used in the YES test. Each sample was measured at seven dilution levels with four technical replicates. Solvent control, process control and E2 standard in seven dilutions were included. At least three independent replicates were performed per sample. The estrogenic activity of the samples was calculated using the E2 standard as reference and was expressed as 17β-estradiol equivalents (ng EEQ/L bile). EEQ values for some of the fish exposed to the sediment (n=8) and water-only control (n=10) were below the LOD. For further statistical analysis, the determined LOD values were used. Moreover, bile samples were tested in the planar-YES assay, which utilizes high performance thin-layer chromatography (HPTLC) plates in combination with the YES assay. The general procedure of the p- YES has previously been described in detail in Chapter 2.5. The calculation of the equivalent concentration of the individual signals detected in the samples was based on the comparison to the proportion of the 5 pg E2 reference compound. Subsequently, values are expressed in ngEEQ/mL of the original sample considering the enrichment factor and application volume. This evaluation method might not allow for an accurate quantification but rather provide a qualitative estimation. 65 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.3.9 Modeling approach Uptake and metabolism of sediment-borne nonylphenol were predicted using a multi-compartment physiologically-based toxicokinetic (PBTK) model for rainbow trout described initially by Nichols et al. (1990), with modifications by Stadnicka et al. (2012) and Brinkmann et al. (2014a). The model was re-implemented in Python language (http://www.python.org) using Jupyter Notebooks (Kluyver et al. 2016) for the present study. The code is available from the corresponding author upon request. Seven different compartments (fat, liver, kidney, richly perfused tissues, poorly perfused tissues, blood, and bile) were explicitly represented in the model. Measured aqueous concentrations of nonylphenol as described in section 2.5, as well as measured temperature, dissolved oxygen concentration, and the weight of each fish, were used as input variables. Hepatic biotransformation of nonylphenol was extrapolated from in vitro intrinsic clearance values published by Nichols et al. (2018) using the spreadsheet model provided by Nichols et al. (2013), and incorporated into the PBTK model as described in Stadnicka-Michalak and Schirmer (2019). Based on a previous model by Brinkmann et al. (2014b), the complete mass of cleared NP was assumed to be instantly secreted into the gallbladder and corrected for the mass difference between NP and hydroxy-NP since the concentrations of metabolites were determined as de-conjugated hydroxy- nonylphenol. Concentrations of hydroxy-NP equivalents in bile liquid were then calculated by dividing the chemical mass with bile volume. Bile volume data from fish > 24h after feeding were obtained from Grosell et al. (2000) and resulted in an average value of 1.92 mL bile per kg fish. Bile flow was assumed to be 74.4 µL h-1 kg-1 based on a study by Gingerich et al. (1978) conducted at 12°C. Resulting modeled biliary hydroxy-NP equivalent concentrations were plotted against measured biliary hydroxy-nonylphenol equivalents, and the regression line, as well as the goodness of fit (R2) and the root mean squared error (RMSE) were calculated. 4.3.10 Vitellogenin analysis The general assay procedure has been described in Chapter 2.9. The lower limit of detection (LLD) and lower limit of quantification were defined as the corresponding concentration of the mean absorbance of the blank standard plus three or seven times the standard deviation and ranged between 66 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 0.005 up to 0.007 ng/mL and 0.007 up to 0.01 ng/mL, respectively. Vtg concentrations were corrected by the total protein concentration in the sample and expressed as ng/mL per mgprotein. 4.3.11 Liver histology Liver sections for histology were prepared and stained by the Immunohistochemistry Facility (IZKF) of Uniklinik RWTH Aachen. 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). Three paraffin sections per individual were mounted on 3-aminopropyltriethoxysilane-coated microscope slides (Sigma-Aldrich, Buchs, Switzerland) and dewaxed. After drying, the slides were strained with Haematoxylin-Eosin (H&E). The sections were examined using light microscopy (Nikon). 4.3.12 RNA extraction, cDNA synthesis and Illumina sequencing Liver samples from three male trout exposed to the treatments of the Luppe sediment, 1:2 dilution of Luppe:sediment control, sediment control, and water-only control were chosen for transcriptome analysis. Frozen liver samples were homogenized in RNAlater using an UltraTurrax homogenizer (IKA, Staufen, Germany) for approx. 10 seconds. Afterward, RNA was isolated using the Maxwell RSC miRNA tissue kit according to the manufacturer’s instructions. RNA samples were assessed for quality (integrity) with the TapeStation 4200 using the Agilent RNA ScreenTape Assay kit (all eRINs were >= 8). Quantification was performed using the Quantus Fluorometer and QuantiFluor RNA Dye (Promega GmbH, Mannheim, Germany). All cDNA libraries were generated from 1 μg total RNA using the TrueSeq Stranded Total RNA Library Preparation Kit with the Ribo-Zero Gold Kit according to the manufacturer´s instructions (Illumina Inc., San Diego, CA, USA). The quality, and quantity of the cDNA libraries were assessed using the 4200 TapeStation (D1000 screen tape assay) and the Quantus Fluorometer, respectively. The libraries were run on an Illumina NextSeq 500 platform using the High Output 150 cycles Kit (paired-end reads, 76 cycles for read 1, 76 cycles for read 2 and 6 cycles for index 1) resulting in a sequencing depth between 41.8 to 52.8 million reads per sample. 67 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.3.13 Statistical analysis and functional annotation of transcriptome results Normality assumptions (Kolmogorov-Smirnov test, Shapiro-Wilk test) were tested prior to parametric analysis. To analyze the significance of variations between experimental data, one-way analysis of variance (ANOVA) was used in combination with Tukey’s post-hoc test. When criteria of normal distribution were not met, data were analyzed using the Kruskal-Wallis test in combination with the Dunn’s post-hoc test. Least squares linear regression was used to evaluate the relationship between concentrations of NP and E1 (E2 and EE2 where applicable) in water, sediment, and rainbow trout bile and vtg. The analysis was done with the open-source program R (R Core Team 2018) and figures were created with GraphPad Prism 6 software (GraphPad Software, San Diego, USA). The statistical significance was determined with a type I error (α) of 0.05. An in-house pipeline embedded in the QuickNGS workflow management system (Wagle et al. 2015) was used to analyze RNA-Seq data. Generation of the fastq files and adapter removal were completed using the Illumina software bcl2fastq (https://support.illumina.com/sequencing/sequencing_software/bcl2fastq-conversion-software.html). Quality control checks of the RNA-Seq data were done with fastqc available online (http://www.bioinformatics.babraham.ac.uk/projects/fastqc). Subsequently, the reads were aligned to the O. mykiss reference genome sequence (Ensembl database version 95) using STAR 2.5.2b (Dobin et al. 2013) with default parameters. In the quantification step, reads were counted with feature Counts (Liao et al. 2014) from the subread package v1.5.1 available online (http://subread.sourceforge.net). Afterwards, differential expression analysis was performed in R using DESeq2 (Love et al. 2014), the local regression parameter, and otherwise default parameters. DESeq2 models generated sequencing reads as negative-binomially distributed and used the Waldtest to determine the significance of coefficients in a negative binomial generalized linear model using previously calculated size factors and dispersion estimates. Benjamini Hochberg correction was applied to account for multiple testing. Differentially expressed hepatic genes in the treatments of both the Luppe and 1:2 dilution in comparison to the water-only or sediment control were analyzed using the Waldtest with a p-value of 0.05 used as the threshold for statistical significance. Venny 2.1 was used to analyze overlaps in differentially expressed genes throughout treatment and controls (Oliveros 2007-2015). Only genes with 68 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout a log2 fold change greater than 2 (or alternatively 1.5) were used to identify the most relevant genes expressed in male trout in response to the exposure to suspended Luppe sediment. Ortholog gene IDs from zebrafish (Danio rerio) were queried from the National Centre for Biotechnology Information (NCBI) genome database and OrthoDB online database (Kriventseva et al. 2019) to identify functional gene ontology (GO) terms for biological processes, molecular function and immune system processes for the encoded proteins in the Cytoscape pathway analysis plugin ClueGO (Bindea et al. 2009; Bindea et al. 2013). GO enrichment analysis was done for differentially expressed genes (DEG) in Cytoscape using the ClueGO plugin. DEG overlapping between Luppe-treated fish and the sediment control in comparison to untreated fish, and DEG only present in Luppe-exposed fish compared to untreated fish were loaded into separate clusters. The GO tree interval ranged from three to eight, and the minimum number of genes per cluster was set to three. P-values were corrected using Bonferroni step-down correction. 4.4 Results 4.4.1 Exposure conditions and bioavailability of EDCs from sediment Exposure conditions were stable throughout the 21 days of exposure and no differences in dissolved oxygen concentration (11 mg/L), temperature (7.7 °C), pH (7.8) and water quality, including nitrate (8.7 mg/L ), nitrite (0.1 mg/L) and ammonium (0.6 mg/L), were observed between the different treatments and controls (see Table A.5). Survival of rainbow trout during the 21-d exposure was > 80 % in controls and treatments. Measured SPM concentration varied in all sediment treatments over time from 0.5 to 2 g/L (Figure 11), and mean concentrations of SPM varied from 0.9 g/L (Luppe treatment) to 2 g/L (sediment control). Both sediments from the Luppe and Rhine ( sediment control) River showed similar characteristics regarding particle size distribution, both characterized as clayey silt, but total organic carbon (TOC) content was slightly higher (9%) in the Luppe sediment compared to the sediment control (5%) (see Table A.2). Sum of PCDD/Fs and dl-PCBs, expressed as WHO 2005 TEQ, as well as concentrations of indicator PCBs, were increased 12-fold and 7-fold in the sediment of the Luppe (109.3 ng TEQ/kg; 158.7 µg/kg) compared to the sediment control (9.7 ng TEQ/kg; 23.9 µg/kg), respectively 69 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout (Table A.3 and A.4). Similarly, concentrations of metals measured in Luppe sediment exceeded those in the sediment control at least by a factor of two (Table A.4). Figure 11: Suspended particles during 21 days of exposure in g/L measured in the different treatments: sediment control, 1:8, 1:4, 1:2; dilution of Luppe sediment: sediment control, LU: Luppe sediment, SC: sediment control. 4.4.1.1 EDCs in sediment Target EDCs (NP, E1, E2 and EE2) were measured in the sediment of all treatments using LC- MS/MS. However, only NP and E1 were detected in sediment samples of all treatments by LC-MS/MS. Concentrations of NP and E1, as well as estrogenic activity expressed as EEQ measured via the YES assay, were highly elevated in Luppe sediment and exceeded those of the sediment control by a factor of 150 (EEQs), 11,050 (NP) and 7 (E1) (Table 6). EEQs and concentrations of E1 and NP in sediment mixtures of Luppe and the sediment control were serially diluted in a 1:2, 1:4, and 1:8 manner, respectively, as confirmed by the YES assay and LC-MS/MS measurements (Table 6). Moreover, a 2 positive relationship (F1,3=345 p=0.0003, r =0.99) between EEQs and concentrations of NP was observed, whereas the relationship between EEQs and concentrations of E1 was not significant (p=0.1). 4.4.1.2 EDCs in aqueous phase Similar to concentrations of EDCs in the sediment, only NP and E1 were detected in Chemcatcher extracts by LC-MS/MS measurement. Using the experimentally determined Rs (Müller et al. 2019), the concentrations in the water of NP and E1 as Cfree were calculated (Table 6). Both NP and E1 in the water of the Luppe treatment exceeded those of the sediment control by a factor of 66 and 15, respectively. In contrast to the corresponding concentrations of NP and E1 in the sediment dilution series, the concentrations of NP and E1 in the water did not proportionally reflect the serial dilution of the sediment 70 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout treatments (1:2; 1:4; 1:8). For instance, Cfree of both NP and E1 measured in the 1:2 dilution of Luppe:sediment control did not differ from concentrations found in the undiluted Luppe treatment. Nevertheless, Cfree of NP and E1 were reduced in the 1:4 and 1:8 treatments (Table 6). E1 was not detected in passive samplers exposed in the water-only control; however, NP was measured at concentrations of 5.6 ng/L, which agrees with reports on NP background levels in other laboratory studies (Loos et al. 2007). Partitioning coefficients (Kp=Csed/Cwat) between sediment and water concentrations normalized to organic carbon fraction (Koc=Kp/foc) were calculated for NP and E1 for each of the sediment treatments (Table 6). Table 6: EDCs were measured from extracts of the passive samplers (Chemcatcher) for the aqueous phase as well as from extracts of the sediment of all treatments using LC-MS/MS. Sediment extracts were additionally analyzed in the Yeast Estrogen Sceen (YES) assay. Two Chemcatcher samplers were included in rainbow trout 21-day exposure experiment with sediments from Luppe and Rhine River (sediment control) and a dilution series 1:2; 1:4; 1:8 of both, and exchanged after 7 days. Concentrations in the water as Cfree were calculated from measured concentrations in the Chemcatchers (n=6) and the respective sampling rate. The fraction of organic carbon (foc) was estimated in the dilutions from parent sediments. Sediment-water partitioning coefficient (Koc) normalized to foc, for E1 and NP were calculated. Concentration in water Concentration in logKoc mean Cfree + SD [ng/L] sediment dry weight Treatment LC-MS/MS foc EEQ E1 NP E1 NP E1 NP [µg/kg] [µg/kg] [mg/kg] LU 6.7 ± 6.1 1045.9 ± 110.5 10.0 22.1 30.4 ± 8.8 0.09 4.2 5.4 1:2 LU:SC 6.5 ± 4.6 925.3 ± 239.3 7.0 8.4 12.1 ± 3.4 0.07 4.2 5.1 1:4 LU:SC 2.8 ± 2.1 748.5 ± 96.4 8.2 4.4 4.8 ± 1.6 0.06 4.7 5.0 1:8 LU:SC 0.9 ± 0.6 592.7 ± 179.8 3.5 0.8 2.9 ± 0.37 0.05 4.9 4.4 SC 0.4 ± 0.1 15.8 ± 12.9 1.4 0.002 0.2 ± 0.03 0.05 4.8 3.4 1.0 4.4.1.3 EDC uptake in rainbow trout Target EDCs were analyzed in plasma and bile samples of trout exposed to Luppe sediment, the 1:2 dilution of Luppe:sediment control, the sediment control, and the water-only control via LC-MS/MS to evaluate uptake and exposure to EDCs (Table 7). Additionally, EDCs and estrogenic activity in bile samples were analyzed using the p-YES and the YES assay. Background concentrations of NP measured with LC-MS/MS in the sediment control ranged from 2 (plasma) to 4 (bile) ng/mL, and thus were >LOD 71 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout due to low detection of the internal deuterated standard. Only NP was detected in both plasma and bile samples by LC-MS/MS measurements. Figure 12: Exemplary results of water-only control (WC), sediment control (SC), 1:2 dilution of Luppe with the sediment control (1:2 LU:SC) and Luppe River sediment (LU ) analyzed by high performance thin-layer chromatography combined with the planar-Yeast Estrogen Screen (p-YES). Different amounts of a mixture consisting of the model compounds estrone (E1, 5 – 100 pg/spot ), ethynylestradiol (EE2, 0.5 – 10 pg/spot), 17β- estradiol (E2, 0.5 – 10 pg/spot), estriol (E3, 0.05 – 1 ng/spot) and nonylphenol (technical mixture; NP, 10 – 100 ng/spot) were separated by a two-step chromatographic development using methanol and chloroform / ethyl acetate / petroleum fraction 55:20:25 (v/v/v). The estrogenic activity was detected after an exposure time of 3 h at 30 °C. The image shows the signal detection with fluorescence-imaging at λexitation = 366 nm. However, in all bile samples, three or four distinct fluorescence signals could be detected after separation by HPTLC followed by subsequent evaluation by the p-YES. The retention times (Rf) of these signals aligned with the standard compounds (Rf NP: 0.819 ± 0.010; E1: 0.681 ± 0.014; EE2: 0.567 ± 0.010; E2: 0.4233 ± 0.0079) (Figure 12). Considering the corresponding Rf, as well as the reporter assay induction, NP, E1, EE2 and E2 were found in all bile samples (Figure 12). Since there was no difference in the concentration of NP, E1, E2, EE2 in bile and plasma detected either via LC-MS/MS or p-YES when examining the influence of fish sex (Kruskal-Wallis; p>0.05), concentrations were pooled for each treatment. Similar mean concentrations of NP in plasma and concentrations of NP, E1, E2, EE2 as well as estrogenic activity in bile of rainbow trout were observed with exposure to the Luppe sediment and the 1:2 dilution treatment (Kruskal-Wallis; p>0.05) (Table 7). Concentrations of NP measured in the plasma of fish exposed to both Luppe treatments either diluted or undiluted significantly exceeded those of the sediment control by a factor of three (Kruskal-Wallis; p<0.0001). Similarly, concentrations of the target EDCs measured via LC-MS/MS or p-YES in fish from both Luppe treatments were 72 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout significantly greater compared to the water-only control, with mean values exceeding those of the water- only control by a factor of 1,000 for NP, 4 for E1, 8 for E2 and 10 for EE2, respectively (Kruskal-Wallis; p<0.0001) (Table 7). The greatest EEQ values were found for rainbow trout exposed to Luppe sediment and were 8.5-fold greater than mean values in fish exposed to the sediment or water-only control. Since no differences were observed between biliary concentrations of NP, E1, E2 or EE2 as well as estrogenic activity in bile between the Luppe and 1:2 dilution treatment, concentrations of NP, E1, E2, EE2 and EEQs were pooled (Kruskal-Wallis; p>0.05). Figure 13: Linear regression of biliary nonylphenol (NP) concentrations [µg/mL] measured by LC-MS/MS and estrogenic activity by in vitro bioassay expressed as 17β-estradiol equivalents (EEQ) in bile of rainbow trout exposed for 21 days to a 1:2 dilution of Luppe:control sediment (orange squares) or the undiluted Luppe sediment (red circle) (the other treatments i.e. 1:4 and 1:8 dilution were not quantified). A positive relationship was observed between EEQs and concentrations of NP, E1, E2 as well as 2 EE2 in bile of fish from the Luppe and 1:2 dilution treatment (NP: F1,23=69.8, p<0.0001, r =0.75; E1: 2 2 2 F1,23=5.8, p<0.02, r =0.20; E2: F1,23=34.8, p<0.0001, r =0.60; EE2: F1,22=32.0, p<0.0001, r =0.59 ) (NP:Figure 13). 73 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Table 7: Concentrations of nonylphenol (NP), 17β-estradiol (E2), 17α-ethynylestradiol (EE2) and estrone (E1) in blood plasma and bile of rainbow trout using LC-MS/MS, p-YES and estrogenic activity measured in the YES assay. Concentrations measured in the p-YES and YES are expressed as E2 equivalents (EEQs) 24 mixed sex trout were exposed for 21-days to sediment suspensions of either Luppe sediment, the sediment control, dilutions of Luppe sediment with the sediment control (1:8; 1:4; 1:2) or the water-only control. The number (n) indicates plasma and bile samples analyzed. Treatment Compound Water-only Sediment 1:8 LU:SC 1:4 LU:SC 1:2 LU:SC Luppe control control LC-MS/MS 34.96 ± 125.53 ± 102.29 ± NP [ng/mL] - 20.17 - - 78.91 59.22 plasma (n=14) (n=16) (n=21) 563.81 ± 984.2 ± 0.52 ± 7.54 ± NP [µg/mL] bile - - 525.1 735.3 0.3 (n=17) 7.3 (n=13) (n=14) (n=12) p-YES E2 5.08 ± 1.9 8.16 ± 3.1 49.98 ± 40.1 40.05 ± 33.2 - - [ngEEQ/mL] bile (n=16) (n=9) (n=13) (n=12) EE2 1.18 ± 1.5 3.29 ± 2.2 - - 11.45 ± 14.2 10.11 ± 8.9 [ngEEQ/mL] bile E1 1.77 ±1.0 2.69 ± 1.4 - - 8.74 ± 10.1 13.83 ± 18.9 [ngEEQ/mL] bile Estrogenic activity YES EEQ [ng/L] REF 6.47 ± 6.43 ± 45.77 ± 55.41 ± - - 0.0006 bile 7.9 (n=17) 4.9 (n=13) 42.2 (n=14) 46.7 (n=12) Concentrations of hydroxy-NP in rainbow trout bile were predicted using the PBTK model based on experimentally determined concentrations of hydroxy-NP in the water. The predicted average concentrations of hydroxy-NP in the bile across treatments were relatively accurate; however, the model was unable to resolve the high variability of up to 90 % standard deviation among individuals (see Figure 14). Figure 14: Relationship between the measured and predicted concentrations of nonylphenol (NP) in bile of rainbow trout (Cbile, n=20-22) exposed for 21 days to suspensions of native sediments from the Luppe River (LU, red dots), a 1:2 dilution of Luppe:sediment control (orange dots), a sediment control from the Rhine River (SC, 74 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout green dots) or fish held under control conditions (water-only control; WC, blue dots). The green line represents the regression line while the shaded area within the grey lines represent the 95% confidence interval. The dashed black line indicates the 1:1 line. Modeled values are based on measured aqueous concentrations of NP and accounted for measured temperature, dissolved oxygen concentration, and the weight of each fish. 4.4.2 Impact of remobilized EDCs form the sediment on rainbow trout 4.4.2.1 Hepatic differentially expressed gene profiles To investigate effects of remobilized, sediment-bound EDCs on the liver transcriptome of male rainbow trout, next-generation sequencing was used to analyze hepatic gene expression profiles in fish exposed for 21 days to suspended sediments of the Luppe River, or the 1:2 dilution of Luppe:sediment control relative to unexposed fish (water-only control) or fish exposed to the sediment control. Of the total 55,525 identified transcripts, fewer than 2 % were significantly differentially expressed (p<0.05) in any of the treatments in comparisons with either of the controls (sediment or water). The majority of transcripts were significantly decreased in abundance in all treatments compared to the unexposed fish (Table 8). In particular, out of the 913 or 215 transcripts of differentially expressed gene (DEG) (p<0.05) identified to be regulated in male trout exposed to suspended sediments of the Luppe River compared to the unexposed fish from the water-only control or fish exposed to the sediment control, respectively, only 17/ 1 and 60/ 0 transcripts were increased or decreased in abundance above a log2 fold change of 2, respectively (Table 8). Similarly, out of the 212 and 140 transcripts found to be differentially regulated in male fish exposed to the 1:2 dilution in relation to unexposed fish or fish exposed to the sediment control, respectively, only 0/4 and 9/ 0 transcripts were increased or decreased in abundance above the threshold of a log2 fold change of 2. Furthermore, 325 DEG in fish were identified by comparing unexposed fish in the water-only control with the fish exposed to the sediment control indicating an altered gene expression in response to suspended sediment exposure with moderate contamination background. Of the DEG identified 6 showed an increase in log2 fold change above 2 whereas 11 showed a decrease exceeding this threshold. 75 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Table 8: Numbers of genes differentially expressed (DEG; p< 0.05) in male rainbow trout (Oncorhynchus mykiss) after exposure to suspended sediment from the Luppe River (LU), a sediment control (SC), a 1:2 dilution of Luppe sediment with the sediment control (1:2 LU:SC) or a water-only control (WC). Vs; versus indicates comparison of two groups. DEG and DEG Comparison DEG log2 fold change >1.5 and log2 fold change >2 - + - + LU vs. WC 913 139 25 60 17 LU vs. SC 215 8 8 0 1 LU vs. 1:2 LU:SC 75 1 3 1 2 1:2 LU:SC vs. WC 212 32 13 9 0 1:2 LU:SC vs. SC 140 4 9 2 4 SC vs. WC 325 85 13 11 6 Figure 15: Venn-diagram (Oliveros 2007-2015) of differentially expressed genes (p<0.05) in male rainbow trout (Oncorhynchus mykiss) exposed to suspended sediments from the Luppe River (LU), a sediment control (SC) and a 1:2 dilution of Luppe with the sediment control (1:2 LU:SC) relative to gene expression in unexposed fish from a water-only control (WC) (left) or relative to the fish exposed to the sediment control (right). Versus: vs. Moreover, 26 DEGs were commonly regulated in fish exposed to Luppe sediment, the 1:2 dilution, as well as the sediment control when compared to unexposed fish from the water-only control (Figure 15). Additional 133 and 26 DEGs of the Luppe treatment and the 1:2 dilution in comparison to the water- only control overlap with transcripts of genes altered in the comparison of the sediment control with the water exposed fish (Figure 15). In contrast, about 6 % (n=67) of the DEGs were commonly regulated in both treatments, the Luppe as well as the 1:2 dilution of Luppe:sediment control, when compared to the water-only control, but were not regulated in the fish exposed to the sediment control (Figure 15). Additionally, exposure to suspended sediments from the Luppe or the 1:2 dilution altered the abundance of 687 and 93 unique transcripts in comparison with the water-only control, respectively. Similarly, the abundance of 176 and 95 transcripts were uniquely altered in treatments of the Luppe or the 1:2 dilution compared to the sediment control (Figure 15). 76 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Orthologous genes of zebrafish were used for annotation of functional gene ontology (GO) terms. Pathway enrichment analysis within the Cytoscape plugin ClueGo and biological processes associated with hepatic genes differentially regulated above a log2 fold change of 2 are indicated within Table 9 and Table 10, respectively. Transcripts of genes (p<0.05 and log2 fold change of 2 in at least one comparison) commonly regulated in male fish exposed to sediment from the Luppe, the 1:2 dilution as well as the sediment control compared to the water-only control were associated with energy metabolism (gatm ↓, gltp ↓), immune response (interferon-induced GTP-binding protein Mx3 ↓, ccr9a;ccr9b ↑), DNA binding (zinc finger protein 239-like ↓), Hsp70 chaperone machinery (sacs ↓) and protein-protein interaction (ankrd37 ↑) (Table 9). Exposure to Luppe sediment as well as the sediment control relative to the water-only control was associated with the enrichment of cell cycle process (p<0.001) including DNA packing (p = 0.03) and mitotic cytokinesis (p<0.001), where the abundance of transcripts was significantly decreased (↓: ccna2, ccnb2, ccnb3, prr11, kmt5ab, tubb1, esco2, ect2) (Table 9; Figure 16). 77 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Table 9: Significantly differentially expressed (p<0.05 and log2 fold change of 2 in at least one comparison) genes in male rainbow trout (n=3) as a response to sediment exposures. Genes were significantly regulated in fish either exposed to sediments from the Luppe River (LU), the sediment control (SC), or a 1:2 dilution of Luppe with the sediment control (1:2 LU:SC) compared to unexposed fish (water-only control). - 3.5 - 3 - 2.5 - 2 - 1.5 0 1.5 2 2.5 3 3.5 log fold change Ortholog 2 1:2 Gene ID Gene description gene symbol LU LU: SC Danio rerio SC metabolic processes LOC110513671 glycine amidinotransferase mitochondrial-like gatm ↓ gltp glycolipid transfer protein gltpa; gltpb ↓ immune response si:dkeyp- LOC110494493 interferon-induced GTP-binding protein Mx3 ↓ 110c12.3 LOC110529980 C-C chemokine receptor type 9-like ccr9a; ccr9b ↑ LOC110519923 neural cell adhesion molecule 1-like ncam1a ↓ DNA-/ metal ion binding LOC110492308 tripartite motif-containing protein 16-like trim16 ↓ LOC110528540 TOX high mobility group box family member 2-like tox2 ↑ LOC110528938 zinc finger protein 239-like ↓ DNA repair LOC110493570 ATPase family AAA domain-containing protein 5-like atad5a ↓ DNA packing LOC110500154 condensin-2 complex subunit H2-like ncaph2 ↓ LOC110493456 centromere protein M-like pane1 ↓ LOC110498601 structural maintenance of chromosomes protein 2-like smc2 ↓ cell cycle process LOC110505148 cyclin-A2-like ccna2 ↓ LOC100136691 cyclin B2 ccnb2 ↓ LOC110510157 G2/mitotic-specific cyclin-B3-like ccnb3 ↓ LOC110533544 proline-rich protein 11-like prr11 ↓ LOC110525699 N-lysine methyltransferase KMT5A-like kmt5ab ↓ LOC110535859 tubulin beta-1 chain tubb1 ↓ LOC110529614 N-acetyltransferase ESCO2-like esco2 ↓ LOC110522232 N-acetyltransferase ESCO2-like esco2 ↓ LOC110524785 abnormal spindle-like microcephaly-associated protein homolog aspm ↓ si:ch211- LOC110525857 cell division cycle-associated protein 2-like ↓ 244o22.2 LOC110529408 serine/threonine-protein kinase Nek2-like nek2 ↓ LOC110527904 protein FAM83D-like fam83D ↓ spc25 SPC25 component of NDC80 kinetochore complex spc25 ↓ LOC110505614 transforming acidic coiled-coil-containing protein 3-like tacc3 ↓ LOC110518846 dual specificity protein kinase Ttk-like ttk ↓ mitotic cytokinesis ect2 epithelial cell transforming sequence 2 ect2 ↓ LOC110494113 citron Rho-interacting kinase-like cita ↓ LOC110492103 rac GTPase-activating protein 1-like racgap1 ↓ Hsp70 chaperone machinery LOC110505631 sacsin-like sacs ↓ protein-protein interaction ankrd37 Ankyrin repeat domain-containing protein 37 ankrd37 ↑ txlnba; LOC110497567 beta-taxilin-like ↑ txlnbb 78 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Only five transcripts of genes were identified to be commonly regulated (p<0.05) in both treatments, Luppe sediment and the 1:2 dilution, either compared to water or sediment control, among which only the abundance of vitelline envelope protein α transcripts was increased above a log2 fold change of 2 (1.6 - 2.4) followed by cyp1a with log2fold change induction of 1.4 -1.8 (Table 10). Exposure to suspended sediments from the Luppe resulted in an estrogenic response and further increased abundance of vitelline envelope protein γ and vitellogenin-like transcripts (Table 10). Moreover, DEG induced in both treatments compared to either of the controls (sediment or water) were associated with response to xenobiotic stimulus (↑: eef2b, cyp1a, gadd45ga, nfe2l2a) or mitochondrial expression (↑: slc25a55a), whereas, DEG deceased in abundance were related to response to hormone (↓: rerg, dio2), DNA-/RNA binding activity (↓: helz2, zbtb16a) integral components of membranes (↓: cmtm8a; cmtm8b, pigr) and proteolysis (↓: st14a, psma2). Exposure to Luppe sediment relative to the water-only control was further associated with enrichment of cell cycle process (p <0.001), where the abundance of all DEG were significantly decreased (↓: tpx2, numa1, bub1, sgo1, ddx4, fam83d, kif20a, cep55, cks1b, knl1, kif11, cenpn, prc1a, espl1, incenp) (Figure 16) except for cyclin-dependent kinase inhibitor 1B-like ↑, which was also increased in abundance in the 1:2 dilution (Table 10). Greatest induction of 5.8 log2 fold change was observed in the Luppe exposed fish relative to the water-only control for transcripts of the transmembrane protein 163-like, which is involved in the tolerance to metal ions. 79 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Table 10: Significantly differentially expressed (p<0.05 and log2 fold change of 2 in at least one comparison) genes in 21-d exposed male rainbow trout (n=3). Genes were significantly regulated in fish either exposed to sediments from the Luppe River (LU) or a 1:2 dilution of Luppe with the sediment control (1:2 LU:SC) compared to unexposed fish (water-only control; WC) or compared to fish exposed to the sediment control (SC). - 3.5 - 3 - 2.5 - 2 - 1.5 0 1.5 2 3 4 5 6 log2 fold change Ortholog gene vs. WC vs. SC Gene ID Gene description symbol Danio 1:2 1:2 rerio LU LU: LU LU: SC SC egg coat formation/ response to estradiol LOC100135906 vitelline envelope protein α zp2/2 ↑ LOC100135907 vitelline envelope protein γ zp3b ↑ LOC110521946 vitellogenin-like vtg1 ↑ response to xenobiotic stimulus LOC110536498 elongation factor 2-like eef2b ↑ LOC110526570 cytochrome P450 1A1 cyp1a ↑ growth arrest and DNA damage-inducible protein ↑ LOC110536341 GADD45 γ-like gadd45ga LOC110527610 nuclear factor erythroid 2-related factor 2-like nfe2l2a ↑ response to hormone ras-related and estrogen-regulated growth ↓ LOC110500247 inhibitor-like rerg dio2 iodothyronine deiodinase 2 dio2 ↓ LOC110505893 type II iodothyronine deiodinase-like dio2 ↓ cell cycle process LOC110532132 targeting protein for Xklp2-like tpx2 ↓ LOC110503183 nuclear mitotic apparatus protein 1-like numa1 ↓ mitotic checkpoint serine/threonine-protein kinase ↓ LOC110522100 BUB1-like bub1 LOC110496596 shugoshin 1-like sgo1 ↓ ddx4 DEAD-box helicase 4 ddx4 ↓ LOC110494562 protein FAM83D-like fam83d ↓ LOC110534112 kinesin-like protein KIF20A kif20a ↓ centrosomal protein of 55 kDa-like transcript ↓ LOC110491408 variant X3 cep55 LOC110521455 cyclin-dependent kinases regulatory subunit 1-like cks1b ↓ knl1 kinetochore scaffold 1 knl1 ↓ LOC110502544 kinesin-like protein KIF11 kif11 ↓ cenpn centromere protein N cenpn ↓ LOC110523805 protein regulator of cytokinesis 1-like prc1a ↓ espl1 extra spindle pole bodies like 1, separase espl1 ↓ LOC110501272 inner centromere protein A-like incenp ↓ LOC110500295 cyclin-dependent kinase inhibitor 1B-like cdkn1ba; cdkn1a ↑ metal ion binding ArfGAP with SH3 domain, ankyrin repeat and PH ↓ asap2 domain 2 asap2a; asap2b LOC110518405 E3 ubiquitin/ISG15 ligase TRIM25-like LOC570079 ↓ fih1 factor inhibiting hypoxia-inducible factor 1 α hif1an ↑ LOC110492610 carbonic anhydrase 6-like ca6 ↑ tmem163a; ↑ LOC110527562 transmembrane protein 163-like tmem163b inflammatory response LOC110494052 ETS-related transcription factor Elf-3-like elf3 ↓ LOC110530070 B-cell CLL/lymphoma 6 member B protein-like bcl6a ↑ LOC110509901 eosinophil peroxidase-like mpx ↑ 80 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout mitochondrial ATPase family AAA domain-containing protein ↓ LOC110532364 3-like atad3 cytochrome c oxidase assembly factor 4 homolog ↑ LOC110533269 mitochondrial si:ch211-215a9.5 mitochondrial glutamate carrier 1-like transcript ↑ LOC110526530 variant X3 slc25a55a proteolysis LOC110538136 suppressor of tumorigenicity 14 protein homolog st14a ↓ LOC110520866 proteasome subunit α type-2-like psma2 ↓ integral component of membrane CKLF-like MARVEL transmembrane domain- ↓ LOC110506736 containing protein 8 cmtm8a; cmtm8b LOC110508040 transmembrane protein 244-like tmem244 ↓ LOC110531070 polymeric immunoglobulin receptor-like pigr ↓ signalling LOC110509650 neuropeptide Y receptor type 1-like npy1r ↑ dlgap5 DLG associated protein 5 dlgap5 ↓ protein phosphorylation; phosphorylation lymphokine-activated killer T-cell-originated ↓ LOC110522233 protein kinase homolog transcript variant X1 pbk lipid metabolic process patatin-like phospholipase domain-containing ↑ LOC110500176 protein 2 pnpla2 negative regulation of circadian rhythm LOC110522462 CLOCK-interacting pacemaker-like cipa ↑ DNA-, RNA binding helz2 helicase with zinc finger 2 helz2 ↓ zinc finger and BTB domain-containing protein ↓ LOC110537609 16-A zbtb16a cell motility iqgap3 IQ motif containing GTPase activating protein 3 iqgap3 ↓ unknown LOC110517725 putative nuclease HARBI1 LOC101884462 ↑ cellular amino acid biosynthetic process branched-chain-amino-acid aminotransferase ↑ LOC110538599 cytosolic-like bcat1 tissue development LOC110531712 netrin-1-like ntn1 ↑ 81 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Figure 16: Pathway enrichment network of significant gene ontology (GO) terms of differentially expressed hepatic genes in male rainbow trout following exposure to suspended sediment from the sediment control as well as the Luppe River (colored in green; cluster 1) when compared to unexposed fish or solely regulated in Luppe exposed fish (colored in red; cluster 2). The size of the circle indicates the significance of that node, with larger circles being more significant. Node color intensity indicates the proportion of genes from each cluster that are associated with the term (darker intensity represents greater association). 4.4.2.2 Vtg induction Vtg was measured in fish mucus prior to exposure and after the 21-d exposure as a biomarker of exposure to EDCs (Kroon et al. 2017). Prior to exposure, vtg (0.98 ± 0.61 ng/mL per mgprotein) was only above the LLD in 27 out of total 144 juvenile rainbow trout, with vtg ranging in those 27 fish from 0.39 up to 3.38 ng/mL per mgprotein. When comparing vtg concentrations before and after exposure, vtg was only significantly induced in 9 out of the 21 surviving fish exposed to the Luppe sediment (6.1 ± 7.8 82 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout ng/mL per mgprotein; Kruskal-Wallis; p=0.0001) (Figure 17), where concentrations of vtg in male (n=5) and female (n=4) trout were 3.5 and 9.5 times greater, respectively. Moreover, there was no difference in the concentration of vtg when examining the influence of fish sex in the Luppe treatment (Mann Whitney; p=1). In contrast, concentrations of vtg in fish exposed for 21 days in the water and sediment control were below the LLD. Likewise, vtg was detected above the LLQ but did not differ from mean vtg concentration measured prior to exposure in fish exposed to the 1:8 (n=2), 1:4 (n=2) and 1:2 (n=6) dilution of Luppe:sediment control (Kruskal-Wallis; p>0.05) (Table 11 and Figure 17). Figure 17: Vitellogenin (vtg) concentration [ng/mL per mgprotein] in mucus samples of mixed-sex rainbow trout after 21 days of exposure to suspension of Luppe sediment (LU) as well as 1:8, 1:4, and 1:2 dilution of Luppe sediment with the sediment control (LU:SC). The blue line indicates the mean concentration of vtg measured prior to exposure. Number of analyzed fish is indicated above bars (n=2-9). All other measured values were below the LLD. Kruskal-Wallis analysis was used to identify statistically significant differences in combination with Dunn’s multiple comparison test and are indicated by asterisks (***). A positive relationship between mean vtg concentrations pooled among male and female trout and 2 the estrogenic activity (EEQs) (F1,3=86, p=0.003, r =0.97) (Figure 18) as well as concentrations of NP 2 (F1,3=77, p=0.003, r =0.96) measured in the sediment of the different treatments was observed. Moreover, a weaker relationship existed between mean concentrations of vtg and concentrations of NP 2 measured in the water of the different treatments (F1,4=9, p=0.04, r =0.69). 83 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Table 11: Summary of general health parameters including Fulton’s condition index (K), gonadosomatic index (GSI) and liver somatic index (LSI), concentrations of vitellogenin (vtg) measured in mucus by ELISA of rainbow trout. 24 fish were exposed for 21-days to sediment suspensions of either Luppe sediment, the sediment control, dilutions of Luppe sediment with the sediment control (1:8; 1:4; 1:2) or the water-only control. The number (n) of fish after the experiment is indicated as the total number after exposure. Treatment Mean ± SD Water-only Sediment 1:8 LU:SC 1:4 LU:SC 1:2 LU:SC Luppe control control Total number after 23 20 22 23 24 21 exposure (n) Male (n) 11 10 11 9 8 12 Female (n) 12 10 11 14 16 9 Fork length [cm] 12.3 ± 1.5 11.5 ± 0.9 12.1 ± 1 .4 11.6 ± 1.2 11.6 ± 0.8 11.8 ± 1.2 Weight [g] 22.2 ± 8.1 17.7 ± 3.4 20.3 ± 7.1 16.8 ± 5.1 18.1 ± 3.8 17.5 ± 5.0 Fulton´s condition 1.1 ± 0.1 1.2 ± 0.1 1.1 ± 0.1 1.1 ± 0.1 1.1 ± 0.1 1.0 ± 0.1 index (K) GSI 0.1 ± 0.1 0.2 ± 0.2 0.2 ± 0.1 0.2 ± 0.2 0.2 ± 0.2 0.2 ± 0.1 LSI 1.2 ± 0.3 1.2 ± 0.3 1.3 ± 0.3 1.2 ± 0.3 1.2 ± 0.2 1.0 ± 0.2 Vtg concentration [ng/mL per mgprotein] Vtg pooled across 0.98 ± 0.61 treatments day 0 (n=27) Vtg day 21 0.45 ± 0.19 1.4 ± 0.3 3.5 ± 4.3 6.1 ± 7.8 Figure 18: Linear regression of mean concentrations of vitellogenin (vtg) in the mucus of 21-day exposed rainbow (LU: n=9; 1:2 LU:SC: n=6; 1:4 LU:SC: n=2; 1:8 LU:SC: n=2; SC: set to lower limit of detection) trout measured by ELISA and estrogenic activity detected in Luppe sediment (LU), control sediment (SC) and 1:8, 1:4, 1:2 dilution of Luppe:sediment control by in vitro bioassay expressed as 17β-estradiol equivalents (EEQ). 84 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.4.2.3 Histology The liver parenchyma of unexposed trout (water-only control) was normal, with a dense network of sinusoids extended between columns of hepatocytes. No external lesions, inflammatory reaction or necrotic foci were observed in liver of fish from the water-only control (Figure 19 A). Energy reserves in form of lipid droplets and glycogen deposits were generally low but varied between individual fish. In contrast, histological alterations were observed in liver from trout exposed to sediments from the Luppe. The structure of the liver parenchyma of those fish was partly disorganized, with local areas of cell lysis (Figure 19 D). Figure 19: Histological alterations in liver of rainbow trout (Oncorhynchus mykiss) of unexposed fish (A: water- only control) compared to fish exposed for 21 days to suspended sediments (5 g/L dry weight) from the Rhine River as sediment control (B), a dilution of this sediment control with a highly EDC contaminated sediment from the Luppe River in a 1:8 manner (C), and the undiluted sediment from the Luppe River (D). Liver sections were stained with hematoxylin and eosin (H&E). Moreover, infiltration of immune cells within the parenchyma was observed. There were small areas where hepatocytes were enlarged and the cytoplasm became more eosinophilic. Several hepatocytes displayed degenerative nuclear alterations, including hyperchromatic changes and irregular nuclear contour. In a few cases, “giant cells” containing multiple nuclei were present. The liver histological 85 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout alterations were found in similar abundance and severity with all Luppe sediment dilutions (1:2; 1:4; 1:8) as well as the sediment control (Figure 19 B-D). 4.4.2.4 General fish health indices Somatic indices including LSI, GSI and Fulton’s conditions K index were analyzed to assess effects on the reproductive and overall fitness of rainbow trout (Kroon et al. 2017) (Table 11). In this study, fish showed an overall good condition as indicated by Fulton’s condition K index above 1.0. However, K was significantly lower in fish exposed to sediment from the Luppe River compared to fish exposed to the water-only control, sediment control and the 1:2 dilution treatment (ANOVA: p=0.003) (Table 11). In contrast, no differences in LSI among treatments and control fish were observed (ANOVA: p=0.23). GSI was approx. 10-times higher in female fish compared to males (Mann Whitney: p=0.0001), but no differences in GSI of male fish between treatments and control were observed. In contrast, GSI was significantly reduced in female fish of the water-only control compared to the sediment control and the 1:4 dilution (Kruskal-Wallis: p=0.01), but this is consistent with ranges of natural variations. 4.5 Discussion 4.5.1 Bioavailability of sediment-bound EDCs One main objective of the present study was to evaluate the bioavailability of sediment-bound EDCs under flood-like conditions towards fish. We used a passive sampling approach to demonstrate that sediment-bound EDCs were bioavailable in terms of concentrations of freely dissolved EDCs in the water phase during the resuspension of the sediment. Moreover, EDCs were readily taken up and metabolized by the fish as indicated by concentration of NP, E1, E2 and EE2 in plasma and bile of fish exposed to the Luppe sediment or its 1:2 dilution with the sediment control. 4.5.1.1 Estrogenic activity in sediments While estrogenic activity measured in sediment of the Luppe River in the present study exceeded that in sediment samples of major German streams including the upper Rhine (5.1 µg EEQ/kg; 4.1 mg/kg NP; Brauch et al. 2001; Schulze-Sylvester et al. 2016), upper Danube (1.3 µg EEQ/kg; 1.4 -9.9 mg/kg NP, 0.2 µg/kg E1; Brauch et al. 2001; Grund et al. 2010b) and Elbe River (1.9 µg EEEQ/kg; 2.2 mg/kg NP; Schmitt et al. 2012) by at least a factor 6, these were within the range of the 1:4 and 1:8 dilution of 86 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Luppe:sediment control of the present study (Table 6). Estrogenic activity was detected using in vitro bioassays, e.g. YES, in sediments of urban, rural and industrialized areas worldwide ranging from 0.03 up to up to 101 µg EEQ/kg (Céspedes et al. 2004; Grund et al. 2010b; Hilscherova et al. 2002; Thomas et al. 2004; Viganò et al. 2008; Zhao et al. 2011), mostly identifying NP and other alkylphenols in combination with E1, E2 and BPA as active compounds. Similarly high EEQs as detected in sediments from the Luppe River, as well as concentrations of EDCs in sediment, were reported for, e.g., the Po River, Italy, (15.6 µg EEQ/kg; Viganò et al. 2008), the Tees estuary in the UK (13 µg EEQ/kg; Thomas et al. 2004), or the Pearl River system, China (up to 101 µg EEQ/kg; Zhao et al. 2011). 4.5.1.2 Partitioning of EDCs between sediment-water interphase during flood-like conditions Partitioning of EDCs between sediment or SPM interfaces and water strongly depends on organic carbon content, where sorption increases with TOC content (Li et al. 2011; Navarro et al. 2009). Sediment water distribution coefficients normalized to organic carbon content (Koc) determined for NP and E1 in the present study for the Luppe sediment and the sediment control (Table 6) agree with reported literature Koc values ranging from 3.6 up to 5.39 for NP and from 4.08 up to 7.09 for E1, respectively (Ding et al. 2019; Ferguson et al. 2001; Ma and Yates 2018). For example, Ferguson et al. (2001) reported a Koc of 5.39 based on the measured concentration of NP in water and SPM in the Jamaica Bay, US, where concentrations of NP ranged from 0.05 up to 30 mg/kg in sediment. Concentrations of NP as well as E1 measured in water in the present study did not proportionately correspond with the serial dilution in the sediment, which might be explained by the short aging period (5 days) (Table 6). Five days may not allow for full equilibration of EDCs in the mixture of Luppe and control sediment. A concentration-dependent partitioning between sediment and water phase for both chemicals would be expected when NP and E1 would have been equally distributed within the mixture of Luppe to control sediment (Schwarzenbach et al. 2002). Although major fractions of NP and E1 were associated with SPM during sediment suspension as indicated by high Koc values, concentrations of NP, for example, were 14 times higher in the water phase during suspension compared to measured concentration in field collected water samples from the Luppe River in a previous study (Müller et al. 2019) and exceeded the environmental quality standard (EQS) of 0.3 µg/L (EU 2008/105) by a factor of 3. Similar results were reported for desorption of NP in 87 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout laboratory resuspension studies with sediment from the Huerva River, Spain (20 mg/kg NP) and the Chaohu Lake, China (0.1 mg/kg NP) (Ding et al. 2019; Weert et al. 2010). Weert et al. (2010), who investigated desorption of NP from resuspended sediment compared to settled sediment under laboratory conditions, found that the mass transfer of NP from the sediment into the water phase was stable around 0.3 µg/L under static conditions, whereas upon resuspension NP concentrations in the water rapidly increased by a factor of 100 up to 12.7 µg/L. While NP contamination in sediment of the Huerva River is comparable to concentrations of NP measured in the sediment from the Luppe River, TOC was three times lower compared to the Luppe sediment, and desorption of NP during resuspension of the Luppe sediment resulted in a slightly lower water concentration around 1 µg/L (Table 6). Weert et al. (2010) concluded that the increased desorption of NP from the sediment resulted from the increased exchange surface area during resuspension, which is supported by the results of the present study. This further agrees with reports of Li et al. (2013), who found that concentrations of NP increased in the water phase by approx. a factor of 3, up to 0.6 µg/L, during a flood event at the Daliao River estuary in China and were positively related to measured SPM concentrations ranging from 0.1 up to 0.4 g/L. SPM measured during the 21 days of exposure in the present study were, on average, higher and ranged from 0.5 up to 2 g/L, but the nominal concentration of 5 g/L was not reached. While SPM concentrations above 5 g/L have been reported for a few flood events worldwide, e.g., at the Yellow River in China (28 g/L; Xia et al. 2006), SPM concentrations reported in the environment as a result of flood events by other studies were lower or similar to those measured in the present study (Li et al. 2013; LHW 2014). For example, SPM concentration ranging from 0.01 up to 0.2 g/L were observed during a flood in 2013 at the Elbe River (LHW 2014). Considering the abovementioned studies and results from the present study, concentrations of EDCs in the water phase are concluded to increase during suspension of sediment either under laboratory conditions or during a natural flood and, thus, indicate that “dormant” sediment-bound contaminants can impact water quality and contaminant mobility. A question addressed in future research is how sediment characteristics influence desorption rates and whether Koc, amount of suspended SPM and thereby surface area exchange capacity for partitioning, or other parameters such as composition of organic matter (e.g. aromaticity) are the major drivers of desorption of EDCs from sediment during suspension. 88 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.5.1.3 Uptake routes of sediment-bound EDCs into rainbow trout Results from the present study demonstrated that EDCs became bioavailable for fish during sediment suspension as indicated by the concentrations of NP, E1, E2 and EE2 measured as uptake in the blood or bile of rainbow trout after exposure to suspended sediments and sediment dilutions from the Luppe River. Concentrations of EDCs, i.e., NP, in plasma and bile measured in the present study in exposed trout (Table 7) are consistent with literature reports on NP tissue distribution and biotransformation in fish. NP was rapidly metabolized in the liver and the hydroxylated NP glucuronides were excreted through the bile and faces (Cravedi et al. 2001; Lewis and Lech 1996; Thibaut et al. 2000). In a previous study where rainbow trout were exposed to suspended sediment contaminated with PAHs, dioxins, and dl-PCBs, Hudjetz et al. (2014) observed accumulation of sediment in the intestine of the fish suggesting an uptake of contaminants through absorption of ingested SPM (Brinkmann et al. 2010; Hudjetz et al. 2014). In contrast, concentrations of EDCs remobilized in the aqueous phase are likely the main route of exposure in the present study as similar patterns were observed between aqueous exposure and uptake in the fish. Specifically, concentration of NP, E1, E2 and EE2 as well as estrogenic activity measured in sediment of the 1:2 dilution of Luppe:sediment control were reduced by a factor of 2 compared to undiluted sediments from the Luppe River, while concentrations of EDCs in water did not differ between the two treatments similar to the concentrations of NP, E1, E2 and EE2 and estrogenic activity measured in bile of exposed fish that also did not differ between the two treatments. Furthermore, the accurate fit of biliary NP concentrations modeled based on water concentrations assessed by passive sampling and experimentally determined concentrations of NP in the bile of exposed trout offers further support that aqueous concentrations of EDCs may be one of the dominant routes of exposure (Figure 14). Passive sampling of EDCs has been reported previously by Vermeirssen et al. (2005) to be a good predictor for biliary concentrations in caged trout and was a useful tool to assess the bioavailability of EDCs in the present study in a sediment-water suspension system. Thus, the combination of passive sampling and PBTK modeling might be a helpful tool used in future risk assessment for sediment-bound EDCs with respect to their potential for remobilization during flood events as demanded in the European parliament´s directive 2007/69/EC regarding the assessment and management of flood risks (EU 2007/60). 89 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.5.2 Impact of sediment-bound EDCs on rainbow trout during a simulated flood event Our study demonstrates that exposure of suspended sediment contaminated with EDCs resulted in an altered hepatic gene expression profile, including the induction of estrogen-responsive genes. An estrogenic response in male fish exposed to these sediments was further observed by a slight induction of vtg on the protein level. Few studies have utilized RNA sequencing in the context of environmental mixtures of chemicals, where it might be difficult to assign causal effect relationships (Ings et al. 2011; Mehinto et al. 2012). However, as investigated with suspended sediments in the present study, it was observed that the liver functions as a major site for biotransformation of xenobiotics, including NP, which is supported by the measured biliary concentrations of NP in fish exposed to both the diluted and undiluted Luppe treatments. 4.5.2.1 EDC responsive hepatic gene expression in rainbow trout Induction of typical estrogenic responsive genes such as vitelline envelope protein α and vitellogenin- like in the liver of male fish exposed to suspended sediments from the Luppe, either 1:2 diluted with sediment control or undiluted, clearly demonstrate that exposure to remobilized sediment-bound EDCs resulted in an estrogenic response on gene level (Table 10). Induction of those genes have been identified in several exposure studies as response to single substances including E2, EE2, E1 and NP mediated through estrogen receptor (ER) activation (Benninghoff and Williams 2008; Feswick et al. 2017; Gunnarsson et al. 2007; Hook et al. 2008; Huff et al. 2019; Hultman et al. 2015; Osachoff et al. 2016; Shelley et al. 2012). Among all analyzed differentially regulated transcripts vitelline envelope protein α was solely identified to be significantly induced in male trout above a log2 fold change of 2 in relation to both treatments (Luppe and 1:2 dilution). Additionally, the abundance of transcripts for vitellogenin- like was increased (log2 fold change of 2.1) in male fish exposed to suspended sediments from the Luppe in comparison to the sediment control. While other genes known to be estrogen regulated such as estrogen receptor α (esr1), vitelline envelope protein γ or zona radiata protein (zp2.3) were differentially regulated (p<0.05) in response to suspended sediment of the Luppe, induction was lower and ranged around a log2 fold change of 1. Induction of vitelline envelope protein or zona radiata protein either on the gene or protein level has been controversially discussed in the literature to be a more sensitive biomarker for EDC exposure compared to vitellogenin (Arukwe et al. 1997; Gunnarsson et al. 2007; 90 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout Huff et al. 2019; Osachoff et al. 2016). Both, Gunnarsson et al. (2007) and Osachoff et al. (2016), reported that zona radiata protein 3 (zp3) or vitelline envelope proteins α, β, γ were induced in the liver of juvenile rainbow trout at lower exposure concentrations of 0.9 or 24 ng/L EE2 or E1, respectively, or on longer time frames compared to conditions required to induce vtg. The fact that transcripts of vitelline envelope protein α were more frequently increased in abundance in relation to the Luppe and 1:2 dilution treatment compared to vitellogenin-like in the present study, where concentrations of E1 and possibly EE2 in the water were similar or lower, supports the findings from the literature. 4.5.2.2 Altered hepatic gene expression profiles in response to sediment exposure Besides the typical set of estrogen responsive genes involved in vitellogenesis (vtg, zp, esr1, vitelline envelope proteins), other studies identified hepatic gene expression profiles in rainbow trout upon exposure to E2, EE2 or NP (Hook et al. 2008; Huff et al. 2019; Shelley et al. 2012). Those were associated with GO terms of transport, cytokinesis, inflammatory response, blood coagulation and immune response (Hook et al. 2008). Additionally, Benninghoff and Williams (2008) and Shelley et al. (2012) observed expression profiles related to stress, including cell cycle, growth and proliferation, lipoprotein biosynthesis, DNA repair and damage response. Several GO terms of DEG related to suspended sediment of the Luppe agree with the literature (Benninghoff and Williams 2008; Hook et al. 2008; Shelley et al. 2012) and were associated with an inflammatory response and lipid metabolic processes as well as cell cycle processes which might indicate a broader estrogenic response in liver of Luppe exposed fish (Table 10). A recent study by Huff et al. (2019), who investigated the non-estrogenic signature of NP on hepatic transcriptome of zebrafish, found that while NP and E2 exposure commonly dysregulated hepatic expression of genes involved in vitellogenesis, metabolic pathways, as well as cellular response, NP uniquely enriched GO terms of DEG associated with immune and inflammatory pathways, the cell´s response to DNA damage, and cell cycle. Enrichment of GO terms of DEGs was closely linked to an enrichment in the metabolism of reactive oxidative species (ROS) (Huff et al. 2019). Besides its endocrine disruptive properties NP also exhibit genotoxic characteristics (Frassinetti et al. 2011); such that ROS formation as well as DNA-damage associated with inflammation and cell death have been observed in fish in a concentration-dependent manner after exposure to NP possibly as a result of biotransformation processes (Abd-Elkareem et al. 2018; Sharma and Chadha 2017; Xu et al. 91 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 2013). Results of Huff et al. (2019) regarding genes involved in inflammatory responses as well as literature on hepatotoxicity of NP agree with regulated transcripts related to inflammatory responses in the present study, where ETS-related transcription factor Elf-3-like involved in epithelial cell differentiation was suppressed. In contrast, the abundance of transcripts of eosinophil peroxidase-like, which are further associated with the response to oxidative stress and the regulation of hydrogen peroxide metabolic process, were increased in fish exposed to suspended sediment from the Luppe River (Table 10). Further, unique for NP exposure was an enrichment in metabolic fatty acid pathways (Huff et al. 2019) corresponding to the induction of patatin-like phospholipase domain-containing protein 2, positively regulating triglyceride catabolic process, in fish exposed to suspended sediment from the Luppe River. Estrogens exhibit carcinogenic characteristics in vertebrates, including fish, mediated through oxidative quinone metabolite formation, ER signaling pathways associated with the inhibition of apoptosis, and increased cell proliferation resulting in tumor induction and formation (Lam et al. 2011). Using transcriptome analysis, Lam and co-workers identified conserved signaling pathways and genes involved in E2 induced carcinogenesis in zebrafish males and human cancer cells. In particular, signaling pathways involved mainly in cell cycle progression as well as DNA damage and repair such as mitotic checkpoint, in which cell cycle-related genes (e.g., human homolog E2F4, CDK2, CCNA, CCNE, CDC6, PLK2, CDC23, CCNB1, PRC1) were all up-regulated (Lam et al. 2011). This is in contrast to observed decreased abundance of transcripts related to cell cycle (e.g., ccna2, ccnb2, ccnb3, esco2, ttk, prc1a, bub1) in the present study, which was not restricted to Luppe exposed fish but was similarly evident in fish exposed to the sediment control when compared to water-only unexposed fish (Figure 16). Central cell cycle proteins such as cyclin A and B as well as cyclin-dependent kinases including G1/S (ccna2, e2f1) and G2/M (ccnb3, prc1, prr1) and spindle assembly checkpoints (esco2, nek2, aurkb) were suppressed in addition to many other mitotic genes indicating a cell cycle arrest at several stages (Table 9 and Table 10). This effect was even more pronounced in Luppe exposed fish, in which an additional large set of cell cycle genes (e.g., bub1, sgo1, fam83d, kif20a, smc4, cdk1, cdk2, aurkb) were uniquely suppressed compared to unexposed fish. Transcriptional down-regulation of many genes essential for progression through the cell cycle has been described in the human context by 92 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout activation of the p53-DREAM pathway (Engeland 2018; Uxa et al. 2019). Although p53 was characterized in rainbow trout, the existence of the p53-DREAM pathway was controversially discussed in the context as an apoptosis inducer in fish and might include alternative mechanisms compared to humans (Liu et al. 2011; Mai et al. 2012). Upon p53 activation the DREAM complex acts as a transcriptional repressor and, thereby, depletes a broad range of regulatory proteins required for cell cycle progression (Engeland 2018). Top 116 genes were identified as p53 targets in multiple genome- wide data sets in the literature including key examples for repressed genes such as CCNA, CCNB1, CCNB2, PLK1, CDK1, CDC20 (Engeland 2018; Fischer 2017). Many orthologous genes were found in the present study in exposed trout (Table 9; Table 10). Interestingly, this pathway requires transcriptional up-regulation of p21/CDKN1A ortholog to cyclin-dependent kinase inhibitor 1B-like in trout which was solely up-regulated among the cell cycle-related genes in Luppe and 1:2 dilution exposed trout in the present study (Table 10). P53 is a key mediator of cell cycle arrest and apoptosis induction in response to genotoxic stress. Global repression of cell cycle genes in control or EDC treatment indicates that sediment exposure itself induces stress on a cellular level. However, with this effect being more pronounced in relation to the Luppe treatment, exposure xenobiotics such as NP or E1 might have enhanced genotoxic cell stress in these fish. Pro-apoptotic transcripts of genes e.g. bcl2-like protein, apopt1, which positively regulates the release of cytochrome c from mitochondria, cytochrome c oxidase assembly factor 4, were increased in abundance in fish exposed to Luppe sediment, indicating that in addition to global cell cycle arrest, apoptosis was triggered in hepatocytes (Table 10). Another interesting finding of the present study was the induction of cytochrome P450 1A1, an aryl hydrocarbon receptor (AhR) mediated biomarker of exposure to dioxins and dl-PCBs (Brinkmann et al. 2016), in relation to Luppe sediment exposure. Although bi-directional cross-talk between AhR and ER exist and up- as well as downregulation of cyp1a1 following exposure to NP or E2 has been reported (Göttel et al. 2014; Shelley et al. 2012; Won et al. 2014), this might indicate that concentrations of dioxins and PCBs measured in the Luppe sediment (Table A.3 and A.4) were remobilized during suspension and were bioavailable for trout. This is further supported by findings of Brinkmann et al. (2015), who found that sediment-bound dioxins, PCBs and PAHs became bioavailable during sediment suspension in a 90 days exposure study with juvenile rainbow trout within the same test set-up as the 93 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout present study. Likewise, the strong induction of transmembrane protein 163-like, which was found to play a role in tolerance to divalent metal ions such as zin (Cuajungco and Kiselyov 2017), in fish exposed to suspended sediment from the Luppe River might suggest additional exposure to metals. Measured concentration of metals including zinc, lead, copper, and cadmium in sediment from the Luppe River were elevated and exceeded the sediment quality criteria according to German water policies (OGewV 2016) (Table A.3 and A.4). In summary, altered gene expression profiles of trout in response to suspended sediment exposure from the Luppe River, either diluted or undiluted, suggest that in addition to the characterized EDCs a number of other contaminants such as dioxins, PCBs and heavy metals were remobilized from sediments during suspension creating a complex chemical mixture situation. However, besides an overall cellular stress response indicated by global cell cycle arrest as well as induction of apoptosis, estrogen responsive genes were the most significantly altered genes in fish in response to Luppe sediment exposure, highlighting the sensitiveness of EDC exposure even under exposure to complex mixtures. 4.5.2.3 Effects on biomarker induction, liver histology and overall fitness Mucus vtg induction in male as well as female trout exposed to suspended sediment of the Luppe River agree with the RNA sequencing results and indicate that exposure to EDCs resulted in alterations on the protein level. While vtg was significantly induced in both sexes of Luppe exposed fish compared to unexposed fish in the present study, the magnitude of induction was very low compared to other literature reports on single substance exposure experiments with juvenile trout and reflected the similar low induction of vtg transcripts on a gene level (Jobling et al. 1996; Thorpe et al. 2003). Jobling et al. (1996) reported a 500-fold vtg induction up to 10 µg/mL in plasma of juvenile male rainbow trout after exposure for 21 days to concentrations of 30 µg/L NP. Moreover, vtg induction was measured in less than half of the Luppe exposed fish and vtg was measured only in few fish exposed to the serial dilution. Assuming that waterborne exposure to remobilized EDCs from the sediment is the major route of exposure, low vtg induction rate might be explained by the fact that measured water concentrations of both NP, E1 and possible E2 and EE2, which were below the LOD of LC-MS/MS measurements, were close to or below the predicted no-effect concentrations (E1: 6 ng/L, E2: 2 ng/L, EE2: 0.1 ng/L) derived for vtg induction in fish (Caldwell et al. 2012). Additive effects of even low doses of EDCs might have 94 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout in sum lead to the observed mild induction of vtg in nine fish exposed to suspended sediment of the Luppe River, while biotransformation capacity or lower internal exposure levels might have been protective enough in exposed fish without vtg induction. Moreover, reduced condition factors (K) of fish exposed to suspended sediment from the Luppe River in the present study might further indicate that the overall stress caused by exposure to a highly contaminated sediment resulted in reduced nutrition status. Similarly, Brinkmann et al. (2015) also observed reduced K after exposure to a highly contaminated sediment from the Elbe River. Additionally, alterations observed in the liver of fish exposed to sediments from the Luppe as well as the sediment control and all dilution treatments indicated that exposure to suspended sediment itself resulted in cellular stress leading to hepatotoxic responses in fish. Hepatotoxic effects of NP exposure in fish have previously been reported in the literature, including various histopathological alterations as hepatic coagulative necrosis, nuclear alterations, in addition to apoptosis of hepatocytes, necrosis of endothelial cells and proliferation of connective tissue (Abd- Elkareem et al. 2018; Sayed and Soliman 2018). However, hepatic alterations observed in the present study were not related to NP (Luppe sediment) treatments but instead to exposure to SPM. This is an important finding, since it indicates that flood events might induce acute liver alterations even at less polluted sites. Moreover, exposure to suspended sediment highly contaminated with EDCs resulted in an estrogenic response on the gene level but only on a mild biomarker induction on the protein level. Future research should investigate the long-term effects of such exposure and whether these alterations might lead to impaired reproductive success in adults or if fish might recover from enhanced EDC exposure after disturbance events such as floods. However, the toxic mode of action of EDCs is not solely restricted to adverse effects on the reproductive fitness but comprise also of carcinogenic properties (Lam et al. 2011) and immunotoxicity leading to an enhanced pathogen susceptibility (Burki et al. 2012; Rehberger et al. 2017; Shelley et al. 2012). It is still unclear what consequences annually occurring flood events have on fish that are exposed to a variety of stressors, including remobilized concentrations of EDCs, and further research is needed to evaluate the impact not only on the reproductive fitness but on the overall health including immunocompetence of wild fish. 95 Chapter 4 – Impacts of remobilized sediment-bound EDCs on trout 4.6 Conclusion Our results demonstrate that sediment-bound EDCs are remobilized and become readily bioavailable for fish when sediments are suspended under conditions similar to those of a flood event. Partitioning of EDCs into the water phase was enhanced during the suspension of the sediment, and the aqueous phase was a major route for uptake of remobilized sediment-bound EDCs into the fish. Passive sampling was a useful tool to assess the bioavailability of sediment-bound EDCs and could be a good indicator of sediment toxicity in a regulatory context. Even with exposure to complex mixtures from the field sediment during suspension, the effects of EDCs were observed with an estrogenic response on the transcript level and partly on protein level was observed in male fish exposed to the Luppe sediment. In addition to the induction of the typical hepatic estrogen-responsive genes, repression of cell cycle genes in combination with the induction of genes involved in apoptosis in male fish exposed to sediment from the Luppe River suggest a broader effect and future research should address how annual flood events affect the overall fitness of wild fish populations. 4.7 Acknowledgments This study was generously supported by a Project House of the Exploratory Research Space (ERS) at RWTH Aachen University, as part of the German Excellence Initiative via the German Research Foundation (DFG). The study was also supported by the SOLUTIONS project (European Union's Seventh Framework Programme for research, technological development and demonstration under Grant Agreement No. 603437). M.B. is currently a faculty member of the Global Water Futures (GWF) program, which is supported by the Canada First Research Excellence Fund (CFREF). We are thankful for the support of the TecoMedical Group for providing help with vtg ELISA training and data evaluation. 96 Chapter 5 5 Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment 97 98 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment 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, 52074 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 This chapter has been developed based on a publication in the following peer-reviewed journal: Müller A-K, Markert N, Leser K, Kämpfer D, Crawford SE, Schäffer A, Segner H, Hollert H (2019) Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment. Environmental Pollution: 113636. Doi: 10.1016/j.envpol.2019.113636 99 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 5.1 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 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 100 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 5.2 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), ethynylestradiol (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. Ethynylestradiol 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. 2010b; 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. 2013b) 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, Nebraska, USA. Whereas, hepatic expression of vtg and estrogen receptor α was reduced in female 101 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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 102 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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 5.3 Material and Methods 5.3.1 Study sites The Luppe River is a tributary of the Saale River located between Leipzig and Merseburg, Germany. It is a slow-moving stream, approximately 50 km long, and is characterized by abundant macrophyte growth and thick layers of sediment (Kammerad et al. 2014) (Figure 5; 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. 2013b). 5.3.2 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. For further details on sediment and water sampling as well as sediment characteristics see Chapter 2.2. 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 in Chapter 2.5. 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 collected, 103 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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) and gonadosomatic index (GSI) and liver somatic index (LSI). Scales of each fish were carefully scraped off at the side between head and dorsal fin for age determination. 5.3.3 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. Detailed extraction method is described in Chapter 2.8. Samples were stored at -20°C until measurement with LC-MS/MS. Method blanks (n=6) were produced following extraction protocol without plasma sample. All measured concentrations of target compounds were blank corrected (see Chapter 2.7 for LC-MS/MS settings). 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 2). 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. 5.3.4 Sediment contact test with the blackworm and Luppe sediment Sediment contact test with the blackworm 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 104 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species and NP in DMSO (< 0.1%) matching concentrations and general physicochemical properties of Luppe sediment (see Table A.2) as a positive control (Table 12); (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 Chapter 2.5. Table 12: 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 Nominal spike Compound concentration [ng/g dry weight] NP 120,000 E1 20 E2 1.5 EE2 0.008 5.3.5 Vitellogenin analysis Vtg was measured using the Cyprinid Vitellogenin ELISA kit (TECOmedical GmbH, Bünde, Germany). All steps were performed in accordance with the instruction manual using the provided materials (Chapter 2.9). For measured absorbances which were below the calibration curve, these concentrations were set at the LOD for the data analysis. 5.3.6 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 105 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species were not used for age estimation. The examinations were repeated in two readings and results were cross-checked between two readers, independently. 5.3.7 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 prepared 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). 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 (Johnson et al. 2010): Female • 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. Male • Juvenile: spermatogonia exclusively, difficult to confirm the sex. • Stage 0 (Undeveloped): entirely immature phases (spermatogonia to spermatids). 106 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species • 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. 5.3.8 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 (R Core Team 2018) 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 107 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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. 5.4 Results and Discussion 5.4.1 EDC exposure of fish under field conditions 5.4.1.1 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 13). 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 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; 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. 5.4.1.2 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 13). 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/ River in 2016 (Müller et al. 2019; 17.5 µg/kg EEQs Luppe; NP 37.3 mg/kg; E1 93.9 µg/kg; E2 0.9 108 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species µg/kg). Accumulation of EDCs in sediments of rural, urban or industrial areas have been reported worldwide, mostly identifying NP, E1, E2 and EE2, but also BPA and other alkylphenols, as the active compounds (Gong et al. 2016; Li et al. 2019; Viganò et al. 2008). In major German rivers, NP concentrations in sediments reached values of approximately 0.4 - 1 mg/kg (Elbe), 1.4 mg/kg (Saale), 4 mg/kg (Rhine) and 9.9 mg/kg (Danube) (Heemken et al. 2001; LUBW 2001; Stachel et al. 2003), which were exceeded by at least 2.5 times for concentrations measured in the Luppe River of the present study. In agreement with Buchinger et al. (2013), the present study demonstrates that the Luppe serves as a “hotspot” for EDCs in sediments (Table 13), exceeding mean estrogenic contamination in the nearby catchment areas of the Saale River. Table 13: Estrogenic activity in sediment and water samples from the Luppe and Laucha Rivers and literature reports worldwide determined by the Yeast Estrogen Screen (YES) in 17β-estradiol equivalents (EEQs) and concentration of target EDCs nonylphenol (NP), estrone (E1) and 17β-estradiol (E2) by LC-MS/MS (or other analytical methods). Compound in sediment [µg/kg] or Sampling site Matrices/ EEQs water [ng/L] (country) Reference [µg/kg] NP E1 E2 61.2 ± Luppe (present Sediment 22342.6 67.3 >LOD 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 109 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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. 5.4.1.4 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 (Table 14). E1, E2 and EE2 were below the limit of detection (LOD; see Table 2) 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. 110 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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 20), 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) (Table 14). 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). 111 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species Table 14: 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. Mean Mean Mean Mean Mean vtg ± SD blood Sampling Gender length weight age ± LSI GSI K [ng/mL plasma NP site (Nb.) [cm] + [g] + SD per [ng/mL] ± SD SD mgprotein] SD Tench juvenile 10.1 ± 21.1 ± 2.0 ± 0.2 ± 2.0 ± 27.0 ± 2 ± 0.9 2.0 ± 0.9 (5) 1.4 8.7 0.2 0.05 0.3 10.3 14.4 ± 61.8 ± 1.8 ± 0.6 ± 1.8 ± 454.2 ± male (6) 3 ± 0.8 21.5 ± 2.8 2.1 28.7 0.2 0.4 0.2 858 female 13.6 ± 53.9 ± 1.8 ± 4.2 ± 2.0 ± 10991.1 3 ± 1.2 22.7 ± 7.0 (9) 2.5 26.1 0.5 3.8 0.3 ± 11490 Roach Luppe juvenile 1 10 15.6 0.9 0.9 1.6 205.8 - (1) 10.4 ± 20.1 ± 1.4 ± 2.8 ± 1.9 ± 162.5 ± male (5) 2 ± 0 - 1.2 5.1 0.2 1.4 0.4 178.4 female 12.6 ± 36.2 ± 2.5 ± 2.1 ± 1.7 ± 165.1 ± 2 ± 0.4 35.5 (n=1) (7) 3.13 28.5 2.0 2.0 0.6 198.0 Pike mixed 2 ± 0 19.7 ± 70.4 ± 1.8 ± 0.1 ± 0.9 ± 49.4 ± n.a. (10) 2.3 21.0 0.5 0.1 0.1 42.0 Tench juvenile 8.6 ± 12.1 ± 1.7 ± 0.1 ± 1.8 ± 21.9 ± 1 ± 0.4 4.6 ± 4.4 (6) 1.1 4.3 0.4 0.08 0.3 11.7 14.6 ± 57.7 ± 1.4 ± 0.2 ± 1.7 ± 15.8 ± 2.1 male (4) 3 ± 0.7 47.7 ± 93 2.4 29.3 0.2 0.08 0.3 (n=3) female 12.5 ± 38.6 ± 1.3 ± 1.2 ± 1.6 ± 21.1 ± 3 ± 1.3 5.2 ± 8.7 Laucha (11) 3.1 25.4 0.2 0.4 0.3 6.1(n=10) Roach juvenile 11.2 ± 21.0 ± 1.3 ± 0.6 ± 1.5 ± 1 5.8 ± 7.7 10.8 ± 3.7 (2) 0.4 1.6 0.3 0.2 0.04 male (0) - - - - - female 11.4 ± 26.7 1.2 ± 1.1 ± 1.6 ± 297.4 ± 13.3 ± 7.2 2 ± 0 (8) 2.7 ±15.7 0.5 0.5 0.2 810.8 (n=7) Tench male 11.1 ± 20.8 ± 1.5 ± 0.8 ± 1.4 ± 1 ± 0.5 1.5 ± 2.9 2.7 (n=2) (20) 2.2 13.7 0.4 0.4 0.1 female 11.4 ± 20.5 ± 1.6 ± 2.3 ± 1.2 ± 19.0 ± 8.1 ±6.7 1 ± 0.5 Cultured (10) 2.2 10.5 0.4 0.9 0.1 38.0 (n=5) fish Roach 10.3 ± 11.2 ± 1.9 ± 1.0 ± 1.0 ± male (5) 2 ± 0 1.5 ± 2.2 - 0.2 1.0 0.8 0.5 0.1 female 11.1 ± 16.6 ± 1.2 ± 2.2 ± 1.1 ± 11.2 ± 2 ± 0 - (14) 1.3 6.5 0.4 0.9 0.1 18.4 - indicates insufficient plasma volume for analysis; n.a.: not analyzed 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, 112 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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. 2019a). 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 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 14). 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 sediment- bound 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. 2019a). 5.4.2 Endocrine disruptive effects in tench and roach 5.4.2.1 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 113 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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 14). Furthermore, cultured roach and roach caught at both rivers were about two years old (Table 14). 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 (Table 14). 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 14). Reduced GSI has been related previously to EDC exposure and occurrence of intersex (Jobling et al. 2002a), but no differences in GSI of tench and roach between sampling sites were observed in the present study (Kruskal-Wallis, p>0.05) (Table 14). 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 20; Table 14). 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 20). 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). 114 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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 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. 115 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species Figure 20: 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. 5.4.2.2 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 22 D -F). Field tench from both sampling sites showed testes in similar maturation stages ranging from immature to mid-development (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 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 116 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species River (Figure 21; Table A.7). This is consistent with the spontaneous prevalence of intersex in fish held under control conditions reported by Wolf (2011). Figure 21: 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. 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). 117 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species Figure 22: 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. 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 118 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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 22 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 22 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 22 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 early- to 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 22 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 119 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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 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. 5.5 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 120 Chapter 5 – Impact of sediment-bound EDCs on freshwater fish species 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. 5.6 Acknowledgments 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. 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). 121 122 Chapter 6 6 Discussion 123 124 Chapter 6 - Discussion 6.1 General discussion The preceding chapters have demonstrated that the bioavailability of sediment-bound EDCs are enhanced when sediments are subjected to disturbances resulting in re-suspension as in a flood-event. This research showed that sediment-bound EDCs, such as NP, have the potential to negatively impact water quality under normal weather conditions (Chapter 3 and 5) but partitioning of EDCs into the water was enhanced during flood-like conditions (Chapter 3 and 4). During suspension of sediment, EDCs became bioavailable and were readily taken up by exposed rainbow trout (Oncorhynchus mykiss). While most of the estrogenic activity remained associated with suspended particles, uptake of dissolved EDCs in the water phase appeared to be the dominant route of exposure during the simulated flood-event to rainbow trout. Passive sampling proved to be a useful tool to assess the potential bioavailability of remobilized sediment-bound EDCs in a water-sediment suspension system and related to internal concentrations of measured EDCs in the fish. Analysis of liver transcriptome as well as histopathology of male trout revealed that exposure to high loads of suspended particles itself induced a stress related global cell cycle arrest and apoptosis of hepatocytes, which was even more pronounced in relation to EDC-contaminated sediment from the Luppe River. Besides induction of typical estrogen responsive genes, the altered hepatic gene expression profiles suggest that, in addition to EDCs, other sediment- bound contaminants such as dioxins and heavy metals became bioavailable to fish during suspension of sediments. Induction of mucus vitellogenin (vtg) as a biomarker of EDC exposure in male roach (Rutilus rutilus) and tench (Tinca tinca) caught at the Luppe River under normal weather conditions as well as mild vtg induction in male rainbow trout after laboratory exposure to the suspended Luppe sediment consistently demonstrated that sediment-bound contaminants have the potential to alter physiological responses in fish. Within this chapter, the relevance of these results and implication for current legislation will be critically evaluated and discussed. 6.1.1 The use of sediment quality standards in sediment risk assessment Within the context of the European environmental control and monitoring program, the Water Framework Directive (WFD), sediment quality assessment plays a minor role and is addressed within the surface water protection (Brils 2008). The main goal of the WFD is to achieve “good ecological” and “good chemical” status in order to protect the aquatic environment and human health. The good 125 Chapter 6 - Discussion ecological status is based on the assessment of the biological community including fish as well as benthic macroinvertebrates and, thus, is indirectly indicative of sediment quality. The assessment of the chemical status in water, sediment and biota is based on the compliance with environmental quality standards (EQS), which were established for the 45 priority substances, including NP, and river basin specific pollutants and concentration thresholds should be protective for human health and the environment (EU 2007/60, 2008/105, 2013/39). While EQS are commonly applied in the context of environmental monitoring of waterbodies under the European WFD, until now no EQS for sediment have been derived for use in the WFD. EQS for sediment have been critically discussed within the scientific community (Brils 2008; Burton 2018; Chapman 2018; Kwok et al. 2014). Many efforts have been made over the past decades towards the development of consensus-based sediment quality criteria and several approaches were proposed (Ahlf et al. 2002; Chapman 1990; Chapman and Hollert 2006; MacDonald et al. 2000). 6.1.1.1 Bioavailability of sediment-bound contaminants Originally, sediment EQS or sediment quality guidelines were based on measurement of bulk chemical concentrations compared to a reference value (Burton 2002). The bioavailability for sediment- bound contaminants was discussed in several reviews (Ahlf et al. 2002; Burton 2018; Chapman and Hollert 2006). One concept to address the bioavailability of sediment-bound substances, is by measuring contaminants in pore water, which is assumed to be bioavailable in terms of ready for uptake in for e.g. benthic macroinvertebrates (Burton and Johnston 2010). Moreover, in vitro bioassays for testing of sediment toxicity under controlled laboratory conditions based on whole organisms were developed since the 1980s, as they are integrative methods and provide information about effects and, thus, indirectly exposure (Ahlf et al. 2002). Data from toxicity tests paired with measurements of contaminants in sediments from the field were used to develop empirically based EQS for sediment, where average threshold levels were later summarized as consensus-based sediment quality standards i.e. guidelines (Cormier et al. 2008; Ingersoll et al. 2001; MacDonald et al. 2000). However, the relation between sediment quality and risk remains complex and several limitations of sediment EQS have been identified (Brack et al. 2018; Burton and Johnston 2010; Chapman 2018). One major concern is related to the bioavailability of sediment-bound contaminants. Bioavailability has been described as the portion 126 Chapter 6 - Discussion of contaminants that is available for uptake by an organism (Rand 1995; Schwarzenbach et al. 2002). Concentrations of freely dissolved organic compounds, such as EDCs, are generally assumed to be more readily available for uptake in an organism and thus bioavailable in water (Rand 1995). Bioavailability of sediment-bound contaminants strongly depends on its physical-chemical properties and the ability to sorb to organic carbon and other sediment properties such as the amount of fine particles (Campana et al. 2013; Yamamoto et al. 2003; Zhou et al. 2007b). Several studies highlighted that not only organic carbon content impacts sorption of EDCs and other organic contaminants to SPM and sediment, but also other mechanisms might be equally important such as aromaticity, π-π interactions between EDCs and colloid organic content, hydrogen bonding or quality of organic matter such as humic versus tannic acid content (reviewed by Ma and Yates 2018). Future research is needed to address how different physical and chemical sediment characteristics influence sorption and desorption rates during suspension of sediment and whether normalization to total organic carbon content (logKoc values) adequately predicts bioavailability of sediment-bound contaminants. The abovementioned approaches to integrate bioavailability into sediment quality assessments, such as measurement of pore water concentrations or ecotoxicity testing, are entirely based on single exposure routes to bedded sediment. Several literature reports including the results of the present thesis emphasized the importance to characterize bioavailability in context of sediment resuspension which is occurring under natural conditions due to extreme weather events like flooding but also in context of dredging or bioturbation (Brinkmann et al. 2010; Burton 2002; Burton and Johnston 2010; Hollert et al. 2007a). Partitioning and, thus, bioavailability of sediment-bound contaminants might be relatively low due to high sorption capacity as indicated by logKoc > 3 when the sediment is settled but can significantly increase during resuspension of the sediment. The results presented in this thesis demonstrate that bioavailability in terms of freely dissolved concentrations of a contaminant in the water phase was enhanced, e.g. for NP by a factor of 250 (in beaker experiment with passive sampling, Chapter 3) or 14 (within the fish exposure study, Chapter 4), during suspension of the sediment compared to measured concentrations in field collected water samples from the Luppe River. Similar results were reported by Weert et al. (2010), who found that concentrations of NP released from the sediment (total concentration of 20 mg/kg d.w. NP) into the 127 Chapter 6 - Discussion water phase was enhanced upon suspension of the sediment from 0.3 µg/L (bedded sediment) up to 20.4 µg/L (suspended sediment). Likewise, calculated mass transfer of NP increased by a factor of 100 from 0.3 µg/d up to 29 µg/d upon suspension of the sediment. Interestingly, Weert et al. (2010) reported that concentrations of NP sharply increased upon suspension of the sediment possibly resulting from the release of pore water concentrations as well as the increased exchange surface area between water and sediment. Furthermore, they observed that as soon as suspension was stopped, and bedded sediment was established again, the mass transfer of NP from the sediment into the water phase decreased back to the original 0.3 µg/L (Weert et al. 2010). Enhanced bioavailability during suspension of sediment either in laboratory experiments or during naturally occurring flood events in the field was further observed for dioxins, PCBs, PAHs as well as heavy metals (Brinkmann et al. 2013; Brinkmann et al. 2015; Schwandt and Hübner 2014). Based on the literature and the results presented in this thesis, it is important to assess bioavailability of sediment- bound contaminants in context of flood events. Even more importantly, the results presented in this thesis demonstrated that remobilized sediment-bound EDCs were readily taken up by exposed fish and, therefore, showed that sediment-bound contaminants are not only potentially hazardous for benthic or bottom-dwelling organisms but also for pelagic living fish species. 6.1.1.2 Routes of exposure to remobilized sediment-bound contaminants Another route of exposure to sediment-bound contaminants during suspension of sediment might be through ingestion of the sediment (Burton and Johnston 2010; Hudjetz et al. 2014). Hudjetz et al. (2014) observed significant amounts of ingested sediment in the guts of rainbow trout exposed to suspended sediment contaminated with PAHs. Hudjetz and co-workers hypothesised that uptake of PAHs was related to sediment ingestion and absorption resulting from solubilization by digestive fluids. This was further supported by other studies including Moermond et al. (2004), who identified that up to 20 % of total contaminant uptake was related to sediment ingestion and absorption in the gastrointestinal tract. In contrast, the results presented in this thesis in Chapter 4 suggest that waterborne exposure to remobilized EDCs during sediment suspension might be the major route of uptake into exposed rainbow trout. Although small amounts of sediment were generally observed in the intestine of rainbow trout during dissection, absorption of EDCs through ingested SPM appeared to be of minor importance: while 128 Chapter 6 - Discussion total concentrations of the target EDCs, i.e. NP and E1, in the Luppe sediment and the 1:2 dilution of Luppe and control sediment differ significantly by a factor of two, average internal concentrations of NP and E1 measured in plasma and bile of fish as well as Cfree of NP in the water phase by passive sampling did not differ significantly between the two treatments. If sediment ingestion additionally enhanced bioavailability of sediment-bound EDCs, we would expect based on the significantly greater concentrations (and total amount of contaminated sediment) of EDCs in Luppe sediment to observe significantly greater concentrations of EDCs in fish exposed to the undiluted Luppe sediment compared to undiluted and controls. However, measured concentrations of i.e. NP in blood plasma and bile of the exposed fish in the Luppe and 1:2 dilution varied up to 90 %, which might account for the fact that exposure to sediment-bound contaminants during suspension is a highly dynamic process itself. Desorption from the sediment into the water as well as adsorption onto SPM occur at the same time resulting in a dynamic exposure scenario. Consequently, uptake of EDCs from the sediment through ingestion and subsequent absorption within the gastrointestinal tract might have occurred but this effect is insignificant due to the overall high variability. Further research is needed to distinguish to what extent ingestion of contaminated sediment might contribute to overall uptake of sediment-bound contaminants during suspension. This might be addressed in future studies where fish could be exposed only to the water phase of a resuspension sediment-water system but where the sediment and SPM are removed i.e. through filtering. 6.1.1.3 Mixture toxicity Sediment quality evaluation defined by single substances as proposed by the usage of EQS does not address additive, antagonistic or synergistic effects of exposure to multiple contaminants accumulating in sediments. This concern has been raised in several critical reviews and is particularly important with regards to EDCs (Burton and Johnston 2010; Chapman 2018), where estrogens but also alkylphenols in mixtures interact in an additive manner (e.g., trigger the induction of vtg in juvenile rainbow trout; Brinkmann et al. 2018; Thorpe et al. 2003). Concentrations of estrogens, e.g. E2 and EE2 and NP, below their individual lowest observed effect concentrations (LOEC) have been shown to be more potent in combination and lead to induction of vtg in vivo (Serra et al. 2019; Silva et al. 2002; Thorpe et al. 2006; Yu et al. 2019). Similarly, the presence of estrogen receptor antagonists and substances that inhibit parts 129 Chapter 6 - Discussion of the estrogen responsive pathway, e.g. through bidirectional cross-talk between AhR and ER or steroid synthesis and metabolism, may alter the estrogenic response in vivo and, therefore could result in a reduction of the estrogenic activity of a mixture (Nagler et al. 2007; Smeets et al. 1999; Thibaut and Porte 2004). While dioxins, such as. 2,3,7,8, - tetrachlordibenzo-p-dioxin (TCDD), are known to exert anti-estrogenic effects via AhR- mediated molecular AhR/ER cross-talk mechanisms at transcriptional level through indirect inhibition of ER (Göttel et al. 2014), hydroxylated PAHs exhibit antiestrogenic as well as estrogenic activities (Hayakawa et al. 2007). In addition to high concentrations of EDCs, dioxins, PCBs and heavy metals were measured in sediments from the Luppe River, possibly being remobilized during suspension in similar manner as observed in a previous study by Brinkmann et al. (2015) or during flooding in the field (Schwandt and Hübner 2014). Results presented in Chapter 4 of this thesis showed that vtg as a biomarker for EDC exposure was induced in male and female fish exposed to suspensions of Luppe sediment. However, induction was only observed in nine out of twenty fish and were below 10 ng/mL per mgprotein. Assuming that waterborne exposure to remobilized sediment-bound EDCs was the major route of uptake, measured concentrations of E1 and NP themselves were close to respective NOEC levels. Furthermore, uptake of substances such as PCBs exhibiting antiestrogenic activity might have lowered the overall estrogenic effect. The RNA sequencing results from the present research (Chapter 4) provides further evidence that modulation of hepatic gene expression profiles of rainbow trout exposed to suspended sediment from the Luppe River result from multiple stressors. In addition to the regulation of typical estrogen responsive genes such as vitelline envelope protein α, transcripts of cyp1a1 AhR mediated biomarker of exposure to dioxins and dl-PCBs were up-regulated in relation to Luppe sediment exposure. This might indicate that dioxins and PCBs found in the Luppe sediment (sum of PCDD/Fs and dl-PCBs of 109 ng toxicity equivalents (TEQ)/kg) were remobilized during suspension and bioavailable for trout. Likewise, the strong induction of transmembrane protein 163-like, which was found to play a role in tolerance to divalent metal ions such as zinc (Cuajungco and Kiselyov 2017), in fish exposed to suspended sediment from the Luppe River might suggest additional exposure to metals. Measured concentration of metals including zinc, lead, copper and cadmium in sediment from the Luppe River were elevated and exceeded the sediment quality criteria according to German water policies (OGewV 2016). 130 Chapter 6 - Discussion 6.1.1.4 Sediment quality assessment as weight of evidence approach Given the fact that various contaminants accumulate in sediments worldwide and are released from these sediments during suspension, sediment quality criteria should be developed from sites where mixtures of toxic contaminants occur, and physiological responses are based on exposure to multiple stressors (Burton and Johnston 2010). Consequently, EQS for sediment should be used as a first indication or initial screening tool to identify sites where a negative impact of sediment-bound substances based on measured concentrations might be expected. EQS, however, might not be robust enough considering the complexity of exposure routes as well as mixture of contaminants for decision making. Instead, it has been suggested that multiple lines of evidence should be conducted including the use of bioassays, investigation of bioavailability during sediment suspension and bioavailability of contaminants in complex mixtures and the assessment of benthic as well as pelagic living organisms (Burton 2018; Burton and Johnston 2010; Chapman 2018; Tarazona et al. 2014). Figure 23: Sediment quality triad as described in Chapman et al. 1990, later complemented with additional lines of evidence in 2006 (Chapman and Hollert 2006) and modified according to results on bioavailability of sediment- bound contaminants observed in the present study. The sediment quality triad, based on the weight of evident approach, utilizes simultaneous observations of sediment chemistry, ecotoxicity testing and in situ measurements of e.g. benthic macroinvertebrate communities has already been recommended for determination of ecological impact of sediment contamination (Chapman 1990). Later additional line of evidences were proposed such as effect/ bioassay directed fractionation (Chapman and Hollert 2006) (Figure 23). Passive sampling, which was reported to be a good predictor of internal concentrations of contaminants in rainbow trout (Vermeirssen et al. 2005), was used in the present research to estimate bioavailability of sediment-bound 131 Chapter 6 - Discussion EDCs during suspension of sediment. Based on the results, it is recommended that aqueous concentrations of concentrations measured via passive sampling should be utilized as an additional line of evidence to investigate the risk of sediment contamination. The results of the present study suggest that waterborne exposure to remobilized sediment-bound EDCs might be the dominant route of uptake for fish, however, this must be confirmed in future investigations. 6.1.2 Implications for sediment risk assessment under the WFD As mentioned above, the major goal of the EU WFD is to achieve the good ecological and chemical status of surface water bodies by 2015, which is now prolonged until 2027 (EU 2000/60). Neither the good ecological nor the good chemical status has been achieved in most European river basins, including Germany, and it has been recognized that this poor quality status is associated with the toxic risk to the aquatic ecosystem based on complex mixtures of chemicals in surface water and sediment (Altenburger et al. 2019; BMUB/UBA 2016; Brack et al. 2017; Brack et al. 2018; Brack et al. 2019a; Brack et al. 2019b; Gils et al. 2019; Hollert et al. 2007b). While the WFD assessment of the chemical status is restricted to monitoring of the 45 defined priority substances (PS) in addition to river basin specific pollutants (RBSP) (EU 2000/60, 2008/105, 2013/39), more than 22,000 chemicals are currently registered alone under the European´s legislation on the regulation evaluation authorisation and restriction of chemicals (REACH) (ECHA webpage 2019). Many thousands of these chemicals, in addition to multiple pesticides and pharmaceuticals, were measured in sediment and water samples across Europe (UBA 2018). Thus, current water as well as sediment quality monitoring under the WFD inadequately reflect the actual chemical contamination load and new hazardous substances of emerging concern, resulting in an underestimation of the risk posed to the aquatic environment and human health (Brack et al. 2018; Gils et al. 2019). The fact that NP (banned from EU countries) was detected in water samples of the Luppe by LC MS/MS in the present thesis demonstrates how sediment contamination might negatively impact water quality. With this year’s revision of the WFD, several recommendations for monitoring, assessing and managing risks of contamination in the aquatic environment have been addressed (Brack et al. 2017; Brack et al. 2018; Escher et al. 2018). Based on the research project SOLUTIONS and the European monitoring network NORMAN, a series of policy briefs and research articles have been released on how to implement solution-oriented monitoring concepts and state-of- 132 Chapter 6 - Discussion the-art science into recommendations for the WFD (Altenburger et al. 2019; Brack et al. 2019a; Brack et al. 2019b; Brack 2019; Faust et al. 2019; Gils et al. 2019; Posthuma et al. 2019; Slobodník et al. 2019). One main recommendation for the improvement of environmental monitoring was to include effect- based methods (EBMs), such as bioassays and biomarkers, as integrative techniques that bridge the gap between chemical contamination and ecological effects (Brack et al. 2019a). Based on the 5-year SOLUTION project a basic bioassay battery has been suggested including in vitro tests for estrogenic activity (Brack et al. 2019a; Escher et al. 2018) (Figure 24). Moreover, based on a review by Kase et al. (2018) as well as an international workshop organized by the Swiss Centre for Applied Ecotoxicology of Eawag/EpFL (Ecotox Centre, Duebendorf, Switzerland) a set of 15 EBMs (five in vivo and ten in vitro) for the detection of EDCs have been recommended (Hecker and Hollert 2011). In addition to EBMs based on receptor mediated endocrine disruption, in vitro test investigating other mode of actions of EDCs such as interference with hormone production i.e. the H295R steroidogenesis assay were selected for monitoring purposes (Hecker and Hollert 2011). Several field and laboratory studies demonstrated the applicability of the EBMs in context of EDC monitoring, particularly the monitoring of EDCs and estrogenic potential in sediments (Grund et al. 2010a; Grund et al. 2010b; Hecker et al. 2002; Hecker et al. 2007; Keiter et al. 2006). The p-YES assay used in the current study was demonstrated to be a robust and fast screening tool for the detection of estrogenic compounds that is applicable to sediment and bile extracts and provides information about the quality, in terms of EEQs, and the quantities of target EDCs in environmental samples. Moreover, data obtained in the present study by the p-YES were comparable to conventional chemical analysis but had an advantage in that they had lower LODs and LOQs. Hence, the p-YES would be a good addition to the suggested EBMs for environmental monitoring as the assay links chemical and bioanalytical analysis and, thereby, provides information on cause-effect relationships. 133 Chapter 6 - Discussion Figure 24: Effect based methods for water monitoring as presented in Brack et al. (2019a) and modified in Altenburger et al. (2019). Furthermore, sediment contamination has been recognized to contribute to the failure of WFD environmental quality goals and recommendations for improvements have included the development of reliable approaches for the determination of the bioavailability with emphasis on resuspended sediment (Brack et al. 2017). Further, Brack et al. (2017) proposed that EQS for sediment should be derived based on the freely dissolved concentration, that can be determined using passive sampling, instead of using total concentrations. This is supported by the findings presented in this thesis, where passive sampling was used to determine the freely dissolved concentrations of EDCs in the water during sediment suspension and was found to correlate with EDC uptake in exposed rainbow trout. However, further research is needed to confirm this relationship and it´s potential for modelling and future prediction of EDC exposure to aquatic organisms based on passive sampling. Two types of passive sampling devices, POCIS and Chemcatcher, were evaluated for their suitability to determine freely dissolved concentrations of EDCs in sediment suspension systems. Both devices had advantages and disadvantages: the Chemcatcher exhibited, as reported by previous studies, higher sampling rates compared to the POCIS, which favors shorter time frames of monitoring. However, our results demonstrated that the Chemcatcher was depleting the target substances from the water phase and 134 Chapter 6 - Discussion that the uptake of the freely dissolved substances in the water into the Chemcatcher in this experimental setup was faster than the dissociation of the substance from the suspended sediment. This effect is important to consider for passive sampling in laboratory studies with a defined amount of sediment but might be negligible under field conditions where the sediment acts as an infinite source. Based on the observed depletion, the POCIS might be more suitable for laboratory investigations. However, NP exhibited no linear uptake into the POCIS, which is supported by similar findings previously described (Vallejo et al. 2013). In addition to the estrogens E1, E2, EE2 as well as NP investigated within this thesis, both passive sampling devices have been shown to sample a large range of polar organic substances from different classes of organic compounds with a log Kow from 2 up to 4 including pesticides, non-ionic detergents, polar pharmaceuticals and sampling rates derived from laboratory calibrations are available for a set of target substances (Ibrahim et al. 2013; Vermeirssen et al. 2012, 2013; Vrana et al. 2005). Moreover, passive sampling combined with non-target or suspect screening could be used to characterize the bioavailability of the complex chemical mixture accumulated in sediments and identify substances with enhanced availability (Altenburger et al. 2019). Future research should evaluate whether passive sampling devices such as the POCIS and Chemcatcher can be applied to assess the bioavailability of other compounds from the sediment during suspension. Furthermore, standardized protocols for the evaluation of bioavailability of sediment-bound compounds during sediment suspension utilizing passive sampling should be derived. In summary, the results presented in this thesis illustrate that a dynamic interaction of the sediment compartment with the overlying water exist and, thus, supports the recommendations made by Brack and co-workers that in order to sufficiently protect the aquatic environment sediment risk assessment should be further implemented within the legal context of the WFD. 6.1.3 Bridging data from laboratory to field – effects of sediment-bound EDCs Naturally, multiple abiotic and biotic factors such as temperature or nutrition are physiologically integrated and influence maturation and the reproductive cycle of fish. Consequently, causal relationships between reduced reproduction resulting from exposure to EDCs are often difficult to establish under field conditions (Segner 2005). Thus, laboratory testing, where confounding factors are minimized, is important to identify stressors related adverse effects. For instance, induction of estrogen 135 Chapter 6 - Discussion responsive genes as well as the biomarker vtg in rainbow trout was related to exposure and uptake of remobilized sediment-bound EDCs presented in Chapter 4 of this thesis. This related to the induction of vtg in field collected tench and roach from the Luppe River. The altered hepatic gene expression profile including the induction of cyp1A1 as dioxin responsive gene of fish exposed to suspended sediment from the Luppe indicated that besides EDCs other contaminants were released and available for uptake. Given the fact that suspension of the Luppe sediment resulted in an exposure to multiple contaminants, the observed estrogenic response either on the gene or vtg biomarker level in rainbow trout indicated that even in the context of exposure to a complex mixture effects on the endocrine system are an important endpoint to assess. While suspension of contaminated sediment such as from the Luppe River resulted in the release of multiple contaminants, other parameters such as nitrate and ammonium concentrations, oxygen, pH or temperature were controlled in an optimal range during the laboratory exposure study. During a flood event under field conditions, however, water parameters significantly vary from normal conditions. Temperature and pH typically decrease during flood events, whereas nitrate and ammonium might increase due to enhanced run-off from agriculture (Schwandt and Hübner 2014). Oxygen concentrations often decline to critical levels for fish below 2 mg/L as a result of resuspension of anoxic sediment or in some cases sapropel (Schwandt and Hübner 2014). As a consequence, river basin wide fish kills were observed during a flood at the Elbe River and its tributaries in 2002 (Reincke 2003). Hence, in addition to contaminant release from the sediment and the exposure to high amounts of suspended sediment other stressors might modulate physiological responses to EDCs and affect the endocrine system. Temperature, hypoxia as well as food supply have been reported to induce alterations of the endocrine system including vtg induction in fish (King et al. 2003; Lai et al. 2019; Thorpe et al. 2000a). These are still unknown confounding factors which were not addressed within the laboratory exposure study presented in this thesis. Future research is needed to evaluate how a flood event under natural conditions, including multiple stressors, ultimately might impact the response of the fish to remobilized sediment- bound EDCs. 136 Chapter 6 - Discussion Moreover, exposure of rainbow trout to suspended control sediment exhibiting low levels of EDC as well as PCB, dioxin and metal contamination lead to an altered hepatic gene expression characterized by repression of multiple cell cycle related genes as well as histological alterations of the liver. This indicates that exposure to “clean” sediment suspensions acts as a stressor towards fish. Physical and physiological factors related to oxygen availability and uptake by fish during sediment suspension might explain observed alterations. Oxygen exchange at the gill surface might be impaired due to physical interactions with suspended particles (Servizi and Martens 1991). This is an important finding, since it indicates that adverse alterations of fish health might occur due to sediment resuspension during flood events even at places with low levels of contamination. Concentrations of suspended sediment used in this study ranging from 0.9 to 2 g/L were higher than commonly observed SPM concentrations varying around 0.2 g/L during flood events (Schwandt and Hübner 2014) and, thus, might represent a worst- case scenario. Induction of hepatic estrogen responsive genes together with the induction of vtg observed in rainbow trout after exposure to suspended sediment as well as in captured male roach and tench from the Luppe River indicate an estrogenic response in relation to the sediment-bound EDCs. Whether or not this observed alterations of the endocrine system might translate into an adverse effect on the reproductive fitness, and thus, is of biological relevance, should be addressed in future studies. However, intersex or other pathologies indicating an endocrine disruptive effect were absence in testis of male tench and roach sampled at the Luppe River. Although tench is a bottom dwelling benthic fish living in close contact with the contaminated sediment, male gonad development appeared to be unaffected. This is in agreement with other studies observing induction of vtg but only low incidences of intersex in male fish inhabiting rivers where EDCs accumulate in sediment (Goncalves et al. 2014; Hecker et al. 2002). Our findings imply that under normal weather condition when the sediment is settled the bioavailability of sediment-bound EDCs is reduced to a level where no adverse effects related to male gonad development and reproduction occur. However, factors such as exposure duration of tench and roach at the sampling sites, migration from less polluted rivers, stocking of juvenile fish from aquaculture, life stage and species sensitivity and mixture effects remain unknown and must be addressed in future research to estimate whether or not sediment contamination under static conditions has no lasting 137 Chapter 6 - Discussion adverse effects on the endocrine system under field conditions. Based on the laboratory studies presented in this thesis, bioavailability of sediment-bound EDCs is expected to be similarly enhanced during a natural flood event at the Luppe River. Long-term effects on reproduction and population fitness resulting from exposure to multiple stressors during flood events should be investigated in future studies. An important unknown to address is whether fish might recover and observed effects on gene and biomarker induction are reversible. Furthermore, EDCs exhibit carcinogenic properties (Lam et al. 2011) and immunotoxicity leading to an enhanced susceptibility to pathogens (Burki et al. 2012; Rehberger et al. 2017; Shelley et al. 2012). It is still unclear what consequences annually occurring flood events where fish are exposed to a great variety of stressors including EDCs might ultimately have on other endpoints and further research is needed to evaluate the impact not only on the reproductive fitness but also on the overall health including immunocompetence of wild fish. Figure 25: Bridging lab to field: Assessing the bioavailability of endocrine disrupting chemicals during flood events by means of passive sampling, in vitro bioassays such as the planar-Yeast Estrogen Screen (YES) and in vivo exposure studies.E2: 17β-estradiol; vtg: vitellogenin; GtH: gonadotropin. Based on the findings of the present thesis, future studies should utilize several lines of evidence in a tiered approach to address the abovementioned research questions (Figure 25). As a first line of evidence passive sampling techniques could be used to assess the bioavailability of sediment-bound EDCs (Figure 25). However, future research should confirm whether uptake of remobilized EDCs from 138 Chapter 6 - Discussion the sediment in fish is mainly through freely dissolved EDCs in the aqueous phase and, thus, passive sampling is representative of internal body concentrations. This could be tested in a laboratory exposure study, in which fish i.e. rainbow trout are exposed to (i) a water-sediment suspension and (ii) only the water phase of the suspension where sediment and SPM are removed. The p-YES assay could be used as an EBMs to investigate EDCs and estrogenic activity in passive sampling extracts as well as bile extracts (Figure 25). Another line of evidence to bridge between field and laboratory data should include a laboratory exposure study investigating the recovery and lasting effects including gonad maturation as well as immunocompetence in fish after exposure to remobilized EDCs from the sediment. Therefore, future exposure studies in which fish are exposed to suspended sediment i.e. form the Luppe should include a recovery phase subsequent to exposure. A third line of evidence should include field measurements of concentrations of EDCs during an actual flood event. Again, passive sampling could be used to monitor concentrations of EDCs during a naturally occurring flood event i.e. at the Luppe River. Together with the abovementioned laboratory investigations, these three lines of evidence would provide further information for the evaluation of the risk posed by remobilization of sediment-bound EDCs during flood events. 6.1.4 Implications for flood management at the Luppe River A flood risk assessment as stated in the Directive 2007/60/EC should provide an assessment of the potential adverse consequences of future floods for human health and the environment including information on sources of environmental pollution as consequence of floods and the identification of polluted areas (EU 2007/60). In agreement with Buchinger et al. (2013b), the results presented in this thesis demonstrates that the Luppe serves as a “hotspot” for EDCs, with contamination in sediments exceeding the mean estrogenic activity found in the nearby catchment area of the Saale River by at least 37 times and concentrations of NP found in major German rivers by at least 2.5 times. Moreover, the results of the present thesis demonstrate that bioavailability of sediment-bound EDCs increase during a flood event at the Luppe River. Hence, the sediment acts as a significant source of pollution during flooding and remobilized EDCs pose a risk to the aquatic environment with unpredictable consequences for fish health and reproductive fitness of populations. Furthermore, the Saale River, as well as adjacent agriculture lands and floodplains, are likely to receive enhanced EDC contamination due to transport 139 Chapter 6 - Discussion and redistribution of suspended sediment from the Luppe River during flood events. Current flood protection measures at the Luppe River include a water gate dividing the natural riverbed of the Luppe into an upper reach (Neue Luppe) and lower reach (Luppe) (Kammerad et al. 2014). Consequently, flooding of adjacent land and floodplains is predicted to occur only in the upper end of the Luppe River including sampling sites one or one to three during a flood event depending on its severity (https://www.geofachdatenserver.de/de/hochwassergefahrenkarte-hq100). The results of this thesis highlight that the Luppe River should be prioritised within the flood risk management plans according to Directive 2007/60/EC and additional flood protection measures should be evaluated in order to minimize the risk posed by sediment-bound contamination. Interestingly, despite the overall high estrogenic activity along the river course determined from the current study, measured activities at sampling site two and five appeared to be lower compared to the other sites. The lower estrogenic activity could be related to differences in physicochemical properties of the sediment among the sites, with the sediment type at these sites having fewer fine particles and more sandy material compared to the other sites. Past restoration measures at sampling site two included dredging of sediment (personal communication with local community) and might explain the observed differences. This indicates that dredging of contaminated sediment may be an effective measure to reduce sediment contaminations within in this river system. The city of Leipzig in collaboration with the Nature And Biodiversity Conservation Union (NABU) and the Helmholz centre for environmental research (UFZ) have put many efforts in the past years to restore functional floodplains and pasture landscapes in the area between the upper reach (Neue Luppe) and the lower reach (Luppe). Part of their restoration measures is the reconnection of several small streams to the Luppe in order to increase the flow and enhance water level fluctuations (NABU 2015). However, these measures might release contaminants from the sediment and negate the efforts to restore a functional ecosystem as they might lead to exposure of sediment-bound EDCs to sensitive species of amphibians or reptiles. 140 Chapter 6 - Discussion 6.2 Conclusion and outlook The present thesis demonstrated that sediments not only function as a sink for EDCs but can turn into a significant source of pollution when sediments are resuspended. This highlights the need for sediment quality criteria considering bioavailability sediment-bound contaminants in context of flood events. The use of passive sampling provided a valuable tool to assess the bioavailability of sediment- bound EDCs during sediment suspension. Two passive samplers, the POCIS and Chemcatcher, were evaluated within the present thesis with respect to their applicability to monitor EDCs in a sediment- water suspension system. Furthermore, a novel screening tool (p-YES), which combines high performance thin-layer chromatography with an in vitro bioassay (YES), was used to identify compounds accounting for the high estrogenic activity in sediments. Hence, the applied methods represent a valuable addition to existing concepts for sediment risk assessment utilizing multiple lines of evidence. Experimental evidence was provided indicating that a) bioavailability of sediment-bound EDCs were enhanced during suspension of the sediment; b) remobilized EDCs were readily taken up by exposed rainbow trout; and c) resulted in an estrogenic response on gene and protein level in male trout. Whole transcriptome analysis provided insights in how exposure to complex environmental mixtures modulate hepatic gene expression and which pathways are particularly important. Moreover, biomarker induction together with EDC plasma concentrations indicate a potential endocrine disruptive effect in wild male fish sampled at a “hotspot” for EDC accumulation in sediment. However, this was not translated to adverse effects on the gonad level and, thus, might not be of relevance for the reproductive success of these populations. Further research is needed to evaluate whether exposure to sediment-bound EDCs during flood events might lead to lasting adverse effects on the reproductive success of fish. This could be addressed in a future laboratory exposure study, in which a recovery phase is included subsequent to the exposure of fish to the suspended sediment. Considering the recent efforts of the scientific community to implement state-of-the-art science such as EBM in the WFD, future monitoring and water policy might shift towards an integrative risk assessment approach where adverse effects from chemical mixtures are acknowledged. Experimental data presented within this thesis provide promising concepts for the future derivation of sediment EQS for EDCs. Finally, this thesis contributed to the understanding of the fate of EDCs in the aquatic environment. 141 Chapter 6 - Discussion 142 References 7 References 2010/63/EU (2010) Richtlinie des Europäischen Parlaments und des Rates vom 22. September 2010 zum Schutz der für wissenschaftliche Zwecke verwendeten Tiere Abd-Elkareem M, Abou Khalil NS, Sayed AH (2018) Hepatotoxic responses of 4-nonylphenol on African catfish (Clarias gariepinus): Antixoidant and histochemical biomarkers. 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GPS coordinates Water parameters Sampling Width Depth Conduct- Date Velocity Temper- site N E [m] [cm] O²[mg/L] pH ivity [m/s] ature [°C] [µS/cm] Luppe 1 03.08.- 51 23 08.0 12 00 33.2 n.a. 0.01 7.04 18.4 1308 17.6 90 05.08.2016 Luppe 2 03.08.- 51 22 0.14 12 02 27.9 n.a. 0.61 7.47 17.7 1230 10 35 05.08.2016 Luppe 3 03.08.- 51 21 51.1 12 05 42.3 n.a. 0.62 7.49 20.6 1118 8.5 52 05.08.2016 Luppe 4 03.08.- 51 21 40.2 12 07 55.0 n.a. 0.3 7.19 18.8 1217 11.9 96 05.08.2016 Luppe 5 03.08.- 51 22 05.2 12 10 34.7 n.a. 0.01 6.86 17.6 1330 16.7 31 05.08.2016 Luppe 1 02.08.2017 51 23 08.0 12 00 33.2 17.6 90 3.8 0.6 7.77 22.8 1460 Laucha 02.08.2017 51 23 47.3 11 59 13.4 1.3 40 2.7 0.2 7.3 22.6 1990 Rhine 26.07.2017 50 21 19.8 7 36 33.56 n.a. n.a. n.a. n.a. n.a. n.a. n.a. n.a. – not analyzed Table A. 2: Physicochemical characteristics of sediment samples from the Luppe, Laucha and Rhine Rivers. Particle size distribution was done with a laser diffraction system and was kindly provided by the Institute of Geology of the RWTH Aachen. Organic carbon content was determined by dry combustion using an PE 2400 Series II CHNS/O Analyzer. Total Total Total Water Sampling sand (2000 - 63 silt (63 - 2 clay (< 2 Type of carbon organic inorganic content site µm) [%] µm) [%] µm) [%] sediment (TC) [%] carbon carbon [%] (TOC) [%] (IC) [%] Luppe 1 n.a. n.a. n.a. 82.5 8.64 79.31 12.05 clayey silt (2017) Luppe 1 39.61 51.87 8.52 clayey silt 9.98 8.98 1.00 72.3 (2016) Luppe 2 53.75 37.90 8.35 loamy 5.61 3.28 2.33 75.2 sand Luppe 3 29.72 62.16 8.12 clayey silt 12.00 8.28 3.72 81.7 Luppe 4 21.85 68.30 9.85 clayey silt 13.83 11.24 2.59 77.5 Luppe 5 63.52 31.13 5.36 silty sand 4.53 3.71 0.82 58.5 Laucha 25.46 61.55 12.99 clayey silt n.a. n.a. n.a. n.a. Rhine 20.97 64.0 15.03 clayey silt 6.33 4.96 1.37 51.6 164 Annex I Figure A. 1: High performance thin-layer chromatography Yeast Estrogen Screen (YES) with Luppe sediment extracts (Luppe1 – 5). Standard mixtures of estrone (E1),17β- estradiol (E2), ethynylestradiol (EE2), estriol (E3) and nonylphenol (technical mixture; NP) in different concentrations and a blank were included. 165 Annex II II. Appendix: Supplementary information to chapter 4 Table A. 3: Concentrations of 17 PCDD/Fs and 12 dl-PCBs in sediments form the Rhine River sampled at Koblenz (Ehrenbreitstein) and the Luppe River. Analysis was done by münster analytical solutions gmbh according to the guideline DIN 38414-24 (DIN 2000). WHO 2005 toxic equivalency quotients (TEQs) were calculated and limit of quantification is indicated (LOQ). Concentration [ng/kg] dry weight Compound LOQ Rhine Luppe 2,3,7,8-TCDD 0.02 0.47 2.58 1,2,3,7,8-PeCDD 0.04 1.61 4.60 1,2,3,4,7,8-HxCDD 0.06 1.56 5.09 1,2,3,6,7,8-HxCDD 0.06 4.33 43.97 1,2,3,7,8,9-HxCDD 0.06 3.34 16.78 1,2,3,4,6,7,8-HpCDD 0.3 52.67 1,074.00 OctaCDD 0.9 717.34 10,288.93 2,3,7,8-TCDF 0.02 5.71 15.14 1,2,3,7,8-PeCDF 0.04 2.47 8.34 2,3,4,7,8-PeCDF 0.04 3.67 11.12 1,2,3,4,7,8-HxCDF 0.06 6.54 15.58 1,2,3,6,7,8-HxCDF 0.06 2.62 8.92 1,2,3,7,8,9-HxCDF 0.06 0.64 2.74 2,3,4,6,7,8-HxCDF 0.06 2.39 11.91 1,2,3,4,6,7,8-HpCDF 0.3 22.53 227.71 1,2,3,4,7,8,9-HpCDF 0.3 2.93 28.04 OctaCDF 0.9 120.47 1,149.47 PCB 77 40 259.03 1,294.56 PCB 81 20 166 Annex II Table A. 4: Concentrations of indicatior/i-PCBs and metals in sediments from the Rhine and Luppe River. PCB analysis was done by münster analytical solutions gmbh. Concentration of metals were analyzed using X-ray fluorescence spectroscopy and were kindly provided by the Institute of Geology of the RWTH Aachen. Units dry Compound Rhine Luppe weight PCB 28 µg/kg 0.92 21.39 PCB 52 µg/kg 1.38 22.68 PCB 101 µg/kg 3.01 37.50 PCB 153 µg/kg 7.48 29.64 PCB 138 µg/kg 5.85 30.28 PCB 180 µg/kg 5.31 17.18 Arsenic [As] mg/kg 6.65 41.05 Cadmium [Cd] mg/kg 1.85 34.10 Chromium [Cr] mg/kg 90.50 693.10 Copper [Cu] mg/kg 46.85 442.50 Lead[Pb] mg/kg 55.25 318.10 Mercury [Hg] mg/kg 1.50 3.25 Nickel [Ni] mg/kg 37.05 235.95 Zinc [Zn] mg/kg 355.85 2,845.00 Σi-PCBs µg/kg 23.95 158.67 Exposure conditions: Table A. 5: Exposure conditions over 21 days exposure Temperature Conductivity Ammonium Hardness Suspended Treatment O [mg/L] pH Nitrate [mg/L] Nitrite [mg/L] 2 [°C] [µS] [mg/L] [°dH] particles [g/L] Luppe 11.3 ± 1.4 7.7 ± 0.5 7.8 ± 0.1 440.7 ± 39.8 0.3 ± 0.2 8.7 ± 3.6 0.07 ± 0.03 7.0 ± 0.5 0.9 ± 0.7 Luppe 1/2 11.4 ± 1.3 7.7 ± 0.4 7.8 ± 0.1 405.4 ± 31.3 0.3 ± 0.2 8.7 ± 3.6 0.1 ± 0.06 7.1 ± 0.3 1.4 ± 0.6 Luppe 1/4 11.5 ± 1.2 7.7 ± 0.4 7.9 ± 0.2 367.1 ± 16.6 0.6 ± 0.2 6.9 ± 2.4 0.1 ± 0.02 7.1 ± 0.3 1.1 ± 0.4 Luppe 1/8 11.4 ± 1.1 7.7 ± 0.4 7.8 ± 0.1 354.2 ± 13.2 0.6 ± 0.2 8 ± 0 0.1 ± 0.01 7.1 ± 0.3 1.6 ± 0.6 Sediment 11.3 ± 1.1 7.7 ± 0.5 7.8 ± 0.05 334.0 ± 6.1 0.5 ± 0.2 8 ± 0 0.1 ± 0.06 7.1 ± 0.3 2.0 ± 0.6 control Water control 11.5 ± 1.0 7.5 ± 0.4 7.8 ± 0.1 336.7 ± 7.0 0.5 ± 0.4 6.2 ± 3.0 0.1 ± 0.02 7.1 ± 0.3 - 167 Annex III III. Appendix: Supplementary information to chapter 5 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 Sampling site Juvenile Stage 0 Stage 1 Stage 2 Stage 3 Stage 4 Stage 5 . age [a] Tench juvenile 6 2 ± 0.9 6 (100 %) male 5 3 ± 0.8 0 0 2 (40 %) 3 (60 %) 0 0 0 female 9 3 ± 1 0 2 (22 %) 2 (22 %) 1 (11 %) 0 4 (44 %) 0 Luppe Roach juvenile 1 1 male 5 2 ± 0 0 0 2 (40 %) 0 3 (60 %) 0 0 female 7 2 ± 0.4 0 2 (29 %) 2 (29 %) 0 3 (43 %) 0 0 Tench juvenile 6 1 ± 0.4 6 (100 %) male 4 3 ± 0.7 0 3 (75 %) 1 (25 %) 0 0 0 0 female 11 3 ± 1.3 0 2 (18 %) 9 (82 %) 0 0 0 0 Laucha Roach juvenile 2 1 2 male 0 ------1 1 female 8 2 ± 0 0 2 (25 %) 4 (50 %) 0 0 (12.5%) (12.5%) Tench juvenile 0 ------male 20 1 ± 0.5 0 2 (10 %) 15 (75%) 3 (15 %) 0 0 0 Culture female 10 1 ± 0.5 0 1 (10 %) 0 9 (90 %) 0 0 0 d fish Roach juvenile 0 ------male 5 2 ± 0 0 5 (100%) 0 0 0 0 0 female 14 2 ± 0 0 5 (36 %) 0 1 (7 %) 7 (50 %) 1 (7 %) 0 168 Annex III 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 Luppe Laucha Cultured Range of severity fish ovary Nb. examinated 9 11 10 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 Tench testis Nb. examinated 5 4 20 inflammation 0 1 (25 %) 0 minimal intersex 1 (20 %) 0 3 (15 %) minimal to moderate ovary 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 Roach testis Nb. examinated 5 0 5 inflammation 1 (20 %) 0 0 minimal 169 Contributions to the published articles and chapters Contributions to the published articles and chapters Chapter 1 This chapter was written by A-KM. Chapter 2 This chapter was written by A-KM. Chapter 3 A-KM has been responsible for the study design and concept of the study, sampling campaigns, drafted the manuscript and performed major parts of the practical work. A-KM conducted the YES assay. A-KM and KL performed the laboratory calibration of the passive samplers and conducted the bioavailability studies. DK contributed to analytical LC-MS/MS measurements, method development and conceptual input regarding passive sampling. KS has contributed to passive sampling study development, data evaluation and manuscript concept. CR performed the p-YES. SB contributed to p-YES bioassay data and evaluation, EV provided guidance and conceptual input for passive sampling. SB, SC and HH were responsible for the study design and supervised the study. All authors read, improved and approved the final manuscript. Chapter 4 A-KM has been responsible for the concept of the study, performed the practical work including sediment sampling, sediment extraction and evaluation in YES bioassay, TOC analysis, passive sampling, fish husbandry, fish exposure experiments, extraction bile and blood, biomarker analysis, RNA-sequencing data evaluation, histological evaluation and drafted the manuscript. NM assisted in fish maintenance, with the exposure study and laboratory biomarker evaluation. A-KM and KL performed the laboratory calibration of the passive samplers and extraction of passive samplers. DK contributed to analytical LC-MS/MS measurements, method development and conceptual input regarding passive sampling. CR and SB did the p-YES measurements and data evaluation. LG, ATA and BD conducted the RNA 170 Contributions to the published articles and chapters sequencing and bioanalytical analysis. MB provided conceptional input to the study design and conducted the PBTK modelling. HS provided guidance and conceptual input for the study design and evaluation of liver histology. SC, SS and HH were responsible for the study design, and supervised the study. All authors read, improved and approved the final manuscript. Chapter 5 A-KM has been responsible for the concept of the study, sampling campaigns, drafted the manuscript and performed major parts of the practical work including sediment, water, biota extraction, LC-MS/MS measurement, YES bioassay, histological analysis, vtg biomarker measurement and evaluation. NM contributed to sampling and performed 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. Chapter 6 This chapter was written by A-KM. A-KM: Anne-Katrin Müller; NM: Nele Markert; KL: Katharina Leser; DK: David Kämpfer; KS: Killian Smith; SS: Sabrina Schiwy; CR: Carolin Riegraf; SB: Sebastian Buchinger; EV: Etienne Vermeiersen; AS: Andreas Schäffer; LG: Lin Gan; ATA: Ali T. Abdallah; BD: Bernd Denecke; HS: Helmut Segner; MB: Markus Brinkmann; SC: Sarah E. Crawford; HH: Henner Hollert 171 Acknowledgement Acknowledgement An dieser Stelle möchte ich mich bei allen von Herzen bedanken, die mich während meiner Promotion unterstützt und damit maßgeblich zum Abschluss dieser Arbeit beigetragen haben. Zuallererst gilt mein Dank Prof. Dr. Henner Hollert für die Möglichkeit, zu diesem spannenden Thema unter seiner Leitung die Doktorarbeit anfertigen zu können. Des Weiteren möchte ich mich für seine stetige positive Unterstützung, das entgegengebrachte Vertrauen, das Einbringen seiner fachlichen Expertise und die stets offenen und motivierenden Gespräche bedanken. Besonders zu wertschätzen wusste ich auch die vielen Tagungsbesuche, Fortbildungen und Kooperationen, die er mir während meiner Doktorarbeit ermöglicht hat. Bei Prof. Dr. Andreas Schäffer möchte ich mich besonders für das eingebrachte Fachwissen bezüglich der chemischen Analytik und umweltchemischen Prozesse und für die bereitwillige Übernahme des Korreferates dieser Doktorarbeit bedanken. Des Weiteren möchte ich mich herzlich bei Prof. Dr. Helmut Segner in seiner Funktion als Drittgutachter für das Interesse an dieser Arbeit bedanken. Besonderen Dank gilt seinen konzeptionellen Vorschlägen und der fachlichen Unterstützung zur Endokrinologie der Fische sowie histologischen Auswertung. Einen besonderen Dank möchte ich an Dr. Sarah Crawford aussprechen, für ihre großartige Betreuung meiner Arbeit, eine bessere Betreuung hätte man sich kaum wünschen können. Vielen Dank für die wertvollen fachlichen Diskussionen, deine positive Art und gute Zusprache in schwierigen Situationen. Natürlich auch vielen Danke dafür, dass du mit mir auf Sedimentprobenahme an die Luppe gefahren bist und mit gemeinsamer Muskelkraft die Sedimente ausgehoben hast. Gleichermaßen möchte ich mich bei Dr. Markus Brinkmann bedanken für seine starke fachliche Unterstützung zu dieser Arbeit. Vielen Dank, dass du mich auch weiter aus Kanada in dieser Arbeit begleitet hast und ich die von dir aufgebauten Fischexpositionsanlagen benutzen durfte. Dr. Killian Smith möchte ich herzlich danken für seine geduldige Unterstützung in der Etablierung der passive sampling Systeme und der Hormonanalytik. Besonders bedanken möchte ich mich für seine stets offene Tür und Ohren für die vielen Fragen und kleineren Probleme des Laboraltages. Einen herzlichen Dank möchte ich auch Dr. Sebastian Buchinger und Carolin Riegraf von der Bundesanstalt für Gewässerkunde in Koblenz und Dr. Etienne Vermeiersen von dem Ökotoxzentrum an der EAWAG in Zürich aussprechen für die positive und gelungene Zusammenarbeit und ihre Auswertung diverser Proben mittels des p-YES bzw. konstante Unterstützung bezüglich der passive sampling Systeme. Zudem danke ich allen Kollegen und Partnern aus dem Projekthaus Wasser für ihre Zusammenarbeit und Unterstützung bei dieser Arbeit. Besonderen Dank gilt hierbei Prof. Dr. Holger Schüttrumpf und dem Team des Institutes für Wasserbau und Wasserwirtschaft für die gute Zusammenarbeit. Und dem Programm Exploratory Research Space (ERS) der RWTH-Aachen Universität sowie der Deutschen Forschungsgemeinschaft als Teil der Deutschen Exzellenzinitiative für die Förderung des Projekthaus Wassers. Timothy Rosenberger, Markus Schmitz, Nele Marker und Katharina Leser möchte ich danken, dass sie ihre Abschlussarbeiten im Rahmen dieses Projektes und in Anlehnung an meine Doktorarbeit gemacht haben. Ganz besonders möchte ich dabei Katharina Leser und Nele Markert danken, dafür, 172 Acknowledgement dass ihr immer zu mir gehalten habt und uns gegenseitig motivieren konnten während der langen Laborarbeitstage und Feldkampanien. Es hat mir viel Freude gemacht mit Euch zusammenzuarbeiten. Bei David Kämpfer bedanke ich mich sehr für seine umfangreiche Hilfe bei der analytischen Methodenentwicklung und Beratung in Sachen passive sampling. Es hat mir stets Freude bereitet, mit dir über die Herausforderungen der Hormonanalytik zu diskutieren. Auch möchte ich mich bei allen Kolleginnen und Kollegen am Institut für Umweltforschung der RWTH Aachen für die gegenseitige Unterstützung im Laboralltag, vor allem den fleißigen Helfern beim Sezieren der vielen Fische, während gemeinsamer Konferenzen und in der Lehre bedanken sowie für die lustige Zeit auf den zahlreichen sozialen Veranstaltungen. Besonderes danken möchte ich für ihre positive Art und tatkräftige Hilfe bei dieser Arbeit Irina Politowski, Dr. Felix Stibany, Dr. Leonie Mueller, Henriette Alert-Meyer, Alexander Bach, Dr. Jake Quellet, Dr. Martina Roß-Nikoll, Dr. Richard Ottermanns und Dr. Sabrina Schiwy. Ein riesiges Dankeschön geht an Peter Hostenbach und sein Team aus der mechanischen Werkstatt für ihren unermüdlichen Einsatz bei kleinen und größeren Überschwemmungen, den vielen Reparaturen und dem Austauschen aller Abflussrohre. Gleicherweise danke ich Uwe Mohnen von der Aquakultur Mohnen in Stolberg für seine Hilfe und Einweisung in die Haltung der Forellen. Ein großes Dankeschön geht an die Mitarbeiter des Sekretariats, Leonie Lubczyk und Dennis Goebele, das IT-Team, Edith Schröder und Dr. Richard Ottermanns, und die Labortechnischen Assistentinnen, Simone Hotz und Brigitta Goffart, dafür, dass ihr mir den Rücken in so vielen Dingen freigehalten habt. Nicht zuletzt möchte ich mich bei meinen Freunden Dr. Nora Niehus, Christiane Herold, Dr. Elena Herweg, Ronja Kossak, Eva Prensler und Brigitta Goffart für die unvergessliche Zeit während des Studiums und der Doktorarbeit bedanken. Auf euch war immer Verlass, ob zum Feiern, Lachen, Spaß haben oder auf eure Hilfe und Unterstützung in schwierigen Zeiten. Meiner Familie, meinen Eltern, Marie Adams und Maximilian Müller möchte ich von ganzem Herzen Danke sagen. Durch euren unerschütterlichen Rückhalt und Glauben an mich ist diese Arbeit erst zustande gekommen. 173 Curriculum vitae Curriculum vitae Personal information Name Anne-Katrin Müller Born September 4th , in Goch, Germany Citizenship Germany Language skills German (native), English (proficiency), Spanish (basic), Latin (basic) Address Fritz-Bauer-Str. 21, D- 72074 Tübingen E-Mail [email protected] Academic education Since April 2016 PhD research fellow at the Department of Ecosystem Analysis, Institute for Environmental Research, RWTH Aachen University January 2015 Graduation in Biology with emphasis on environmental research and ecotoxicology (MSc.). Topic of the Master thesis: Ultrastructural alterations and genotoxicity in yellow perch (Perca flavescens) from a biological mercury hotspot in Nova Scotia, Canada. Supported by ERASMUS Placement and PROMOS grants. Supervised by Prof. Dr. Henner Hollert an Prof. Dr. Karen Kidd at the Institute for Environmental Research, RWTH Aachen University and Canadian Rivers Institute, University of New Brunswick. September 2012 – January 2015 Studies of Biology with emphasis on environmental research, ecotoxicology and plant physiology. July 2011 Graduation in Biology with emphasis on environmental research (BSc.). Topic of the Master thesis: Assessing the impact of genetically engineered MON810 maize pollen on European lepidopteran species: Aspects of a risk assessment. Supervised by Dr. Stephan Rauschen and Prof. Alan Slusarenkow. September 2008 – July 2011 Studies of Batchelor Biology with emphasis on plant physiology Professional experience August 2015 – April 2016 Research scientist at the Federal Institute for Risk Assessment, Berlin, Germany. Topic: Data availability and quality assessment under REACH 174 Curriculum vitae March 2015 – August 2015 Research scientist at Gaiac, Research Institute for Ecosystem analysis and assessment. Topic: aquatic mesocosm studies (GLP), benthic invertebrate assessment (WFD) October 2012 – August 2013 Research assistant at the Institute for Environmental Research and the Institute of Plant Physiology, RWTH-Aachen University. Tasks: Bioassay for detection of endocrine disruptors, Molecular biological practice: RNA, DNA isolation, qPCR, genotyping October 2011 – March 2012 Research Internship in the department of biodiversity and environmental management at the federal research facility Agroscope Reckenholz-Tänikon (ART), Zurich, Switzerland. August 2011 – Oktober 2011 Research Internship at the Canadian Rivers Institute, University of New Brunswick, New Brunswick, Canada, funded by the Undergraduates Research Opportunities Program (UROP) of the RWTH Aachen University, supervised by Prof. Dr. Karen Kidd. April – September 2010/ March – April 2011 Undergraduate research assistant at the Institute of Plant Physiology RWTH-Aachen University. Tasks: Field mapping, GIS, care of several insect breeding Recognitions Travel and registration grant of the Society of Environmental Toxicology and Chemistry (SETAC) Europe for participating at the 29th annual meeting in Helsinki DAAD founding for participating at the 27th annual meeting of SETAC Europe in Brussels Student Travel Award of the Society of Environmental Toxicology and Chemistry (SETAC) North America to attend the 35th SETAC annual meeting in Vancouver Erasmus placement for a research stay at the University of Bern, Switzerland, as part of the master thesis UROP-abroad grant for the project „Effects of human activities on aquatic ecosystems“ at the Canadian Rivers Institute, Canada German corn committee’s junior scientists award 2011 for the Bachelor thesis „Assessing the impact of genetically engineered MON810 maize pollen on European lepidopteran species: Aspects of a risk assessment“, October 2011 UROP grant for the undergraduate research project „Functional analysis of cell wall mutants in the non-host interaction between Arabidopsis thaliana and Phakopsora pachyrhizi”, Februar 2011 Memberships 175 Curriculum vitae Since 2017 Associated member of the student´s advisory council of the Society of Environmental Toxicology and Chemistry Since 2017 DIN working group for bioassay development for endocrine disruptors Since 2017 NORMAN working group II on Bioassays Since 2014 Society of Environmental Toxicology and Chemistry Scientific contributions Research articles in international peer-reviewed journals *Publications contributing to this thesis are highlighted with asterisks. *Müller A-K, Markert N, Leser K, Kämpfer D, Crawford S E, Schäffer A, Segner H, Hollert H (2019) Assessing endocrine disruption in freshwater fish species from a “hotspot” for estrogenic activity in sediment. Environmental Pollution. Doi: 10.1016/j.envpol.2019.113636 Aumeier BM, Graula H, Müller A-K, Lackmann C, Wünsch R, Wintgens T, Hollert H, Wessling M (2020) Self-sustaining and versatile decentralized water treatment: process and piloting. Water Research. Doi: 10.1016/j.watres.2019.115338 *Müller A-K, Leser K, Kämpfer D, Riegraf C, Crawford SE, Smith K, Vermeirssen E, Buchinger S, Hollert H (2019) Bioavailability of estrogenic compounds from sediment in the context of flood events evaluated by passive sampling. Water Research. doi: 10.1016/j.watres.2019.06.020 Deutschmann B., Müller A-K., Hollert H., Brinkmann M. (2019) Assessing the fate of brown trout (Salmo trutta) environmental DNA in a natural stream using a sensitive and specific dual-labelled probe. Sci Total Environ 655:321–327. doi: 10.1016/j.scitotenv.2018.11.247 Crawford S. E., Cofalla C., Aumeier B., Brinkmann M., Classen E., Esser V., Ganal C., Kaip E., Häussling R., Lehmkuhl F., Letmathe P., Müller A-K, Rabinovitch I., Reicherter K., Schwarzbauer J., Schmitt M., Stauch G., Wessling M., Yüce S., Hecker M., Kidd K., Altenburger R., Brack W., Schüttrumpf H., Hollert H. (2017) Project house water: a novel interdisciplinary framework to assess the environmental and socioeconomic consequences of flood-related impacts. Environmental Sciences Europe 29:23. doi: 10.1186/s12302-017-0121-1 Müller A-K., Brinkmann M., Baumann L., Stoffel M., Segner H., K. Kidd and H. Hollert (2015) Morphological Alterations in the Liver of Yellow Perch (Perca flavescens) from a Biological Mercury Hotspot. Environmental Science and Pollution Research 05/2015, DOI:10.1007/s11356-015-4177-4 Schuppener M., Mühlhause J., Müller A-K., Rauschen S. (2012) Environmental risk assessment for the small tortoiseshell Aglais urticae and a stacked Bt-maize with combined resistances against Lepidoptera and Chrysomelidae in central European agrarian landscapes. Molecular Ecology. (18):4646-62 Other journals: Angelika Oertel, Katrin Maul, Jakob Menz, Anna Lena Kronsbein, Dana Sittner, Andrea Springer, Müller A-K, Uta Herbst, Kerstin Schlegel, Agnes Schulte REACH Compliance: Data availability in REACH registrations Part 2: Evaluation of data waiving and adaptations for chemicals ≥ 1000 tpa. Umweltbundesamt. ISSN 1862-4804 Müller A-K., Schuppener M., Rauschen S. (2012) Assessing the impact of Cry1Ab expressing corn pollen on larvae of Aglais urticae in a laboratory bioassay. IOBC/wprs Bulletin Vol. 73, 2012 pp. 55- 60 176 Curriculum vitae Müller A-K., Schuppener M., Rauschen S. (2012) Schädigt MON810 den Kleinen Fuchs? Mais. Ausgabe 02/2012, pp. 86-87 Platform presentations Schmitz M, Müller A-K, Markert N, Ganal C, Crawford SE, Brinkmann M, Schiwy S, Schüttrumpf H, Hollert H (2019) Mobilization of estrogenic compounds from sediment during a simulated flood like event. SETAC NA Annual Meeting 2019, Toronto, Canada. Müller A-K, Markert N, Leser K, Crawford SE, Kämpfer D, Schiwy S, Gan L, Abdallah A, Denecke B, Schüttrumpf H, Segner H, Brinkmann M, Hollert H (2019) Toxic floods Impact s of re mobilized endocrine disruptors from sediments in rain bow trout (Oncorhynchus mykiss). SETAC German Language Brand Annual Meeting 2019, Landau, Germany. Markert N, Müller A-K, Segner H, Hollert H. (2019) Effects of Endocrine Disruptors in Sediments of the River Luppe assessed by Biomarker and Histopathological Analysis of Freshwater Fish. SETAC German Language Brand Annual Meeting 2019, Landau, Germany. Müller A-K, Markert N, Leser K, Crawford SE, Schüttrumpf H, Segner H, Brinkmann M, Hollert H (2019) Toxic floods Impact s of re mobilized endocrine disruptors from sediments in rain bow trout (Oncorhynchus mykiss). SETAC Europe Annual Meeting 2019, Helsinki, Finnland. Müller A-K, Markert N, Segner H, Hollert H (2019) Effects of Endocrine Disruptors in Sediments of the River Luppe assessed by Biomarker and Histopathological Analysis of Freshwater Fish. Young Environmental Scientist Meeting, Ghent, Belgium. Schmitz M, Müller A-K, Markert N, Ganal C, Crawford SE, Brinkmann M, Schiwy S, Schüttrumpf H, Hollert H (2019) Mobilization of estrogenic compounds from sediment during a simulated flood like event. Young Environmental Scientist Meeting, Ghent, Belgium. Markert N, Müller A-K, Segner H, Hollert H. (2017) Effects of Endocrine Disruptors in Sediments of the River Luppe assessed by Biomarker and Histopathological Analysis of Freshwater Fish. Young Water Researchers Symposium, Aachen, Germany. Leser, K, Müller A-K, Crawford SE, Kämpfer D, Markert N, Smith K, Vermeirssen E, Hollert H (2017) Assessment of Bioavailability and Effects of Endocrine Disruptors in Sediments of the River Luppe on Organisms in the Sediment and Water-Phase. Young Water Researchers Symposium, Aachen, Germany. Müller A-K, Brinkmann M., Baumann L., Stoffel M., Segner H., K. Kidd and H. Hollert (2015) Ultrastructural alterations in yellow perch (Perca flavescens) from a biological mercury hotspot in Nova Scotia, Canada. Young Environmental Scientist Meeting, Petnica, Serbia. Poster presentations Schmitz M, Müller A-K, Crawford SE, Ganal C, Crawford SE, Brinkmann M, Schiwy S, Schüttrumpf H, Hollert H (2018) Floodhydrotox – an interdisciplinary approach to assess the endocrine disrupting potential of sediments during flood events. SETAC German Language Brand Annual Meeting 2018, Münster, Germany. Markert N, Müller A-K, Segner H, Hollert H. (2018) Effects of Endocrine Disruptors in Sediments of the River Luppe assessed by Biomarker and Histopathological Analysis of Freshwater Fish. SETAC German Language Brand Annual Meeting 2018, Münster, Germany. Markert N, Müller A-K, Segner H, Hollert H. (2017) Effects of Endocrine Disruptors in Sediments of the River Luppe assessed by Biomarker and Histopathological Analysis of Freshwater Fish. SETAC German Language Brand Annual Meeting 2017, Neustadt an der Weinstraße, Germany. 177 Curriculum vitae Leser, K, Müller A.- K, Crawford SE, Kämpfer D, Markert N, Smith K, Vermeirssen E, Hollert H (2017) Assessment of Bioavailability and Effects of Endocrine Disruptors in Sediments of the River Luppe on Organisms in the Sediment and Water-Phase. SETAC German Language Brand Annual Meeting 2017, Neustadt an der Weinstraße, Germany. Müller A-K, Brinkmann M, Crawford SE, Cofalla C, Schüttrumpf H, Lehmkuhl F, Schulze T, Brack W, Hollert H (2017) Floodhydrotox – an interdisciplinary approach to assess the endocrine disrupting potential of sediments during flood events. SETAC Europe Annual Meeting 2017, Brussels, Belgium. Müller A-K, Brinkmann M, Crawford SE, Cofalla C, Schüttrumpf H, Lehmkuhl F, Schulze T, Brack W, Hollert H (2017) Floodhydrotox – an interdisciplinary approach to assess the endocrine disrupting potential of sediments during flood events. Young Environmental Scientist Meeting, Stockholm, Sweden. Müller A-K, Brinkmann M, Crawford SE, Cofalla C, Schüttrumpf H, Lehmkuhl F, Schulze T, Brack W, Hollert H (2016) Floodhydrotox – an interdisciplinary approach to assess the endocrine disrupting potential of sediments during flood events. SETAC German Language Brand Annual Meeting 2016, Tübingen, Germany. Oertel A, Müller A-K, Springer A, Sittner D, Herbst U, Schulte A (2016) REACH Compliance: Data availability in REACH registrations. SETAC Europe Annual Meeting 2016, Nantes, France. Müller A-K, Oertel A, Springer A, Sittner D, Herbst U, Schulte A (2016) REACH Compliance: Data availability in REACH registrations. German Society of experimental and clinic Pharmacology and Toxicology (DGPT), Berlin, Germany. Müller A-K, Brinkmann M., Baumann L., Stoffel M., Segner H., K. Kidd and H. Hollert (2014) Ultrastructural alterations in yellow perch (Perca flavescens) from a biological mercury hotspot in Nova Scotia, Canada. SETAC NA Annual Meeting 2014 in Vancouver, Canada. Müller A-K, Brinkmann M., Baumann L., Stoffel M., Segner H., K. Kidd and H. Hollert (2014) Ultrastructural alterations in yellow perch (Perca flavescens) from a biological mercury hotspot in Nova Scotia, Canada. SETAC Europe Annual Meeting 2014 in Basel, Switzerland. Müller A-K, Brinkmann M., Baumann L., Stoffel M., Segner H., K. Kidd and H. Hollert (2014) Ultrastructural alterations in yellow perch (Perca flavescens) from a biological mercury hotspot in Nova Scotia, Canada. SETAC German Language Brand Annual Meeting 2014 in Gießen, Germany. Müller A-K, Schuppener M, Rauschen S (2011) Assessing the impact of Cry1Ab expressing corn pollen on larvae of Aglais urticae in a laboratory bioassay. Environmental Impact of Genetically Modified Crops: European Experience, České Budějovice, Check Republic. 178