Diss. ETH No. 15560

Effects of Treated Wastewater on Trout: A Case Study of a Swiss River

A dissertation submitted to the

SWISS FEDERAL INSTITUTE OF TECHNOLOGY, ZURICH

for the degree of

Doctor of Natural Sciences

presented by

BERND KOBLER

Dipl. Biol., Universität Bern

born March 13, 1972

citizen of Aegerten (BE), Switzerland

accepted on the recommendation of

Prof. Dr. Alexander Zehnder, examiner

Prof. Dr. H. Segner, co-examiner

Dr. A. Peter, co-examiner

Luzern, 2004 Copyright © 2004 by Bernd Kobler, EAWAG Kastanienbaum

All rights reserved. No part of this book may be reproduced, stored in a retrieval system or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, without the prior written permission of the copyright holder.

First Edition 2004 Für Nicole

Chapter 5 of the thesis has been published:

Aerni, H.R., B. Kobler, B.V. Rutishauser, F.E. Wettstein, R. Fischer, W. Giger, A. Hungerbühler, M.D. Marazuela, A. Peter, R. Schönenberger, A.C. Vögeli, M.J.-F Suter & R.I.L. Eggen, 2004. Combined biological and chemical assessment of estrogenic activities in wastewater treatment plant effluents. Anal. Bioanal. Chem. 378: 688-696.

3 manuscripts have been submitted for publication:

Kobler, B., F. Wettstein & A. Peter. How is the quality of brown trout (Salmo trutta Fario L.) in a wastewater polluted Swiss river? Submitted to Hydrobiologia. (Chapter 2)

Kobler, B., T. Luckenbach & A. Peter. Mortality of brown trout (Salmo trutta Fario L.) embryos in a wastewater polluted Swiss river. Submitted to Aquatic Toxicology. (Chapter 3)

Kobler, B. & A. Peter. Stress response of a brown trout population in the river Wyna to treated wastewater inputs. Submitted to Hydrobiologia. (Chapter 4)

Table of contents

Summary...... 1

Zusammenfassung...... 3

1 INTRODUCTION...... 5

2 HOW IS THE QUALITY OF EGG OF BROWN TROUT (SALMO TRUTTA

FARIO L.) IN A WASTEWATER-POLLUTED SWISS RIVER?...... 13

3 MORTALITY OF BROWN TROUT (SALMO TRUTTA FARIO L.) EMBRYOS IN

A POLLUTED SWISS RIVER...... 29

4 STRESS RESPONSE OF A BROWN TROUT POPULATION IN THE RIVER

WYNA TO TREATED WASTEWATER INPUTS...... 45

5 COMBINED BIOLOGICAL AND CHEMICAL ASSESSMENT OF ESTROGENIC

ACTIVITIES IN WASTEWATER TREATMENT PLANT EFFLUENTS...... 69

6 SYNOPSIS...... 89

Danksagung...... 97

Curriculum Vitae...... 99

1

Summary

Today, wastewater from treatment plants is discharged in almost every large river in Switzerland. Wastewater can cause significant problems to fish such as impairing their health and survival. High pollutant load even threatens the existence of entire fish populations.

- + Thus, the present thesis focused on conventional toxicity (NO2, NH4 ) and hormone activity of treated wastewater and its effects on fish. Effects from habitat degradation due to siltation with fine sediments were considered, too.

Between 1999 and 2000, a case study on wild brown trout (Salmon trutta fario L.) was conducted. As a study river, the Wyna located in central Switzerland was chosen. Its headwater section is not contaminated with wastewater. Further downstream, effluents from four wastewater treatment plants affect the chemical composition of the water. At low flow, wastewater addition can reach as much as 50% of the water volume at the stream’s mouth. In order to analyze the effect of treated wastewater on trout individuals and their entire population, differently polluted river sections were investigated and compared. Results show that the population’s recruitment was lowest in the most polluted river section (Chapter 4). The low recruitment success also influenced the structure and density of the entire population. Due to missing young fish, the structure was over-aged and the density low. The low density in its turn increased the availability of resources such as food for trout. Since food intake regulates the growth of fish, the excess in food could explain the fast growth and the high fecundity of trout from the polluted sites.

Either the survival of early stages in the life cycle, or the reproduction of adult species were suspected to create a crucial bottleneck in the life cycle of Wyna trout. Surveys and experiments in the river showed that not a low reproductive ability of adult females (Chapter 2), but the low survival of young trout (Chapter 3), was the major cause for this recruitment failure. The river water was found to be acutely toxic for fish due to temporarily high concentrations of ammonia. In addition, measurements in the river indicated a strong estrogenic activity with a potential for damaging eggs and embryos. Concurrently, the quality of spawning habitat for trout was low particularly due to high siltation rates of fine sediments. Poor water quality and siltation thus were found to cause the death of more than 95% of eggs and embryos, which explains the low recruitment in the polluted section of River Wyna. Surveys and experiments on the population-level and individual-level presented in this thesis show that fish should be used to evaluate anthropogenic impacts such as water pollution and degradation of spawning habitats, since they are integrative indicator of complex stress. But, 2 recruitment failure - such as found for the Wyna population in the wastewater polluted river section - would result in a collapse of the population. Most likely due to immigration from upstream sites and regular stocking with trout, such a collapse of trout populations was prevented.

Another topic addressed in this thesis is a possible estrogenic contamination from treatment plants, since endocrine disruption is discussed as a possible cause of declines in fish populations worldwide. In chapter 5, the estrogenic activity of several wastewater treatment plant effluents in Switzerland and France were assessed. Chemical analysis and biological effect analysis were combined, in vitro as well as in vivo. This investigation was part of a European community program of research (COMPREHEND) into environmental hormones and endocrine disruptors and was conducted as a cooperation between different EAWAG research departments (fish ecology, chemistry and microbiology). Since wastewater is a complex mixture of different chemicals it is not clear which one might cause significant estrogenicity in wildlife. The work showed that natural and synthetic steroid hormones are major contributors to estrogenicity in the effluents tested.

From the results of this thesis, it is concluded that treated wastewater impairs fish from the individual up to the population-level and should be seen as a major burden on river systems. 3

Zusammenfassung

Viele grosse Fliessgewässer der Schweiz werden mit Abwasser aus Kläranlagen belastet. Abwasser kann den Gesundheitszustand und das Überleben von Fischen beeinträchtigen. Eine starke Belastung kann so das Überleben der gesamten Population gefährden. Die vorliegende Arbeit untersuchte deshalb Auswirkungen von gereinigtem Abwasser auf Fische.

- + Es wurde die Wirkung der Toxizität (NO2, NH4 ) und Hormonaktivität erfasst, zusätzlich auch die Zerstörung von Lebensräumen durch die Ablagerung von Feinsedimenten untersucht.

Zwischen 1999 und 2000 wurde eine Fallstudie mit Bachforellen (Salmon trutta fario L.) realisiert. Die Studie wurde in der Wyna, einem Fluss im Zentrum der Schweiz, durchgeführt. Die Wyna ist im Oberlauf nicht belastet, weiter flussabwärts hingegen werden Abwässer aus insgesamt 4 Kläranlagen eingeleitet. Bei niedrigem Abfluss kann der Anteil im Mündungsbereich bis zu 50% erreichen. Um Auswirkungen von Abwasser auf Bachforellen und deren Population zu beurteilen, wurden unterschiedlich belastete Flussabschnitte untersucht und verglichen. Die Resultate zeigen, dass der Fortpflanzungserfolg der Population im stark belasteten Flussabschnitt am niedrigsten war (Kapitel 4). Der geringe Fortpflanzungserfolg seinerseits beeinflusste die gesamte Population. Die Dichte war niedrig und der Altersaufbau überaltert. Die geringe Dichte wiederrum verbesserte die Verfügbarkeit der Nahrung. Mit zunehmender Futteraufnahme wachsen Fische schneller. So kann der Futterüberschuss die hohe Wachstumsrate und Fruchtbarkeit von Forellen im belasteten Flussabschnitt so wie es in der Untersuchung festgestellt wurde, erklären.

Die frühe Embryonalentwicklung und die Fortpflanzung geschlechtsreifer Tiere wurden verdächtigt, kritische Phasen im Lebenszyklus der untersuchten Forellen zu sein. Untersuchungen und Experimente im Fluss zeigten aber, dass nicht eine geringe reproduktive Leistung der geschlechtsreifen Weibchen (Kapitel 2), sondern die niedrigen Überlebensraten junger Forellen (Kapitel 3) hauptverantwortlich für den Reproduktionsausfall waren. Aufgrund zeitweise hoher Ammoniak-Konzentrationen war das Wasser der Wyna akut fischtoxisch. Weitere Messungen im Fluss lieferten auch Hinweise einer starken Belastung mit östrogen-aktiven Substanzen, welche möglicherweise ebenfalls Eier und Embryos schädigten. Daneben war aufgrund hoher Sedimentationsraten die Qualität der Laichhabitate für Bachforellen stark beeinträchtigt. Beides, die schlechte Wasserqualität und die Sedimentation, verursachten den Tod von rund 95% der experimentell exponierten Eier und Embryos. 4

Ein Reproduktionsausfall, so wie er für die untersuchte Population im abwasserbelasteten Abschnitt festgestellt wurde, hätte deren Zusammenbruch verursachen müssen. Der Zusammenbruch der Forellenpopulation wurde wahrscheinlich durch regelmässigem Besatz und Immigration von Bachforellen aus dem Oberlauf verhindert. Beobachtungen und Experimente der vorliegenden Arbeit zeigen, dass man Fische dazu benutzen kann, den Einfluss von Wasserverschmutzung und Lebensraumzerstörung zu erfassen.

Kapitel 5 befasst sich mit der Belastung durch östrogen-aktive Stoffe aus Kläranlagen. Solche Stoffe werden weltweit als eine mögliche Ursache von Fischrückgängen diskutiert. Es wurde die östrogene Aktivität von Abwasser aus verschiedenen Kläranlagen in der Schweiz und in Frankreich untersucht. Chemische und biologische Analysen wurden in vitro als auch in vivo angewendet. Diese Untersuchungen waren Teil eines Europäischen Forschungsprogramms über Hormone in der Umwelt und östrogen-aktive Stoffe (COMPREHEND) und wurden als Kooperation verschiedener EAWAG-Forschungsgruppen (Fischökologie, Chemie und Mikrobiologie) durchgeführt. Die Arbeit zeigt, dass natürliche und synthetische Steroidhormone einen grossen Beitrag zum östrogenen Potential getesteter Abwässer beisteuern und so auch eine grosse Belastung für Fliessgewässer darstellen können. 5

1 INTRODUCTION

Water pollution and degradation of aquatic fish habitats

Water pollution (Kendall et al., 1998; Rao, 1999; Adams, 1989; Clement, 2000) and degradation of aquatic habitats (Elliott, 1994) impair fish populations. For trout, water pollution can increase mortality rates and decrease the reproductive success (Kime, 1998). The anthropogenic destruction of aquatic environments can also explain adverse effects on fish (Elliott, 1994). Thus, both pollution and loss of fish habitats endanger the existence of entire trout populations. Because of multiple stressors, conclusions regarding the human effects on fish populations are difficult to achieve. In addition, trout normally produce many more offspring than needed to guarantee the recruitment and survival of the population. If the population’s density is high, a large number of offspring die early due to the limitation of resources such as food or habitat. In cases of low population density and no resource limitation, the overproduction of young trout can compensate adverse environmental effects on fish survival. It remains unclear so far, what levels of anthropogenic stress can be compensated and which critical limits mark a situation which endangers the existence of the entire population. Empirical data from wild populations are required to understand the response to such stress on the level of a population.

Therefore, the major objective of this thesis was to collect data of different trout populations, on structure, size, and recruitment success in order to analyze effects from human activities like discharge of treated wastewater on trout and to contribute thereby to an understanding of the critical limits in anthropogenically impacted fish populations.

Fish population data as integrative indicators of environmental stress

Since fish are an economically important resource, many societies share an interest in protecting them. Fish are studied on multiple organizational scales from the individual up to the population level, in order to identify adverse environmental stress factors. The interpretation of information on the population level is an especially important, but particularly difficult, task because the complexity of biological systems increases with its organizational scale. Colby (1984) developed an approach to overcome this problem. He measured diagnostic population-variables such as the population’s density, biomass, mean age, recruitment of young of the year, as well as the individual’s growth rate, age at maturity, and fecundity. All these variables strongly depend on the size of the population. The population 6 size depends on the amount of available resources, also called carrying capacity of a habitat. The size of a population normally stands in a balance with the carrying capacity. An increase of the population size relative to the carrying capacity decreases the availability of resources for individuals. Fish can respond in characteristic manners to such changes in resource availability per individual. Increased growth rates and fecundity are stimulated by increased resource availability with a decrease in resources causing opposite effects.

Using this information on the population’s responses, Colby (1984) was able to define patterns of population variables which are characteristic for problems on the population level such as decreased population size due to recruitment failure or food limitation. Gibbons & Munkittrick (1994) later extended Colby’s approach and developed a conceptual framework to diagnose specific population problems or responses, such as exploitation, multiple stressors, food limitation, and recruitment failure to analyze the level of stress in fish. This had already been successfully applied to fish living in different aquatic habitats such as lakes, reservoirs (e.g. Gibbons & Munkittrick, 1994), and large (McMaster et al. 1996) and small rivers (Fitzgerald et al., 1999), in order to analyze adverse anthropogenic impacts on the population level. Based on these studies, fish became an integrative indicator for the analysis of ecological problems in aquatic systems.

Nevertheless, this approach has serious limits for detecting the origin of a population’s stress, because the information based on the population level does not allow to diagnose the effect of single stressors. Population data are useful indicators to screen for general symptoms of stress. For more specific information investigations or experiments in the field and laboratory addressing physiological characteristics are required. The network declining fish yields Switzerland (Fischnetz) concentrated on such specific information from fish (Projekt Fischnetz, 2004). As a complement analyzing and understanding crucial phases in the fish life cycle offers a promising approach.

The life cycle of trout

In autumn, the life cycle of trout starts with the reproduction and spawning of eggs into the gravel of the riverbed. During winter the embryos develop until they hatch in spring and leave the riverbed (emergence). In spring and summer until late autumn, trout feed and grow. After two to three growing seasons, trout can reach sexual maturity. With the reproduction of the new generation, the cycle closes again (Figure 1). 7 ·~ Sub-adult Juvenile

Figure 1: Hatching, emergence, dispersal, growth, sexual maturity and reproduction characterize the life cycle of brown trout. Early survival and fertility of mature trout as sensitive phases to stress were studied in this dissertation (chapters 3 & 4).

In this cycle, a number of different phases can be investigated. The survival of early stages and the reproductive success of adult spawners are most informative indicators for the sustaining of the population (e.g. Elliott, 1994 ). Shuter (1990), thus proposes to concentrate on these early and late phases in the fish life cycle when investigating adverse effects on populations due to human activities. Such investigations of sensitive life stages have the potential to detect effects on the population level as well as diagnose specific causative stress factors.

Effects of wastewater discharge

Wastewater is a complex mixture of more than 100'000 different chemicals and therefore is believed to cause major stress on fish (Sumpter, 1997). Several chemicals, such as ammonia and nitrite, are toxic to fish. Others, such as natural and synthetic hormones from 8 human metabolic waste, medicine or industry, can 1) accumulate in fish, 2) disrupt hormonal control and 3) change the synthesis of sex-specific proteins (Bätscher et al., 1999; Cleveland et al., 1999). Such fundamental physiological changes in fish may lead to reproductive dysfunction (Kime, 1995, 1999). Concurrently, wastewater can cause toxic effects and thus initiate high mortality in young trout (McKim, 1977; Von Westernhagen, 1988; Crisp, 1989; Miller, 1993; Brooks et al., 1997). The question then is which one - reproduction or mortality of early life stages - is more sensitive to wastewater and thus represents the bottleneck of the population’s life cycle, thereby causing stress on the population level. Furthermore, fine solids can infiltrate stream gravels and may seriously affect development of trout early life stages. Such a degradation of spawning habitats also needs to be considered, because a loss of fish habitats and spawning areas may be a major draw-back on the population’s reproductive success, structure and density (Lanka et al., 1987; Kozel & Hubert, 1989; Riley & Fausch, 1995).

Key questions and objectives of the thesis

The objective of this dissertation was to diagnose the effects of treated wastewater and habitat degradation on trout individuals and populations. Key questions were:

Are the characteristics of a fish population helpful in diagnosing anthropogenic stress due to wastewater discharge? How important is the water quality for fish individuals and populations? How important is the quality of habitats for fish individuals and populations?

Studies of trout individuals and populations in the Wyna River (Chapters 2–4) The main surveys and studies in this thesis (Chapter 2–4) were conducted in the wastewater-receiving River Wyna located in central Switzerland. It was chosen for two reasons:

Recent studies in the Wyna revealed an impaired trout health status due to discharged wastewater (Escher, 1999; Bernet, 2000; Wahli et al., 1998). Unpolluted control sites can be found upstream of wastewater discharges. Such control sites are important for a comparison with river sites that are impacted by habitat degradation and wastewater. 9

Developments late and early in the life cycle were studied in two separate chapters in this thesis. Chapter 2 describes the reproductive potential of female trout from the River Wyna. The goal was to analyze if wastewater from treatment plants may cause a reproductive dysfunction in trout. Chapter 3 looks at the survival of trout eggs and larvae in the River Wyna. Here, the goal was to analyze if wastewater impairs early survival. The objective in both chapters was to obtain a better idea of whether wastewater may be the cause of a substantial bottleneck in the life cycle of Wyna trout.

In chapter 4, brown trout (Salmo trutta fario L.) populations at river sites with different levels of pollution and habitat degradation are characterized. The objective was to compare population characteristics between different study sites in order to evaluate the impact of wastewater on the population level. The results from Wyna trout populations were critically evaluated against other studies of stressed fish populations, described by Gibbons and Munkittrick (1994).

Estrogen potential of wastewater on individual fish (Chapter 5) The effects of the estrogen potential of wastewater on individual rainbow trout were investigated (e.g. Purdom et al., 1994). This part of the dissertation is not directly linked to the studies conducted in River Wyna. The experiments were part of a European community program of research into environmental hormones and endocrine disruptors (COMPREHEND, 2001) and were conducted as a cooperation between different EAWAG research departments (chemistry, microbiology and fish ecology). The investigation included:

the in vivo (trout exposure) endocrine activity, the in vitro (yeast estrogen screen) endocrine activity and the chemical composition (steroids and nonylphenolethoxylates) of wastewater from five WWTPs in Switzerland and France. The objective was to correlate the chemical pollution between an in vitro test system and effects on fish to find chemicals responsible for endocrine disruption in fish.

Synopsis (Chapter 6) The synopsis discusses the effects of wastewater and fine solid infiltration on trout. Population characteristics were found to be suitable to diagnose adverse effects of various ecological problems due to human activity. The synopsis provides suggestions for improving the quality of the aquatic habitat of River Wyna. 10

References

Adams, S.M., K.L. Shepard, M.S. Greely, B.D. Jimenez, M.G. Ryon, L.R. Shugart & J.F. McCarthy, 1989. The use of bioindicators for assessing the effects of pollutant stress on fish. Marine Environmental Research 28: 459–464. Bätscher, R., C. Studer & K. Fent, 1999. Stoffe mit endokriner Wirkung in der Umwelt. Swiss Agency of the Environment, Forests and Landscape (SAEFL), Bern, No. 308, 257 pp. Bernet, D., 2000. Synthese zu verschiedenen Untersuchungen an Fischen in Vorflutern von 41 Kläranlagen. Fischnetz-Info 4: 7–8. Brooks, S., C.R. Tyler & J.P. Sumpter, 1997. Egg quality in fish: What makes a good egg? Rev. Fish Biol. Fish. 7: 387–416. Clement, W.H., 2000. Integrating effects of contaminants across levels of biological organization: an overview. J. Aquat. Stress. Recovery 7: 113–116. Cleveland, L., J.F. Fairchild & E.E. Little, 1999. Biomonitoring and ecotoxicology: Fish as indicators of pollution- induced stress in aquatic systems. Environmental Science Forum 96: 195–232. Colby, P.J., 1984. Appraising the status of fisheries: Rehabilitation techniques. In: Cairns, V.W, P.V. Hodson & J.O. Nriagu (eds). Contaminant effects on Fisheries. Vol. 16. Adv. Envir. Sci. Technol.: 233–257. COMPREHEND, 2001. Homepage: www.ife.ac.uk/comprehend. Crisp, D.T., 1989. Some impacts of human activities on trout, Salmo trutta, populations. Freshwat. Biol. 21: 21–33. Elliott, J.M., 1994. Quantitative ecology and the brown trout. Oxford University Press, Oxford, New York, Tokyo, 286 pp. Escher, M., 1999. Einfluss von Abwassereinleitungen aus Kläranlagen auf Fischbestände und Bachforelleneier. Swiss Agency of the Environment, Forests and Landscape (SAEFL), Mitteilungen zur Fischerei No. 61, 202 pp. Projekt Fischnetz, 2004. Fischnetz Schlussbericht. Swiss Federal Institute for Environmental Science and Technology (EAWAG), Dübendorf, 178 pp. Fitzgerald, D.G., R.P. Lanno & D.G. Dixon, 1999. A comparison of a sentinel species evaluation using creek chub (Semotilus atromaculatus Mitchill) to a fish community evaluation for the initial identification of environmental stressors in small streams. Ecotoxicology 8: 33–48. Gibbons, W.N. & K.R. Munkittrick, 1994. A sentinel monitoring framework for identifying fish population responses to industrial discharges. J. Aquat. Ecosys. Health 3: 227–237. Kendall, R.J., R.L. Dickerson, J.P. Giesy & W.A. Suk, 1998. Principals and processes for evaluating endocrine disruption in wildlife. Proceedings from principles and processes for evaluating endocrine disruption in wildlife, March 1996, Kiawah Island SC. Pensacola FL, Soc. Environ. Tox. Chem. (SETAC), 515 p. Kime, D.E., 1995. The effect of pollution on reproduction in fish. Rev. Fish Biol. Fish. 5: 52–96. Kime, D.E., 1998. Endocrine disruption in fish. Kluwer Academic Publishers, Boston, Dordrecht, London, 396 pp. Kime, D.E., 1999. A strategy for assessing the effect of xenobiotics on fish reproduction. Sci. Total. Environ. 225: 3–11. Kozel, S.J. & W.A. Hubert, 1989. Testing of habitat assessment models for small trout streams in the Medicine Bow National Forest, Wyoming. North Am. J. Fish Manage. 9: 458–464. Lanka, R.P., W.A. Hubert & T.A. Wesche, 1987. Relations of geomorphology to stream habitat and trout standing stock in small Rocky Mountain streams. Trans. Am. Fish. Soc. 116: 21–28. McKim, J.M., 1977. Evaluation of tests with early live stages of fish for predicting long-term toxicity. J. Fish. Res. Board Can. 34: 1148–1154. 11

McMaster, M.E., G.J. Vanderkraak & K.R. Munkittrick, 1996. Exposure to bleached kraft pulp-mill effluent reduces the steroid biosynthetic capacity of white sucker ovarian follicles. Comp. Biochem. Physiol. 112C: 169–178. Miller, M.A., 1993. Maternal transfer of organochlorine compounds in salmonides to their eggs. Can. J. Fish. Aquat. Sci. 50: 1405–1413. Purdom, C., P. Hardiman, V. Bye, N. Eno, C. Tyler & J. Sumpter, 1994. Estrogenic effects of effluents from sewage treatment works. Chem Ecol. 8: 275–285. Rao, S.S., 1999. Impact assessment of hazardous aquatic contaminants. Concepts and approaches. Lewis Publishers Boca Raton, London, New York, 219 pp. Riley, S.C. & K.D. Fausch, 1995. Trout population response to habitat enhancement in six northern Colorado streams. Can. J. Fish. Aquat Sci. 52: 34–53. Shuter, B.J., 1990. Population-level indicators as stress. Am. Fish. Soc. Symp. 8: 145–166. Sumpter, J.P., 1997. Environmental control of fish reproduction: a different perspective. Fish Physiol. Biochem. 17: 25-31. von Westernhagen, H., 1988. Sublethal effects of pollutants on fish eggs and larvae. In: Hoar, W.S. & D.J. Randall (eds.), Fish Physiology. Vol. 11. Part A. Academic Press, San Diego: 253–346. Wahli T., W. Meier, H. Segner, & P. Burkhardt-Holm, 1998. Immunohistological detection of vitellogenin in male brown trout of Swiss rivers. Histochem. J. 30: 753–758. 12 13

2 HOW IS THE QUALITY OF EGG OF BROWN TROUT (SALMO TRUTTA FARIO L.) IN A WASTEWATER- POLLUTED SWISS RIVER?

ABSTRACT

Water pollution can impair egg quality in fish. In this study, we investigated egg quality of female brown trout (Salmo trutta fario L.) from a river in which wastewater discharge is known to cause adverse effects to the health status of trout. We measured the egg quality as fertilization and hatching success.

Results show that egg quality was not impaired (fertilization and hatching success > 93%), although the investigated females came from a river that contains about 50% treated wastewater from treatment plants. In contrast, female trout from the unpolluted headwater of the river had a low egg quality (fertilization success 30 - 100% and hatching success 0 - 72%).

A recent study on brown trout from the river revealed a low feeding status in the headwater and a high feeding status in the downstream reach. Therefore, we suspect that the feeding status of trout is the main factor explaining our results, rather than the degree of water pollution.

We also experimentally exposed eggs to polluted river water for a short duration between spawning and egg hardening. No effect on fertilization and hatching success was observed, but a clear reduction in hatchling size (p < 0.05). Polluted river water used in the experiments exhibited strong estrogenic activity as measured by Rutishauser (pers. comment) in vitro. 14

Introduction

Water pollution can impair the reproductive function in fish. As reproduction is even more sensitive to pollution than survival, reproductive health can be a key factor for the survival of a fish population (Mayer et al., 1986).

Pollutants can accumulate in the gonads during oocyte growth or later in the egg after spawning and cause problems by impairing egg fertilization or acting as an embryo-toxin (Miller, 1993; Brooks et al., 1997; Kime, 1998; Von Westernhagen et al., 1988). Additionally, pollutants can damage the health of fish, causing toxic effects or disrupting endocrine control. Such stress can directly impair oogenesis in female fish (Kime, 1998). Treated wastewater, a complex mixture of many different pollutants, can affect reproduction in fish (Kime, 1998). Thus, discharged wastewater from wastewater treatment plants (WWTP) may constitute a potential risk for fish populations in flowing waters.

We used the Wyna river system as a study site. It is located in the center of Switzerland. The average load of wastewater in its downstream reach is 50%, and this loading was recently found to impair the health status of brown trout (Escher, 1999) and disrupt the endocrine function in male trout (Wahli et al., 1998).

The objective of our study was to evaluate whether the reproductive function of female brown trout from the Wyna is impaired. We investigated also if polluted river water from the downstream reach can affect the quality of trout eggs during the short transition stage between spawning and egg hardening. We measured egg quality as percentages of fertilized eggs and viable hatching embryos. The size of hatchlings was also analyzed as an indicator of egg quality.

Material and Methods

Origin of trout Egg quality was investigated in female brown trout (Salmo trutta fario L.) from the Wyna River located in central Switzerland (Figure 1). From its headwater to its downstream reach, the Wyna receives wastewater from four WWTPs (Figure 1). Artificial barriers with a height of more than 1.2 m prevent the movement of fish upstream. On November 20, 2000, we caught female trout in the spawning season using a smooth-DC stationary electrofishing unit (EFKO 7.5KW). Sampling sites were located in the downstream reach with an average load of 50% wastewater and at a headwater reference site free of wastewater. Male trout were caught in 15

the headwater. Trout were held in oxygenated tanks (<2 h, 5-6 °C) and later anesthetized with a solution of neutralized MS-222 (100 mg/L, 3-aminobenzoic acid ethyl ester, Fluka, Buchs). Total length (± 1 mm) and body weight (± 1 g) were recorded. Scales were collected for age determination according to Baglinièr et al. (1985).

Figure 1: Sampling sites in the headwater and downstream reach of the River Wyna. From its headwater to its downstream reach, the Wyna receives wastewater from four wastewater treatment plants (WWTP). Several large barriers (white bars) separate the studied sampling sites. 16

Treatment and fertilization of eggs According to Elster (1949) ripe females were dried with a towel and eggs were striped into receptacles. Sperm from several (≥ 3) males was added and carefully mixed with the eggs (144 - 1200 eggs/batch). Afterwards, we split the batch of eggs from each female into two equal sized groups (Figure 2). In trout, fresh water activates the sperm (Elster, 1949). Adding river water from either the headwater and the downstream, initiated egg fertilization in our experiment.

Swelling and hardening of eggs occurred when the eggs were transported to the laboratory (3 h, 5-6 °C, 6 - 8 mg oxygen/L). To measure egg diameter (± 0.01 mm) and weight (48 h dried in a lyophylisator; ± 0.0001 g) at the beginning of development, sub-samples of ten eggs per experimental unit were taken.

Figure 2: Experimental design: Eggs came from females in the downstream reach and the headwater. We split the batch of eggs from each female into two equal sized groups. During fertilization and swelling, eggs were treated with test water sampled in the headwater and downstream reach. 17

Incubation in the laboratory under controlled conditions At the EAWAG station in Kastanienbaum, eggs were incubated under controlled conditions with constant temperature and water flow. Temperatures were recorded in one-hour intervals with loggers (VEMCO, Ltd, Shad Bay, Nova Scotia, Canada, precision ± 0.07 °C). Input water came from Lake Lucerne, a clean alpine lake. To avoid light-induced effects on embryonic development, eggs were kept in darkness (Hamdorf, 1960).

Fertilization success was measured as the number of “green” unfertilized eggs. Between fertilization and hatching, eggs were daily controlled for coagulated “white” eggs that were removed and recorded. Hatching success was measured as the number of hatching eggs.

After hatching, the number of deformed hatchlings was recorded. Eighteen days after hatching, random sub-samples of hatchlings were collected (n = 7 - 10) to measure the total length ( ± 0.01 mm).

Data analysis Eggs from each female were treated pair-wise with test water from the headwater and downstream reach. To analyze the effect of different test-waters on fertilization and hatching success, a non-parametric paired Kruskal-Wallis-Test was used. Differences in hatchling length were analyzed with a two-way analysis of variance (ANOVA). The factor “parent” considered the variation in the hatchling size due to eggs from different females. The factor “treatment” considered the variation due to egg treatment with test water from either the headwater and the downstream reach.

Physio-chemical characteristics of the test water used for egg treatment In the test waters for egg treatment, we measured temperature, conductivity, pH, and dissolved oxygen using a digital meter (Multi Line P4, WTW, Weilheim, Germany). Ammonia, nitrite, nitrate and orthophosphate were measured using standard methods (DEW, 1996). We used the concentration of ammonia, pH, and temperature to calculate the concentration of un-ionized ammonia (NH4OH-N) according to Emerson et al. (1975).

In addition, we measured concentrations of Nonylphenol (NP), Nonylphenolmono- (NP1EO) and -diethoxylate (NP2EO), as they can disrupt the reproduction and endocrine system in fish (Kime, 1998). Estrogenic activity (expressed as estradiol equivalency E2-EQ calc. in ng/L) was measured in vitro by B. Rutishauser with yeast cells (YES), stably transfected with the human estrogen receptor gene. Detailed methods for measurements of nonylphenol (-ethoxylates) and estrogenic activity are described in Aerni et al. (2004). 18

Results

Fertilization and hatching success We investigated the egg quality of seven ripe females from the downstream reach and four from the headwater (Table 1). Females from the downstream reach were older, larger and heavier in comparison with headwater females. Downstream females had also larger and heavier eggs.

The percentage of fertilized eggs from downstream females varied between 94 and 100% (Figure 3). With an average of 60% (range: 30% - 100%), the fertilization success of eggs

from headwater females was significantly lower (W-test: n1 = 4, n2 = 7, p < 0.05). Incubated in the laboratory at 6.7 ± 0.46 °C, eggs hatched 59 to 60 days after fertilization. The percentage of deformed hatchlings per experimental unit was less than 2%.

Between 96% and 100% of the eggs from downstream females hatched (Figure 4). This was

significantly more (W-test: n1 = 4, n2 = 7, p < 0.05) than the hatching success found for eggs from headwater females (average 28%, range between 0 - 72%).

Table 1: Measured variables of female brown trout used to investigate the egg quality. Egg diameter and dry weight is given as a mean and corresponding standard deviation (n=10).

Female trout Eggs

Origina Age Length Weight Condition Diameter Dry weight (years) (mm) (g) (100 weight / length 3) (mm) ** (mg) ** DR 4+ 398 626 0.99 5.8 (0.19)* 31.0 (0.8)

DR 3+ 325 428 1.25 5.8 (0.22) 32.8 (1.0)

DR 3+ 319 404 1.24 5.7 (0.19) 32.5 (0.7)

DR 3+ 310 375 1.26 5.6 (0.22) 28.0 (0.8)

DR 3+ 296 317 1.22 5.4 (0.14) 27.3 (1.0)

DR 3+ 286 277 1.18 5.2 (0.12) 25.1 (1.5)

DR 3+ 284 294 1.28 5.3 (0.10) 26.6 (0.6)

HW 3+ 258 226 1.32 4.8 (0.18) 21.7 (0.9)

HW 2+ 212 105 1.10 4.7 (0.25) 19.9 (0.8)

HW 2+ 186 77 1.20 No data No data

HW 2+ 173 64 1.24 No data No data a Trout caught in the downstream reach (DR) and in the headwater (HW) at the River Wyna * Standard-deviation is given in parentheses ** Sample-size is 10 eggs (n=10). 19

100 Eggs treated with water from: D Headwater -~ - Downstream reach ~ C/) 75 C/) Cl) (.) (.) ::J C/) c 0 ~ 50 N t Cl) -Cl) Cl ...... co Cl) 25 > <(

0 -+--~--~ Eggs from downstream Eggs from headwater females (n=7) females (n=4)

Figure 3: Fertilization success of eggs from headwater and downstream female trout. Eggs were pair-wise treated with test water from the downstream reach (black bar) and from the headwater (white bar) as a reference.

100 Eggs treated with water from: CJ Headwater • Downstream reach -~ ~ 75 C/) C/) Cl) (.) (.) ::J C/) Clc E 50 (.) iii ..c: Cl) Cl ...... co Cl) 25 ~

Eggs from downstream Eggs from headwater females (n=7) females (n=4)

Figure 4: Hatching success of eggs from headwater and downstream female trout. Eggs were pair-wise treated with test water from the downstream reach (black bar) and from the headwater (white bar) as a reference. 20

Egg treatment with test water from the River Wyna Different test-water did not influence fertilization (paired-W-test: n = 11, p = 0.10) and hatching (paired-W-test n = 11, p = 0.79) success. However, treatment with test water from the downstream reach caused a significantly smaller size of hatchlings (Table 2, p< 0.0001).

Table 2: Hatchling size from eggs treated with test water from the downstream reach and headwater. Given are the average of 10 hatchlings each and the standard deviation.

Origina Eggs treated with test-water from the

Downstream reach Headwater reach

DR 21.7 (0.24) 21.9 (0.50)

DR 21.5 (0.41) 21.8 (0.21)

DR 21.5 (0.23) 21.9 (0.43)

DR 21.8 (0.31) 22.3 (0.30)

DR 20.7 (0.57) 22.0 (0.31)

DR 22.1 (0.26) 22.4 (0.36)

HW 20.9 (0.27) 20.9 (0.27)

HW 20.7 (0.30) 20.9 (0.46)

a Trout caught in the downstream reach (DR) and in the headwater (HW) at the River Wyna * Standard-deviation is given in parentheses

Chemistry of test water from the River Wyna - Test water from the downstream reach contained nitrate (NO3), nitrite (NO2), ammonia + 3- (NH4 ), unionized-Ammonia (NH4OH), orthophosphate (PO4 -P), Nonylphenol (NP), Nonylphenolmono- (NP1EO), and -diethoxylate (NP2EO) (Table 3). Test water from the downstream reach had a high estrogenic activity. The test water from the headwater was uncontaminated and served as a control. 21

Table 3: Physiochemical parameters of the test water from the headwater and downstream reach used for egg treatment.

Water from Water from Parameter Headwater Downstream Reach Temperature (°C) 8.0 8.1

pH 8.2 8.7

Conductivity (at 25 °C mS/cm) 603 642

Oxygen (mg O2/L) 11.0 11.8

+ Ammonia (mg NH4 -N/L) 0.005 0.253

Ammonium (mg NH4OH-N/L) 0.0001 0.0188

- Nitrite (mg NO2N/L) 0.0027 0.0380

- Nitrate (mg NO3N/L) 3.0 3.9

3- Orthophosphate (mg PO4 -P/L) 0.026 0.060

Nonylphenol (µg NP/L) <0.09* 0.20

Nonylphenole-mono-ethoxylate (µg NP1EO/L) <0.08* 0.31

Nonylphenole-di-ethoxylate (µg NP2EO/L) <0.08* 0.38

Estradiol equivalent (ng E2-EQ/L) <0.20* 1.90±0.08

*Concentration below detection limit

Discussion

Egg quality of trout from the Wyna A federal monitoring program measured several physicochemical parameters in the downstream reach of the Wyna River (Table 4). Due to high loads of wastewater, high

+ - concentrations of ammonia and nitrite up to 1.2 mg NH4 -N/L and 0.32 mg NO2-N/L were th + - typical. 80 percentiles of ammonia (>0.2 mg NH4 -N/L) and nitrite (> 0.03 mg NO2-N/L) exceeded maximum concentrations proposed for running waters (EIFAC, 1973; EIFAC, 1984).

Wastewater discharged in the Wyna River causes adverse effects in wild trout (Table 5). Escher (2000) evaluated the health status of trout by investigating histological changes in gill, liver and kidney and by calculating histopathological-indexes (Hp-I). He showed that the health status of trout from the wastewater loaded downstream reach (Hp-I = 53) was significantly lower compared to trout from the wastewater free headwater (Hp-I = 21). In addition, recent studies give evidence that the reproductive health of trout from the downstream reach is impaired. This was indicated by the existence of female trout with 22

degenerative deformations in their gonads (Escher, 1999, 2000). Furthermore, Wahli et al. (1998) noticed abnormally high concentrations of vitellogenin in male trout. Vitellogenin is a yolk precursor for egg ripening and used as an indicator for endocrine active wastewater when measured in male fish (Sumpter & Jobling, 1995).

Stress, and in particular endocrine disruption, can impair reproductive function in trout (Campbell et al., 1992, 1994; Kime, 1998;). Based on this information, we hypothesized a low egg quality of free-living female trout from the wastewater loaded downstream reach.

However, the opposite was found. We found an excellent egg quality of female trout from the downstream reach. More than 94% of their eggs became fertilized and survived until hatching. In this respect, our result differs with other studies that show sensitivity of reproduction to pollution (Kime, 1998).

Table 4: Physiochemical parameters in the downstream reach of the river Wyna, measured monthly between 1998 and 2001 (n = 36). Data come from the federal chemical monitoring program in rivers of the canton Aargau (Station Wyna-Suhr, Aargau, HydroPro).

50%- 80%- Parameter Mean Minimum quantile quantile Maximum

Temperature (°C) 10.3 2.6 10.7 15.1 18.4

pH 8.5 8.1 8.5 8.6 8.8

Conductivity (at 25 °C mS/cm) 560 429 573 626 666

Suspended Solids (mg/L) 13 1 6 15 132

Oxygen (mg O2/lL 11.1 8.2 10.9 12.3 14.0

+ Ammonia (mg NH4 -N/L) 0.242 0.012 0.210 0.300 1.200

Ammonium (mg NH4OH-N/L) 0.0129 0.0011 0.0106 0.0186 0.0577

- Nitrite (mg NO2-N/L) 0.073 0.017 0.060 0.100 0.320

- Nitrate (mg NO3-N/L) 3.9 2.5 4.1 4.8 6.6

3- Total Phosphor (mg PO4 tot-P/L) 0.11 0.03 0.10 0.16 0.36

3- Phosphate (mg PO4 -P/L) 0.06 0.02 0.05 0.09 0.14

Sulfate (mg SO4/L) 15.4 8.1 16.6 18.0 30.0

Chloride (mg Cl-/L) 16.7 6.6 16.3 23.4 33.2

Biol. Oxygen Demand (mg BOD5/L) 3.1 1.1 2.7 3.3 8.1

Dissolved Org. Comp. (mg DOC/L) 3.0 1.7 3.0 3.4 6.2 23

It is possible that trout from less polluted river sites may have immigrated into our sampling site in the downstream reach. As a result, we may have sampled immigrant trout with a short pollution exposure history and a better egg quality. However, based on a recently conducted mark-recapture experiment, we know that between 50 to 100% of adult Wyna trout are stationary (Kobler & Peter, in preparation). Consequently, the chance is less than 1% that all seven investigated females from the downstream reach were non-stationary. Therefore, we have evidence that the egg quality of female trout in the downstream reach was not impaired. It appears that female trout can tolerate the pollution in the downstream reach of the Wyna River.

Table 5: Indication of the health status and reproductive health of Wyna brown trout upstream (UPST) and downstream (DWST) of the wastewater treatment plant (WWTP) Mittleres Wynental. Data come from five recently conducted studies in the River Wyna.

Year of study UPST WWTP DWST WWTP Reference

Health status

Proliferative kidney disease (PKD) in brown trout 1999 100% (2)* 100% (2) Wahli et al. 2002

Histopathology-Index (Hp-I) based on liver, gill and kidney according the method of Bernet et al. (1999)** 1999 21 Hp-I(8) 53 Hp-I(8) Escher, 2000

Reproductive health

Degenerative deformations in female gonads 1997 0% (4) 67% (3) Escher, 1999

Degenerative deformations in female gonads 1999 9% (11) 38% (8) Escher, 2000

Elevated Vitellogenin (Vtg) levels in male trout as an indicator for endocrine disruption 1997 0% (5) 100% (2) Wahli et al., 1998

* Parenthesis provides the number of investigated trout from the Wyna (n) ** A high histopathology-Index (Hp-I) indicates an impaired health status. Median (52%) Hp-I of brown trout from more than hundred Swiss rivers is near 20 Hp-I (H. Segner, Fischnetz meeting the 9th of Sept. 2003, Switzerland) 24

Brooks et al. (1997) and Bagenal (1969) note that when more food is available trout can grow faster and produce better eggs. As we have evidence that the feeding status and growth performance of downstream trout is excellent (Kobler & Peter, in preparation), such a relationship between feeding status and egg quality might be relevant for Wyna trout. In contrast, the situation in the headwater was the opposite. Feeding status and growth performance of headwater trout was low (Kobler & Peter, in preparation). Therefore, we suspect that the availability of food in the Wyna could be a key factor explaining egg qualities of trout from the headwater and downstream reach.

Fertilization and first water uptake of trout eggs During early transition stage between spawning and egg swelling, eggs can be very sensitive to water pollution (Brooks et al., 1997; Kime, 1998). We treated eggs with water from the Wyna to test its effect on this early step in the life cycle of trout. The results show that treatment with water from the downstream reach did not impair the fertilization success of eggs. This result was surprising, as a similar experiment with water from a wastewater polluted river in Germany revealed a clear negative effect on the fertilization success of trout eggs (Luckenbach et al., 2001). We conclude that the polluted water used in our study does not reduce fertilization success.

After fertilization, trout eggs take up water from the surroundings and increase in volume. Later, the eggshell hardens and the development of the embryo begins. Toxic and endocrine-active substances can be taken into the egg with the water, and these compounds can potentially stress the developing embryos (Kime, 1998). Egg treatment with water from the Wyna downstream reach did not influence hatching success. However, we noticed a significantly reduced hatchling size due to exposure of the eggs to polluted water. Since Von Westernhagen (1988) shows that sub-lethal stress due to chemical contamination can result in a retarded embryonic development and finally in a smaller hatchling size, we suspect that polluted Wyna water retarded the embryonic development. A small hatchling size also indicates a lower chance to survive (Elliott, 1994). Thus, a small hatchling size might indicate an impaired offspring survival and consequently an impaired reproductive success of trout.

Test waters used for egg treatment Water sampled in the downstream reach reflected the median physiochemical characteristics in the Wyna River measured between 1998 and 2000 (Table 3 and 4). High concentrations of ammonia and nitrite may have a negative impact of treated wastewater. High concentrations of ammonia and nitrite are toxic to fish (Alabaster & Lloyd, 1980). Therefore it was expected 25 that these problematic high concentrations would cause immediate effects on our experiment. But such short-term effects were not observed; instead we noticed a long-term effect on embryo development. Our in vitro measurements with yeast cells (YES) revealed a strong estrogenic activity in the test water from the Wyna downstream reach. Schwaiger et al. (2002) showed that nonylphenol (-ethoxylates) can be estrogenic. However, concentrations of NP, NP1EO and

NP2EO in the water were below 0.4 µg/L and thus too low to explain the estrogenic activity. On the other hand, steroid hormones commonly found in domestic wastewaters such as ethinylestradiol and estradiol are estrogenically active at concentrations as low as 0.1 ng/L and 1-10 ng/L, respectively (Purdom et al., 1994; Routledge et al. 1998). It is possible that steroid hormones were responsible for the estrogenic activity of Wyna River water from the wastewater loaded downstream reach. We suspect that the estrogenic test water impaired the endocrine function of developing embryos and thus retarded their development.

Conclusion The reproductive health of fish is essential for the health of the whole fish population. We investigated female brown trout (Salmo trutta fario L.) from a wastewater polluted river. Although recent studies reveal an impaired health status in these trout, our results show that egg quality is not necessarily impaired. We have indications that a good feeding status is a dominant factor explaining good egg qualities in the wastewater polluted river site. However, we cannot preclude that reproduction is a limiting factor for the Wyna trout population, as we have indication from another recent study that the reproductive health of male trout is impaired (Kobler, 2003).

We treated trout eggs during fertilization and first water uptake with polluted river water to test the sensitivity of eggs after spawning. Measurements of the estrogen activity in the water indicated a high potential to disrupt the endocrine regulation of developing embryos. We found no effects on fertilization and hatching success, but we observed a smaller size of yolk-sac embryos after hatching. Since undersized fry are less able to survive in a natural environment, such an effect of Wyna water on trout embryos may endanger the successful recruitment of the whole population.

We hypothese that parent females can protect their reproductive ability from water pollution in the Wyna River. Early life stages cannot and are more sensitive. Reproduction and early survival are key factors for the survival of wild fish populations. 26

References

Aerni, H.R., B. Kobler, B.V. Rutishauser, F.E. Wettstein, R. Fischer, W. Giger, A. Hungerbühler, M.D. Marazuela, A. Peter, R. Schönenberger, A.C. Vögeli, M.J.-F Suter & R.I.L. Eggen, 2004. Combined biological and chemical assessment of estrogenic activities in wastewater treatment plant effluents. Anal. Bioanal. Chem. 378: 688-696. Alabaster, J.S. & R. Lloyd, 1980. Ammonia. In: Water quality criteria for freshwater fish. Butterworths, London, England, pp. 85-102. Bagenal, T.B., 1969. The relationship between food supply and fecundity in brown trout (Salmon trutta L.). J. Fish Biol. 1: 167-182. Baglinièr, J.L., P. Bomassi, B. Bousquet, F. Chancerel, H. De Pontual, J. Dumas, G. Euzenat, G. Fontenelle, F. Fournel, F. Gayou, J.F. Luquet, G. Maisse Martin, J.A. Ventura, A. Marty, A. Nihouarn, J.P. Porcher, E. Prevost, P. Prouzet, G. Pustelnik, A. Richard & H. Troadec, 1985. La détermination de l’age par scalimétrie chez le saumon atlantique (Salmo salar) dans son aire de réparation méridionale: utilisation pratique et difficultés de la Méthode. Bull. Fr. Pêche Piscic. 298: 69-105. Bernet, D., H. Schmidt, W. Meier, P. Burkhardt-Holm & T. Wahli, 1999. Histopathology in fish: proposal for a protocol to assess aquatic pollution. J. Fish Diseases 22: 25-34. Brooks, S., C.R. Tyler & J.P. Sumpter, 1997. Egg quality in fish: What makes a good egg? Rev. Fish Biol. Fish. 7: 387-416. Campbell, P.M., T.G. Pottinger & J.P. Sumpter, 1992. Stress reduces the quality of gametes produced by rainbow trout. Biol. Reprod. 47: 1140-1151. Campbell, P.M., T.G. Pottinger & J.P. Sumpter, 1994. Preliminary evidence that chronic confinement stress reduces the quality of gametes produced by brown and rainbow-trout. Aquaculture 120: 151-169. DEW, 1996. Deutsche Einheitsverfahren zur Wasseruntersuchung, Bd. II, VCH Weinheim, New York, Cambridge, Tokyo, Berlin, Wien, Zurich. EIFAC, 1973. Water quality criteria for European freshwater fish. Report on ammonia and inland fisheries. Water Res. 7: 1011-1022. EIFAC, 1984. Water quality criteria for European freshwater fish. Report on nitrite. European Inland Fisheries Advisory Commission of FAO, EIFAC Tech. Pap. 46. Rome. Elliott, J.M., 1994. Quantitative ecology and the brown trout. Oxford University Press, New York, 286 pp. Elster, H.J., 1949. Über die Vorgänge bei der natürlichen und künstlichen Befruchtung der Fische. Die Binnenfischerei 2: 10-14. Emerson, K., R.C. Russo, R.E. Lund & R.V. Thurston, 1975. Aqueous ammonia equilibrium calculations: Effects of pH and temperature. J. Fish. Res. Board Can. 32: 2379-2383. Escher, M., 1999. Einfluss von Abwassereinleitungen aus Kläranlagen auf Fischbestände und Bachforelleneier. Swiss Agency of the Environment, Forests and Landscape (SAEFL), Mitteilungen zur Fischerei No. 61, 202 pp. Escher, M., 2000. Fischereibiologische Untersuchung im Bereich der ARA Teufenthal, Im Auftrag der Fischereiverwaltung des Kantons AG (unpublished), 15 pp. Hamdorf, K., 1960. Die Beeinflussung der Embryonal- und Larvenentwicklung der Regenbogenforelle (Salmo irideus Gibb.) durch Strahlung im sichtbaren Bereich. Zeitschrift für vergleichende Physiologie 42: 525- 565. Kime, D.E., 1998. Endocrine disruption in fish. Kluwer Academic Publishers, Boston, London, 396 pp. 27

Kobler, B. 2003. Hormonelle Störungen freilebender Bachforellen im Fluss Wyna ober- und unterhalb von Abwasserreinigungsanlagen. Bericht im Auftrag des Amts für Umweltschutz des Kantons Luzern, Oktober 2003, 19 Seiten. Kobler, B. & A. Peter. Stress response pattern of a brown trout population in River Wyna with treated wastewater inputs. In preparation. Luckenbach, T., M. Kilian, R. Triebskorn, A & Oberemm, 2001. Fish early life stages as a tool to assess embryotoxic potentials in small streams. J. Aquatic Ecosystem Stress and Recovery 8: 355-370. Mayer, L.L., Jr. K.S. Mayer & M.R. Ellersieck, 1986. Relation of survival to other end points in chronic toxicity tests with fish. Environ. Tox. Chem. 5: 737-748. Miller, M.A., 1993. Maternal transfer of organochlorine compounds in salmonines to their eggs. Can. J. Fish. Aquat. Sci. 50: 1405-1413. Purdom C.E., P.A. Hardimann, V.J. Bye, N.C. Eno, C.R. Tyler & J.P. Sumpter. 1994. Estrogenic sffects of effluents from sewage treatment works. Chem. Ecol. 8: 275-285. Routledge E.J., D. Sheahan, C. Desbrow, G.C. Brighty, M. Waldock & J.P. Sumpter. 1998. . Identification of estrogenic chemicals in STW effluent. II : In vivo responses in trout and roach. Environ. Sci. Technol. 32: 1559-1565. Schwaiger, J., U. Mallow, H. Ferling, S. Knoerr, T. Braunbeck, W. Kalbfus & R.D. Negele, 2002. How estrogenic is nonylphenol? A transgenerational study using rainbow trout (Oncorhynchus mykiss) as a test organism. Aquat. Tox. 59: 177-189. Sumpter, J.P. & S. Jobling, 1995. Vitellogenesis as a biomarker for estrogenic contamination of aquatic environments. Environ. Health Perspect. 103: 173-178. von Westernhagen, H., 1988. Sublethal effects of pollutants on fish eggs and larvae. In: Hoar, W.S. & D.J. Randall (eds.), Fish Physiology. Vol. 11. Part A. Academic Press, San Diego: 253-346. Wahli, T., R. Knuesel, D. Bernet, H. Segner, D. Pugovkin, P. Burkhardt-Holm, M. Escher & H. Schmidt-Posthaus, 2002. Proliferative kidney disease in Switzerland: current state of knowledge. J. Fish Diseases 25: 491- 500. Wahli, T., W. Meier, H. Segner & P. Burkhardt-Holm, 1998. Immunohistological detection of vitellogenin in male brown trout of Swiss rivers. Histochem. J. 30: 753-758. 28 29

3 MORTALITY OF BROWN TROUT (SALMO TRUTTA FARIO L.) EMBRYOS IN A POLLUTED SWISS RIVER

Abstract

Exposure experiments with brown trout (Salmo trutta fario L.) early life stages were used to assess the effect of pollution on recruitment. Exposure sites were located in the downstream reach and headwater of a river in Switzerland. Exposures in the laboratory served as a control. The river’s downstream reach is polluted by sewage from treatment plants. Furthermore, silt degrades the quality of trout spawning sites in the river.

We used two different incubator systems to expose embryos: (1) Vibert boxes filled with eggs and buried “in-gravel”; (2) Incubators fixed “on-gravel”, which allowed frequent examination of the embryos during development.

The embryo mortality was highest in the downstream reach. Since silt covered the in-gravel embryos, we suspect that silt either impaired gas exchange or was toxic. On-gravel incubators were not affected by siltation. Nevertheless, more than 90% of embryos exposed in the downstream reach died. The highest mortality rates were noticed during hatching. Hatching, therefore, was suspected to represent a crucial bottleneck for early development in the river. The study shows that early life stages were seriously stressed. This indicates that the reproductive success is impaired for trout resident in the downstream reach. We conclude that sewage and fine solids infiltrating the riverbed may seriously impair early development of brown trout. 30

Introduction

Early life stages of salmonids develop in stream gravel until the embryos hatch and emerge. During this phase, early stages are sensitive to various abiotic and biotic factors, such as low oxygen, poor water chemistry, the occurrence of mechanical shock, predation, and parasitism (Crisp, 1989, 2000; Luckenbach et al., 2001, 2003).

In particular, anthropogenic influence can disrupt early development. Especially for brown trout it has been shown that water pollution and excessive loads of silt cause stress (Crisp, 1989; McKim, 1977; von Westernhagen, 1988). Embryo-toxic substances in the water and silt infiltrating the streambed can impair physiochemical conditions for early development in gravel (Crisp, 1989). Such anthropogenically induced stress can impair the population’s natural recruitment resulting in negative effects on the population structure and abundance (Harris, 1973; Turnpenny and Williams, 1980; Hulsman et al., 1983; Segner et al., 1988; Ingendahl & Neumann, 1996).

Here, we present exposure experiments with brown trout early life stages which were performed in a representative Swiss river in two different polluted sections. Field studies on the structure of a brown trout population resident in this river revealed a lack of early life stages in the more polluted section, whereas these stages were found in high abundance in the less polluted part (Kobler, unpublished). The objective of this paper is to study if the development of early life stages is impaired in the more polluted river section. An additional objective is to determine the responsible environmental factors.

Fry traps or the excavation of redds are direct approaches to estimate early survival of salmonids under natural conditions. However, their disadvantage is the relatively high uncertainty of the results obtained since many unknown variables may be involved (e.g., quality of spawned eggs, disappearance of dead eggs due to decomposition) (Rubin, 1995; Rubin and Glimsäter, 1996). A more controlled approach is to expose salmonid eggs in the field in artificial redds or boxes filled with gravel (e.g., Harries, 1973; Turnpenny and Williams, 1980). The advantage of such in-gravel exposures is that they provide information on the early developmental success under near natural conditions. A disadvantage is the limitation of examinations during the exposure. As consequence temporal events cannot be determined. Luckenbach et al. (2003) therefore developed an exposure device with identical chambers for a single embryo exposure in the field. This device can be fixed on the gravel and thus is henceforth called “on-gravel” exposure. Its advantage is that it supports examinations with 31 little disturbance of exposed embryos, which allows the observation of development and stage-typical endpoints over time. Its disadvantage is that environmental factors occurring in the gravel where trout eggs normally develop are underestimated (e.g., siltation, lack of oxygen, decomposition, toxic silt) (Luckenbach et al., 2003). In the present study, therefore, both in-gravel and on-gravel egg exposure methods were used to provide independent information about the external factors impairing the development in early life stages.

The study stream and sites Test river was the Wyna located in central Switzerland (Figure 1). This river is 25 km long and flows through an area of intensely used farmland and developed industry. Its watershed area is 119 km2 with a population density of 195 persons/km2. The average flow in the downstream reach is about 1.76 m3/s.

From its headwater (HW) to its downstream reach (DR), the river receives wastewater from four wastewater treatment plants (WWTP). In its downstream reach, the average sewage load is about 50%. In the downstream section, the concentration of nutrients is high and bio- indicators, such as diatom and macro-invertebrate abundance surveyed between 1996 and 1997, classified it as “heavily influenced by organic pollutants” (Swiss water quality class: II- III) (Stöckli & Schmid, 1999). In particular, ammonia and nitrite close to acute fish-toxic concentrations were suspected to be problematic. Further investigations additionally reveal that river water in the downstream reach can be estrogenic for trout (Wahli et al., 2002, Kobler, 2003). Responsible chemicals have not yet been identified, however it is suspected that natural and synthetic estrogens from sewage are abundant in biologically relevant concentrations.

Test sites were located in the downstream reach, one km downstream of the last WWTP, and in the less polluted headwater, upstream of the first WWTP (Figure 1). Abiotic conditions concerning oxygen, pH and water hardness are similar in the headwater and downstream reach. The width of the streambed is between 5.8 and 8.6m in the downstream and between 4.0 and 7.0 m in the headwater reach. At low flow, the average depth is 63 cm in the downstream and 32 cm in the headwater reach. Flow rates are between 11 and 74 cm/s in the downstream and between 14 and 38 cm/s in the headwater reach.

Streambeds consist of coarse gravel and rock at both test sites. Due to infiltration of suspended solids, streambeds in the downstream reach are embedded. This is not the case for streambeds in the headwater. The average water temperature in the downstream reach is about 1.1 °C higher than in the headwater (Kobler, unpublished). 32

Material and methods

Egg exposure On the December 14, 2000, freshly fertilized eggs of brown trout (Salmo trutta fario L.) from a local hatchery (Waser, Andelfingen, Switzerland) were exposed in the Wyna River and in the laboratory. In the field, eggs were exposed both in-gravel and on-gravel.

In-gravel exposures were conducted with Vibert boxes (perforated plastic boxes 7 x 4.5 x 5 cm; Vibert, 1949) each filled with 25 eggs and gravel. Burial depth of boxes in the riverbed was 10 cm, which is a common depth for naturally spawned trout eggs (Chapman, 1988; Crisp and Carling, 1989; Gorst et al., 1991; De Vries, 1997). At each experimental site, 15 boxes were buried. Shortly before hatching, after 62 to 68 exposure days, the boxes were removed, and living and dead embryos were counted.

For the on-gravel exposure, egg incubators as described by Luckenbach et al. (2003) were fixed on the riverbed. At each study site, 7 units containing 60 embryos each were installed. The devices contained incubation chambers, allowing an individual exposure of fish embryos, which prevented inter-individual effects. Furthermore, the devices allowed multiple surveys of developing embryos. The embryos were examined every 1 to 5 weeks, and dead embryos were removed. During eying, which is defined as the first visible appearance of eye pigment within the embryo (Crisp, 1988) and during hatching, the embryos were examined every 1 to 3 days. On-gravel exposures were stopped after 111 – 114 days.

As a control (C), 820 embryos of the same batch as used in the field were kept in an incubator for trout rearing (Vetterli-Brutschrank; Metric 14P; GEC Elliot Process Instr. Ldt. England). The water for the control came from a clean alpine lake (Lake of Lucerne). The temperature regime was regularly adjusted to conditions measured in the field. Until hatching, incubated embryos were surveyed every 1 to 2 days.

During the experiment, water temperatures in the downstream reach, headwater and laboratory were recorded continuously (1 h interval) by thermistor loggers (VEMCO, Ltd, Shad Bay, Nova Scotia, Canada, precision ± 0.07 °C). 33

Figure 1: The Wyna River and the location of study sites in the headwater (less polluted) and downstream reach (polluted). 34

Early life stage (ELS) test The endpoints of the ELS test were mortality rate and developmental rate. Mortality rates were calculated on the basis of the total number of eggs at the beginning of the exposure. Time points when eggs were eyed and later when embryos hatched were used to document developmental rates. Eying was defined as the first visible appearance of eye pigment in the embryo and was determined by an experienced observer.

Time points when 5%, 50% and 95% of embryos hatched were modeled using logistic regression: log (pi / 1-pi) = a + b xi; pi = ratio hatched embryos; xi = number of exposure days (Stahel, 1999). The influence of temperature on developmental rate was assessed by predicting the duration until median hatch using mean water temperature at each site. The following equation was used: D = a T-b; a = 308; b = 0.83; D = duration until median hatch; T = temperature (Embody, 1934).

Statistical analysis For each treatment, we summarized numbers of dead and hatched embryos and determined total cumulative mortality/hatching rates for treatment and time point. Significant differences in mortality rates between different treatments were evaluated by the c2 test and subsequent Bonferroni adaptation of p-values (Stahel, 1999).

Physico-chemical water analysis and silt Water from the river and from the laboratory was sampled every 2 to 4 days to measure its physiochemical condition. Dissolved oxygen (DO), conductivity and pH were measured using

+ a digital meter (Multi Line P4, WTW, Weilheim, Germany). Ammonium (NH4 ), nitrate (NO3),

+ 3- 2- - nitrite (NO2 ), total phosphorus (Ptot), orthophosphate (PO4 ) sulfate (SO4 ), chloride (Cl) biological oxygen demand (BOD5) and dissolved organic carbon (DOC) were determined with standard methods in the laboratory (DEW, 1996). Ammonia (NH3) was calculated on the

+ basis of ammonium (NH4 ), pH and temperature (Emerson et al., 1975). Silt which accumulated in the Vibert boxes during exposure was collected and dried (80 °C, 72 h). Its grain size distribution (fractions: 2000 - 1000, 1000 - 500, 500 - 250, 250 - 125, 125 - 65, < 65 mm) was determined using a sedi-graph (Rentsch Analysen-Siebmaschine TYP AS 200 “BASIC”). 35

Results

Survival We retrieved 10 Vibert boxes from downstream and 9 boxes from headwater exposure sites. In the downstream reach, 70% of 250 in-gravel exposed embryos died, whereas in the headwater, 21% of 225 embryos died (Table 1). This difference in mortality was statistically significant (p < 0.001).

The accumulation rate of silt in the Vibert boxes differed between downstream and headwater exposure sites. In the downstream reach, a large mass (> 80%) of silt with a grain-size between 125 and 250 mm had silted in the boxes and covered the embryos. In the headwater, no siltation occurred.

Table 1: Mortality of in-gravel exposed embryos (25 per box).

Dead eggs Downstream reach absolute number relative number

Box 1 24 96 %

Box 2 12 48 %

Box 3 14 56 %

Box 4 24 96 %

Box 5 22 88 %

Box 6 23 92 %

Box 7 16 64 %

Box 8 11 44 %

Box 9 15 60 %

Headwater

Box 1 1 4 %

Box 2 8 32 %

Box 3 1 4 %

Box 4 4 16 %

Box 5 2 8 %

Box 6 16 64 %

Box 7 5 20 %

Box 8 5 20 %

Box 9 7 28 %

Box 10 11 44 % 36

On-gravel exposed embryos were not affected by siltation. In the downstream reach, 29% of 420 embryos died between fertilization and hatching, which was significantly lower than the mortality of in-gravel exposed embryos (p > 0.001 ). During hatching (days 67 - 81 ), the mortality rate increased and an additional 35% of embryos died. At the end of exposure, the total mortality in the downstream reach was 93%, which exceeded the field and laboratory control by 37% and 57%, respectively (Figure 2). Mortality rates after eying (DR, 41 days; HW, 48 days; C, 39 days), after hatching (DR, 82 days, HW, 89 days, C, 75 days) and at the end of the embryonic period (DR, 114 days, HW, 111 days, C, 111 days), were significantly higher in both DR and HW embryos compared to the control (p < 0.001 ). Mortality was significantly lower in the HW compared to the DR after eying (p < 0.001) and after hatching (p < 0.001).

Average concentrations of ammonia (NH/-N), ammonium (NH3-N) and nitrite (N02-N) in the downstream reach exceeded values measured in the headwater and laboratory by a factor of more than ten (Table 2). Concentrations of ammonia and ammonium varied and reached values up to 2.2 NH/-N mg/I and 0.08 NH3-N mg/I. 3 Other parameters such as nitrate (N03+), total phosphorus (P101), phosphate (P04-), sulfate 2 (So-4), chloride (Ci-), dissolved oxygen (DO), biological oxygen demand (BOD5), dissolved organic compounds (DOC), conductivity and pH were constant and were similar in the field and laboratory.

1 00 -io=~~~~~~~~=;-~~~~~~~~~~~~~~~~~--. -.-Downstream Reach -e-Headweater -A-control 80 ...µ....~~~~~~~~--'-~~~~~~~~~~~~_,,,,£.~~~~--I

-;:R. ~ 60 ~ !ii t:: 0 E 40 > ~ :; E 20 ::i (.)

0 10 20 30 40 50 60 70 80 90 100 110 120 exposure time (d) Figure 2: Cumulative mortality percentage during incubation at the test sites and in the laboratory control (E:) eying 50%; (H:) hatching 50%. n=420 for DR and HW sites, n=820 for control. 37

Table 2: Physiochemical measurements during the period of experiment in the downstream reach, headwater and laboratory control, as average ± standard deviation (SD) and concentration range.

Downstream Reach (DR) Headwater (HW) Laboratory

Mean ± SD Mean ± SD Mean ± SD n* n n (Range) (Range) (Range)

Conductivity [mS cm-1] 7 646 ± 116 7 651 ± 914 7 265 ± 56 (499 - 883) (477 - 935) (226 - 375)

pH 7 8.5 ± 0.1 7 8.3 ± 0.3 7 7.7 ± 0.2 (8.3 - 8.7) (8.0 - 8.9) (7.5 - 8.3)

DO [mg/l] 7 11.6 ± 0.6 6 11.6 ± 0.5 7 10.3 ± 0.9 (10.7 - 12.2) (10.6 - 11.9) (8.7 - 11.5)

+ NO3 - N [mg/l] 7 4.07 ± 1.01 7 3.40 ± 0.35 7 0.64 ± 0.04 (2.90 - 6.10) (2.82 - 3.84) (0.57 - 0.70)

+ NO2 - N [mg/l] 7 0.029 ± 0.017 7 0.003 ± 0.002 7 <0.002 (0.010 - 0.059) (0.002 - 0.004)

+ NH4 - N [mg/l] 7 0.553 ± 0.723 7 0.015 ± 0.007 7 0.009 ± 0.001 (0.164 - 2.169) (0.004 - 0.025) (0.008 - 0.010)

NH3 - N [mg/l] 7 0.022 ± 0.026 7 <0.001 7 <0.001 (0.005 - 0.079)

3- PO4 - P [mg/l] 7 0.044 ± 0.015 7 0.021 ± 0.016 7 <0.002 (0.023 - 0.065) (0.004 - 0.050)

Ptot - P [mg/l] 5 0.138 ± 0.063 4 0.160 ± 0.125 5 0.007 ± 0.005 (0.091 - 0.240) (0.068 - 0.339) (0.005 - 0.016)

2- SO 4 [mg/l] 5 13.9 ± 3.6 4 10.8 ± 2.7 5 13.3 ± 0.5 (9.8 - 19.6) (7.0 - 13.4) (12.5 - 13.6)

Cl- [mg/l] 5 12.6 ± 4.6 4 12.0 ± 4.4 5 1.8 ± 0.2 (6.8 - 19.4) (7.4 - 18.0) (1.7 - 2.0)

BOD5 [mg/l] 5 3.9 ± 1.3 4 3.1 ± 2.7 4 1.8 ± 1.1 (2.4 - 6.0) (1.2 - 7.0) (1.1 - 3.4)

DOC [mg/l] 5 2.9 ± 0.7 4 4.8 ± 1.6 5 1.1 ± 0.2 (2.2 - 4.0) (2.6 - 6.5) (0.9 - 1.3) * n=number of samples

Time points of median hatch were 80 days at the downstream reach, 85 days at the field control, and 70 days in the laboratory control (Figure 3). 38

90 • Median hatching ·- - - -·predicted duration until hatch _85 ...... • ~ .. .. Cl .. c Headwater... - ...... :.c ...... ~ 80 ...... I ~ .. .. l. c ...... Downstream .!ll ...... "O ...... Reach Q) ...... cE 75 -...... ·a ...... a...... Q) .. .. E ...... :;::: 70 1• ... control 65 4.7 4.9 5.1 5.3 5.5 5.7 5.9 average exposure temperature (°C)

Figure 3 : Time until 5% (lower:-). 50% (•). and 95% (upper:-) of embryos hatched at test sites in the Wyna River and in the laboratory control in relation to the average exposure temperature. n=150 for DR; n=340 for HW; n=753 for laboratory control. The dotted line indicates the predicted time until hatching.

For laboratory and the headwater, these values were similar (control, 4.8 days; HW, 5.6 days) with time until hatching predicted by mean temperature from the beginning of exposure until hatching (DR, 5.5°C; HW, 4.8°C; control, S.8°C). DR was underestimated by 5 days. Temperature regimes in the DR, HW and laboratory control were similar (Figure 4). The average temperature in the DR until hatching was O.?°C warmer than measured in the HW and 0.3°C cooler than measured in the laboratory. The average temperature during the whole exposure period was 6.3°C in the DR with minimum of 2.0°C and maximum of 11.8°C. The average temperature in the HW and laboratory control were 5.5°C (minimum, 1.1 °C; maximum, 11.4 °C) and 5.9°C (minimum, 4.3 °C; maximum, 9.0 °C).

Discussion

On-gravel exposure On-gravel exposures with brown trout embryos focused on effects initiated by water pollution. Embryo mortality was taken as a sensitive endpoint and revealed clear differences between the two exposure sites in the field and laboratory control. Remarkable was the drastic increase in mortality during hatching in the downstream reach, which indicated stress on the embryos caused by environmental factors. 39

12 -i-.-~~~~~~~~~.....-~~~~~~~~~~~~~~~~~~----, - Downstream Reach - - -Headwater - control

ro ~ 4 +-~----t-r------\l-f--~~tftf--~~--1.--f--~~~~--t\---h-=-#-~~~~~~~~---j ro Q) E

0 20 40 60 80 100 120 14/12/2000 exposure time (d)

Figure 4: Water temperature during the period of the experiment.

Hatching is a sensitive moment during early development which requires physiological changes in fish embryos (von Westernhagen, 1988). Such physiological changes in combination with other environmental factors such as water pollution could cause serious stress for early stages. Thus, it is likely that water pollution in the studied river weakened the exposed embryos which resulted in a crucial bottleneck during hatching.

In contrast, the mortality before hatching was not very high. Nevertheless, there is evidence that exposed embryos were still seriously stressed. We observed that the time point of hatching in the downstream reach was retarded. Von Westenhagen (1988) mentioned that a retarded time point of hatching could indicate physiological stress and a reduced viability. Oxygen deficiency or biological active chemicals can cause such changes during early development. Oxygen deficiency was excluded as a possible cause because embryos were exposed on-gravel in the free-flowing water and, in contrast to Luckenbach et al. (2003), siltation in the exposure chambers was negligible. Therefore, we believe that other biologically active substances influenced the experimentally exposed early life stages. The resulting stress during the development before hatching may have damaged the embryos and thus may explain the high mortality during hatching.

Characteristic for the water pollution in the downstream reach were high concentrations of ammonia, ammonium, and nitrite. In particular, ammonia reached an average concentration 40

shown in literature to cause biological effects in salmonids (0.025 mg NH3/l, Calamari et al., 1977). The maximal measured concentration of ammonia even were close to known acute toxicity concentrations for salmonids (LC5096h = 0.2 mg NH3/l; Alabaster and Lloyd, 1980). Therefore, we suspect that ammonia, probably in combination with nitrite among other substances, increased the mortality and reduced the viability of early life stages in the downstream reach of River Wyna. Recent surveys on wild trout from the Wyna indicated biological effects from xenobiotic substances. Vitellogenin (yolk protein) was detected in the blood of male Wyna trout indicating potential endocrine disruption. Two previous studies (Wahli et al., 2002; Kobler, 2003) have shown that river water in the downstream reach exhibit estrogenic effects in trout.

In-gravel exposure In-gravel exposures with brown trout embryos were used to monitor mortality rates under conditions similar to those of natural spawning. The comparison between in-gravel and on- gravel exposed embryos revealed clear differences. The mortality of in-gravel exposed embryos was much higher. This result indicated that water quality alone played only a part in explaining the embryo mortality in the DR. Since siltation covered in-gravel exposed embryos, it is likely that sediments contributed to the observed high mortality. Turnpenny and Williams (1980) also demonstrated the negative effects of silt on intragravel stages of trout. They showed that silt can reduce intragravel flow and consequently cause oxygen deficiencies, which these authors suspected to be the main cause for a drastically increased mortality of trout embryos. For the present study it is also likely that a limited supply of oxygen due to siltation initiated high mortalities of embryos in the Wyna. In addition, silt is not just inert material, but also a carrier of chemicals associated with it, which may desorb and affect embryonic development (Crisp, 2000).

Garrett and Bennett (1996) showed that siltation rates in Vibert boxes are comparable to rates in the surrounding streambed. Therefore, we assume that high siltation rates in our exposed Vibert boxes indicate a general degradation of the Wyna riverbeds by silt. Escher (1999) showed that the riverbed in the downstream reach was clogged with silt. Furthermore, he noticed crusts of iron sulfide, indicating the absence of oxygen in the riverbed.

Silt was identified here as a major factor disrupting early survival of brown trout in the Wyna. The origin of silt could be sewage or soil erosion due to agriculture activities in the watershed. 41

Conclusion The experimental design using early life stages of brown trout allowed a monitoring of streambed quality and water quality.

Lethal and sublethal effects were recognized at DR and were potentially caused by fine sediments if exposed in gravel and also by ammonia and nitrite if exposed on gravel. Because survival at in gravel exposed eggs was already reduced very early during development it is expected that the poor riverbed quality is a major limiting factor for healthy brown trout recruitment at DR. On the other hand, the water quality had also negative impacts on survival and developmental rates.

Thus both, the poor streambed quality and the poor water quality together are expected to reduce the natural recruitment success of DR trout population. This problematic is even increased by the fact, that many impassable physical barriers fragment the stream. Avoidance behavior of trout, such as temporal upstream migration or migration into side streams during spawning, is impossible. Thus the observed negative impacts on early life stages reflect directly the risk for a poor natural recruitment success of wild trout populations located DR. 42

References

Alabaster, J.S., Lloyd. R., 1980. Ammonia. In: Water quality criteria for freshwater fish. Butterworths, London, England, pp. 85-102. Calamari, D., Marchetti, R., Vailati, G., 1977. Effetti di trattamenti prolungati con ammoniaca su stadi di sviluppo del Salmo gairdneri. Novi Annali Ig. Microbiol. 28, 333-345. Chapman, D.W., 1988. Critical review of variables used to define effects of fines in redds of large salmonids. Trans. Am. Fish. Soc. 117, 1-21. Crisp, D.T., 1989. Some impacts of human activities on trout, Salmo trutta, populations. Freshwat. Biol. 21, 21-33. Crisp, D.T., Carling, P.A., 1989. Observation on siting, dimensions and structure of salmonid redds. J. Fish Biol. 34, 119-134. Crisp, D.T., 2000. Trout and Salmon; Ecology, Conservation and Rehabilitation. Blackwell Science, Oxford: 212 pp. De Vries, P., 1997. Riverine salmonid egg burial depths: review of published data and implications for scour studies. Can. J. Fish. Aquat. Sci. 54, 1685-1698. DEW, 1996. Deutsche Einheitsverfahren zur Wasseruntersuchung, Bd. II, VCH Weinheim, New York, Basel, Cambridge, Tokyo; Beuth Berlin, Wien, Zurich. Embody, G.C., 1934. Relation of temperature to incubation periods of egg of four species of trout. Trans. Am. Fish. Soc. 64, 281-292. Emerson, K., Russo, R.C., Lund, R.E., Thurston, R.V., 1975. Aqueous ammonia equilibrium calculations: Effects of pH and temperature. J. Fish. Res. Board Can. 32, 2379-2383. Escher, M., 1999. Einfluss von Abwassereinleitungen aus Kläranlagen auf Fischbestände und Bachforelleneier . Bundesamt für Umwelt, Wald und Landschaft (BUWAL), Mitteilung zur Fischerei 61, p. 210. Garrett, J.W., Bennett, D.H., 1996. Evaluation of fine sediment intrusion into Whitlock-Vibert boxes. North American Journal of Fisheries Management 16, 448-452. Grost, R.T., Hubert, W.A., Wesche, T.A., 1991. Description of brown trout redds in mountain stream. Trans. Am. Fish. Soc. 120, 582-588. Harris, G.S., 1973. A simple egg box planting technique to estimate the survival of eggs deposited in stream gravel. J. Fish Biol. 5, 85-88. Hulsman, P.F., Powles, P.M., Gunn, J.M., 1983. Mortality of walleye eggs and rainbow trout yolksac larvae in low- pH waters of the LaCloche mountain area, Ontario. Trans. Am. Fish. Soc. 112, 680-688. Kobler, B, A. Peter, in preparation. Stress response of a brown trout population in river Wyna to treated wastewater inputs. Kobler, B. 2003. Hormonelle Störungen freilebender Bachforellen im Fluss Wyna ober- und unterhalb von Abwasserreinigungsanlagen. Report on behalf of the environmental protection agency of canton Lucerne, Switzerland, p.19. Luckenbach, T., Kilian, M., Triebskorn, R., Oberemm A., 2003. Assessment of the developmental success of brown trout (Samla trutta f. fario L.) embryos in two differently polluted streams in Germany. Hydrobiologia 490, 53-62. Luckenbach, T., Triebskorn, R., Müller, E., Oberemm, A., 2001. Toxicity of waters from two streams to early life stages of brown trout (Salmo trutta f. fario L.), tested under semi-field conditions. Chemosphere 45, 571- 579. McKim, J.M., 1977. Evaluation of tests with early live stages of fish for predicting long-term toxicity. J. Fish. Res. Board Can. 34, 1148-1154. Rubin, J.-F., 1995. Estimating the success of natural spawning of salmonids in streams. J. Fish Biol. 46, 603-622. 43

Rubin, J.-F., Glimsäter, C., 1996. Egg-to-fry survival of the sea trout in some streams of Gotland. J. Fish Biol. 48, 585-606. Segner, H., Marthaler, R., Linnenbach, M., 1988. Growth, aluminium uptake and mucous cell morphometrics of early life stages of brown trout, Salmo trutta, in low pH water. Environ. Biol. Fishes 21, 153-159. Stahel, A.W., 1999. Statistische Datenanalyse. Eine Einführung für Naturwissenschaftler, 2. Auflage, Vieweg Verlag, p. 382. Stöckli. A., Schmid, M., 1999. Zustand der aarauischen Fliessgewässer 1996/97. Bericht über die Wasserqualität. Sondernummer aus der Reihe UMWELT AARAU. p. 32. Ingendahl, D., Neumann, D., 1996. Possibilities for successful reproduction of reintroduced salmon in tributaries of the River Rhine. Arch. Hydrobiol. 113, 333-337. Turnpenny, A.W.H., Williams, R., 1980. Effects of sedimentation on the gravels of an industrial river system. J. Fish Biol. 17, 681-693. Vibert, R., 1949. Du repeuplement en truites et saumons par enfouissement de “boîtes d’alevinage” garnies d’œufs dans les graviers. Bulletin Français de Pisciculture 153, 125-150. von Westernhagen, H., 1988. Sublethal effects of pollutants on fish eggs and larvea. In: Hoar, W.S., Randall, D.J. (Eds.), Fish Physiology, vol. 11, Part A. Academic Press, San Diego, pp. 253-346. Wahli T., W. Meier, H. Segner & P. Burkhardt-Holm, 1998. Immunohistological detection of vitellogenin in male brown trout of Swiss rivers. Histochem. J. 30: 753-758. 44 45

4 STRESS RESPONSE OF A BROWN TROUT POPULATION IN THE RIVER WYNA TO TREATED WASTEWATER INPUTS

ABSTRACT

In 2000 and 2001, we surveyed brown trout (Salmo trutta fario L.) populations inhabiting the Wyna River. Barriers separate this river into three stretches. Loads of treated wastewater to the stretches were none, moderate, and high for the upper, middle, and lower sections, respectively. To characterize trout populations resident in these different river stretches, we measured density, biomass, age-structure, recruitment, Fulton’s condition, and growth rate. The health of the population in the highly polluted lower area was impaired; the population density was lowest and the population structure was over-aged. Trout exhibited a high growth rate and a high Fulton’s condition factor, indicating a population response pattern due to chronic recruitment failure that was additionally confirmed by the low natural recruitment of young of the year (<100 YOY/ha). There is evidence that other factors, such as riverbed quality and the interruption of the river connectivity also influenced recruitment. We conclude that a combination of multiple stressors including water pollution, habitat degradation and habitat fragmentation impaired trout populations in the Wyna River. 46

Introduction

Many streams and rivers are degraded by anthropogenic impacts such as pollution and habitat destruction. To understand the ecological consequences of such degradations, conceptual frameworks have been developed. Colby (1984) published a conceptual framework to characterize the status of fish populations which was later extended by Gibbons & Munkittrick (1994) to use fish as an integrative indicator for ecosystem health. As diagnostic variables for fish populations and individuals, measurement of density, mean age, recruitment of young of the year, growth rate, age at maturity, and fecundity were proposed. Since fish populations respond to environmental stress in a varied, but often characteristic manner, the pattern of diagnostic variables can give information about the response of a fish population to environmental stressors (Cleveland et al., 1999; Power, 1997).

Several empirical studies on fish indicated characteristic response patterns. They allow to distinguish between exploitation, recruitment failure, multiple stressors, food limitation or niche shift (Gibbons & Munkittrick, 1994). Response patterns were analyzed for fish populations in lakes, reservoirs (e.g. Munkittrick et al., 1991; Munkittrick, 1992; Gibbons & Munkittrick, 1994) and large (McMaster et al. 1996) and small streams (Fitzgerald, 1999). These analyses were made to evaluate the impacts of water pollution and habitat destruction.

Water from wastewater treatment plants (WWTP) can be an important source of pollution in flowing waters, being toxic or endocrine active for aquatic life (e.g. Clement, 2000; Sumpter, 1997). Consequences for fish are impaired health status, survival or reproductive success (e.g. Adams et al., 1989; Kime, 1995). Such effects may create bottlenecks in the fish life cycle with adverse consequences for the entire population (Kime, 1995; Sumpter, 1997). Nevertheless, little is known about the population’s response to such stress. How and to what extent a fish can compensate adverse effects from wastewater are open questions. In particular, empirical data from wild populations can answer these ecologically complex problems. For the present study, we collected empirical data from a free-living brown trout population influenced by water from treatment plants.

As a study site, we chose the Wyna River located in the center of Switzerland, which receives wastewater from four treatment plants. Water pollution in this river was found to cause adverse effects on the health status of trout (Escher, 1999). Other investigations documented effects on the hormone system of male trout, indicating complex pollution with endocrine active chemicals (Wahli et al., 1998; Kobler, 2004). Effects on the entire 47 population were expected especially because recent studies have shown that early survival was impaired, indicating problems concerning the population’s recruitment.

We studied trout populations in polluted downstream reaches and, as a control, in the less polluted upstream reaches. The goal was to characterize the density, mean age and recruitment success of the population in both river reaches to analyze the population’s stress response triggered by water pollution and habitat degradation. To identify the specific bottleneck for reproduction the results of this study river were compared with findings from other stressed fish populations described in the literature.

Material and Methods

Study areas in the Wyna River The Wyna River is located in the central part of Switzerland in a densely populated valley. The size of its watershed is 119 km2 and its land is used for agriculture (50%), forestry (32%), and residential purposes (14%). From the headwater (elevation: 640m, mean flow: 0.2 m3/s, slope: 1.4%) to the downstream reach (elevation: 430m, mean flow: 1.76 m3/s, slope: 0.8%) four wastewater treatment plants (WWTP) discharge into the river.

We selected six different study sections, each between 128 and 168 m long. Two sections were located in the headwater (HW), two in the middle reach (MR) and two in the downstream reach (DR) (Figure 1). Artificial barriers with a height of more than 1.2 meter prevent an upstream movement of fish between separate selected study sites.

The two sections in the headwater are not influenced by wastewater. Downstream of three WWTPs in the middle reach, the river contains about 42% wastewater. A forth wastewater input increases the wastewater portion to 50% in the downstream reach (Stöckli, pers. comm.).

Brown trout is the main game fish for anglers in the Wyna. As shown in Figure 1, the Wyna River is subdivided into three fishery sectors (leases). Sector A, 5.8 km long (2.9 ha) and sector C, 9.6 km long (6.7 ha), are normally stocked in summer with 0+ trout (A: 2’600 trout/ha; C: 4’500 trout/ha). Sector B, 9.8 km long (5.0 ha), is normally stocked in late autumn with 0+ trout (640 trout/ha) and 1+ trout (54 trout/ha). In 2001, the stocking in sectors A and C was halted to obtain a better estimate of the natural recruitment. 48

Figure 1: Study sites located in the headwater (HW), middle reach (MR) and downstream reach (DR) of the river Wyna. Four wastewater treatment plants discharge into the river: 1) WWTP Winon, 2) WWTP Oberes- Wynental, 3) WWTP Gontenschwil and 4) WWTP Mittleres-Wynental. The bar on the left side indicates the three fishery sections (A, B and C) in the river. River barriers more than 1.2 meter in height are shown as white bars. 49

The legal catch size of trout is > 24 cm total length. Between 1995 and 1999, an average of 31 trout/ha per year were caught in sector A, and 50 trout/ha per year were caught in sector B and C (Cantonal angling statistics).

Fish-habitat characteristics, temperature and flow regime We determined habitat parameters at all study sites at low flow in October 2001 and December 2001. A transect method was applied to record maximum depth and river width. Transects were 10 meters apart within each study site. Relative pool, riffle and glide habitats were quantified (Bisson et al., 1982). In addition, the relative area of cover for fish was estimated as described by Peter (1992).

Substrate size was visually measured using a view tube with a transparent bottom. The substrate type was classified using the modified Wentworth scale, and the average diameter of gravel was calculated (Heggenens et al., 1991). The “kick test” was used to estimate riverbed quality to describe the riverbed clogging at three to five randomly selected points. In addition, in December 2001 after the completion of spawning, redds were visually counted for three long sections (1.4 - 2.3 km long) in the headwater, middle and downstream reaches to estimate spawning activity.

The temperature regime was recorded during the study period in two hour intervals in the headwater, middle and downstream reaches using data loggers (VEMCO, Ltd, Shad Bay, Nova Scotia, Canada, precision ± 0.07 °C). Stream flow data were available for the downstream reach near Suhr since 1980 (Swiss National Hydrological and Geological Survey, Bern).

Population survey of brown trout We surveyed trout populations at all study sites in November 2000 and April, October and December 2001. The abundance was estimated by three-pass electrofishing removals with a stationary unit (smooth-DC, EFKO 7.5KW). Fish were held in aerated tanks and anesthetized with a solution of clove oil (0.03 ml/L, Hänseler AG, Switzerland; Anderson et al., 1997) or neutralized MS 222 (100 mg/L, 3-aminobenzoic acid ethyl ester, Fluka, Buchs).

Total length and body weight were measured to the nearest millimeter and gram, respectively. During spawning season (Nov. 2000, Dec. 2001) we also recorded sex and maturity of trout. Scales were sampled for age determination. All trout larger than 80 mm in size (n=3107) were dye-marked by Alcian Blue injected with a Panjet (Wright Health Group Ldt, Dundee) on the ventral location (Hart & Pitcher, 1969). In 50 addition, 433 trout larger than 180 mm in size were individually tagged with visible-implanted tags (Haw et al., 1990). After recovery from the anaesthetic, the fish were released.

Age determination According to the Petersen method, we separated 0+ (n=976) and 1+ (n=1632) age classes with length frequency histograms (Bagenal, 1978). Frequent population estimates (4 within a 13-month period) increased reliability as growth of different age classes was traced. The age of trout older than one year (>1+) was determined either by scale analysis (n=166), by the Petersen method (n=369), or by mark-recapture (n=71).

Scale analyses were conducted as described by Bagenal (1978). Scales were cleaned (ultrasound; Elgasonic EC3, Elga SA, Basel) and fixed on a microscope slide (glue: Lesso 1487, Lesser, Erlinsbach). Slower growth in winter was visible as an annulus on the scales. The numbers of annuli were determined using a projection microscope (Reichert, Austria), and the method was standardized with scales from marked and recaptured trout with a known age (Beamish & McFarlane, 1983).

Variables measured for trout and trout populations Population density and sampling variance were calculated with a maximum-likelihood estimator for the case of three removals (Ricker, 1975). For these calculations we used the program Microfish 3.0 (Van Deventer & Platts, 1989). The mean weight, biomass and variance of the standing stock were estimated according to Newman & Martin (1983). We also calculated the mean age of the trout population.

Egg densities were calculated from the biomass of mature females. We used a fecundity index (4.2 eggs/g trout) that was previously estimated from brown trout in the nearby Wigger River (Peter, 1987).

Sex ratios were recorded as numbers of phenotypic males and females in November 2000 and December 2001. Length (l) and weight (w) of trout was used to calculate the Fulton’s condition factor (FCF = 100w / l 3). Growth was measured for marked and recaptured trout.

Available were the initial size (Lt1) when caught the first time (t1) and the terminal size (Lt2) when recaptured (t2). Growth rates (GR) were calculated according to Bagenal (1978):

Model 1: GR = 100 ln ((Lt2-Lt1)/(t2-t1))

The ratio of immigrated and stationary trout were considered as numbers of unmarked and marked, recaptured trout. 51

Data analysis

The condition (Ki) of trout resident at different locations (Ai; Levels HW, MR, DR) was analyzed with a one-way analysis of variance (ANOVA). The full model corrected for effects of age whereas cosine and sine terms considered seasonal variation (Model 2). The significance of seasonal variation was analyzed as a model comparison between the full model and the reduced model without cosine and sine terms.

Ê2p ˆ Ê2p ˆ Model 2: AK ii i sinx Á ⋅++= i˜+cosx Á ⋅xi˜ (T = 1) ËT ¯ ËT ¯

Growth data from individual marked and recaptured trout were analyzed with a three-way ANOVA. We considered growth in length during summer (April-October) and during autumn (October-December) as factor “Season”, and trout age (young trout: ≤1+, older trout: >1+) as factor “Age”. Factor “Location” considered the location where trout were resident (HW, MR, DR).

ANOVA’s were calculated using the statistical program package S-Plus (Version 2000, Lucent Technologies Inc.). Data were checked for the assumption of normality and variance homogeneity. P-values of multiple comparisons were corrected according to Bonferroni (Stahel, 1999).

Results

Fish habitat characteristics, temperature and flow regime Table 1 summarizes the habitat variables measured at all studied sites. The maximum depth and width in middle (MR) and downstream reach (DR) study sites were similar, whereas at headwater (HW) sites they were significantly smaller. All sites revealed variable maximum depths (CV: 27 - 61%) and width (CV: 10 - 32%).

Pools, riffles, and glides were present at most studied sites. Riffle habitats were dominant (>70%) particularly in the headwater sections. All study sites offered cover area for large fish, which covered between 6 and 29 % of their surface.

Riverbeds of all studied sections were clogged and their substrate compositions were dominated by middle-sized (2 – 5 cm) and coarse (> 5 cm) gravel. These riverbeds were used by trout for spawning. The observed redd densities were highest in the HW area (38 redds/ha) and low in the DR (6 redds/ha) and MR area (10 redds/ha). 52

Table 1: Habitat characteristics measured at all six study sites located in the river Wyna.

Area of hydraulic habitats Riverbed Analysis of transects and cover characterises

Mean width Mean max. depth Mean grain Clogged in cm (CV*) in cm (CV*) Pools RifflesGlides Cover size in cm riverbed Headwater-1 550 32 6% 86% 8% 6% 2.9 Yes (11.8) (31.2)

Headwater-2 480 35 17% 72% 11% 16% 3.7 Yes (31.8) (60.8)

Middle 740 51 0% 53% 47% 8% 3.7 Yes Reach-1 (10.0) (26.9)

Middle 720 65 29% 0% 71% 29% 3.5 Yes Reach-2 (12.9) (40.1)

Downstream 750 65 22% 44% 34% 28% 3.8 Yes Reach-1 (22.2) (50.6)

Downstream 730 63 9% 49% 42% 19% 3.0 Yes Reach-2 (12.7) (28.9) *Coefficient of variation (100 s/ average of x)

Figure 2 shows the temperature and discharge in the Wyna during the study period. Temperatures in the DR and MR were similar (MR - DR; -0.1 ± 0.40 °C), whereas the HW temperature was colder (HW - DR; -1.1 ± 1.05 °C). Maximum values were about 20 °C. High discharge occurred in February and June 2001 with a maximum of 32 m3/s, which was smaller than 19 other flood events documented since 1980. Maximum discharge since 1980 was 63 m3/s.

Characteristics of trout population We caught between three and five fish species at the MR and the DR sites, but brown trout was dominant in terms of fish biomass (Table 2). At HW sites, brown trout was the only resident fish species.

Figure 3 illustrates that the trout density was lowest at downstream sites. In particular, young trout were missing. As a result, the mean age in the population was highest, and downstream sites revealed an over-aged population structure (Table 3).

The biomass of trout varied spatially and seasonally (Table 3). The highest biomass was found at one study site in the middle reach (MR-2), whereas all other sites revealed a similar, but lower biomass. Due to seasonal variation, the biomass decreased in winter (Nov. 2000 - 53

Apr. 2001), increased in summer (Apr. 2001 - Oct. 2001), and decreased in autumn (Oct. 2001 - Dec. 2001). As exceptions MR-1 and DR-2 sites did not reveal such a pattern. At these sites we found an increase of biomass in winter and a decrease in summer.

Table 2: Fish assemblage at study sites during investigation. Biomass of brown trout relative to the total biomass of fish.

Headwater-1 Headwater-2 Middle Middle Downstream Downstream Reach-1 Reach-2 Reach-1 Reach-2

Order of Brown trout Brown trout Bullhead1 Stone loach2 Stone loach2 Stone loach2 abundance Brown trout Brown trout Brown trout Brown trout Stone loach2 Bullhead1 Bullhead1* Chub4* Perch5* Bullhead1* Rainbow trout3*

Proportion of trout to total 100% 100% 69% 92% 79% 70% fish biomass

1 Cottus gobio, 2 Barbatula barbatula, 3 Oncorhynchus mykiss, 4 Leuciscus cephalus, 5 Perca fluviatilis *Single or rare individuals

During spawning seasons, we observed mature male and female trout at all study sites. At downstream study sites, the sex ratios were consistently female biased. At all other study sites we noticed a male biased sex ratio (Table 4).

Table 4 shows that female trout from the downstream reach were larger and heavier than females from sites in the middle reach and the headwater. Considering the biomass, we calculated a large egg-stock from females at downstream reach (DR) study sites.

Despite large egg-stocks at DR sites, the resulting recruitment success, as young of the year, was low (< 100 YOY/ha in Oct. 2001). Recruitment in the middle reach and headwater was significantly higher (Table 4).

Immigration of trout into studied sections Table 3 presents the immigration of trout into the studied sites. We found a high proportion trout that immigrated between November 2000 and April 2001 into MR and DR sites (Table 3). This result was dominated by immigrating 1+ (age class 1999) and 2+ (age class 2000) trout. Older trout (> 2+) were more stationary and between 50% and 100% of these trout were recaptured. Following surveys later in the year revealed clearly lower immigration rates at study sites in the middle and downstream reach. 54

Figure 2: Upper graph: temperature regime (°C) in headwaters (dotted line) and in the downstream reach (solid line) during study as daily mean values. Lower graph: Hourly mean flow rate in the downstream reach near Suhr. Dotted lines show median and percentiles (50%, 95%, 99%, 99.9%) of a twenty-year discharge period measured at this site. The maximum measured flow rate was 63 m3/s. 55

Figure 3: Population density of brown trout at six study sites located in the headwater (HW), middle (MR) and downstream (DR) reaches of the river Wyna. Density is given for all sampling events in November 2000, April 2001, October 2001 and December 2001. Upper 95% confidence limits are shown as horizontal bars. Different age classes are distinguished by different colors.

56

2.3 2.3

1.9 1.9

newly immigrated immigrated newly

of of

1.8 1.8

1.6 1.6 1.9

years years

in in

structure structure

0.6 0.6 1.9 2.5

0.7 0.7 2

0.6 0.6

0.6 0.6 1.5 1.9

0.4 0.4 1.6 1.8

April October December December October April

2001 2001 2001 2001

Population Population

as mean age age mean as

2 2

1.7 1.7

1.5 1.5

1.5 1.5

1.2 1.2 0.5

1.2 1.2

30% 30%

10% 10%

at all studied sites. In addition, the proportion proportion the addition, In sites. studied all at

December 2000 December

2001 2001

newly newly

trout trout

of of

57% 57%

October October

immigrated immigrated

Proportion Proportion

and December December and

96% 96% 71% 45%

85% 85% 52%

40% 40% 47% 18%

70% 70% 83% 9%

100% 100%

100% 100% 59% 40%

2001 2001

71) 71)

61) 61)

-94) -94)

45 45

69 69

93 93

58 58

69 69

150 150

2001 2001 April

(45-

(58 -63) -63) (58

(69-

(93 -163) -163) (93

(150-154) (150-154)

77 77

102 102

(118-122) (118-122)

biomass biomass

as kg/ha kg/ha as

November 2000, April 2001, October October 2001, April 2000, November

Trout Trout

77 77 82

67 67 54

97 97 118

in in

144 144 187

2001 2001 2001

(50 -51) (76 -79) -79) (76 -51) (50

(97-101) (97-101)

(142 -149) (186 -195) -195) (186 -149) (142

parentheses parentheses

in in

67)a 67)a

-163) -163)

trout caught caught trout

59 59

70 70 59

39 39

64 64 50

149 149

159 159

2000 2000

of of

(39 -80) -80) (39

(69 -71) (59 -60) (102 -103) -103) (102 -60) (59 -71) (69

{63 - {63

November April October December November- April- October- November November October- April- November- December October April November

(158 (158

Reach-2 (59-115) (66 -74) (54 -58) -58) (54 -74) (66 (59-115) Reach-2

Reach-1 (147 -168) (77 -83) (82 -89) (65 -89) (82 -83) (77 -168) (147 Reach-1

Downstream Downstream

Downstream Downstream

trout and the population structure as mean age are given. given. are age mean as structure population the and trout

Headwater-2 Headwater-2

Headwater-1 Headwater-1

Middle Reach-2 Reach-2 Middle

Middle Reach-1 Reach-1 Middle

95%-cofidence intervals are given given are intervals 95%-cofidence

a a Table 3: Total biomass biomass Total 3: Table

57

80'000 80'000

66'000 66'000

37'000 37'000

45'000 45'000

105'000 105'000

109'000 109'000

Eggs/ha Eggs/ha

Egg-stock Egg-stock

Calculated Calculated

54 54

121 121

± 17 17 ±

62±12 62±12

75 75

g±S.D g±S.D

2001 2001

186 ± ± 186

312 ± 138 138 ± 312

October October

91) 91)

in in

45 45 ± 425

80 80

84 84

173 173

117 117

(45 -64) -64) (45

(80 -94) -94) (80

(84-

Reproductive potential potential Reproductive

(102-105) (102-105)

(173-181) (173-181)

(117-130) (117-130)

the Mature Mean Weight of of Weight Mean Mature the

in in

57% 57%

75% 75%

38% 38% 50 ± 102 254

38% 38%

52% 52%

29% 29%

population population

reproductive females females females females reproductive

2001 2001

90 90

80 80

370 370

620 620

853 853

661 661

(80-86) (80-86)

(72-185) (72-185)

(318-479) (318-479)

success success

(599-667) (599-667)

(853-879) (853-879)

(661-675) (661-675)

Reproductive Reproductive

density and mean weight of mature female trout. The reproductive success is given as as given is success reproductive The trout. female mature of weight mean and density

as as

2001 2001

'000 '000

-stock in October October in -stock

59'000 59'000

34'000 34'000

281

Eggs/ha Trout/ha Ratio of females Trout/ha Trout/ha females of Eggs/ha Ratio Trout/ha

Egg

Calculated Young trout of the year Sex ratio ratio Sex year the of trout Young Calculated

±44 ±44

±34 ±34

76 76

88 88

g±S.D g±S.D

283 ± 65 65 ± 283

365 ± 136 289'000 136 289'000 ± 365

492 ± 198 153'000 198 153'000 ± 492

8 8

November 2000 and October October and 2000 November

-90) -90)

November 2000 2000 November

81 81

53 235 ± 59 48'000 59 ± 48'000 53 235

173 173

117 117

in in

in in

(81 (81

(53 -76) -76) (53

(205 -222) -222) (205

(257 -265) -265) (257

(173 -179) -179) (173

(117 -130) (117

Reproductive potential potential Reproductive

the Mature Mean Weight Weight Mean Mature the

in in

study sites sites study

parentheses parentheses

in in

in in

population population

reproductive females of females females of females reproductive

Sex ratio ratio Sex

Ratio of females Trout/ha Trout/ha females of Ratio

the year. year. the

of of

Headwater-2 50% Headwater-2

Headwater-1 33% Headwater-1

Middle Reach-2 64% Reach-2 257 Middle

Middle Reach-1 44% Reach-1 Middle

young young

Downstream Reach-2 75% Reach-2 Downstream

Downstream Reach-1 58% Reach-1 205 Downstream

95%-cofidence intervals are given given are intervals 95%-cofidence

a a Table 4: Reproductive potential potential Reproductive 4: Table 58

Fulton’s condition Figure 4 illustrates the Fulton’s condition of trout in November 2000 and April, October and December 2001. Trout from the HW sites generally revealed a median condition below 1, in contrast to trout from the MR and DR sites. The median level of condition varied seasonally. In summer, levels were higher than in autumn and winter.

Figure 4: Fulton’s condition of trout at six study sites located in the headwater, middle and downstream reaches of the river Wyna. Condition is given for all samplings for November 2000, April 2001, October 2001 and December 2001. Box plots indicate the median, 75th and 95th percentiles. 59

The applied one-way ANOVA explains 26% of the total variance of trout condition. Model reduction confirms the significance of seasonal variation (F = 290; p < 0.001). Trout age as a

co-variable was significant (xi: F = 77; p < 0.001). Multiple comparisons show that the condition of HW trout was significantly (p < 0.05) lower than trout at the MR and HW sites.

Growth Growth rate in length was measured for 62 marked and recaptured resident trout from the HW (n = 14), MR (n = 25) and DR (n = 23) sites. The full three-way ANOVA explained 43% of the observed variance in growth, and interactions of factors (location x age x season) were not significant. As shown in Table 5 all three factors (location, age, season) significantly influenced the variance of growth.

Growth in summer (April - October) was faster than in autumn (October - December) and young trout (≤ 1+) grew faster than older trout (> 1+). Multiple comparisons reveal that resident trout in the DR grew significantly faster (p < 0.05) than resident trout in the MR and HW sites.

Table 5: Three-way ANOVA of length-growth rates of individually marked and recaptured trout in the river Wyna. Factors are: 1) the age of the trout (≤1+, >1+), 2) the season when growth took place (summer, autumn), and 3) the location where trout resided (HW, MR, DR). The table gives degrees of freedom, Sum of Squares, Mean of Squares, F-Value and P-Value of each factor.

Factors Degrees of Sum of Squares Mean of F-Value P-Value freedom Squares Age 1 0.028 0.028 22.0 < 0.001 Season 1 0.014 0.014 10.7 < 0,01 Location 2 0.013 0.006 5.0 < 0.05 Residuals 57 0.073 0.001

Discussion

Habitat and trout management We investigated trout populations inhabiting six different sites in the headwater (HW), middle (MR) and downstream reach (DR) to evaluate effects of wastewater discharge and habitat degradation. We compared population characteristics in differently impacted river reaches (HW, MR, DR). 60

Colby (1984) used a similar approach to compare stressed versus unstressed populations in lakes and reservoirs. Doing so, he mentioned the importance of controlling for habitat characteristics, since they are important factors influencing the density and structure of fish populations (e.g. Scarnecchia & Bergersen, 1987; Kozel & Hubert, 1989; Modde et al., 1991; Riley & Fausch, 1995). Our habitat measurements reveal similar habitats at study sites in the middle and downstream reach, offering large areas of pool and cover habitat. According to Jungwirth & Winkler (1983) such habitat characteristics indicate good environments for large trout. Furthermore, these sites also revealed a high similarity in riverbed structure and flow and temperature regime. On the other hand, habitat characteristics in the headwater sites were different from those at the middle and downstream reaches. The temperature was cooler and the habitat was dominated by riffles, which provided good habitat for juvenile trout. Because of these differences, the headwater sites were not perfect control sites for the polluted downstream reaches.

Besides habitat characteristics, fish management and trout immigration are also important factors to be controlled. The cantonal angling statistics for the Wyna River indicate that few fish were caught (31 trout/ha) in comparison to the density of legal-sized trout (> 24 cm) in MR (228 trout/ha) and DR sites (196 trout/ha). Therefore, angling only slightly influenced the populations in our study.

However, we noticed that stocking of trout caused significant effects on population structure and density. In December 2000, parr (640 trout-0+/ha) and yearling (54 trout-1+/ha) were introduced into sector B at the Wyna (Figure 1). Although this sector was located several kilometers upstream of the MR and DR study sites, it appears that stocking increased the density of the corresponding two age classes (1999 and 2000) (Figure 5). Effects immediately after stocking were noticed as increased density, increased biomass, and decreased population mean age. This effect can be seen as a single event, because there was no further stocking at the Wyna River.

The study sites in the Wyna River were separated by a number of barriers over 1.2 meter high which prevented upstream movement of trout in the river system. On the other hand, movement downstream is possible and thus would tend to equalize the population characteristics at downstream, middle and headwater reaches. Our results show that a large ratio of trout individuals was stationary. Furthermore, we found significant differences in the population characteristics at different polluted study sites indicating that trout movement had no large effects on our results. 61

600 ::R 0 •2000 .£ (/) Q) (/) 01999 (/) ro u 400 Q) Cl ro 0 -,..._ ...... -..._c: N -c: 200 0 ...... 0

Q) Cl c: ro .!: u 0 ~ (/) c: Q) HW-1 HW-2 MR-1 MR-2 DR-1 DR-2 0

-200

Figure 5: Relative change in population density between November 2000 and April 2001. Age classes 1999 (black bars) and 2000 (white bars) of trout are distinguished.

Population characteristics in polluted river reaches (DR) Populations in downstream and middle reaches were living in similar environments. Only the load with wastewater differed between these two stream reaches. Instead of headwater reaches, we therefore selected the middle reaches as controls for evaluating the effect of water pollution on trout. Characteristics such as density, mean age, growth performance and recruitment success were compared between populations from more and less wastewater­ polluted stream areas. We found that the population in the more polluted area was less dense, the recruitment success was lower and the population structure was over-aged. Furthermore, the comparison indicates that trout from the polluted DR grew faster. Table 6 summarizes this characteristic pattern as a difference between control (MR) and test (DR) populations.

Different authors report on specific patterns that are characteristic for specific population deficits. This information made it possible to compare the downstream population's pattern with other fish populations studied (Colby 1984; Munkittrick & Dixon, 1989a, 1989b; Gibbons & Munkittrick, 1994 ). This comparison shows that the downstream population's pattern is similar to patterns indicating a "chronic recruitment failure". Such a recruitment failure in fish 62 I I 0 0 0 I I I the river river the male male male female female 0 0 0 0 0 0 0 0 0 0 0 0 0 0 male male in in to to to to to to Sex ratio ratio Sex Shift to female female to Shift Shift Shift Shift Shift 0 0 0 Shift to to 0 Shift + + 0 0 0 0 0 + + + + - age age -10 -10 headwater. headwater. and and 0 0 0 0 + + - - - - ND ND Of+ Of+ -10 -10 change. change. reach reach Biomass Mean Mean Biomass significant significant 0 0 - 0 0 0 0 + + - - - -10 -10 size size -10 -10 -10 -10 -10 no a a the downstream downstream the in in b b identifies identifies c c "O" "O" - - - QC QC oc oc - NDd NDd and and +IOI- Munkittrick, 1994) compared to those identified for brown trout trout brown for identified those to compared 1994) Munkittrick, & & Fecundity Fecundity - Diagnostic Variable Variable Diagnostic decrease, decrease, 0 0 + + ------+ + + + a a Individual Population Population Population Population Individual identifies identifies at at "-" "-" 0 0 0 0 + + + + + + + - - ND ND ND ND ND a control for populations resident resident populations for control a Age Age maturity maturity interest. interest. as as (MR), (MR), of of taken taken 0 0 0 0 0 0 0 0 - + + 0 0 0 0 - - population population Factor Factor and and population population "O" "O" the the reference reference + + - - - - - + + + + + within within rate rate +10 +10 Growth Condition Condition Growth the the to to alteration alteration to to to to relative relative 0 0 0 0 0 0 0 0 0 0 0 0 the the Age Age older older older older older older older older of of Shift Shift Shift Shift Shift to to Shift Shift to to Shift Shift to to Shift younger younger younger younger structure structure increase increase an an the middle reach were set to to set were reach middle the magnitude magnitude in in the the on on "O") "O") identifies identifies IV) IV) II) II) on on a"+" a"+" Type Type Type Type with with depending depending to to limitation) limitation) vary vary recruitment recruitment the Wyna River River Wyna the on on in in variables variables may may (Chronic recruitment failure) failure) recruitment (Chronic (Metabolic redistribution) Shift to to Shift redistribution) (Metabolic (Food (Food (Multiple stressors) stressors) (Multiple (Recruitment failure) failure) (Recruitment (Niche shift) shift) (Niche response) response) Comparison of stress response patterns identified for white sucker (adapted from Gibbons Gibbons from (adapted sucker white for identified patterns response stress of Comparison (Exploitation) (Exploitation) V: V: VI: VI: Ill: Ill: IV: lia lia lib: I I Wyna. Populations resident resident Populations Wyna. 6: 6: based based (Reference was set set was (Reference (No (No (most similar to to similar (most determined determined DR DR MR MR HW (most similar similar (most HW "O" "O" Type Type Type Reaches Reaches Type Type Type Type Type Type Type Type not not Response pattern pattern Response Response Response Data Data ct ct 0 0 b b •Diagnostic •Diagnostic Table Table 63 can be the result of a reproductive dysfunction due to stress or pollution (Kime, 1995, 1998). To obtain information on the reproductive health, we investigated the egg quality of trout from the river Wyna to evaluate the population’s egg stock quality (Kobler et al., in prep.). We found that trout from the downstream reach had excellent egg quality (> 98% viable ova). Also densities of produced eggs in the downstream reach were high (80’000-105’000 Eggs/ha), indicating a large and qualitatively good egg stock.

Like growth, fecundity of fish is strongly dependent on the availability of food (Weatherley & Gill, 1987; Baccante & Reid, 1988; Haugen, 2000). The more food, the faster fish can grow and the more eggs they can produce. Fish in the downstream reach grew fast (Table 5). An excess of food could explain these growth rates and probably also the good egg stocks in the Wyna River downstream reach. Therefore egg stocks in the downstream reach were not limited and thus do not explain the observed reproductive bottleneck. Instead, later phases in the life cycle must be crucial for the population’s recruitment failure.

Factors impairing the population’s recruitment success Crisp (1989) highlighted the importance of good riverbed quality to reproductive success of trout. Our investigation of the Wyna riverbed in the downstream reach revealed an ideal substrate size (0.5 - 7.6 cm) for the spawning of trout (Reiser & Bjornn, 1979). However, the number of redds indicate that spawning activity was low although there were many potential spawners. The riverbed in the downstream reach was clogged with fine sediments, thus degrading the riverbed quality for spawning.

A recent study on the survival of in-gravel trout eggs shows that indeed fine sediments are a major disruptive factor in the Wyna River (Kobler & Peter, 2002). Fine sediments were suspected of being toxic and reducing in-gravel transport of oxygen and metabolic waste, causing egg/embryo mortality. Such problems for trout were recently reported in the German Körsch River. Comparable with the Wyna River, Luckenbach et al. (2001) identified fine sediments as a major impairment of early development in this German river.

Additionally, water quality in the Wyna River lower reach is poor due to the high wastewater loading. Long-term chemical measurements confirm that ammonia frequently exceeded tolerable levels (>0.2 mg/L ammonia) in the downstream reach (Stöckli & Schmid, 1997; Kobler et al. in prep.), in particular ammonia and probably endocrine active chemicals impaired early survival of trout in the Wyna River (Kobler & Peter, 2002). 64

Because fine sediments and ammonia impaired the early survival and, as a consequence, may have caused a “chronic recruitment failure”, the downstream reach was not suitable for reproduction. The population’s characteristics support such a causality.

Stream fragmentation and fish diseases Due to many impassable barriers in the river system, spawning trout cannot reach suitable spawning sites in less contaminated upstream reaches or tributaries. This river fragmentation causes populations in the downstream reach to be trapped in unsuitable environmental conditions for spawning, although their potential egg-stocks are good. Thus, local environmental pollution and degraded spawning habitats in combination with habitat fragmentation could endanger trout in the Wyna River.

Wahli & Escher (2000) found a high proportion of trout in the Wyna to be infected with proliferative kidney disease (PKD). At water temperatures above 15 °C, PKD can cause high mortalities particularly of young trout (Hedrick et al., 1993). The same authors also found that water pollution can increase mortality rates due to PKD. Thus, it is possible that wastewater and PKD together may have increased the mortality of studied populations. However, we have no detailed information about mortality rates of trout to give safe conclusions about the influence of PKD in the Wyna River.

Conclusion Trout in the headwaters were living in different habitats compared with trout in middle and downstream reaches. Differences were also noticed in the population density, mean age and recruitment success, summarized in Table 6 (difference between MR and DR population). Most characteristic for headwater trout was the low Fulton’s condition factors. According to Weatherley & Gill (1987) and Jones (1989) low condition factors can indicate a low feeding status due to food limitation. Food limitation can reduce growth and reproductive potential of trout (Elliott, 1994). Both growth performance and fecundity were limited for headwater trout and thus indicated that food in the headwaters was limited relative to middle and downstream reaches. The reason could be that population densities in the headwaters were high relative to the quantity of available food. Such competition for limited resources can indicate a population size which has reached the carrying capacity of the habitat.

In the middle and downstream reaches, fast growth rates and high fecundity indicated a better availability of food and thus a population density lower than the carrying capacity of their habitats. 65

The present study indicates that anthropogenic impacts such as habitat degradation and water pollution limited the density of trout populations. In particular, recruitment failure due to low early survival was a major problem for Wyna trout populations.

According to Colby (1984) recruitment failure normally results in a collapse of the population. Due to regular stocking with trout in the context of the Swiss trout-managing program, such a collapse of trout populations in the Wyna River was prevented.

References

Adams, S.M., K.L. Shepard, M.S. Greely, B.D. Jimenez, M.G. Ryon, L.R. Shugart & J.F. McCarthy, 1989. The use of bioindicators for assessing the effects of pollutant stress on fish. Marine Environmental Research 28: 459-464. Anderson, W.G., R.S. McKinley & M. Colavecchia, 1997. The use of clove oil as an anesthetic for rainbow trout and its effects on swimming performance. North Am. J. Fish. Manage. 17: 301-307. Baccante, D.A. & D.M Reid, 1988. Fecundity changes in two exploited walleye populations. North Am. J. Fish. Manage. 8: 199-209. Bagenal, T.B., 1978. Methods for the assessment of fish production in fresh waters. IBP Handbook, 3rd edition, Blackwell Scientific publications, Oxford, London, 365 pp. Beamish, R.J. & G.A. McFarlane, 1983. The forgotten requirement for age validation in fisheries biology. Trans. Am. Fish. Soc. 112:735-743. Bisson, P.A., J.L. Nielsen, R.A. Palmason & L.E. Grove, 1982. A system of naming habitat types in small streams, with examples of habitat utilization by salmonids during low streamflow. In Armantrout, N.B. (ed.), Acquisition and utilization of aquatic habitat inventory information. American Fisheries Society, Western Division Hagen, 62-73. Clement, W.H., 2000. Integrating effects of contaminants across levels of biological organization: an overview. J. Aquat. Stress. Recovery 7: 113-116. Cleveland, L., J.F. Fairchild & E.E. Little, 1999. Biomonitoring and ecotoxicology: Fish as indicators of pollution- induced stress in aquatic systems. Environmental Science Forum 96: 195-232. Colby P.J. & S.J. Nepszy, 1981. Variation among stocks of walleye (Stizostedion vitreum vitreum): Management implications. Can. J. Fish Aquat. Sci. 38: 1814-1831. Colby, P.J., 1984. Appraising the status of fisheries: Rehabilitation techniques. In: Cairns, V.W, P.V. Hodson & J.O. Nriagu (eds). Contaminant effects on Fisheries. Vol. 16. Adv. Envir. Sci. Technol.: 233-257. Crisp, D.T., 1989. Some impacts of human activities on trout, Salmo trutta, populations. Freshwat. Biol. 21: 21-33. Elliott, J.M., 1994. Quantitative ecology and the brown trout. Oxford University Press, Oxford, New York, Tokyo, 286 pp. Escher, M., 1999. Einfluss von Abwassereinleitungen aus Kläranlagen auf Fischbestände und Bachforelleneier. Swiss Agency of the Environment, Forests and Landscape (SAEFL), Mitteilungen zur Fischerei No. 61: 210 pp. Fitzgerald, D.G., R.P. Lanno & D.G. Dixon, 1999. A comparison of a sentinel species evaluation using creek chub (Semotilus atromaculatus Mitchill) to a fish community evaluation for the initial identification of environmental stressors in small streams. Ecotoxicology 8: 33-48. 66

Gibbons, W.N. & K.R. Munkittrick, 1994. A sentil monitoring framework for identifying fish population responses to industrial discharges. J. Aquat. Ecosys. Health 3: 227-237. Hart, P.J.B. & T.J. Pitcher, 1969. Field trails of fish marking using a jet inoculator. J. Fish Biol. 1: 383-385. Haugen, T.O., 2000. Growth and survival effects on maturation pattern in populations of grayling with recent common ancestors. Oikos 90: 107-118. Haw, F., P.K. Bergman, R.D. Fralick, R.M. Buckley & H.L. Blankenship 1990. Visible Implanted. In: Parker, N.C, A.E. Giorgi, R.C. Heidinger, D.B. Jester, E.D. Prince & G.A. Winans, (eds), Fish-Marking Techniques. Am. Fish. Soc. Symp. 7, Bethesda, Maryland: 311-315. Hedrick, R.P., D. Monge, P. De Kinkelin, 1993. Proliferative kidney disease of salmonid fish. Annual Rev. Fish Diseases 3: 277-290. Heggenens, J., T.G. Northcote & A. Peter, 1991. Spatial stability of cutthroat trout (Oncorhynchus clarki) in a small, coastal stream. Can. J. Fish. Aquat. Sci. 48: 757-762. Jones, R., 1989. Towards a general theory of population regulation in marine teleosts. Journal de Conseil, Conseil International pour l'Exploration de la Mer 45: 176-189. Jungwirth, M & H. Winkler, 1983. Die Bedeutung der Flussbettstruktur für Fischgemeinschaften. Österreichs Wasserwirtschaft 35: 229-234. Kime, D.E., 1995. The effect of pollution on reproduction in fish. Rev. Fish Biol. Fish. 5: 52-96. Kime, D.E., 1998. Endocrine disruption in fish. Kluwer Academic Publishers, Boston, London, 396 pp. Kobler, B., F. Wettstein & A. Peter, in prep. Is the egg quality of Brown trout (Salmo trutta fario L.) in a polluted stream reduced? Kobler, B & A. Peter. 2002. Veränderungen bei Bachforelleneiern und -brütlingen in einem belasteten Fluss. fischnetz-Info 9: 15-17. Kozel, S.J. & W.A. Hubert, 1989. Testing of habitat assessment models for small trout streams in the Medicine Bow National Forest, Wyoming. North Am. J. Fish Manage. 9: 458-464. Luckenbach, T., M. Kilian, R. Triebskorn & A. Oberemm, 2001. Fish early life stage tests as a tool to assess embryotoxic potentials in a small stream. J. Aquat. Ecosyst. Stress & Recovery 8: 355-370. McMaster, M.E., G.J. Vanderkraak & K.R. Munkittrick, 1996. Exposure to bleached kraft pulp-mill effluent reduces the steroid biosynthetic capacity of white sucker ovarian follicles. Comp. Biochem. Physiol. 112C: 169-178. Modde, T., R.C. Ford & M.G. Parsons, 1991.Use of habitat based stream classification system for categorizing trout biomass. North Am. J. Fish. Manage. 11: 305-311. Munkittrick, K.R. & D.G. Dixon, 1989a. A holistic approach to ecosystem health assessment using fish population characteristics. Hydrobiologia 188-189: 123-135. Munkittrick, K.R. & D.G. Dixon, 1989b. Use of white sucker (Catostomus commersoni) populations to assess the health of aquatic ecosystems exposed to low-level contaminant stress. Can. J. Fish. Aquat. Sci. 46, 1455- 1462. Munkittrick, K.R., P.A. Miller, D.R. Barton & D.G. Dixon, 1991. Altered performance of white sucker populations in the Manitouwadge chain of lakes is associated with changes in betic macroinvertebrate communities as a result of copper and zinc contamination. Ecotoxicol. Environ. Safety 21: 318-326. Munkittrick, K.R, 1992. A review and evaluation of study design considerations for site-specifically assessing the health of fish populations. J. Aquat. Ecosys. Health 1: 283-293. Newman, R.M. & F.B. Martin, 1983. Estimation of fish production rates and associated variances. Can. J. Fish. Aquat. Sci. 40: 1729-1736. 67

Peter, A., 1987. Untersuchungen über die Populationsdynamik der Bachforelle (Salmo trutta fario) im System der Wigger mit besonderer Berücksichtigung der Besatzproblematik. Ph.D. Thesis, ETH Diss. Nr. 8307, Swiss Fed. Inst. Of Technol., Zurich, Switzerland, 247 pp. Peter, A., 1992. Analyse von Fischmikrohabitaten zur Beurteilung der strukturellen Komplexität eines Fliessgewässers. Swiss Federal Institute for Environmental Science and Technology (EAWAG), Annual Report, 60-61. Power, M., 1997. Assessing the effects of environmental stressors on fish populations. Aquat. Tox. 39: 151-169. Reiser, D.W. & Bjornn, 1979. Habitat requirements for anadromous salmonids. Influence of forest and rangeland management on anadromous fish habitat in the western united states and Canada, Rep. No. PNW96. USDA Forest Service Technical Report. Ricker, W.E., 1975. Computation and interpretation of biological statistics of fish populations. Bull. Fish. Res. Bd Can. Rep, No. 191: 1- 382. Riley, S.C. & K.D. Fausch, 1995. Trout population response to habitat enhancement in six northern Colorado streams. Can. J. Fish. Aquat Sci. 52: 34- 53. Scarnecchia, D.L. & E.P. Bergersen, 1987. Trout production and standing crop in Colorado's small streams, as related to environmental feature. North Am. J. Fish. Manage. 7: 315-330. Stahel, W., 1999. Statistische Datenanalyse. Eine Einführung für Naturwissenschaftler, second edition, Vieweg Verlag, Zürich, 377 pp. Sumpter, J.P., 1997. Environmental control of fish reproduction: a different perspective. Fish Physiol. Biochem. 17: 25-31. Stöckli, A. & M. Schmid. 1997. Zustand der aargauischen Fliessgewässer 1996/97. Bericht über die Wasserqualität. Sondernummer aus der Reihe UMWELT AARGAU, Redaktion: [email protected]: 31 pp. Van Deventer, J.S. & W.S. Platts, 1989. Microcomputer software system for generating population statistics from electrofishing data - user's guide for MicroFish 3.0. U.S. For. Ser. Gen. Tech. Rep. INT-254. Wahli T., W. Meier, H. Segner & P. Burkhardt-Holm, 1998. Immunohistological detection of vitellogenin in male brown trout of Swiss rivers. Histochem. J. 30: 753-758. Wahli, T. & M. Escher, 2000. Verbreitung der Proliferativen Nierenkrankheit (PKD) in der Schweiz, Petri Heil, 17: 23-25. Weatherley, A.H. & H.S. Gill, 1987. The biology of fish growth. Academic Press, New York, 443 pp. 68 69

5 Combined biological and chemical assessment of estrogenic activities in wastewater treatment plant effluents

Hans-Rudolf Aerni†§, Bernd Kobler†§, Barbara V. Rutishauser†§, Felix E. Wettstein†§,

René Fischer‡, Walter Giger†, Andreas Hungerbühler†, M. Dolores Marazuela#, Armin Peter†,

René Schönenberger†, A. Christiane Vögeli†, Marc J.-F. Suter†*, Rik I.L. Eggen†

†Swiss Federal Institute for Environmental Science and Technology (EAWAG), CH-8600

Duebendorf, Switzerland

‡Swiss Federal Institute of Technology (ETH), CH-8092 Zurich, Switzerland

#Complutense University, E-28040 Madrid, Spain

§Authors marked with a § equally contributed to the manuscript and should be considered as first

authors.

*Corresponding author

Marc J.-F. Suter, Swiss Federal Institute for Environmental Science and Technology,

Ueberlandstr. 133, P.O. Box 611, CH-8600 Dubendorf, Switzerland

Phone: +411 823 54 79, Fax: +411 823 53 11, e-mail: [email protected]

Published in Analytical and Bioanalytical Chemistry 70

Abstract

Five wastewater treatment plant effluents were analyzed for known endocrine disrupters and estrogenicity. Estrogenicity was determined by using the yeast estrogen screen (YES) and by measuring the blood plasma vitellogenin (VTG) concentrations in exposed male rainbow trout (Oncorhynchus mykiss). While all wastewater treatment plant effluents contained measurable concentrations of estrogens and gave a positive response with the YES, only at two sites did the male fish have significantly increased VTG blood plasma concentrations after the exposure, compared to pre-exposure concentrations. Estrone (E1) concentrations ranged up to 51 ng L-1, estradiol (E2) up to 6 ng L-1, and ethinylestradiol (EE2) up to 2 ng L-1 in the 90 samples analyzed. Alkylphenols, alkylphenolmono- and -diethoxylates, even though found at µg L-1 concentrations in effluents from wastewater treatment plants with a significant industrial content, did not contribute much to the overall estrogenicity of the samples taken, due to their low relative potency. Expected estrogenicities were calculated from the chemical data for each sample using the principle of concentration additivity and relative potencies of the various chemicals as determined with the yeast estrogen screen. Measured and calculated estradiol equivalents gave the same order of magnitude and correlated rather well (R2 = 0.6).

Keywords Endocrine disruption · Fish exposure · Yeast estrogen screen · Vitellogenin · Steroid hormones · Nonylphenol ethoxylates 71

Introduction

Endocrine-disrupting chemicals (EDCs) in the aquatic environment have become an increasingly important issue for scientists and regulatory bodies. As early as 1970, Tabak and Bunch were investigating the risk posed by natural urinary and especially synthetic ovulation-inhibiting hormones (1; 2), work that was later taken up by other groups (3; 4; 5; 6; 7). The impact of the steroid hormones on the environment is highlighted by the fact that concentrations as low as 0.1 ng L-1 of ethinylestradiol (8), and between 1 and 10 ng L-1 of estradiol (9) induce vitellogenesis in male rainbow trout. This high potency at low concentrations necessitates understanding of the fate and effects of these compounds in the aquatic environment. In the early 1990s various man-made chemicals also were identified as disrupters of the endocrine system (10; 11) and their fate and behavior became a major research topic (12). Since then, accumulating evidence suggests that these chemicals also may affect the health and possibly the fertility of humans and wildlife (13; 14; 15).

To predict environmental concentrations of steroid hormones in river waters, Johnson and co-workers used an exposure assessment model (16; 17; 18). They expected concentrations below 0.4 ng L-1 for estradiol and estrone, and 10 times less for ethinylestradiol. This expectation agrees well with values in the ng L-1 range reported by various groups for wastewater treatment plant effluents when the dilution of the effluent in the receiving rivers is taken into account (5; 7; 19; 20; 21; 22; 23; 24).

The very low environmental concentrations expected for natural and synthetic steroid hormones require sophisticated analytical techniques for their measurement, and to assess their effect on aquatic organisms. Biological tools are needed to measure effects, which are integrated responses to all estrogenic chemicals present, including those that might be missed by chemical analysis. Hence, in vitro testing of environmental samples and chemicals has become an important complementary technique to classical chemical analysis (25; 26; 27; 28). It allows targeting known and unknown chemicals and mixtures of chemicals that cause specific effects at the cellular or molecular level, and it can be used for high- throughput screening. By using the yeast estrogen screen (YES) as an in vitro test system, effects can be measured as a result of additivity in a mixture of EDCs, even if the concentration of individual EDCs is below the no-effect concentration (29). However, in vitro tests represent an artificial system, and their responses should be validated with in vivo experiments (30). For this reason, the induction of vitellogenin (VTG) in juvenile or male fish in vivo has become an accepted biomarker of exposure to estrogenic compounds (8; 31). 72

In the work presented here we have applied a combination of chemical analysis and various bio-analytical tools (in vitro and in vivo) for assessing estrogenicity in wastewater treatment plant (WWTP) effluents and their receiving waters. We have used this approach to correlate chemical data with biological effects and could show that: i) natural and synthetic steroid hormones are the major contributors to estrogenicity in WWTP effluents, ii) the measured (YES) and calculated (based on chemical analysis) estrogenicity correlated well, and iii) no false negative results were obtained in vitro, since the VTG induction in exposed male fish was matched by the response of the YES assay.

Materials and methods

Chemicals The standards b-estradiol-17-acetate, 17-a-ethinylestradiol, 17-b-estradiol, estriol, estrone, and 2,4,6-trimethylphenol were from Sigma-Aldrich (Buchs, Switzerland). E2 and EE2 used in the YES were from Fluka (Buchs, Switzerland), 1,4-dithioerythritol (DTE, >99 %) was from VWR Int. AG (Dietikon, Switzerland), 1-(trimethylsilyl)imidazole (T(M)SIM, >98 %) and N- methyl-N-trimethylsilyltrifluoroacetamid (MSTFA, ~99 %) from Fluka (Buchs, Switzerland). All solvents were puriss p.a., or HPLC-grade.

Sample collection for chemical and in vitro analysis Three-day composite water samples were collected using cooled (5 °C) mobile sampling devices (ISCO Model 6700 Standard, IG Instrumenten-Gesellschaft, Zurich, Switzerland) fitted with acetone-rinsed 4-L glass or aluminum (France1 and France2) bottles. Sampling locations were WWTP effluents, the river upstream of the point of discharge, and Lake Lucerne at Kastanienbaum. To prevent sample degradation, samples were stabilized by adding 10 mL 36 % formaldehyde solution (NPnEO analysis, n = 0-2), or 10 mL methanol (steroid analysis and YES) per liter. The samples were stored at 4 °C and processed within 48 h (steroid analysis and YES) or 14 d (NPnEO analysis, n = 0-2), or frozen at -18 °C if enrichment was delayed. An internal standard (25 ng L-1) was added before the enrichment for steroid (E2-acetate) and NPnEO chemical analysis (2,4,6-trimethylphenol).

Enrichment for steroid analysis and in vitro testing Solid-phase extraction (SPE) was performed according to Ternes et al. (22). Briefly, 1 L of sample was vacuum filtered through a glass fiber filter (GF/F, 90 mm, Whatman International Ltd, Kent, UK) and the pH of the filtrate was adjusted to 3 with 16 % aqueous HCl. The 73 enrichment step was carried out using 3 mL self-packed polypropylene SPE tubes (Supelco, Buchs, Switzerland) containing a mixed solid phase (LiChrolut RP18 and LiChrolutEN; VWR Int. AG, Dietikon, Switzerland). After drying of the solid phase, analytes were eluted successively with 4x1 mL acetone. For steroid analysis, the solvent was reduced to 200 µL under a gentle stream of N2. The extract was then further cleaned on a 6 mL SPE column containing 1 g of silica gel (Sigma-Aldrich, Buchs, Switzerland), deactivated with 1.5 % H20 (w/w). Conditioning and elution were done using hexane/acetone 6:4 (v/v). The eluate was subsequently derivatized for GC/MS analysis (see below). For the in vitro testing, the acetone was evaporated completely and the residues were then taken up in 500 µL of absolute ethanol. Blank samples were prepared from Dubendorf groundwater (chemical analysis) or deionized water (YES) and extracted and treated in the same way as the environmental samples.

GC/MS analysis of steroid hormones The solid-phase extracts were derivatized according to the method reported by Ternes et al. (22), using MSTFA/TSIM/DTE, 1’000:2:2 (v:v:w) and separated on a GC Fisons 8’000 equipped with a PTV injector OPTIC 2 (ATAS, The Netherlands). We injected 10 ml volumes split-less onto a plug of deactivated fused silica (Restek, BGB Analytik AG, Anwil, Switzerland). The carrier gas was He (Carbagas, Zurich, Switzerland). The column was a XTI-5, 30-m, 0.25-mm ID, 0.25-mm film thickness (Restek) attached to a 2-m pre-column (Siltek, 0.53 mm ID, Restek). GC temperature program: 110 °C (4 min), 17 °C to 238 °C, 1.7 °C to 260 °C, 10 °C to 300°C, 300 °C isothermal (7 min).

High-resolution mass spectrometry (resolution 8000 at 5 % valley) with positive electron ionization on an Autospec-Q (Micromass, Manchester, UK) was used to detect steroid hormones. The transfer lines were set to 300 °C, and the source temperature to 280 °C. Optimized +EI conditions for steroid hormones resulted in an electron energy of 54 eV and a trap current of 700 mA. The detector was set to 375 V. Low boiling PFK (Fluka, Buchs, Switzerland) was used for calibration and lock masses. Single-ion monitoring was used during data acquisition.

Recoveries determined from two independent datasets (January 2000/November 2001) for groundwater samples spiked with 30 ng L-1 of the standard steroid mixture prior to or after the enrichment, were 114/106 % for E1, 117/100 % for E2, 119/100 % for EE2, and 40/25 % for estriol E3 (n = 8). For effluent samples, recoveries were 160/136 % for E1, 122/80 % for E2, 141/104 % for EE2, and 166/4 % for E3 (n = 8). The corresponding standard 74 uncertainties for groundwater samples were 15 % for E1, 22 % for E2, 25 % for EE2, and 173 % for estriol E3 (n = 8). For effluent samples, they were 60 % for E1, 23 % for E2, 42 % for EE2, and 116 % for E3 (n = 8). The highly varying recoveries and corresponding high uncertainties for E3 can be explained by the relatively high polarity of the compound, which means that recoveries are affected strongly by the activity of the silica column. Only slight changes in the activity can lead to large losses. Limits of quantitation at signal-to-noise ratios of 10:1 were strongly matrix-dependent and had to be determined separately for each experiment.

Enrichment and analysis of alkylphenol(ethoxylate)s The analytical procedure for determining nonylphenol (NP), nonylphenolmono- (NP1EO) and -diethoxylate (NP2EO) in environmental samples is based on the work published by Ahel et al. (2000). We enriched 100-mL samples by liquid-liquid extraction with 3x2 mL hexane. The hexane phase was dried with anhydrous sodium sulfate, and the volume reduced to 500 µl under a gentle stream of N2. The analysis was performed using normal-phase HPLC on a HP 1090 Series II liquid chromatograph (HPLC) equipped with a photodiode array detector (HP 1090) and a programmable fluorescence detector (HP 1046) from Agilent Technologies AG (Basel, Switzerland).

In vitro analysis (YES) The yeast cells were stably transfected with the human estrogen receptor gene and an expression plasmid containing estrogen-responsive elements (ERE) which control the ß- galactosidase encoding reporter gene lacZ. The translated ß-galactosidase is secreted into the medium, and its activity is determined by measuring the absorbance at 540 nm. Culture and exposure of the yeast cells were performed as described by Routledge and Sumpter (1996). Inhibition of yeast cell growth was regarded as an acute toxic effect of a tested sample or compound; this was observed as reduction of absorbance at 620 nm, compared to reference wells. Dilution series of E1, E2, EE2, E3, NP, and the WWTP effluent extracts were prepared in ethanol. Ten µl of each dilution of the standard compounds and 20 µl of the dilutions of the effluent samples were added to 96-well microtiter plates. The ethanol was evaporated and the yeast cells were added in growth medium. At least four replicate wells per dilution were used. 75

Calculated and measured estradiol equivalents (E2-EQ) Estrogenic activity (expressed as estradiol equivalency E2-EQ calc in ng L-1) calculated from steroid and nonylphenol(ethoxylate)s chemical data was based on potencies relative to E2 as determined with the YES (Table 1). Associated uncertainties were calculated based on the uncertainties of the chemical determination and relative potencies, following the rules of error propagation (33).

Estradiol equivalency of samples measured with the YES was determined by interpolation from E2 standard curves (33). Standard deviations of replicate measurements with the YES are shown.

Table 1: Relative estrogenic potencies (REP) determined with the yeast estrogen screen (33)

E1 E2 E3 EE2 NP Relative estrogenic 0.38 1 2.4x10-3 1.19 2.50x10-5 Potencies Lower and upper 95% 0.36-0.40 0.97-1.04 2.3x10-3-2.5x10-3 1.15-1.24 2.4x10-5-2.7x10-5 confidence limits

Fish exposure experiments Five exposure experiments were performed with fish at three Swiss and two French WWTPs. For reasons of confidentiality, the French WWTPs will be called France1 and France2. The WWTPs chosen for this study received either mostly urban (Glatt, France1) or a mixture of urban and industrial influent (Rontal, Surental, France2; see Table 2). The five plants had activated sludge treatment and were operating under nitrifying conditions. The actual population in the catchment area ranged from 11'000 to 88'000 people.

Before the start of the exposure, mostly male adult rainbow trout (Oncorhynchus mykiss) purchased from a hatchery (Teddy Waser, Andelfingen, Switzerland) were allowed to acclimate in the laboratory for two weeks. During this time they were fed ad libitum with commercially available pellets (Hokovit, Silver Cup, Hofmann AG, Butzberg, Switzerland). The fish were anesthetized in a solution containing MS222 (100 mg L-1 3-aminobenzoic acid ethyl ester, Redmont, USA) and individually marked with numbered jaw tags (34) for future identification. Body length and weight were measured and blood samples were collected using lithium-heparinized monovettes (Sarstedt, Numbrecht, Germany).

After acclimation, the fish were exposed to 100 % WWTP effluent, or 50 % WWTP effluent mixed with upstream river water (Surental, France1, and France2 only), or upstream river water (0 % WWTP effluent), or Lake Lucerne water (laboratory control) in black 400 L polyethylene (HD) tanks. The accuracy of the 50 % mixing was verified by comparing the 76 expected with the actual measured temperature (R2 values for the linear regressions were 0.95, 0.99 and 0.85 for Surental, France1 and 2, respectively). During each experiment the fish were kept in the dark without feeding. Each treatment was done in three replicates, using three tanks with an average of 10 fish. When an experiment ended, the fish were killed with a blow to the head and processed immediately. Length and total weight were measured, blood samples collected, and the sex determined. The fish were then dissected and the liver and gonads weighed. During the exposures the following standard parameters were measured: temperature, oxygen concentration, pH, conductivity, ammonium, nitrate, nitrite, and phosphate.

Determination of VTG blood plasma concentrations Blood was collected from the caudal vein of each fish in every treatment. After centrifugation at 2’000 g and 4 °C, blood serum was collected in Eppendorf tubes, snap-frozen in liquid nitrogen and stored at -80 °C until further analysis. VTG levels were analyzed in the blood samples from the fish exposed at all five sites using a new homologous, polyclonal anti- rainbow trout vitellogenin antibody in a newly established sandwich ELISA protocol. For the production of polyclonal anti-VTG antibody, New Zealand white rabbits (M. Moerter, Uerschhausen, Switzerland) were immunized by subcutaneous injections of 30 µg rainbow trout VTG (Biosense, Norway) in complete Freunds adjuvant (day 0) followed by 2 booster injections of 20 µg VTG each in incomplete Freunds adjuvant on days 28 and 56 (both from Sigma-Aldrich, Buchs, Switzerland). Test bleedings were done on days 42 and 70 by puncture of the arteria auricularis. After 12 weeks, the animals were terminally bled by heart puncture under narcosis.

ELISA protocol Plates were coated with 3’000xdiluted primary monoclonal antibodies BN-5 (Biosense, Norway) in blocking buffer (PBS, 0.05 % Tween 20, 2 % milk powder) and washed with washing buffer (PBS, 0.05 % Tween 20). The plates were subsequently incubated with 100µl of 750-fold diluted rainbow trout plasma for 2 h, washed with washing buffer, and further incubated overnight at 4 °C with polyclonal anti-VTG antibodies, then diluted 1:7’500 in blocking buffer. After rinsing with washing buffer, the plates were incubated with goat anti- rabbit horseradish peroxidase (diluted 1:3’000 in blocking buffer) for 2 h at room temperature and rinsed with washing buffer. Horseradish peroxidase activity was quantified using chemiluminescence substrate according to the manufacturer’s instructions (Promega, Mannheim, Germany).

77

-372 -372

weeks weeks

162-308 162-308

152-347 152-347

177-295 177-295

286

range (g) (g) range

Weight Weight

(%) (%)

97 97

58 58

96 96

89 89

89 89

45 45

97 97

33 33

64 64

74 74

53 53 450-700

100 100

100 100

100 100

100 100

100 100

100 100

100 100

100 100

100 100

100 100

stop the experiment after two two after experiment the stop

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30 30

29 29

29 29

24 24

30 30

17 17

29 29

29 29

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14 14

16 16

12 12

16 16

12 12

11 11

10 10

11 11

2) 2)

9/10/00 (2) (2) 9/10/00

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10-24/4/01 (2) 18 (2) 10-24/4/01

16-30/5/01 ( 16-30/5/01

5-26/10/00 (3) 24 (3) 5-26/10/00

(3) (3)

(2) (2)

(3) (3)

18/10-7/11/99 (3) 18 18 (3) 18/10-7/11/99

(2) (2)

(weeks) fish (number) fish fish (number) (weeks) fish

Surental 100% effluent effluent 100% Surental

he he

t

in in

L-1) L-1)

n/a n/a

n/a n/a

3.6 3.6 5

2.5 18-31/10/99 (2) 12 12 (2) 2.5 18-31/10/99

(g (g

Sludge Sampling date Male surviving Survival all all Survival surviving Male date Sampling Sludge

oxygen demand demand oxygen

) )

mortality mortality

1

1 1

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"Biological "Biological

n/a n/a

n/a

1.6 1.6

(mg L- (mg

BODd BODd

6.6 6.6

7.5 7.5

9.4 9.4

11.7 11.7

16.6 16.6

17.9 17.9

19.8 19.8

12.6 12.6

14.9 14.9

13.3 13.3

12.2 12.2

12.8 12.8

19.8 19.8

12.4 12.4

13.0 13.0

17.5 17.5

17.4 4.7 4.7 17.4

19.8 19.8

(oC) (oC)

AvgT' AvgT'

10 16.3 10 16.3

6-10 6-10

14-18 16.4 16.4 14-18

4-6 4-6

(d) (d)

3 20.7 9.0 1.5 3-16/9/99 (2) (2) 3 9.0 20.7 1.5 3-16/9/99

SRT' SRT'

T' T'

41 41

12 12

12 12

(h) (h)

HR

d-1) d-1)

6,000 6,000

8,100 8,100

8,200 7 8,200

(m3 (m3

Flow Flow

all the fish exposures exposures fish the all

of of

population: 11,000 11,000 population:

industrial: 14,000-19 .000 .000 14,000-19 industrial:

industrial/urban industrial/urban

population: 25,000 25,000 population:

industrial: 13,000 13,000 industrial:

industrial/urban 10 15,000 industrial/urban

approx. 20% industrial industrial 20% approx.

population: 27,000 27,000 population:

WWTP WWTP

time time

retention time time retention

on on

2 Summary Summary 2

aulic aulic

R

dge retention retention dge

dr

er er

bSlu

aHy Lake Lucerne Lucerne Lake

River France2 France2 River

France2 50% 50% France2

cAverage temperature temperature cAverage

Prance2 100% 100% Prance2

Lake Lucerne Lucerne Lake

River France I I France River

France! 50% 50% France!

France 1 100% urban population: 30,000 30,000 population: urban 100% 1 France

Lake Lucerne Lucerne Lake

River Sure Sure River

Surental 50% 50% Surental

Lake Lucerne Lucerne Lake

River Glatt Glatt River

Surental I 00% 00% I Surental

Lake Lucerne Lucerne Lake

Glatt 100% urban population: 88,000 45,000 45,000 88,000 population: 100% urban Glatt

Riv

Rontal 100% industrial/urban industrial/urban 100% Rontal

Site Site Table Table 78

Results and discussion

Chemical analysis Chemical data for steroid hormones and nonylphenol(ethoxylate)s acquired during all the exposure experiments, together with the estradiol equivalents (E2-EQ) determined with YES, are shown in Table 3. Maximal values obtained in the 90 samples analyzed were 50.5ng L-1 estrone (Rontal 100 % effluent), 9.8 ng L-1 estradiol (France2 100 % effluent), 17.5 ng L-1 estriol (Rontal 100 % effluent), and 2.8 ng L-1 ethinylestradiol (Rontal 100 % effluent) and a measured estrogenicity of 53.0 ng L-1 (YES E2-EQ, Rontal 100 % effluent). The 50.5 ng L-1 maximal estrone concentration determined in the Rontal effluent is well matched by the measured estrogenicity as determined with the YES (53.0 ng L-1) and two other in vitro tests ([33], data not shown), but this value clearly deviates from the rest, which did not exceed 5.5 ng L-1 for measured estrogenicity (YES E2-EQ; Glatt 100 % effluent). Since this was a singular event, it was not included in the data set used for the correlation analysis, for reasons of clarity.

The E2 values determined during the Surental exposure (100 % effluent: 0.8 – 3.2 ng L-1, 50% effluent: 1.0 – 3.8 ng L-1, and river: 0.8 – 3.8 ng L-1) were not considered for further analysis, because except for the Lake Lucerne concentrations (

LOQs at the ng L-1 level are influenced greatly by the matrix of the sample. Hence, LOQs must be determined independently for each environmental sample. In 100 % WWTP effluent ranged from 0.7 to 1.0 ng L-1 for E1, 0.5 to 0.9 ng L-1 for E2, 1.0 to 1.5 ng L-1 for E3, 0.7 to 1.0 ng L-1 for EE2. LOQs in Lake Lucerne water were 0.1 ng L-1 for E1, 0.3 ng L-1 for E2, 0.3 ng L- 1 for E3, and 0.2 ng L-1 for EE2. The LOQs for the alkylphenols were 0.09 µg L-1 for NP, and 0.08 µg L-1 for NP1/2EO in all matrices.

In light of the ongoing discussion about effects of mixtures of causative agents at individual concentrations below effect levels (29i), there is clearly need for further lowering the LOQs for the determination of steroidal estrogens. Affinity techniques as well as the use of immunosorbents have been shown to be promising selective enrichment or clean-up tools (24; 35; 36).

79

calc calc

)" )"

1.1 1.1

7.2 7.2

-12.2 -12.2

0.5-5.5 0.5-5.5

1.9-6.4 1.9-6.4

(l.6--4.8) (l.6--4.8)

(1.6--4.9) (1.6--4.9)

1.9-

E2-Eg E2-Eg

(ng L- (ng

L-) L-)

-2.9 -2.9

2.0

0.5-2.2 0.5-2.2

0.4-(53.0) 2.1-(29.0) 2.1-(29.0) 0.4-(53.0)

(ng (ng

E2-Eg E2-Eg

EE2 (ngL-1) (ngL-1) EE2

17/J-estradiol (E2), estriol (E3), 17a-elhi- (E3), estriol (E2), 17/J-estradiol

) )

1

(El), (El),

,2.0 ,2.0

L-

(ng (ng

2.3-17.5

3.3-11.0 3.3-11.0 E3 E3

) )

1

1.1 1.1

-1.0 -1.0

.0 .0

1.3-4

(l.0...:3.8) (l.0...:3.8)

(0.8-3.2) (0.8-3.2)

E2 (ng L- (ng E2

.2 .2

) )

1

L-

-10

{ng {ng

1.9-6.5 3.2-9.8

1.6-3.7 1.6-3.7

4.9-18.8

El El

L-1) L-1)

OQ OQ

OQ-0.22 2.0-5.9 2.0-5.9 OQ-0.22

µg µg

0.41-2.04 0.41-2.04

0.24-0.58 1.2-7.1

0.43-0.84

(

NP2EO NP2EO

L-1) L-1)

0.48-1.09 0.48-1.09

(µg (µg

0.31-0.65 0.41-0.85 0.41-0.85 0.31-0.65

NPlEO NPlEO

) )

1

nonylphenohnonoethoxylate (NPIEO), nonylphenoldiethoxylate (NP2EO), estrone estrone (NP2EO), nonylphenoldiethoxylate (NPIEO), nonylphenohnonoethoxylate

-0.28 -0.28

(NP), (NP),

0.20-0.88 0.20-0.88

0.39-0.88 0.36-0.71 0.36-0.71 0.39-0.88

0.11-0.19 0.10-0.16 0.10-0.16 0.11-0.19

NP (µg L- (µg NP

0.17

0/2000 0.09-0.19 0.08-0.17 0.08-0.17 0.09-0.19 0/2000

5i9/1999 0.89-1.61 0.90-2.58 1.08-3.61 4.0-(50.5) 4.0-(50.5) 1.08-3.61 0.90-2.58 0.89-1.61 5i9/1999

-1

18+23/4/2001 18+23/4/2001

11-15/4/2001 11-15/4/2001

22+ 29/5/2001 29/5/2001 22+

5-25/10/2000

5-25/l 5-25/l

5-19/10/2000 5-19/10/2000

18/10-8/11/l999 18/10-8/11/l999

Date Date

of of

(n=4) (n=4)

posure posure

3 Chemical data for nonylphenol nonylphenol for data Chemical 3

Lucerne Lucerne

Ex

(n=2) (n=2)

(=4) (=4)

•values below LOQ were considered to be zero for E2-EQ calc calc E2-EQ for zero be to considered were LOQ below •values

(n=4) (n=4)

Lake Lake

50% Effluent Effluent 50%

(n=2) (n=2)

River Prance2 Prance2 River

Effluent Effluent

Exposure100% Exposure100%

Lake Lucerne Lucerne Lake

(n=5) (n=5)

50% Effluent Effluent 50%

France2 11-24/4/2001 0.49-1.74 1.10-2.52 0.81-3.97 0.81-3.97 1.10-2.52 0.49-1.74 11-24/4/2001 France2

(n=5) (n=5)

River Francel Francel River

(n=7) (n=7)

Effluent (n=S) (n=S) Effluent

Prancel 16-30/5/2001 0.66-1.27 0.49-1.03 0.49-1.03 0.66-1.27 16-30/5/2001 Prancel

Exposure I 00% 00% I Exposure

River Sure Sure River

50% Effluent Effluent 50%

(n=1) (n=1)

(n=1) (n=1)

Lake Lucerne 5-25/10/2000 5-25/10/2000 Lucerne Lake

Effluent (n=5) (n=5) Effluent

Exposure 100% 100% Exposure

(n=4) (n=4)

Surental Surental

100% Effluent Effluent 100%

Lake Lucerne Lucerne Lake

River Ron (n=4) (n=4) Ron River

Rontal Rontal 3

Lake Lucerne Lucerne Lake

(n=7) (n=7)

(n=7) (n=7)

(n=7) (n=7)

Effluent (n=4) (n=4) Effluent

Exposurel00% Exposurel00%

River Glatt Glatt River

samples) samples)

Glatt Glatt

Site (number (number Site

nylestradiol, estrogenicity measured with the YES (E2-EQ) and calculated (E2-EQ calc) calc) (E2-EQ calculated and (E2-EQ) YES the with measured estrogenicity nylestradiol, Table Table 80

Maximal concentrations in all 90 samples analyzed were up to 1.74 µg L-1 for NP, up to 2.58 µg L-1 for NP1EO, and up to 3.97 µg L-1 for NP2EO.

Effluent concentrations similar to those reported here were found throughout Europe and correspond well to predictions made by Johnson et al. (17; 37). Generally, the highest concentrations are found for E1, followed by E3 and E2. These relative concentrations agree with the relative amounts found in the urine of pre-menopausal women. Post-menopausal women excrete more E2 than E3, but the absolute hormone concentrations are lower (38).

However, chemical data on steroidal estrogens have to be considered with the corresponding uncertainties in mind. Standard uncertainties for effluent samples ranged between 23 % for E2 and 116 % for E3. Even if this seems rather high, it clearly reflects the problems encountered in ultra-trace analysis in complex sample matrices. Albert and Horwitz analyzed almost 10’000 intercalibration exercises and derived the following relationship between the relative standard deviation (RSD; %) and sample concentration (c; mass ratio in g/g-1): RSD = 2^[1-0.5 x log(c)] (39). Thus, at a concentration of 1 ng L-1, the expected relative standard deviation is 130 %, without accounting for the physical and chemical properties of the analytes. Estriol, for example, is the most polar of the steroid metabolites investigated and for this reason recoveries for E3 depend strongly on the activity of the silica used in the clean-up step. Furthermore, E2-acetate proposed by Ternes et al. (22) is not the ideal internal standard for this polar compound, but since estrogenic activity was also measured with the YES, we preferred not to use deuterated standards. This explains the large uncertainty associated with the chemical data for E3, which however still is well within the values predicted by Albert and Horwitz (39). Since the relative estrogenic potency and the measured concentrations are relatively low (Tables 1 and 3), E3 does not greatly affect the total calculated estrogenicity.

Calculated estrogenicity versus estrogenicity measured with the YES The relative potencies determined with the yeast estrogen screen cover a range of five orders of magnitude, with the natural and synthetic steroid hormones being the most potent (Table 1). The relative estrogenic potency of E3 is comparable with previously published results using a yeast-based estrogenic assay (40), but is much lower when compared to cases in which binding affinities with the ER have been determined (40, 41). The estrogenic activity is calculated based on the concept of concentration additivity, which applies since all investigated EDCs target the same receptor (29; 43; 44). The contribution of nonylphenol(ethoxylate)s to the total estrogenicity of the effluents measured is only minor. 81

For instance, even the highest NP concentration measured (1'740 ng L-1, France2), corresponds to an estrogenicity of 0.04 ng L-1 E2, which only accounts for 0.8 % of the total estrogenic activity of 4.9 ng L-1 measured in this sample with the YES. Similarly insignificant contributions are obtained for NP1EO, and NP2EO.

Data acquired during the five exposure experiments (with the exception of the Surental dilution and the singular event in Rontal mentioned above) has been used to analyze the correlation between measured estrogenicity (YES) and calculated E2 equivalents . Figure 1 lists E2-EQs calculated (E2-EQ calc) based on the chemical data as a function of the measured estrogenicity (YES) for each sample. The error bars reflect the standard error of the measured estrogenicity (replicate measurements) and E2-EQ calculated for each sample. The error bars for the calculated E2-EQs incorporate the uncertainties of the chemical measurements (up to 116 % for E3) and the relative estrogenic potencies. Since a traditional least square estimation would be biased due to these uncertainties, we calculated the geometric mean functional relationship (45). A positive correlation (R2 = 0.60) was observed between these two parameters, indicating that the natural and synthetic steroid hormones represent the major contributors to the measured estrogenicity. However, alkylphenols and their ethoxylates, and other xenoestrogens might have to be monitored in special cases, for instance when the WWTP influent has a high industrial component. The graph also shows that the calculation tends to overestimate the estrogenicity. However, due to the measurement uncertainties involved, this should by no means be over interpreted.

Figure 1: Estrogenicity determined with the yeast estrogen screen (YES) versus calculated estradiol equivalents, based on chemical data (E2-EQ calculated). 82

Fish exposure experiments Most exposed individuals (>95 %) were male except for the WWTP Rontal experiment (51 % male). Individual VTG data from before the experiment was not available for the Rontal exposure due to loss of tags. The fish were exposed for a period of two weeks for all treatment experiments, except for a three-week exposure at Surental, and one of the two Glatt experiments (Table 2). The three exposures in Switzerland were done in fall (first week of September to first week of November), while the French exposures were undertaken in spring (end of April and end of May).

Survival of the fish in the experiments conducted at the WWTPs Glatt and France1 were high (>95 %). Rainbow trout used for the exposure at the Rontal WWTP had a fungal skin infection (Saprolegnia) and had higher mortality in all treatments. Fish survival during the Surental effluent treatment was low (33 %) for unknown reasons, while the observed increased mortality in the river-water treatment was caused by an accidental over-saturation with oxygen. The high mortality found for the effluent-treated trout at the WWTP France2 was caused by very high ammonia concentrations of up to 26 mg L-1 on exposure day 7. This concentration was much higher than the reported acute toxic concentration of 200 µg L-1, for non-ionized ammonia (NH3) (46; 47). To avoid total loss in the effluent and mixed water treatment groups, the treatments were interrupted for 12 hours and run with river water, in order to allow the trout to recover from the ammonia peak. HSI (hepatosomatic index) and GSI (gonadosomatic index) were calculated for each single fish. However, in none of the experiments changes in HSI and GSI were significant between treatment and control.

In vitro versus VTG concentrations in blood plasma Significant estrogenic activities were measured in vitro with the YES in all WWTP effluents analyzed. The measured values obtained with the YES correlated well with the estradiol equivalents calculated based on the chemical data, and most importantly did not give false negative results (Figure 1). This clearly supports the idea of using YES as an in vitro tool for a first screening of environmental samples. YES is an easy, relatively fast and robust in vitro assay that has recently been shown to give results similar to those of other in vitro tests, based either on the use of primary fish hepatocytes or fish cell lines (33). Potential hazards can thus be determined in a sensitive way and relatively fast. In case of a positive in vitro response, a further level of biological effect integration can be envisaged using in vivo effect analysis, which also takes into account toxicokinetic and toxicodynamic effects. The importance of this validation step has recently been examined by Legler et al. (48). 83 to Lake Lucerne water. Lucerne Lake to Figure 2: Blood plasma VTG concentrations determined in male rainbow trout before and after exposure to dilutions of WWTP effluents (100 % Effluent, 50 % Effluent), to River and River to Effluent), % 50 Effluent, % (100 effluents WWTP of dilutions to exposure after and before trout rainbow male in determined concentrations VTG plasma Blood 2: Figure 84

They showed that effects seen in vitro do not necessarily mean observable in vivo effects.

In the work presented here we observed a clearly induced vitellogenesis in male rainbow trout exposed to sewage effluents from the WWTPs Glatt and Rontal. This is visualized by an increased number of fish having higher concentrations of VTG in Figure 2. These results indicate estrogenic activity in both effluents. Interestingly, no significant increase of VTG blood plasma concentrations were found at the Surental, France1, and France2 sites (see Figure 2). We do not understand the different responses in vitro versus in vivo. The hERa is used in YES, while rtERa is present in the in vivo test. Differential binding for the two estrogen receptors for steroid hormones and for xenoestrogens has been described controversially (41; 49; 50; 51). This differential binding can, however, not explain why no in vivo responses have been observed. Due to the annual cycle of reproductive hormones in both females and males (52), and, in consequence, inducibility of VTG, the seasonally different exposure schemes (autumn in exposures in Switzerland; spring in exposures in France), could partly have contributed to the different response patterns. The negative in vivo response in estrogenic effluents could be due either to the presence of anti-estrogens in the effluents of these treatment plants, or hepatotoxicity-induced reduction of VTG synthesis, or diet. However, anti-estrogenicity could have been detected with the YES assay. The observed reduction of VTG levels during exposure (Surental, France1, and France2) is most likely caused by estrogenic compounds in the pellets used for feeding at the hatchery before the experiment. Indeed, estrogenicity was measured with the YES (0.12ng g-1 dry weight) in an aqueous extract of the feeding pellets, which very possibly accounts for the elevated VTG levels at the onset of the experiments. No steroid hormones were detected either in aqueous, or hexane extracts of the feeding pellets (A.C. Vögeli - unpublished results). We have no explanation for the different in vivo response in the various effluents, but the results clearly illustrate the difficulty of predicting a propagation of effects onto a higher level of biological organization. A positive effect seen in vitro should hence clearly be defined as a potential hazard, rather than a proof for a biological effect in higher organisms. 85

Conclusion In this work we have assessed the estrogenic activity of wastewater treatment plant effluents and combined chemical analysis and biological effect analysis, in vitro as well as in vivo. We showed that the natural and synthetic steroid hormones are the major contributors to estrogenicity in the effluents tested. We can, however, not exclude that xenoestrogens might have a significant contribution in other effluents. The calculated (based on chemical analysis, and using the principle of concentration additivity) and measured (YES) correlated well (R2 = 0.6). No false negative results were obtained in vitro, supporting the idea to use YES as an in vitro tool for first screening purposes. Effects seen in vitro must not necessarily mean that in vivo effects will occur. The fact that three out of five estrogenic effluents did not induce vitellogenesis, clearly shows that prediction of effects in vivo from in vitro data is and will be extremely difficult. It also shows that we are still very far away from understanding and predicting endocrine disruption, particularly if the effects of complex environmental mixtures are being studied.

Acknowledgement - This work was part of the European “Community programme of research on environmental hormones and endocrine disruptors – COMPREHEND”, supported by the Swiss Federal Office for Education and Science (BBW), grant Nr. 98.0090. The authors would like to thank John Sumpter (Brunel University, Uxbridge, UK) who kindly provided the recombinant yeast estrogen screen, Patricia Holm (EAWAG) and Helmut Segner (Centre for Fish and Wildlife Health, University of Berne, Switzerland), as well as all COMPREHEND consortium partners for fruitful and lively discussions. Karl Fent is acknowledged for his contributions in the initial phase of the project. Mark Borsuk and Beate Escher (EAWAG) are greatly acknowledged for their help with the statistical treatment of the data shown in Fig. 1, as are Michel Gibert and David Benanou (Vivendi Water, Paris, France) and all WWTP personnel for their on-site help during the exposures at the five WWTPs. Our special thanks go to Alan Pickering (Windermere, UK) for his outstanding coordination of COMPREHEND, and Alexander Zehnder (EAWAG) for his continuous support during the whole project. 86

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6 SYNOPSIS

Introduction

Effluent from wastewater treatment plants (WWTP) can cause problems for fish. Numerous studies have revealed adverse effects on molecular and physiological processes (e.g. Adams et al., 1989; Triebskorn et al., 2001). Others have shown adverse effects on higher biological organization scales such as impaired survival, health and reproduction (e.g. Kime, 1995; Escher, 1999; Rao, 1999). Nevertheless, as fish can produce gametes in excess (which may compensate for a certain level of damage), the final consequence for fish populations is uncertain (Colby, 1984; Power, 1997; Clement, 2000). Therefore, we need further studies to better understand the adverse effects of wastewater on wild fish populations and to clarify their limits of compensation on the population level.

Brown trout resident in the River Wyna

As a study area, we chose the Wyna River. The headwater section is not contaminated with wastewater. Further downstream, effluents from four wastewater treatment plants affected the chemical composition of the water. At low flow, wastewater addition could reach as much as 50% of the water volume at the stream’s mouth. The wastewater-contaminated part was divided into two sections, the first including the upper three wastewater inlets and the second (close to the stream’s mouth) including the water from all four treatment plants.

Comparison of the main diagnostic population variables, such as growth, mean age, population density and recruitment between the three sections, revealed major differences. Growth and mean age were clearly higher in the lowest section, and fish density and population recruitment success were significantly impaired and, thus, lower. The low density allowed the remaining population to grow faster because there was little competition for food. Age was higher since the recruitment success of the population was low. The observed pattern was similar to the general response pattern (GRP) type II reported by Munkittrick & Dixon (1989), which is a strong indication of recruitment failure. In the lowest river section, which contained the highest wastewater load, the population recruitment success was lowest. There are statistical analysis providing evidence that wastewater in particular impaired the population’s recruitment. Like water pollution, the structural quality of fish habitats can strongly influence the population’s characteristics and recruitment success (Jungwirth & Winkler, 1983). Since different river sections surveyed in the Wyna revealed 90 comparable fish habitats, such habitat effects were assumed to be negligible. What was remarkable, however, is that the population resident in the lowest section was not sustained by natural reproduction, but by trout from upstream sites.

This recruitment bottleneck was not the result of reproductive dysfunction in female trout. Trout from the lowest section produced fertilizable eggs of excellent quality with high survival rates in the laboratory. However, it was noted that in the field the early survival of trout was impaired. During spawning time in autumn, trout need suitable riverbeds with gravel for depositing their eggs. Later during winter, in-gravel deposited trout eggs develop. During this period, several stress factors can impair early survival (Crisp, 1989). Egg exposure experiments showed that this early development was a crucial period for trout in the Wyna River. First, the river water was embryo-toxic, and second, fine sediments clogged the riverbed. Both resulted in a very low survival rate for early life stages, which could explain the observed recruitment bottleneck. The absence of suitable spawning habitats and the poor water quality were identified to be a dominant problem for trout resident in the lowest river section.

Estrogen-active wastewater

In blood of male brown trout from the Wyna River, Kobler (2003) found high concentrations of vitellogenin (Vtg), which is a yolk protein and important for egg ripening. Normally, Vtg is found in ripe females but not in males. When Vtg is found in male trout, it can indicate water contamination with estrogen active substances (Jobling, 1998). Vtg-levels in trout from sites upstream and downstream of the WWTP Mittleres Wynental -which is located just upstream from the lowest river section- were significantly different. Therefore, this WWTP was identified to be a major source of the estrogenic contamination in the Wyna River.

Kime (1998) summarized several studies that showed how an estrogenic water contamination can impair the reproductive success of fish. Thus, the question arises as to whether the estrogenic contamination in the Wyna River also impairs the reproduction of trout and subsequently the recruitment success of the population. Since the reproductive success of male Wyna trout was not investigated in detail, it remains speculative to suggest that the estrogenic contamination caused a local recruitment failure of the trout population in the lowest river section.

To obtain more information about the estrogenic effect of wastewater effluents from five WWTP located in Switzerland and France were analyzed for known endocrine disrupters and 91 estrogenicity (Chapter 5). Estrogenicity was determined using the yeast estrogen screen (YES) and by measuring the blood vitellogenin (Vtg) concentrations in exposed male rainbow trout.

While all wastewater treatment plant effluents contained measurable concentrations of estrogens and gave a positive response with the YES, only at two sites did the male fish have significantly increased Vtg blood plasma concentrations after the exposure. Effects seen in vitro do not necessarily mean that in vivo effects will occur. However, the fact that two out of five effluents did induce vitellogenesis shows that effluents from Swiss WWTPs can cause effects on fish. Also wastewaters from numerous treatment plants in England, Sweden, and Germany were reported to exhibit estrogenic effects (Harries et al., 1999; Svenson, 2002; Dietrich, 2003, University of Konstanz, Germany, personal communication). Estrogen-active wastewater thus could be a widespread phenomenon in Europe. It is accepted that contamination with estrogenic substances such as natural and synthetic hormones or hormone mimics from human metabolic waste, medicine or industry can impair the reproductive success in fish (Bätscher et al., 1999). The present thesis showed that natural and synthetic steroid hormones are the major contributors to estrogenicity in the effluents tested.

The situation in the Wyna River

Figure 1 illustrates the situation for trout resident in the Wyna River. Wastewater, lack of spawning habitat, and barriers in the river caused a recruitment failure of the resident population, which resulted in a low population density and an over-aged population structure. The fast growth and good egg quality of trout can be seen as a response to the low population density and the associated high availability of resources. Egg and larvae were most sensitive to water pollution and habitat degradation.

Suggestions to improve river quality

Reduce the impact of Wastewater

In the Wyna River, the treated wastewater load is high. In particular, ammonia (NH3) and - nitrite (NO2) frequently exceed tolerable concentrations proposed by Swiss federal law (Stöckli & Schmid, 1997). Own measurements during the survey reveal that the average concentration of ammonia matched a level that is shown to cause biological effects for salmonids (0.025 mg NH3/l, Calamari et al., 1977). The maximal measured concentration 92

even reached levels that are close to acute toxicity (LC5096h= 0.2 mg NH3/l, Alabaster and Lloyd, 1980). Ammonia was suspected to cause serious damage in the sensitive early life stages of trout. Furthermore, wastewater discharged into the river was found to be estrogenic to adult fish, which could disrupt the reproductive function in males.

About 50% of the water flowing in the Wyna at low discharge has its origin in wastewater. A recent guideline proposes loads of no more than 10% wastewater in a river (Stadelmann, 2003, agency of environmental protection, canton Lucerne, personal communication). The Wyna river is overloaded with wastewater.

There are two feasible solutions for the Wyna. First, to collect all treated wastewater and discharge into the Aare, the nearest large river in the same catchment. The second possibility is to reduce the impact of wastewater, i.e. to improve the treatment efficiency of the existing plants and to prevent stormwater from the sewers to enter Wyna. Both solutions are costly, and time will show which one will be chosen.

Create spawning habitats for trout Normally, flood events can destabilize the riverbed and move the gravel, as well as wash out accumulated sediments. Such wash out events are known to be essential for restoring trout spawning habitats (Crisp & Carling, 1989; Gorst et al., 1991).

Obstructions, barriers, and gravel collectors which were built to stabilize the Wyna River may have reduced the riverbed dynamics and thus increased the accumulation of sediments in the riverbed. However, no information was collected in this respect. Conclusions could only be drawn when comparing siltation rate and riverbed stability between differently impacted river systems.

Results of the dissertation show that the poor quality of spawning habitats was a key factor impairing the population’s reproductive success. Thus, steps are needed to improve the quality of the Wyna riverbed for spawning trout. One method to achieve this could be to remove obstructions, barriers and gravel collectors, with the goal of increasing the dynamics of the riverbed. Another, more technical method would be to create artificial spawning habitats by adding sediment-free gravel to the river. However, restoring spawning sites in heavy polluted river sites is not effective and durable. The work should be concentrated on the less polluted sites in the middle reach and the headwaters of the Wyna River. The accessibility of these habitats -also for fish from the polluted downstream reach- is an important factor to consider. 93

Figure 1: Simplified schematic of stress factors and their effect on trout resident in the Wyna River.

Barriers over 10 cm significantly impair the movement of most fish species (Peter, 1998). The Wyna River is interrupted by a number of barriers with a height over 1.2 m, which are impassable even for large trout. Such barriers worsen the situation for trout in the lowest river section, because they are trapped in a very polluted area and cannot reach suitable spawning habitats located upstream or in the tributaries. Removing the barriers or constructing “fish passes” could be a suitable solution. Stocking of 0+-fish can only be regarded as a short term solution. The improvement of spawning habitat, longitudinal connectivity, and water quality should be the appropriate management action. 94

References

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Danksagung

Leiter und Betreuer der vorliegenden Arbeit war Herr Dr. Armin Peter. Er hat mich fachlich geprägt und persönlich unterstützt, wofür ich ihm sehr danke.

Herr Prof. Dr. Alexander J.B. Zehnder übernahm die Funktion des Doktorvaters. Ihm danke ich herzlich für seine kritischen und konstruktiven Feedbacks.

Für seinen Einsatz als Koreferent danke ich Herrn Prof. Dr. Helmut Segner. Seine fachlichen Hinweise waren wichtig für die Qualität dieser Arbeit.

Ganz herzlich danke ich der gesamten Abteilung der Fischereiwissenschaften, Kastanienbaum. Ihre Unterstützung bot mir ein ideales Arbeitsumfeld. Namentlich gilt dieser Dank Eva Schager, Erwin Schäfer, Mbwenemo Bia Mampasi und Brigitte Germann.

Ein besonderer Dank geht an meine Bürokollegen Dr. Daniel McGinnis und Dr. Andreas Lorke.

Für fachliche Diskussionen und Feedbacks geht mein Dank an Prof. Dr. Bernhard Wehrli, Dr. Till Luckenbach, Dr. Marc Borsuk, Dr. Marc Beutel, Patrick Faller, Prof. Dr. Patricia Holm und Georg Holzer.

Auch das Amt für Umwelt und Energie des Kantons Luzern und das Baudepartement des Kantons Aargau standen meiner Arbeit wohlwollend gegenüber und ermöglichten die zahlreichen Untersuchungen. Mein Dank geht an die vielen Verwaltungsangestellten, die mich bei meiner Arbeit selbstlos unterstützten.

Meinen Eltern Lydia Kolaja und Josef Kobler danke ich für ihren Beistand.

Meiner Partnerin Nicole Scheider danke ich für ihre Unterstützung in allen Phasen meiner Arbeit. 98 99

Curriculum vitae

Bernd Kobler Geboren am 13. März 1972 in Biel

Schulbildung

1979 - 1988 Primarschule, Biel 1988 - 1989 Weiterbildungsklasse, Biel 1989 - 1994 Wirtschaftsgymnasium, Biel

Studium und Promotion

1994 - 1998 Studium der Biologie, Bern Lizentiatsarbeit am Zoologischen Institut der Universität Bern, Abteilung: Verhaltensökologie

1999 – 2003 Dissertation am Institut für Angewandte Gewässerökologie der EAWAG/ETH, Zürich, Fischereiwissenschaften, Kastanienbaum

Anschrift Bernd Kobler Stollbergstrasse 30 CH-6003 Luzern