Investigations on the Impact of Toxic Cyanobacteria on Fish as exemplified by the Coregonids in Lake Ammersee

I NVESTIGATIONS ON THE IMPACT OF TOXIC CYANOBACTERIA ON FISH

- AS EXEMPLIFIED BY THE COREGONIDS IN LAKE A MMERSEE -

D ISSERTATION

Zur Erlangung des akademischen Grades des Doktors der Naturwissenschaften

an der Universität Konstanz Fachbereich Biologie

Vorgelegt von

B ^ERNHARD E ^RNST

Tag der mündlichen Prüfung: 05. Nov. 2008

Referent: Prof. Dr. Daniel Dietrich Referent: Prof. Dr. Karl-Otto Rothhaupt

Referent: Prof. Dr. Alexander Bürkle

Konstanzer Online-Publikations-System (KOPS)

»Erst seit gestern und nur für einen Tag auf diesem Planeten weilend, können wir nur hoffen, einen Blick auf das Wissen zu erhaschen, das wir vermutlich nie erlangen werden«

Horace-Bénédict de Saussure (1740-1799)

Pionier der modernen Alpenforschung & Wegbereiter des Alpinismus

Z USAMMENFASSUNG

Giftige Cyanobakterien beeinträchtigen Organismen verschiedenster Entwicklungsstufen und trophischer Ebenen. Besonders bedroht sind aquatische Organismen, weil sie von Cyanobakterien sehr vielfältig beeinflussbar sind und ihnen zudem oft nur sehr begrenzt ausweichen können. Zu den toxinreichsten Cyanobakterien gehören Arten der Gattung Planktothrix. Hierzu zählt auch die Burgunderblutalge Planktothrix rubescens, eine Cyanobakterienart die über die letzten Jahrzehnte im Besonderen in den Seen der Voralpenregionen zunehmend an Bedeutung gewonnen hat.

An einigen dieser Voralpenseen treten seit dem Erstarken von P. rubescens existenzielle, fischereiwirtschaftliche Probleme auf, die wesentlich auf markante Wachstumseinbrüche bei den Coregonenbeständen (Coregonus sp.; i.e. Renken, Felchen, etc.) zurückzuführen sind. So auch am Ammersee, wo die beschriebenen Wachstumseinbrüche vermeintlich sogar regelmäßig zum vorzeitigen Verenden bestimmter Coregonenjahrgänge führen. Interessanterweise hatten die Coregonen im Ammersee wiederholt einen außergewöhnlich blau gefärbten Darminhalt. Diese auffällige Färbung wird vermutlich von cyanobakteriellen Farbpigmenten verursacht und deutet darauf hin, dass die Coregonen im Ammersee mit Cyanobakterien in Kontakt kommen. Es scheint daher grundsätzlich nicht abwegig, dass die Schwierigkeiten der Ammersee-Coregonen in kausalem Zusammenhang zum Auftreten giftiger P. rubescens Filamente stehen könnten. Ziel des Dissertationsprojektes war es daher,

• das Aufkommen, die Verteilung und die Toxizität von P. rubescens im Ammersee über einen aussagekräftigen Zeitraum detailliert zu erfassen,

• in an die natürlichen Verhältnisse angelehnten Laborexperimente zu untersuchen, ob die vorgefundenen P. rubescens-Dichten Coregonenpopulationen beeinträchtigen und gesundheitlich schädigen können und schließlich,

• zu prüfen, ob es Hinweise auf P. rubescens Expositionen und dadurch verursachte Schädigungen wildlebender Coregonen im Ammersee gibt.

Zur Bestimmung der Planktothrix-Dichte im Ammersee war es zunächst notwendig, ein Verfahren zu etablieren, bei dem die in Wasserproben enthaltenen Planktothrix-Filamente auf Filtern mittels Fluoreszenzmikroskopie und digitaler Bildverarbeitung einfach, schnell und akkurat quantifiziert werden können. Dieses Verfahren ermöglichte eine aufwendige Beprobung und damit eine aussagekräftige Beschreibung des zeitlichen und räumlichen Verteilungsmusters von P. rubescens im See.

Bei den von April 1999 bis September 2004 durchgeführten Beprobungen zeigte sich dann, dass P. rubescens-Filamente im Ammersee – wenn auch in unterschiedlicher Dichte – durchgehend vorhanden waren. Der Bereich maximaler Zelldichten korrelierte jeweils von Mai bis Oktober

signifikant mit der unteren Grenze der euphotischen Tiefe und dem Beginn des Metalimnion.

P. rubescens erreichte regelmäßig im Sommer maximale Dichten (zeitweilig bis zu 75.000 Zellen/ml). Darüber hinaus konnte P. rubescens auch während der winterlichen Vollzirkulation über den Wasserkörper verteilt in Zelldichten von bis zu 15.000 Zellen/ml nachgewiesen werden.

In welchem Ausmaß sich P. rubescens entwickeln kann, scheint wesentlich von der Illumination des Metalimnion und damit von der Transparenz des Wassers abzuhängen. Zudem scheint P. rubescens auch von regelmäßigen Phosphat-Auszehrung und den hohen Stickstoffkonzentrationen in dem re-oligotrophierten See zu profitieren.

In 27 bzw. 38 von 54 Planktonproben aus verschiedenen Monaten konnten mittels HPLC und ELISA Toxinanalyse diverse fischgiftige Microcystine (in erster Linie Microcystin-RR Varianten) nachgewiesen werden. Nahezu konstante Microcystin/Phycoerythrin Verhältnisse verdeutlichten, dass die von P. rubescens produzierten Microcystinmengen weitgehend unveränderlich sind. Dies bedeutet, dass bei einem Aufkommen von P. rubescens auch von einem Auftreten messbarer Microcystin-Belastungen auszugehen ist.

Die Auswirkungen von P. rubescens auf Coregonen wurden in Laborexperimenten untersucht, wobei die im Ammersee erfassten P. rubescens Dichten und verschiedene Expositionsformen berücksichtigt wurden. Vorversuche verdeutlichten, dass aufgrund einer kovalenten Bindung des Microcystins im Gewebe eine aussagekräftige Quantifizierung geringer Microcystinmengen in Fischgeweben generell schwierig und via ELISA-, HPLC- und PPAssay-Toxinanalytik unmöglich ist. Microcystine konnten im Fischgewebe aber qualitativ, mittels immunhistochemischer Anfärbung mit anti-Microcystin Antikörpern lokalisiert werden.

Die exponierten Renken reagierten mit auffälligem Verhalten, einer gesteigerten Atemfrequenz und wiesen erhöhte Serumglukose-Konzentrationen auf. Merkmale, die in ihrer Gesamtheit als eindeutige Stressindikatoren zu bewerten sind. Pathologische Auffälligkeiten in der Leber, Niere und im Gastrointestinaltrakt verdeutlichen zudem beträchtliche Organschäden die auf nachhaltige Auswirkungen auf Organfunktionen schließen lassen. Die Tatsache, dass die geschädigten Gewebebereiche gemäß immunhistologischer Anfärbung vielfach auch Microcystin enthielten, veranschaulicht einen kausalen Zusammenhang von Gewebeschäden und der offensichtlichen Aufnahme von Microcystin. Eine erhöhte Empfindlichkeit gegenüber Ektoparasiten und eine erhöhte Mortalitätsrate deuten weiter daraufhin, dass durch die Wirkung von P. rubescens letztendlich auch die Kondition der experimentell exponierten Coregonen beeinträchtigt wurde.

Insgesamt betrachtet war die Symptomatik der Auswirkungen in den verschiedenen Expositionsansätzen vergleichbar – die Intensität der Wirkung war hingegen dosisabhängig. Dies verdeutlicht, dass sich mit zunehmender P. rubescens-Zelldichte entsprechende Effekte früher und deutlicher ausprägen. Nichtsdestoweniger zeigte sich auch bei vergleichsweise geringer P. rubescens Dichten (≈1500 Zellen/ml) eine fischgiftige Wirkung.

P. rubescens Zelldichten von mindestens 1500 Zellen/ml waren im Ammersee in etwa zur Hälfte der 261 Wochen andauernden Beobachtungsperiode festzustellen, was verdeutlicht, dass die Coregonen im Ammersee tatsächlich regelmäßig mit schädlichen P. rubescens Zelldichten konfrontiert sind. Dies scheint vor allem dann problematisch, wenn die Coregonenpopulation einer P. rubescens-Exposition nicht aktiv ausweichen kann (z.B. bei P. rubescens Entwicklungen die den gesamten Wasserkörper umfassen).

In der Tat, auch im See selbst ergaben sich Anhaltspunkte für eine P. rubescens Exposition der Coregonen. So konnte gezeigt werden, dass die Fische regelmäßig P. rubescens Filamente inkorporieren und verdauen. Dadurch werden im Darm der Fische die in den Filamenten enthaltenen Metabolite freigesetzt. Die Freisetzung von Phycocyanin bewirkt die auffällige Blaufärbung des Darminhalts und die Freisetzung der Microcystine verursacht eine Microcystin- Exposition der Coregonen. Da Microcystin stichprobenartig zudem auch in Leberhomogenaten von Ammersee Coregonen nachzuweisen war, ist wahrscheinlich, dass das im Darm freigesetzte Microcystin über das Darmepithel in den Organismus gelangen und sich entsprechend nachhaltig auf den Gesundheitszustand und die physiologische Kondition der Coregonen auswirken kann.

Man kann daher davon auszugehen, dass die in den Laborexperimenten aufgezeigten Microcystin-Schäden auch in den Fischen im See auftreten und sich dauerhafte P. rubescens Vorkommen entsprechend substanziell auf die Coregonen auswirken. Hinzu kommt, dass neben den experimentell aufgezeigten unmittelbaren Schädigungen weiter auch indirekte Einflüsse (z.B. P. rubescens bedingte Veränderungen in der Umwelt der Coregonen) eine entscheidende Rolle spielen können.

Insgesamt betrachtet unterstützen die bisherigen Erkenntnisse damit die Vermutung, dass das anhaltende Aufkommen von P. rubescens eine wesentliche Ursache für den Wachstumseinbruch und die schlechte physiologische Kondition der Coregonen in Voralpenseen wie dem Ammersee ist.

S ^UMMARY

Toxic cyanobacteria affect organisms of almost all stages of development and trophic levels.

Especially threatened are aquatic organisms such as fish, as they can be affected by toxic cyanobacteria via multiple routes, and their options for exposure avoidance in waters containing toxic cyanobacteria are limited. Among the most toxic cyanobacteria are species of the genus Planktothrix, including Planktothrix rubescens. During the last decades, P. rubescens has become one of the predominant species of the phytoplankton community in several lakes in the pre-alpine regions. In some of those lakes (e.g. Lake Ammersee) the rise of P. rubescens has been observed to coincide with pronounced slumps in fishery yields bringing the professional fishery into existential difficulties. These slumps have primarily been characterised by prominent growth reduction of coregonids (Coregonus sp.), resulting in reduced fish fitness which appears to be associated with the regular disappearance of specific age groups of coregonid. As Lake Ammersee coregonids have repeatedly displayed blue coloured gut contents indicating coregonid contact with cyanobacteria, it appeared plausible that the challenge to this coregonid population might be causally related to the occurrence of toxic P. rubescens. The aim of the study was therefore

• to characterise the density, distribution and toxicity of P. rubescens in Lake Ammersee,

• to investigate environmental observations in controlled experimental exposure studies with respect to possible detrimental effects on coregonids and finally,

• to assess the evidence linking P. rubescens exposure and adverse effects on feral coregonids in Lake Ammersee

A prerequisite for the proposed P. rubescens observations was the need for a rapid and precise method for the quantification of P. rubescens cell densities. Thus, initial work focused on the validation of an image processing system, which automatically measures fluorescing P. rubescens filaments on illuminated filters. Subsequent field studies using this system demonstrated that P. rubescens was present during the entire observation period from 1999-2004, albeit at varying cell densities. Filaments were regularly distributed over the entire water column during winter and stratified in distinct metalimnic layers during summer, reaching cell densities of up to 15,000 cells/ml and 75,000 cells/ml, respectively. P. rubescens mass occurrence was demonstrated to be strongly influenced by water transparency, i.e. illumination in the metalimnion.

Microcystins (predominantly MC-RR variants) were detectable in 27 and 38 of 54 monthly seston samples via HPLC and ELISA measurements, respectively. These analyses further suggest that microcystin production by P. rubescens is consistent and consequently, that the appearance of P. rubescens coincides with measurable microcystin levels.

The impact of P. rubescens on coregonids was examined in experimental exposure studies, where the environmentally observed P. rubescens cell densities and various forms of application were

within tissues the quantification of low microcystin concentrations in fish tissue is difficult if not impossible via ELISA-, HPLC- and PPAssay analyses. Consequently, the presence and localisation of microcystin was determined immunohistochemically, using anti-microcystin antibodies.

Coregonids exposed to P. rubescens showed abnormal behaviour, increased ventilation rates and elevated plasma glucose levels, presumably representing a behavioural and physiological stress response. Histopathological alterations in liver, gastrointestinal tract and kidney suggested substantial tissue damage and therefore sustained alteration in normal organ function. The fact that these alterations were also immunopositive for microcystin further indicated an uptake of microcystins and causality of tissue damage and the presence of microcystin. In addition, susceptibility to ectoparasitic infestation and increased mortality in exposed fish suggested these P. rubescens associated effects to impair fish fitness. The pathology and stress response of exposed coregonids was comparable across the different exposure experiments. Although, even low cell densities (≈1500 cells/ml) resulted in significant injury, the progression and severity of the observed adverse effects occurred in a dose-dependent manner, indicating that the higher the P. rubescens cell densities and hence microcystin concentrations, the more pronounced and earlier the onset of the adverse effects.

P. rubescens cell densities greater than 1500 cells/ml were demonstrated to occur in Lake Ammersee during 47% of the 261 weeks observed, thus suggesting that Lake Ammersee coregonids are indeed regularly confronted with detrimental P. rubescens exposure situations.

This is corroborated by field observations demonstrating P. rubescens filaments in gut contents of Lake Ammersee coregonids. This additionally gives evidence that feral coregonids actually ingest P. rubescens. These field investigations further demonstrated this exposure to result in an accumulation of P. rubescens components within the coregonid intestine, as the investigated fish showed a significant accumulation of cyanobacterial biliproteins explaining the prominent blue colouration of gut contents and implying possible coregonid exposure to P. rubescens toxins.

Indeed, from coregonids sampled during P. rubescens bloom episodes in 1998 and 1999, four out of ten fish contained significant microcystin accumulation in the gut content, unambiguously demonstrating microcystin exposure of feral coregonids in Lake Ammersee. The detection of covalently-bound microcystin in liver tissue of Lake Ammersee coregonids furthermore demonstrates microcystins to traverse the ileal membrane and to accumulate in the liver. As corroborated by the experimental exposure studies, this makes substantial detrimental effects on the coregonids appear inevitable.

In conclusion, prolonged occurrence of toxic P. rubescens can thus be expected to substantially affect feral coregonids. In addition to the direct detrimental effects outlined above also indirect effects, such as P. rubescens-induced environmental changes are likely. The current investigation hence substantiates the initial hypothesis of a causal relationship between mass occurrences of P. rubescens and challenged coregonid populations in pre-alpine lakes such as Lake Ammersee.

P UBLICATIONS & PRESENTATIONS

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First author:

• Ernst, B., Hoeger, S.J, O´Brien, E. & Dietrich, D.R.: Abundance and toxicity of Planktothrix rubescens in the pre-alpine Lake Ammersee, Germany. Submitted for publication in Harmful Algae.

• Ernst, B., Hoeger, S.J, O´Brien, E. & Dietrich, D.R. (2007): Physiological stress and pathology in European whitefish (Coregonus lavaretus) induced by subchronic exposure to environmentally relevant densities of Planktothrix rubescens. Aquatic Toxicology 82, 15-26.

• Ernst, B., Hoeger, S.J, O´Brien, E. & Dietrich, D.R. (2006): Oral toxicity of the microcystin- containing cyanobacterium Planktothrix rubescens in European whitefish (Coregonus lavaretus). Aquatic Toxicology 79, 31-40.

• Ernst, B., Neser, S., O´Brien, E., Hoeger, S.J. & Dietrich, D.R. (2006): Determination of filamentous cyanobacteria Planktothrix rubescens in environmental water samples using an image processing system. Harmful Algae 5, 181-189.

• Ernst, B., Dietz, L., Hoeger, S.J. & Dietrich, D.R. (2005): Recovery of MC-LR in fish liver tissue. Environmental Toxicology 20, 449-458.

• Ernst, B., Hitzfeld, B.C. & Dietrich, D.R. (2001): Presence of Planktothrix sp. and cyanobacterial toxins in Lake Ammersee, Germany and their impact on whitefish (Coregonus lavaretus L.) Environmental Toxicology 16, 483-388.

Co-author:

• Hoeger, S.J., Schmid, D., Blom, J., Ernst, B. & Dietrich, D.R. (2007): Specifics of microcystin- RR variants: consequences for analytical procedures and risk assessment. Environmental Science and Technology 41, 2609-2616.

M

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• Ernst, B. & Dietrich, D.R. (2001): Cyanobakterien auf dem Vormarsch. Geoskop, GEO, 12, 217-219.

O

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• Ernst, B. & Dietrich, D.R.: Blaugrünes Alsterwasser, Roter Chiemsee und fragwürdige Nahrungsergänzungsmittel: zu den fatalen Gefahren von Cyanobakterien.

3. interdisziplinäres Forum der Arthur und Aenne Feindt Stiftung, 19. April 2006, Hamburg, Germany.

• Ernst, B. & Dietrich, D.R.: Toxic Planktothrix rubescens and its subchronic impact on European whitefish (Coregonus lavaretus). 6^th International Conference on Toxic Cyanobacteria, 21. - 26. Jun. 2004, Bergen, Norway.

• Ernst, B.: Toxins of Planktothrix rubescens and their subchronic impact on European whitefish (Coregonus lavaretus L.). Second Late Summer Workshop der GDCh, 29. Sep. - 01. Oct. 2003, Maurach, Germany.

• Ernst, B.: Auswirkungen toxischer Cyanobakterien auf Fische am Beispiel der Renken (Coregonus lavaretus L.) im Ammersee. 5^th EESL Statuskolloquium 19. - 20. Nov. 2001, Konstanz.

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First author:

• B. Ernst & Dietrich, D.R.: Regular exposure of coregonids (Coregonus sp.) to Planktothrix rubescens may be causal for reduced fish weight and fitness and hence recurrent slumps in fishery yields in pre-alpine lakes. 7^th International Conference on Toxic Cyanobacteria, 05.-10.

Aug. 2007, Rio das Pedras, Brazil.

• Ernst, B., Neser, S., Hitzfeld, B.C. & Dietrich, D.R.: Determination of filamentous cyanobacteria in water samples using the image processing system Visiometrics IPS.

10^th International Conference on Harmful Algal Blooms, 21. - 25. Oct. 2002, St. Petersburg, Florida, USA.

• Ernst, B., Hitzfeld, B.C. & Dietrich, D.R.: Microcystin contamination of fish from a European pre-alpine lake via chronic exposition to Planktothrix sp.: A human health hazard?

10^th International Conference on Harmful Algal Blooms, 21. - 25. Oct. 2002, St. Petersburg, Florida, USA.

• Ernst, B., Neser, S. & Dietrich, D.R.: Bestimmung der Dichte fädiger Cyanobakterien in Wasserproben mit dem digitalen Bildverarbeitungssystem Visiometrics IPS 5^th EESL Statuskolloquium 19. - 20. Nov. 2001, Konstanz, Germany.

• Ernst, B., Hitzfeld, B.C. & Dietrich, D.R.: Presence of Planktothrix sp. and cyanobacterial toxins in Lake Ammersee, Germany and their impact on whitefish (Coregonus lavaretus L.) 5^th International Conference on Toxic Cyanobacteria, 15. - 20. Jul. 2001, Noosa, Queensland, Australia.

• Ernst, B., Hitzfeld, B. & Dietrich, D.R.: Detection of cyanobacterial toxins in whitefish (Coregonus lavaretus L.) from lake Ammersee. SOT 39^th Annual Meeting, 19. - 23. Mar. 2000, Philadelphia, PA, USA.

Co-author

• D.R. Dietrich, B. Ernst & Day, B.W.: Human consumer death and algal supplement consumption: A post mortem assessment of potential microcystin-intoxication via microcystin immunohistochemical (MC-IHC) analyses. 7^th International Conference on Toxic Cyanobacteria, 05.-10. Aug. 2007, Rio das Pedras, Brazil.

• A. Fischer, S. J. Hoeger, D. Feurstein, B. Ernst & Dietrich, D. R.: Importance of organic anion transporting polypeptides (OATPs) for the toxicity of single microcystin congeners in vitro.

7^th International Conference on Toxic Cyanobacteria, 05.-10. Aug. 2007, Rio das Pedras, Brazil.

• Dietz, L., Ernst, B., Höger, S.J. & Dietrich, D.R.: Recovery of MC-LR in fish liver.

6^th International Conference on Toxic Cyanobacteria, 21. - 26. Jun. 2004, Bergen, Norway.

• Schmid, D., Ernst, B., Höger, S.J. & Dietrich, D.R.: Characterization and differentiation of two microcystins from Planktothrix spec. isolated from a pre-alpine lake in Europe.

6^th International Conference on Toxic Cyanobacteria, 21. - 26. Jun. 2004, Bergen, Norway.

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&

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• Environment Award of the Environment and Living Foundation, University of Konstanz, Germany (2007).

• Ph.D. scholarship by the Arthur & Aenne Feindt Foundation, Hamburg, Germany.

• Congress grant by the Environment and Living Foundation, University of Konstanz (2004).

• Congress grant by the German Research Foundation, DFG (2007).

C ^ONTENTS

1. I

1.1. C

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1.2. C

OLIGOPEPTIDES ……… 20

ALKALOIDS ……… 24

OTHER CYANOBACTERIAL TOXINS ……… 27

CYANOBACTERIAL TOXINS – COMPARISON OF TOXIC POTENTIALS ……… 29

1.3. C

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CYANOBACTERIA INDUCED FISH KILLS ……… 32

THE ICHTHYOTOXICITY OF MICROCYSTIN ……… 34

ICHTHYOTOXICITY OF OTHER CYANOBACTERIAL TOXINS ……… 49

1.4. T

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2. M

2.1. D

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A^BSTRACT ……… 57

INTRODUCTION ……… 58

M^ATERIAL & METHODS ……… 59

R^ESULTS ……… 63

DISCUSSION ……… 65

ACKNOWLEDGEMENTS ……… 67

2.2. R

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A^BSTRACT ……… 68

INTRODUCTION ……… 69

M^ATERIAL & METHODS ……… 70

RESULTS ……… 73

DISCUSSION ……… 76

ACKNOWLEDGEMENTS ……… 80

3. E

3.1. O

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)

ABSTRACT ……… 81

INTRODUCTION ……… 82

M^ATERIAL & METHODS ……… 83

RESULTS ……… 87

DISCUSSION ……… 93

ACKNOWLEDGEMENTS ……… 95

3.2. P

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ENVIRONMENTALLY RELEVANT DENSITIES OF P

.

ABSTRACT ……… 96

INTRODUCTION ……… 97

MATERIAL & METHODS ……… 98

R^ESULTS ……… 102

DISCUSSION ……… 109

ACKNOWLEDGEMENTS ……… 112

4. F

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4.1. A

,

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ABSTRACT ……… 113

INTRODUCTION ……… 114

M^ATERIAL & METHODS ……… 115

RESULTS ……… 120

DISCUSSION ……… 125

ACKNOWLEDGEMENTS ……… 133

4.2. T

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ABSTRACT ……… 134

INTRODUCTION ……… 135

MATERIAL & METHODS ……… 136

R^ESULTS ……… 137

DISCUSSION ……… 139

5. G

5.1. A

L

A

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INITIAL SITUATION ……… 142

THE TOXICITY OF P. RUBESCENS IN LAKE AMMERSEE - BASIC CONSIDERATIONS 144 DIRECT IMPACT OF TOXIC P. RUBESCENS ON LAKE AMMERSEE COREGONIDS … 144

P. ^RUBESCENS-INDUCED ENVIRONMENTAL CHANGES: INDIRECT EFFECTS ON COREGONIDS ……… 147

CONSEQUENCES ……… 149

CONCLUDING ASSESSMENT ……… 150

5.2. A

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I^RRITATION & ACCIDENTAL INTOXICATION DURING RECREATIONAL WATER ACTIVITIES ……… 151

INTOXICATION VIA UPTAKE OF CONTAMINATED DRINKING WATER AND FOOD 152

6. A

7. R

8.

A

1. I NTRODUCTION

1.1.

C

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“LAKE ALBERT IN WAGGA WAGGA HAS BEEN CLOSED BECAUSE OF TOXIC LEVELS OF BLUE-GREEN ALGAE …” (ABC News, Australia 25.01.07), “ALGAE OUTBREAK PROMPTS WATER WARNING IN

QUEBEC –HUNDREDS OF QUEBECERS ARE BEING WARNED NOT TO DRINK THEIR WATER BECAUSE OF AN OUTBREAK OF BLUE-GREEN ALGAE …” (TheStar.com, Canada 06.07.07), “WATER SPORTS HAVE BEEN BANNED AT A MANCHESTER LAKE FOLLOWING REVELATIONS THAT TOXIC BLUE-GREEN ALGAE HAS BLOOMED IN THE WATER …” (BBC News, England 23.08.07). Headlines like these, regularly show people worldwide quite plainly that in a lot of waters a distinct problem is growing:

Cyanobacteria (synonyms: blue-greens, blue-green algae, cyanoprokaryotes, cyanophyceans, cyanophytes, myxophyceans, etc.) – a few micrometer small organisms with a promising potential for the future, but also a source of considerable nuisance and hazard for animal and human health.

The variety of names highlights the important position that cyanobacteria occupied in the scientific past. Since their earliest characterisation by Linné (1753), they have been a matter of interest for many scientists of various disciplines, including botanists (Geitler, 1932; Vaucher, 1803), microbiologists (Forti, 1907), limnologists (Lampert, 1981; Skulberg, 1964), biochemists (Singh et al., 2005) and toxicologists (Falconer et al., 1981).

Cyanobacteria provide an extraordinary contribution to human affairs in every day life, they are important primary producers (Whitton & Potts, 2000 and references therein) and they contribute globally to soil and water fertility in rice fields and other agricultural wetlands (Whitton, 2000).

Cyanobacteria produce an incredible number of metabolites and extraordinary pigments, making them interesting to the pharmaceutical, cosmetics and colours & dyes industries (Singh et al., 2005) and even in the utilisation of alternative energies (Skulberg, 1994).

Cyanobacterial mass occurrences however also present a considerable nuisance for the management of water bodies. Surface scum, water colouration, unpleasant odour and particularly the production and release of highly potent toxins affect aquaculture, drinking water treatment, crop irrigation and recreational water use (Dietrich & Hoeger, 2005; Hitzfeld et al., 2000). The water quality problems caused by dense cyanobacterial populations are intricate and have great health and economic impacts. Thus, public concern is rising and with it, the interest in cyanobacteria and cyanobacterial toxins.

What are cyanobacteria, where do they occur and where do they come from? Cyanobacteria are ancient gram-negative prokaryotes. Fossil occurrences of cyanobacteria are thought to date to

earliest life forms. Fossil cyanobacteria are primarily isolated from stromatolites¹. Despite their long evolutionary history, the fossil forms resemble very closely the currently occurring species (Schopf, 2000).

Cyanobacterial cells are encased by a cell wall and lipopolysaccharides. A few species are additionally covered by exopolysaccharides, which protect against digestion by potential grazers (Friedland et al., 2005; Kolmakov & Gladyshev, 2003; Lewin et al., 2003). Cyanobacterial cells include no nucleus² and no other membrane covered cellular organelles (e.g. golgi, mitochondria, endoplasmatic reticulum and chloroplasts). Their ribosomes, synthesising their proteins, are of the bacterial type. Thus, the only class (classis: cyanobacteria) of the Cyanophyceae is taxonomically categorised as bacteria (summarised in Adams & Duggan, 1999; Mur et al., 1999 and van den Hoek & Jahns, 2002).

The morphology of cyanobacteria is very diverse comprising unicellular, colonial as well as multicellular filamentous forms. They include species with or without specialised cells (i.e.

heterocysts and akinets) and possess many more characteristic specialities. Based on primarily morphological characteristics, cyanobacteria are systematically classified into five orders:

Chroococcales (including the toxic genera Microcystis), Nostocales (including the toxic genera Anabaena, Aphanizomenon, Cylindrospermopsis & Nodularia), Oscillatoriales (including the toxic genera Planktothrix and Lyngbya), Pleurocapsales and Stigonematales (summarised in Whitton

& Potts, 2000 and van den Hoek & Jahns, 2002). According to the International Code of Botanical Nomenclature there are 150 genera including about 2000 species (Mur et al., 1999; van den Hoek

& Jahns, 2002). Not even 50 of these species have yet been shown to produce toxins however, those toxic species are very widespread, leading to the actuality that approximately 75 % of waters containing cyanobacteria also contain cyanobacterial toxins (Sivonen & Jones, 1999).

Their diversity enables cyanobacteria to colonise a tremendous variety of ecosystems, including both largely barren and infertile, as well as nutrient rich conditions. Cyanobacteria have been isolated from hot springs (e.g. Yellowstone National Park; see Ward & Castenholz, 2000 and references therein), from extremely halophilic (e.g. the Dead Sea, Israel; see Oren, 2000 and references therein) and alkaline waters (e.g. Lake Bogoria, Kenya; see Krienitz et al., 2003), as well as from cold glacial- and polar lakes (Vincent, 2000 and references therein; Jungblut et al., 2005). They can be found in terrestrial and arid habitats, where they grow epi- and endolithic on rocks and walls (summarised in Wynn-Williams, 2000 and Pentecost & Whitton, 2000). The majority however arise in aquatic environments, i.e. marshland and waters, including salt- brackish- and in particular freshwater (Mur et al., 1999; van den Hoek & Jahns, 2002).

Altogether, cyanobacteria are ubiquitous and thus it is not surprising that also cyanobacterial toxins can be detected worldwide (Sivonen & Jones, 1999).

1 Stromatolites are sedimentary growth structures formed via trapping, binding, and cementation of sedimentary grains by microorganisms, especially cyanobacteria (Stal, 2000)

Fig. 1.1: Cyanobacterial bloom development is based on a complex interaction of various factors which are influenced by each other and moreover by daily, seasonal and long-ranging alterations (e.g. climatic changes).

The reproduction of cyanobacteria is asexual, via cell division, filament (i.e. trichome-) fragmentation or by the formation of special hormogonia³ (Mur et al., 1999; van den Hoek &

Jahns, 2002 and references therein). A few species can additionally form akinetes, resting cells which develop from vegetative cells and can endure several years of unfavourable environment (Adams & Duggan, 1999).

Whenever conditions are favourable, cyanobacteria can proliferate rapidly, resulting in the formation of cyanobacterial blooms⁴. Those blooms can reach densities of up to 10⁹ cells/ml (Zohary & Madeira, 1990) and normally accumulate at the surface or in the near surface stratum of a lake (Oliver & Ganf, 2000). A few species (e.g. Planktothrix rubescens, Cylindrospermopsis raciborskii) however can also accumulate in distinct layers below the surface stratum, for example in the metalimnion of stratified lakes (Blikstad-Halstvedt et al., 2007; Falconer, 2005;

Hoeger et al., 2004; Jacquet et al., 2005).

Cyanobacterial blooms occur especially in regions with elevated nutrient input into water bodies (i.e. eutrophication), which are consequential from either natural circumstances or anthropogenic influences for example a lack of concomitant sewage treatment (Bartram et al., 1999). This resulted in the assumption that cyanobacterial mass occurrence serves as a unique indicator of eutrophication. However, this must be reviewed as recent scientific findings clearly demonstrate, that cyanobacterial bloom development is based on a complex interaction of numerous factors (summarised in Mur et al., 1999; Oliver & Ganf, 2000 and Falconer, 2005), including water body morphology, water column stability, the temperature and light regime and weather conditions (Fig. 1.1). Thus it has become evident that the reduction of external loading (i.e. re- oligotrophication) alone does not guarantee the disappearance of cyanobacteria from a lake (Jacquet et al., 2005; Morabito et al., 2002).

3 Filament fragments that detach by active gliding motion and gradually develop into new filaments (van den Hoek & Jahns, 2002)

4 A rapid increase in the population of algae/cyanobacteria to densities as to render it visible to the human eye (Vollenweider,

COMPETITION WITH OTHER SPECIES CYANOBACTERIAL

B

CONDITIONS

WATER COLUMN STABILITY PRE

-

-

POPULATIONS MORPHO

- &

TEMPERATURE

,

WEATHER GRAZING

Cyanobacteria are aerobic photoautotrophic organisms – meaning that their principal mode of energy metabolism is photosynthesis and their life processes require beside inorganic nutrients primarily water, carbon dioxide and light (Mur et al., 1999; van den Hoek & Jahns, 2002). It has been demonstrated that certain cyanobacterial species are able to survive long periods in complete darkness (Micheletti et al., 1998) and furthermore that approximately 50% of the existing species show a distinct ability for facultative heterotrophic nutrition⁵ (Adams & Duggan, 1999 and references therein). The ability to colonise those diverse and extreme habitats requires further particular qualities in addition to the basic physiological properties described above.

Indeed, cyanobacteria possess a few remarkable specialisations allowing the occupation of ecological niches and a competitive advantage compared to other photoautotrophic organisms.

The photosynthetic pigments of cyanobacteria include besides chlorophyll a and carotinoids, the accessory pigments phycocyanin (providing their characteristic blue-green colour), allophycocyanin and occasionally phycoerythrin (van den Hoek & Jahns, 2002). These biliproteins are combined in phycobilisomes and enable maximum light utilisation, from the spectrum between 400 and 700 nm wavelengths, also including the green light range (550-650 nm wavelength), which is largely inaccessible to the majority of green algae and plants (Oliver &

Ganf, 2000; van den Hoek & Jahns, 2002). Most cyanobacterial species additionally perform chromatic adaptation (Mur et al., 1999; Oliver & Ganf, 2000). This means that their synthesis of photosynthetic pigments is particularly susceptible to light quality and environmental influences and in consequence that species are able to produce the accessory pigment needed to absorb light most efficiently. Those extraordinary pigments and abilities allow cyanobacteria to exist under low light conditions and thus in ecological niches which are unreachable for most eukaryotic algae, e.g. under ice cover or in the metalimnion of lakes (Blikstad-Halstvedt et al., 2007).

Cyanobacteria furthermore have a remarkable ability to store essential nutrients and metabolites (e.g. carbohydrate granules, lipid globules, cyanophycin granules, polyphosphate bodies, carboxysomes, etc.) enabling them to endure temporary nutritional poverty (summarised in Mur et al., 1999; Oliver & Ganf, 2000 and van den Hoek & Jahns, 2002). Several species moreover have the capability for nitrogen fixation putting them at an advantage during conditions with limited availability of inorganic nitrogen (Adams & Duggan, 1999; Oliver & Ganf, 2000). Nitrogen fixing species use the enzyme nitrogenase to convert molecular nitrogen directly into utilisable ammonium. This usually occurs in heterocysts⁶.

As another speciality, many cyanobacterial species possess gas vesicles - gas filled, cytoplasmatic inclusions enclosed by a semi-permeable membrane and acting as a buoy. In most cyanobacterial species the buoyancy resulting from those vesicles can be regulated, by cellular inclusion of photosynthetic products (corresponding to cellular ballast and changing the cellular turgor pressure which may end in the collapse of vesicles), or via the generation and degradation of gas vesicles (Oliver & Ganf, 2000 and references therein). This buoyancy regulation allows planktonic

5 Heterotrophic species consume and assimilate organic substrates (e.g. glycogen granules, lipid globules, etc.) in order to get carbon and energy for growth and development (Adams & Duggan, 1999)

6 Specialised, nitrogen fixing cells with a thickened cell wall and without photosynthetic activity, so that an anoxic cellular

cyanobacteria to stratify and move effectively in the water column, adapting to the best position within vertical gradients of chemical and physical factors (Dokulil & Teubner, 2000; Mur et al., 1999; Oliver & Ganf, 2000).

Finally, most cyanobacteria species attain maximum growth rates by temperatures above 25°C (Dokulil & Teubner, 2000 and references therein). Thus, at high temperatures cyanobacteria have an obvious growth advantage over green algae and diatoms, which prefer much lower temperatures.

These physiological properties however vary between different cyanobacterial species. Thus, not all species are adapted to the various environmental conditions in the same way. This allows the emergence of diverse ecostrategists⁷ (Mur et al., 1999).

Scum-forming ecostrategists (e.g. Microcystis sp., Anabaena sp., Aphanizomenon sp., etc.) attain best growth conditions and simultaneously protection from detrimental high light intensities at the water surface through continuous movement throughout the near-surface layer (Falconer, 2005). When buoyancy regulation is disturbed (e.g. due to changing weather conditions), these species accumulate in unpleasant-smelling scums at the water surface. Homogenously distributed ecostrategists (e.g. Planktothrix agardhii, Limnothrix redekei, etc.) develop in circulating water bodies where they are distributed over the whole water body (Briand et al., 2002). In contrast, stratifying ecostrategists (e.g. Planktothrix rubescens, Cylindrospermopsis raciborskii) develop distinct layers primarily in the metalimnion of thermally stratified lakes (Blikstad-Halstvedt et al., 2007; Falconer, 2005). These species are adapted to low light conditions and benefit from the nutrient rich metalimnic situation. Benthic ecostrategists (e.g. Oscillatoria limnosa, Phormidium sp., etc.) form coherent mats on the bottom sediments of water bodies that are sufficiently clear to allow light penetration to the ground (Mez et al., 1997; Wood et al., 2007).

Benthic ecostrategists can thus also occur in oligotrophic lakes. In contrast, the mass development of nitrogen fixing ecostrategists (e.g. Anabaena sp., Aphanizomenon sp., Cylindrospermopsis sp., Nodularia sp., etc.) can be mainly related to periodic nitrogen limitation occurring in both shallow and deep water bodies mostly exhibiting eutrophic and hypertrophic conditions (Mur et al., 1999; Oliver & Ganf, 2000).

Based on the described morphological and physiological fundaments, cyanobacteria have managed to survive millions of years. Moreover, they temporarily even developed burgeoning growth - during the Precambrian, also called » the age of cyanobacteria «, 2.5 to 0.5 million years ago. Due to oxygen release associated with their photosynthetic activity, this age culminated in the accumulation of oxygen in the biosphere (Adams & Duggan, 1999). Thus, cyanobacteria actually cleared the way for the predominantly aerobic life on earth as exists today – all the more astonishing that they now increasingly attract attention because of the risks they pose to human and animal life due to the production of highly toxic metabolites.

7 Classification of ecotypes, according to their behaviour in the water column and according to their adaptation for specific

1.2.

C

Cyanobacterial mass occurrences and detrimental effects on human and animal life have been associated at least since the early Middle Ages. As noticed by Bartram et al. (1999), there are reports from the Chinese Han dynasty of loosing troops from poisoning as soldiers drank from a river that was green in colour, approximately thousand years ago. In 1189, Gerald of Wales documented Lake Llangorse, to “turn bright green” and to “became scarlet” throughout his Journey through Wales (Belov et al., 1999). Although largely anecdotal and not documented on scientific analyses as we know it today, these and other reports suggest that not only cyanobacteria, but also cyanobacterial toxins might have accompanied life on earth for hundreds of years.

The first observation giving scientific evidence for poisonings caused by cyanobacterial mass occurrence was given by George Francis (1878). He documented a bloom of Nodularia spumigena

“… forming a thick scum like green oil paint … as thick and pasty as porridge” in Lake Alexandria, Australia, and demonstrated this bloom to act poisonously, and to cause rapidly death of sheep, horses, dogs and pigs subsequently to symptoms including stupor, unconsciousness, as well as convulsions and spasm.

Since this time, the awareness of toxic cyanobacteria has increased continuously. At first, by a rising number of scientific documentations of fish kills and livestock mortalities of animals living in and drinking from lakes and ponds containing toxic cyanobacteria (summarised in Falconer, 2005 and Kuiper-Goodman et al., 1999), and since the nineteen-eighties, by the chemical and toxicological characterisation of numerous toxic metabolites which were isolated from cyanobacterial blooms (Dow & Swoboda, 2000; Sivonen & Jones, 1999).

Cyanobacteria produce a variety of unusual substances which are considered as secondary metabolites and thus appear not essential for the organism’s metabolic pathways by itself (Bentley, 1999). Most of them serve yet unclear functions, however, some have been associated with various bioactive, as well as toxic attributes (summarised in Dow & Swoboda, 2000; Kreitlow et al., 1999 and Singh et al., 2005). The latter belong to diverse groups of natural toxins regarding both, their chemical and toxicological characteristics, including alkaloids, lipopolysaccharids and multifarious oligopeptides that cause neurotoxic, hepatotoxic, cytotoxic, as well as dermatotoxic and irritant effects.

Almost all cyanobacterial toxins are intracellular toxins. Therefore, they are primarily released into the water by natural bloom senescence or anthropogenically induced cell death connected with water treatment (e.g. chlorination, algaecide application) and less by a continuous toxin

Z X D-Measp

Mdha

D-Glu

D-Ala Adda

excretion. Organisms may be exposed either from a direct uptake of cyanobacterial toxins (dissolved in water or accumulated in contaminated food) or from the uptake of toxin containing cyanobacterial cells and their subsequent digestion resulting in a toxin release within the digestive tract (Kuiper-Goodman et al., 1999; Landsberg, 2002).

O

Globally the most frequently found cyanobacterial toxins are oligopeptide toxins (Sivonen &

Jones, 1999). They primarily occur in blooms in fresh and brackish waters and are produced by more than 30 genera belonging to all taxonomical classis (i.e. Oscillatoriales, Nostocales, Chroococcales, Stigonematales and Pleurocapsales). More than 600 peptides have been characterised including 80 structural archetypes of compounds (Welker & Döhren, 2006). While most of them (e.g. aeruginosins, microginins, anabaenopeptins, cyanopeptolins, microviridins, cyclamides) are considered to provide negligible or at most slight toxicity (e.g. inhibiting chymotrypsin, plasmin and proteases)(Chorus, 2006; Grach-Pogrebinsky et al., 2003; Ishida et al., 1997; Itou et al., 1999; Rohrlack et al., 2003; Sano & Kaya, 1995; Sano & Kaya, 1996, etc.), nodularins and especially microcystins have been associated with numerous severe animal and human intoxications all over the world (Falconer, 2005; Kuiper-Goodman et al., 1999).

Microcystin & Nodularin

Microcystins and structural related nodularins are cyclic peptides named after the organism from which they were first isolated: Microcystis aeruginosa (Botes et al., 1982; Botes et al., 1985) and Nodularia spumigena (Rinehart et al., 1988). Microcystins have since also been isolated from various Anabaena sp., Planktothrix sp. and less frequently from Nostoc sp., Anabaenopsis sp. and Hapalosiphon sp. In contrast, nodularin production is restricted to Nodularia sp. yet (Sivonen &

Jones, 1999).

Both, microcystins and nodularins are produced non-ribosomally, by multi enzyme complexes and peptide synthetase genes (summarised in Falconer, 2005). They contain either seven (microcystins; Fig. 1.2), or five (nodularins; Fig. 1.3) amino acids, building a cyclic structure.

Fig. 1.2: Chemical structure of microcystins (MW: 900-1,100 D), i.e. cyclo-(D-Alanine-X-D-MeAsp-Z-Adda-D-

D-Glu

Mdhb

L-Arg Adda

D-Measp

Fig. 1.3: Chemical structure of nodularin (MW: 824 D), i.e. cyclo-(D-Alanine-X-D-MeAsp-Z-Adda-D- glutamate-Mdha)

Both are primarily composed of D-amino acids and collectively contain the unusual Adda⁸ as well as D-Glutamic and D-Methylaspartic acid (Measp). The microcystin molecule moreover contains D-Alanine, N-Methyldehydroalanine (Mdha) and two L-amino acids. Structural variations in microcystins have been characterised for all seven amino acids, but most frequently with substitution of the L-amino acids and demethylation of the Mdha- and Measp-residue. This variability results in more than 75 naturally occurring microcystin congeners (Spoof, 2004). The nodularin molecule is completed by N-Methyldehydrobutyrine (Mdhb) and D-Alanine. In contrast to microcystins, structural variations of nodularin are restricted to limited demethylations. Thus, only seven naturally occurring nodularin congeners have been characterised to date (summarised in Craig & Holmes, 2000 and Sivonen & Jones, 1999).

Microcystins and nodularins are extremely stable and persistent once released into the water.

Rapid chemical hydrolysis has only been obtained under environmentally irrelevant conditions including exposure to 6 M hydrochloric acid and high temperatures (Harada, 1996). In natural environments microcystins and nodularins may persist for weeks until their photochemical breakdown and isomerisation. Photochemical degradation can be accelerated through photosensitisation due to the presence of photopigments and humic substances (summarised in Sivonen & Jones, 1999 and Welker et al., 2001). Despite their chemical stability, microcystins have been shown to be susceptible to breakdown by aquatic bacteria. After an initial lag phase, which may persist for several days, the biodegradation process commences and microcystin removal of up to 90 percent can be achieved within 2-10 days depending on environmental conditions, e.g. water temperature, water body hydrology, initial microcystin concentration, etc.

(Jones et al., 1994; Lahti et al., 1997; Rapala et al., 2005; Welker et al., 2001).

Although the bacterial and photosensitised degradation is assumed to be much more efficient than non-catalysed breakdown, it appears most probable that high microcystin and nodularin concentrations - which mostly occur locally, within and around a senescent cyanobacterial bloom - initially decrease due to dilution within the water body rather than via chemical or biological degradation (Jones & Orr, 1994). Such dilution may prevent acute intoxication however it broadens contamination and thus may increase the number of exposed organisms.

The uptake of microcystin and nodularin by exposed organisms from either water or ingested cyanobacterial cells is dominated by the polar characteristic of the cyclic peptides. Except a few microcystin congeners containing hydrophobic amino acids (e.g. microcystin-LW and microcystin- LF), microcystins and nodularins are hydrophilic and therefore soluble in water and hence also in blood of exposed organisms (Botes et al., 1982; Harada, 1996). Due to their hydrophilic character, nodularins and most microcystins however are unable to penetrate lipid membranes via passive diffusion (Falconer, 2005; Vesterkvist & Meriluoto, 2003). Thus, the majority of those peptides ingested are unable to pass the epithelium of the ileum and consequently remain in the digestive tract, from where the toxins are most likely excreted via the faeces (Fujiki et al., 1996). However, ingested microcystins and nodularins have been shown to be at least fractionally transported across the ileum into the venous bloodstream and also from the portal vein into hepatocytes via bile acid membrane transporters (e.g. organic anion transporters (OATPs)) (Eriksson et al., 1990;

Fischer et al., 2005; Runnegar et al., 1991, etc.). Presumably due to the first pass effect⁹ and the high density of transporting peptides in hepatocytes, the liver is the main target organ for both accumulation and detoxification of microcystins and nodularins. This categorises them as primarily hepatotoxic (Falconer et al., 1986; Kaya, 1996; Robinson et al., 1991, etc.). Microcystins are however also detectable in other OATP-expressing organs (i.e. intestine, kidney, brain, lung, heart, etc.), albeit to a considerably lesser extent (Ito et al., 2000; Robinson et al., 1989; Robinson et al., 1991; Spoof et al., 2003a, etc.). This indicates that toxicity in these organs is also likely.

The toxicity of microcystins and nodularins is mainly mediated via a strong inhibition of protein phosphatases (PP). This inhibition results from a non-covalent interaction of the peptide’s Adda- glutamate domain with the catalytic site of especially the serine/threonin phosphatases PP1 and PP2A (Honkanen et al., 1990; MacKintosh et al., 1990; Yoshizawa et al., 1990). As the Adda- glutamate domain is well conserved across various microcystins and nodularins, the toxicity of almost all structural variants shows low variation (LD50: ≥50 µg/kg bw; mouse toxicity (i.p.)¹⁰), except for a few congeners containing considerable modifications to the Adda-glutamate region and thus showing no or negligible PP-inhibition (summarised in Sivonen & Jones, 1999). The initial binding (formed within minutes) can be followed by the formation of an additional irreversible covalent bond between the Mdha-residue of microcystins and a cysteine residue of the

9 Molecules which were absorbed via the digestive system enter the hepatic system prior to other organs due to the flow direction of the bloodstream leaving the digestive system

10 Single-dose level that will cause death in 50 per cent of exposed animals within 7-14 days. The LD depends, besides the

PP-molecule (Craig et al., 1996; Goldberg et al., 1995). The formation of the covalent bond requires several hours and indeed strengthens the interactions between microcystins and PPs, however it does not strengthen PP-inhibition (Bagu et al., 1995; Maynes et al., 2006). A covalent bond cannot arise with nodularin due to a lack of Mdha. Correspondingly, it is also impeded in conjunction with microcystin congeners that include significant alterations at the Mdha-residue (Craig et al., 1996; Hastie et al., 2005).

As protein phosphatases are key cellular enzymes which regulate and control a variety of cellular functions and processes via the dephosphorylation of proteins (Barford et al., 1998; Cohen, 1989;

Janssens & Goris, 2001), the inhibition of PPs by microcystins and nodularins results in a hyperphosphorylation of phosphate regulated proteins (Falconer & Yeung, 1992; Ohta et al., 1992; Wickstrom et al., 1995; Yoshizawa et al., 1990). Such hyperphosphorylation causes a tremendous cascade of consequences in hepatocytes, including disintegration of cellular structure, loss of cell-cell adhesion, disruption of cellular metabolism, interference with signal transduction and disturbance of cell cycle control (summarised in Craig & Holmes, 2000; Falconer, 2005;

Gehringer et al., 2004; Kuiper-Goodman et al., 1999).

In addition to the described consequences of PP-inhibition, further damage may result from a microcystin induced generation of reactive oxygen species (ROS), which also affects cellular activities (summarised in Gehringer et al., 2004) and furthermore, from binding to the ATP synthetase which represents a trigger of mitochondrial apoptotic signalling (Mikhailov et al., 2003).

Any or all of these alterations may result in severe liver damage (e.g. apoptosis, necrosis, loss of lobular architecture, intrahepatic haemorrhage, hepatic insufficiency, etc.) and subsequently death in the case of acute intoxications, as well as increased cellular proliferation and thus tumour promotion and/or initiation following chronic exposure. As PPs are present in various tissues and due to the organotropism of nodularin and microcystin, comparable cellular alterations and tissue damage also occur in organs other than liver, however these are usually less severe (Falconer et al., 1992; Khan et al., 1995; Nobre et al., 1999; Nobre et al., 2004).

For assessing the toxicity of a substance, not only are uptake, distribution and modes of action fundamental, but detoxification and degradation must also be considered. Microcystins and nodularins are highly resistant to eukaryotic peptidases. This is probably due their cyclic structure and their amino acid composition predominantly consisting of D-amino acids (Harada, 1996). The persistence of microcystins is additionally elevated by their covalent and thus irreversible binding to protein phosphatases (summarised in Falconer, 2005 and Kuiper-Goodman et al., 1999).

A

Despite their chemical relationship, cyanobacterial alkaloid toxins differ considerably in their structure and stability, their mode of action and thus also in their toxicological consequences.

Cyanobacterial alkaloids include dermatotoxic skin irritants (e.g. lyngbyatoxin and aplysiatoxin) with inflammatory and presumably even skin tumour promoting activity. They further include cytotoxic (e.g. cylindrospermopsin) and neurotoxic (e.g. PSPtoxins (saxitoxins) and anatoxins), metabolites that regularly cause animal and human poisonings.

Cylindrospermopsin

Cylindrospermopsin came to attention as a result of severe gastroenteritis outbreak among children drinking from a water supply in Queensland, Australia, containing a bloom of Cylindrospermopsis raciborskii (Byth, 1980; Hawkins et al., 1985). Cylindrospermopsin is primarily detected in tropical and subtropical waters although increasing occurrence has also been observed in temperate climates, for example in New Zealand, Europe and South and North America (Falconer, 2005 and references therein). It is mainly produced by Cylindrospermopsis raciborskii. However, also other cyanobacterial species (e.g. Umezakia natans, Anabaena bergii, Anabaena lapponica, Raphidiopsis curvata and some Aphanizomenon sp.) have meanwhile been shown to produce cylindrospermopsin (Falconer, 2005 and references therein, Rucker et al., 2007;

Spoof et al., 2006).

The toxin was first isolated and characterised by Ohtani et al. (1992). It comprises a tricyclic guanidine combined with hydroxymethyl uracil and is a highly water-soluble molecule (Fig. 1.4). Cylindrospermopsin is stable in darkness and shows slow breakdown in pure water under environmentally relevant conditions. Similar to microcystins, the photochemical breakdown of cylindrospermopsin is accelerated (≥90% degradation within 2-3 days) via photosensitisation in the presence of photosensitive pigments (Chriswell et al., 1999).

Cylindrospermopsin has a flat, tricyclic molecule structure with rotational bonds within the two main components and can thus intercalate into the DNA double helix causing chromosome breaks and irreversible inhibition of protein synthesis (Falconer, 2005). Additional toxic effects have been assumed. Consequently, cylindrospermopsin causes a time- and dose-dependent cytotoxicity and presumably possesses mutagenic, clastogenic and even carcinogenic activity (Falconer, 2005).

In consequence of the inhibition of protein synthesis, cylindrospermopsin primarily injures tissues presenting a high synthesis/turnover of proteins, i.e. liver and kidney, intestine, and the immune system including spleen and thymus. Poisonings may lead to decreased functionality of these organs and subsequently even death, depending on the duration and dose of exposure (Briand et al., 2003; Falconer, 2005 and references therein).

PSPtoxins (Saxitoxins)

Saxitoxins are primarily known in conjunction with paralytic shellfish poisoning (PSP). They were originally isolated from shellfish, where they may accumulate due to filter-feeding of toxic marine dinoflagellates (Anderson, 1994). As a consequence of the consumption of contaminated shellfish, PSPtoxins become available for higher trophic levels and have been associated with animal, livestock and human mortalities (Briand et al., 2003; Kuiper-Goodman et al., 1999;

Landsberg, 2002).

In cyanobacteria, PSPtoxins have been found in Aphanizomenon flos-aquae, various Anabaena sp. (predominantly A. circinalis), Cylindrospermopsis raciborskii (Sivonen & Jones, 1999; van Apeldoorn et al., 2007) as well as in Lyngbya wollei and Planktothrix sp. (Carmichael et al., 1997;

Onodera et al., 1997b; Pomati et al., 2000).

PSPtoxins are carbamate alkaloid toxins with a tricyclic structure of hydropurine rings (Fig. 1.5).

There are at least 21 derivates, which differ among others predominantly in the substitution of five variable positions (Oshima, 1995; Sivonen & Jones, 1999; van Apeldoorn et al., 2007 and references therein).

Similar to microcystin and cylindrospermopsin, PSPtoxins are also rather persistent toxins. Their chemical breakdown in darkness and room temperature is slow, often requiring more than three months for ≥90 % breakdown (Jones & Negri, 1997). However, no detailed studies have as yet been carried out on its breakdown in sunlight and it is unknown if this can be accelerated via photosensitisation.

The severity of toxicity differs considerably among diverse PSPtoxin variants, with saxitoxin being the most toxic. The toxicity of all congeners is attributable to a generalised blockade of sodium channels in the excitable membranes in nerve axons (Briand et al., 2003; Kuiper- Goodman et al., 1999 and references therein). This results in a dose-dependent, partial or complete inhibition of action potential transmission in peripheral nerves and skeletal muscles.

Thus, PSPtoxins cause general nerve dysfunction, culminating in paralysis, respiratory depression, and in the case of an acute dosage death due to respiratory failure.

Anatoxin-a

Presumably the most widespread cyanobacterial alkaloid is anatoxin-a. Anatoxin-a occurrences and episodes, including animal and human poisonings, have been reported from all over the world (Kuiper-Goodman et al., 1999 and references therein; Edwards et al., 1992; Gugger et al., 2004;

Gunn et al., 1992; Wood et al., 2007). Anatoxin-a was first isolated from Anabaena flos-aquae (Devlin et al., 1977) and later from various other Anabaena species as well as from species of Aphanizomenon, Planktothrix, and Cylindrospermopsis (Sivonen & Jones, 1999; van Apeldoorn et al., 2007 and references therein).

It is a low weight secondary amine (Fig. 1.6) and whilst the toxin appears relatively stable in darkness, in contrast to saxitoxin, it undergoes a rapid photochemical degradation in sunlight and even more in alkaline conditions. The toxin moreover appears to be readily degraded by bacteria that are associated with cyanobacterial filaments (summarised in Sivonen & Jones, 1999).

Anatoxin-a is a post-synaptic nicotinic agonist and thus potent neurotoxin. It binds to neuronal nicotinic acetylcholine receptors at neuromuscular junctions, mimicking the effect of acetylcholine. This causes an influx of Na⁺ and subsequently, depolarisation, which results in the opening of voltage sensitive Ca⁺⁺ and Na⁺ channels. In contrast to acetylcholine, anatoxin-a is not susceptible to enzymatic hydrolysis by acetylcholinesterases and consequently causes a prolonged stimulus blocking subsequent electrical transmission and potentially leading to muscular paralysis. Symptoms of anatoxin-a toxicity hence include muscle fasciculation, loss of coordination, staggering, gasping, convulsion and subsequent acute intoxication death by respiratory paralysis (summarised in Briand et al., 2003; Kuiper-Goodman et al., 1999 and Metcalf & Codd, 2005).

The anatoxin-a homologue homoanatoxin-a is produced by certain Phormidium, Anabaena and Raphidiopsis species (Furey et al., 2003; Namikoshi et al., 2003; Skulberg et al., 1992).

Homoanatoxin-a is a methyl derivate of anatoxin-a with slightly lower potency (based on the LD50

in mice following intraperitoneal application). It has been shown to block muscular contraction by neurostimulation in phrenic nerve-hemidiaphragm. Homoanatoxin-a toxicity is based on an enhanced Ca⁺⁺ flux in the cholinergic nerve terminals. Comparable to anatoxin-a it acts very quickly and may cause death due to respiratory arrest within minutes (Lilleheil et al., 1997; van Apeldoorn et al., 2007 and references therein).

Fig. 1.7: Chemical structure of anatoxin-a(s) (MW: 252 D).

Anatoxin-a(s)

Anatoxin-a(s) is a naturally occurring organophosphate. It is a unique phosphate ester of cyclic N- hydroxyguanine (Fig. 1.7) which decomposes rapidly in alkaline settings but is stable under neutral and acidic conditions (Matsunaga et al., 1989). Anatoxin-a(s) has been identified in blooms of Anabaena flos-aquae and A. lemmermannii (Henriksen et al., 1997; Onodera et al., 1997a).

Symptoms of acute anatoxin-a(s) toxicology include muscle weakness, convulsion, respiratory distress and death due to respiratory failure. As these neurotoxic symptoms are very similar to those of anatoxin-a, anatoxin-a(s) was initially considered to be an anatoxin-a homologue.

However, in contrast to anatoxin-a, anatoxin-a(s) is a acetylcholinesterase inhibitor (Cook et al., 1989), blocking the acetylcholine regenerating enzyme in a manner similar to organophosphate insecticides (e.g.malathion, parathion) and poison gases (i.e. tabun and sarin).

O

Only a few cyanobacterial toxins can neither be categorised to oligopeptides or alkaloids. The most prominent are Lipopolysaccharids (LPS), which are located at the outer cyanobacterial cell wall and thus are no intracellular toxins.

Lipopolysaccharides (LPS)

Cyanobacterial LPS were first isolated from Anacystis nidulans (Weise et al., 1970). LPS are endotoxins and thus generally an integral component of the cell wall of Gram negative bacteria, which include cyanobacteria. They are condensed products of sugar and fatty acids, whose composition shows considerable variation among bacteria and also among cyanobacteria (Sivonen

& Jones, 1999).

LPS have pyrogenic activity and can act as irritants mainly due to their fatty acid component (Fig. 1.8). Exposure may also elicit an allergenic response, causing fever and induce gastroenteritis. However, cyanobacterial LPS are considerably less potent than LPS from pathogenic bacteria such as Salmonella (summarised in Briand et al., 2003; Kuiper-Goodman et al., 1999).

Fig. 1.8: Chemical structure of the toxic component of lipopolysaccharides.

Beta-N-methylamino-L-alanine (BMAA)

The unique non-protein amino acid BMAA is structurally similar to methylated alanine (Fig. 1.9).

It is previously known from the seeds of the Guam cycad Cycas micronesica (Vega & Bell, 1967).

BMAA has lately gained attention as it has been shown to biomagnify in the Guam ecosystem and to be associated with an increased occurrence of amytrophic lateral sclerosis/Parkinsonism dementia complex in the indigenous Chamorro people (Spencer et al., 1987). Recent investigations demonstrate the toxicity of the cycad seeds to result from cyanobacteria, living as endosymbionts in the coralloid roots (Cox et al., 2003). Meanwhile, BMAA has been shown to be produced by all known groups of cyanobacteria (i.e. Chroococcales, Nostocales, Oscillatoriales, Pleurocapsales and Stigonematales) in various environments (Cox et al., 2005).

Once ingested, BMAA can be bound by proteins within the body, functioning as an endogenous neurotoxic reservoir that slowly releases toxin directly into the cerebral tissue through protein metabolism (Murch et al., 2004). The toxicity of BMAA is primarily based on the potent interaction as an agonist of glutamate AMPA/kainate receptors (excitotoxin) at glutamergic synapses in the brain and spinal cord (Weiss et al., 1989), whereas additional toxic effects have been assumed (Murch et al., 2004).