# Chapter 32: Cyanobacterial toxins: a qualitative **meta–analysis of concentrations, dosage and**

**effects in freshwater, est uarine and marine biota** 

Bas W Ibelings 1 

# , Karl E Havens 2

1 

# Netherlands Institute of Ecology (NIOO–KNAW) – Centre for Limnology, Rijksstraatweg 6, 3631 AC, Nieuwersluis, The Netherlands; e–mail b.ibelings@nioo.knaw.nl; 2 

# University of Florida, Department of Fisheries and Aquatic Sciences, Gainesville, FL 33458, USA



# Abstract

This paper reviews the rapidly expa nding literature on the ecological effects of cyanobacterial toxins. The study employs a qualitative meta–analysis from the literature examining results from a large number of independent studies and extracts general patt erns from the literature or signals contradictions. The meta–analysis is set up by putting together two large tables – embodying a large and represen tative part of the literature (seeAppendix A). The first table (Table A.1) reviews the presence (concentrations) of different cyanobacterial toxins in the tissues of various groups of aquatic biota after exposure via different routes, experimentally in the labor via natural routes in the environment. The second table (Table A.2) reviews the dose dependent effect of toxins on biota. The great majority of studies deal with the presence and effects of microcystin, especially of the MC–LR congener. Although this may par tly be justified – MC–LR is an abundant and highly toxic protein – our review also emphasizes what is known about (i) other MC congeners (a number of studies showed a preferred accumulation of the less toxic variant MC–RR in animal tissues), (ii) nodularin (data on a range of biota from studies on the Baltic Sea), (iii) neurotoxins like anatoxin–a(s), which are conspicuously often present attimes when mass mortalities of birds occur, (iv) a few studies on the presence and effects of cylindrospermposin, as well as (v) the first examples of

# 676 B.W. Ibelings and K.E. Havens



# ecological effects of newly identified bioactive compounds, like microviridin–J. Data were reorganized to asses s to what extent bioconcentration (uptake and concentration of toxins from the water) or biomagnification (uptake and concentration via the food) of cyanobacterial toxins occurs in ecosystems. There is little support for the occurrence of biomagnification, and this reduces the risk for biota at higher trophic levels. Rather than biomagnification biodilution seems to o ccur in the foodweb with toxins being subject to degradation and excretion at every level. Nevertheless toxins were present at all tropic levels, indi cating that some vectorial transport must take place, and in sufficient qua ntities for effects to possibly occur.Feeding seemed to be the most important route for exposure of aquatic biota to cyanobacterial toxins. A fair number of studies focus on dissolved toxins, but in those studies purified toxin typically is used, and biota do not appear very sensitive to this form of exposure. More effects are found when crude cyanobacterial cell lysates are used, indicating that there may be synergistic effects between diffe rent bioactive compounds. Aquatic biota are by no means defenseless against toxic cyanobacteria. Several studies indicate that those species that are most frequently exposed to toxins in

their natural environment are also the most tolerant. Protection includes

behavioral mechanisms, detoxicati on of MC and NODLN by conjugation with glutathione, and fairly rapid depuration and excretion. A commontheme in much of the ecological studies is that of modulating factors. Effects are seldom straightforward, but are dependent on factors like the(feeding) condition of the animals, environmental conditions and the history of exposure (acclimation and adap tation to toxic cyanobacteria). This makes it harder to generalize on what is known about ecological effects of cyanobacterial toxins. The paper conc ludes by summarizing the risks for birds, fish, macroinvertebrates and zooplankton. Although acute (lethal)

effects are mentioned in the literature , mass mortalities of – especially – fish are more likely to be the result of multiple stress factors that co–occur during cyanobacterial blooms. Bivalves a ppear remarkably resistant, whilst the harmful effects of cyanobacteria on zooplankton vary widely and the specific contribution of toxins is hard to evaluate.

# Chapter 32: Cyanobacterial Toxins 677

List of abbreviations.

Abbreviation Definition AchE Acetyl Choline Esterase

ANA(a)(as) Anatoxin–a; anatoxin–a(s)

BCF Bioconcentration Facor (concentration of toxic compound in an organism as % of that in water)

BMF Biomagnification Factor (concen tration of toxic compound in an organism as % of that in its diet)

CAT Catalase (one of the antioxidative enzymes)

CYN Cylindrospermopsin

GI Gastrointestinal tract

GPx Glutathione Peroxidase (one of the antioxidative enzymes)

GR Glutathione Reductase (one of the antioxidative enzymes)

GSH Glutathione

GST Glutathione–S–Ttransferase (catalyst of the formation of MC– GSH conjugates in detoxication)

H 2 

# O 2 

# Hydrogen peroxide (one of the ROS formed during oxidative stress)

HP Hepatopancreas IP Intraperitoneal injection

LC 50 

# Concentration at which 50 % of the test animals die from exposure to the toxin

LOEC Lowest Observable Effect Concentration

LPO Lipid Peroxidation (outcome of oxidative stress)

LPS Lipopolysacharides

MC Microcystin

N ODLN N odularin

PP Protein phosphatases (inhibition of PP by MC results in hyperphosphorilation of proteins)

PST Paralytic Shellfish Toxin

ROS Reactive Oxygen Species (for med during oxidative stress)

SAX Saxitoxin

SOD Super Oxide Dismutase (one of the antioxidative enzymes)

TDI Tolerable Daily Intake (of a toxin like MC)

TEH Total Extractable Hepatotoxins (sum of toxins and their biotransformation products)

TOSC Total Oxygen Scavenging Capacity

# 6



# Introduction

Until just a few years ago statements like “traditionally research has focused on the acute toxicity of microcystin–LR to laboratory mammals” and “there is a general lack of researc h involving aquatic organisms which may be exposed to toxin producing cyanobact erial blooms in their natural environment” (Zurawell et al. 1999) were fully justified. In contrast, there now exists a large literature from observa tional and experimental studies dealing with cyanobacterial toxins in aquatic systems. Yet major information gaps remain, and our ability to unders tand effects is limited by certain attributes of those studies. The aim of this paper is to review the extant literature, identify general patterns in results, and identify key areas where

additional research is warranted. This assessment is complex because in addition to microcystin–LR there are many other toxins produced by cyanobacteria. Some of these toxins, such as nodularin, are closely related to microcystin, while others are quite different (e.g., the neurotoxins anatoxin–a and a(s) and saxitoxin and theprotease inhibitor cylindrospermopsin). There also exists an ever increasing list of bioactive compounds produced by cyanobacteria, some of which have been shown to be toxic to selected aquatic biota (like microviridin–J

for Daphnia (Rohrlack et al. 2004), but many of which have not been studied in any detail. Furthermore, chr onic and sub–chronic effects (see Havens et al elsewhere in this volume for definitions) may be more relevant to study than acute lethal effects. Exposure of biota in lakes supportingcyanobacterial blooms is likely to be repetitive and over much of an organism’s lifespan. Although cyanobacterial toxins have been claimed to play a

role in acute events like mass mortalities of fish and birds, there is usually

insufficient evidence to link fish and bird kills directly to these toxins. This does not mean that there are no important sub–lethal effects resulting from chronic exposure in the aquatic ecosystem. This study employs qualitative meta–ana lysis of data from the literature, examining results from a large number of independent studies and synthesizing summaries and conclusions addressing the issue of toxic effects. Meta–analysis aims to utilize the increased power of pooled data to clarifythe state of knowledge on that issue, and may include quantitative statistical analyses when the data are consistent with that approach – this is not the case here. We set up the meta–analysis by putting together two large tables (see Appendix A). The first table (Table A.1) reviews the presence(concentrations) of cyanobacterial toxins in the tissues of aquatic biota after exposure via different routes, be it experimental or via natural routes in the field. The second table (Table A.2) reviews the dose dependent effects

78 B.W. Ibelings and K.E. Havens

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# of toxins on biota. While Tables A.1 and A.2 do not include 100% of the published papers on these subjects, th ey do embody a large and representative part of the literature.



# Methodology

The literature was queried using the ISI–Web of Science. Results of the literature search are given in Havens et al (this volume). To assemble Table A.1 – concentrations in biota – the following data were extracted from the literature: (i) biota involved, four groups are distinguished: birds, fish, macroinvertebrates and zooplankton); ( ii) type of cyanobacterial toxinstudied; (iii) exposure route; (iv) toxin concentrations in biota; (v) concentration in the source (i.e. this could be dissolved purified toxin in an experimental setting, cells from cultures of toxic cyanobacteria or natural seston containing toxic cyanobacteria); and (vi) analytical analysis method that was used to quantif y the toxin. In the discussion, data from Table A.1 will be organized in such a way that biomagnification (accumulation of the toxins in biota via consumption of food that contains the toxins) can be quantified. Bioaccumulation of cyanobacter ial toxins is often speculated to increase the risk of exposure for aquatic biota (especially at higher trophic levels), but bioaccumulation has seldo m been analyzed correctly (see Havens et al, this volume for definitions). Especially for MC it is well known that the analytical method may have a marked effect on the concentration

that is measured. This is even more so when MC is measured in biota, rather than in the toxin producing cyanobacteria. Standard MeOH extraction does not include covalently bound MC and analysis using ELISA suffers from cross reactivity between MC or NODLN and their GSH conjugates formed in detoxication. The cons equences of this for interpretation of the concentrations given in Table A.1 are discussed. For Table A.2 the following data were compiled: (i) biota involved; (ii) exposure route; and (iii) dose and effect. In many studies it is not possible

to establish true dose–effect relati onships because organisms are exposed to only one or two different dosages, and Table A.2 will indicate in which cases sufficient data have been gathered to establish a relationship. Ideally the unit for dose would be units of toxin administered per unit of body weight and per unit of time (for comparison TDI for humans equals a dose

of 0.04 μ g kg bw -1 

# ). It is more common however to find toxin contents of the source expressed in μ g L -1 

# . Differences in units hamper interpretation across different studies. What will also emerge from Table A.2 is that theexposure route has a strong influence on biological effects. In fish IP injecChapter 32: Cyanobacterial Toxins

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# tion of MC is often fatal however oral dosage hardly ever results in mortality. As will be discussed, fish–kills in lakes seem to be a consequence of multi–stress factors during blooms of toxi c cyanobacteria rather than of direct intoxication. Tables A.1 and A.2 are included as an appendix of this

paper.



# Results and Discussion

The last 5 years have included a steady increase in the number of papersinvestigating cyanobacterial toxins in aquatic biota. Whether this increasingly large body of literature is sufficiently broad (in terms of the toxins / bioactive compounds and aquatic biota covered by the studies) and deep (in terms of yielding a detailed unders tanding of the effects these toxins have on aquatic biota) is another matte r. Below we formulate an answer to this question – which is central to th is paper – by analyzing the assembled literature presented in the tables.

**Toxins in biota**

**Toxin producing species and their toxins**



# This review takes a somewhat unusual pe rspective in that the focus is not on toxin producing cyanobacteria but rather on the biota in the aquatic ecosystem that may be affected by the toxins. Cyanobacteria that (frequently) appear in the tables are the MC producing genera Microcystis (species M.

aeruginosa

, M. flos aquae and M. viridis ) and Planktothrix ( P. agardhii

and P. rubescens ), the main NODLN producing species Nodularia spumigena and the anatoxin–a and a(s) producing genus Anabaena. Although species within a genus may differ greatly in their ecology ( P. agardhii for instance is commonly found in hypertro phic, turbid shallow lakes, whereas

P. rubescens is typically found in clear, deep alpine lakes), we do not stress differences between toxin producers at the species level since often we know little about differences in toxin production within a genus like for

instance Microcystis

. If there is one thing immediately striking about the tables it is the prominence of MC, especially of MC–LR (not all studies specify which MC congeners were present, although most studies express total MC as MC–LR equivalents since often MC–LR is the only standard used in HPLC analysis). This focus on the presence and effects of MC is probably(partly) justified because surveys of cyanobacterial toxins in several coun80 B.W. Ibelings and K.E. Havens

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# tries have indeed shown that microcystins are prominently present (e.g. in Denmark – Henriksen et al. 1997). The spotlight on MC–LR (one variety in a family of around 70 different MC congeners) may be more biased, and seems influenced by the laboratory work on mammals. MC–LR is relatively toxic, its LD 50 

# in mice is 50 μ g kg -1 

# , considerably lower than for instance the LD 50 

# of MC–RR (600 μ g kg -1 

# ) (Spoof 2005). Yet as the data inTable A.1 indicate there are a number of studies demonstrating that especially the less toxic MC–RR is taken up into tissues of aquatic biota, for instance in silver carp (Jang et al. 2004); or in freshwater snails (only MC– RR present in foot of B.aeruginosa (Chen and Xie 2005) and hepatopancreas of S.histrica (Ozawa et al. 2003). On the basis of their observations on the ratio of MC–LR:MC–RR in different tissues and organs, Xie et al. (2004) suggested that MC–LR may be actively degraded during digestion, whereas MC–RR is transported across the intestines and embedded into body tissues. At the same time Xie su ggested that MC–RR is not acutely toxic to the carp, since no mortality was observed despite the uptake of MC–RR into various organs. In a study using the Thamnocephalus bioassay LC 50 

# of MC–RR (and of MC–YR) was actually very close to the LC 50 

# of MC–LR. The real difference in th is invertebrate was a much gentler slope of MC–RR compared to MC–LR (and YR) when LC 10 

# and LC 90 

# values were included (Blom et al. 2001).

A further relevant distinction within the microcystins is that between the methyldehydroalanine–containing micr ocystins which covalently bind to

PP in the cell and MC that contains dehydrobutyirine and – like NODLN – that do not bind covalently to PP. For NODLN it has been suggested that because NODLN does not bind covalently its transfer in the food web isfacilitated (Kankaanpaa et al. 2001). The same would be true for the dehydrobutyirine containing microcystins, but since concentrations of these have not been analyzed in biota there is no evidence for this. Overall it is clear that studies on MC–LR cover just a small part of the total complexity of interactions between microcystin producing cyanobacteria and aquatic biota, so that the bias indicated by Tables A.1 and A.2 is unjustified. Moreresearch is needed on the ecological effects of the whole spectrum of bioactive – potentially harmful – compounds produced by cyanobacteria, including MC congeners other than LR.Some toxins other than MC have been analyzed in biota. NODLN features fairly prominently in the tables, due to considerable work that has

been done on the Baltic sea. This is one of the few regions in the world where there are sufficient data to actually follow concentrations ofNODLN and its effects throughout much of the food web (more in discussion on ‘bioaccumulation’). Neurotoxins are not a notable group in the table. Interestingly anatoxins (anat oxin–a and/or anatoxin–a(s)) often seem Chapter 32: Cyanobacterial Toxins

# 6



# to play a role when toxic cyanobacteria have been implicated in the death of waterfowl (Henriksen et al. 1997; Krienitz et al 2003). Neurotoxicosis

can be seen as convulsed extremities a nd arched back necks (Codd et al. 2005). The death of the lesser flamingos in Africa’s Rift valley lakes is an interesting example where multiple cy anobactertial toxins (neurotoxins and hepatoxins) may play a simultaneou s role and have synergistic effects (a more accurate description of the in teraction between di fferent toxins or toxins and other stress factors may be additive rather than synergistic effects). This is another area where very little is known today.

**Biota involved and organs affected**



# Tables A.1 and A.2 distinguish four groups of aquatic organisms: waterfowl, fish, macroinvertebrates (i.a. bivalves, crabs, prawns, snails) and zooplankton. There are only a handful of studies involving birds (see above). There are many more studies on fish, the other group of aquatic

vertebrates where mass mortalities have been attributed to blooms of toxiccyanobacteria. Toxins in fish have been analyzed after exposure via different routes, but the majority of work actually involves studies of fish caught

in lake or sea, i.e. fish that has b een exposed to toxins via natural routes (exposure through the food web or to dissolved toxins after lysis of blooms). In the ecosystem the feeding gu ild of a fish species appears to be a primary determinant of exposure to toxins. Phytoplanktivorous fish likesilver carp (e.g. Jang et al. 2004) directly consume cyanobacteria. Zooplanktivorous fish like sticklebacks and smelt (Ibelings et al. 2005) feed on zooplankton that directly consume cy anobacteria. Likewise a species like flounder predates on filter feeding blue–mussels. Piscivorous fish that preyon zooplanktivorous fish are one step further removed from the toxin producing cyanobacteria. It may be expect ed that in the absence of biomagnification (but see discussion below) MC concentrations decrease in the order phytoplanktivorous > zooplanktivorous > piscivorous fish . Omnvirous fish may fit in anywhere. Indeed fish caught in the IJsselmeer,The Netherlands showed an increase in MC in fish–liver moving from larger perch (predatory) to ruffe (benth ic) and zooplanktivorous smelt (Ibelings et al. 2005). In contrast Xie et al. (2005) found that MC in various tissues and organs varied as carnivorous > omnvivorous > phytoplanktivorous fish. Fischer and Dietrich (2000) explain that there are

several differences in GI tract between a carnivorous fish like rainbow trout and planktivorous and herbivorous cyprinids like carp. Cyprinids (as well as cichlids) possess a much longer ileum with larger surface area and higher resorption capacities, so that carnivorous fish would accumulate less MC, i.e. the opposite of what was found in Chinese lakes. In contrast

82 B.W. Ibelings and K.E. Havens

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# Carbis et al. (1997) explain that the ne utral or slightly basic conditions in the GI of carp limit absorption of MC, since cells of cyanobacteria would only be digested in an acid environment. Overall no relationship between feedi ng guild and toxin concentrations in fish can be pulled out from the data in Table A.1. Data from different studies are hard to compare because toxin concentrations, exposure routes

and a host of other biotic and abiotic factors differ between sites and studies. What is clear, and this is found across studies on fish , is that concentrations of MC are mainly present in the gut and liver, to a somewhat lesser extend in kidneys and gonads, and mu ch less in muscle tissue (e.g. Kankaanpaa et al. 2005a; Li et al. 2004; Soares et al. 2004; Malbrouck et al. 2003; Xie et al. 2005). Microcystins are also found in fish faeces in substantial amounts (Jang et al. 2004; Xie et al. 2005) and in pseudofaeces

of Dreissena (Babcock–Jackson et al. 2002; Pires et al. 2004). This could expose the benthic community to cyanob acterial toxins produced in the pelagic zone. There is a surprisingly large number of studies on the presence of toxinsin macroinvertebrates, especially in bivalves (mussels and clams). In contrast there are very few studies that describe effects of cyanobacterial toxins on these animals, perhaps because ge nerally they seem insensitive (e.g. Saker et al. 2004). A common theme in the studies on macroinvertebrates is the analysis of time courses for accumulation and depuration of toxins.

Accumulation is often found to be tim e dependent and proceeds in an orderly manner. Time lagged acccumulation occurred at deeper sites in the Baltic sea. Mussels at deep sites are primarily exposed to toxic Nodularia towards the end of the bloom period, when filaments sink to the sediment, where they over winter. Although mussels at deeper sites contained muchless toxin, they did accumulate some NODLN (Sipia et al. 2002). Depuration from mussels is almost always found to be biphasic (Ozawa et al.2003; Sipia et al. 2001b; Vasconcelos et al. 1995), sometimes concentrations of toxin even increase in the first phase of depuration (Amorim and Vasconcelos 1999). It has been suggested that this is a consequence of dynamics in production and degradation of PP to which the MC are bound (Vasconcelos et al. 2001), but more research is needed (Ozawa et al. 2003). Although depuration is commonly judged to be rapid (e.g. Sipia et al. 2002; Kankaanpaa et al. 2005b; Pereira et al. 2004) it is equally clear that depuration is incomplete even after a considerable period of time. Depuration is temperature dependent and slows down in winter, so that toxins may even be carried on to the next spring (Ozawa et al. 2003). In for instance Lake IJsselmeer this has consequences for thousands of diving ducks that arrive in autumn. Although summer Microcystis blooms have dispersed, the mussels (food for the duc ks) still contain traces of toxins. Chapter 32: Cyanobacterial Toxins

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# Thus the mussels may be considered a vector that prolongs the time whentoxins are able to exert negative effects in that lake ecosystem. In macroinvertebrates hepatotoxins, but also CYN, were mainly found in the haemolymph and hepatopancreas, to a lesser extend also in gonads and muscle tissue (foot).With respect to zooplankton we see the opposite of the macroinvertebrates papers: there are a large number of studies on effects of cyanobacterial toxins on zooplankton but much less on concentrations of toxins in the animals. The available results suggest that concentrations are relatively high in indiscriminately filter feeding taxa such as Daphnia (Ibelings et al. 2005; Kotak et al. 1996a; Thostrup and Christoffersen 1999), and perhaps lower in copepods, but again there is a general lack of data. Toxins seem to be taken up into the body of zooplankt on, concentrations of toxin cannot be explained solely by the presence of toxic cyanobacteria in the gut.

**Analytical analysis methods**



# The standard method for analysis of hepatotoxins (MC and NODLN) is HPLC, coupled to diode array UV det ection (see Table A.1). ELISA is frequently used because of its high sensitivity (see Spoof 2005) for pros and cons of different methods. A drawback of ELISA is cross reactivity with detoxication metabolites, like the conjugates of MC and GSH. These conjugates have been shown to have a much lower toxicity (Metcalf et al. 2000). Conjugates can be detected using LC–MS, but this is still rarely undertaken (however see Karlsson et al . 2003 and Sipia et al. 2002). Because ELISA suffers from cross reativity some studies on biota in the Baltic Sea

have introduced the term TEH – tota l extractable hepatotoxins, which includes the biotransformation products. TEH almost invariably exceeds the concentrations of untransformed hepa totoxins in biota – see for instance the comparison of NODLN (analyzed on LC–MS) and TEH (ELISA) in Kankaanpaa et al. (2005a), which differ by an order of magnitude, or Lehtonen et al. (2003) where NODLN was < 5 % of TEH in Baltic clams. Another important analytical issue is that of extraction of the toxins. A large number of MC congeners – thos e that contain methyldehydroalanine – covalently bind to PP in plant a nd animal cells; these covalently bound MC are not extracted using standard MeOH extraction. The handful of

studies that have used Lemieux oxidation – a method that does extract covalently bound MC – demonstrated that a large part of the MC in biota is covalently bound (Williams et al. 1997; Pires et al. 2004) (see Table 2 in Havens et al, this issue). This means that almost all of the concentrations given in Table A.1 seriously underestimate the total amount of MC present

in biota. What is unknown – and this is important – is whether all these

4 B.W. Ibelings and K.E. Havens

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# studies also underestimate the bioavailability and toxicity of MC. Is covalently bound MC in Daphnia still (equally) toxic to the fish that swallows

cladocerans?

**Bioaccumulation**



# Many studies have suggested that cy anobacterial toxins bioaccumulate in aquatic biota and that this may enhance the risk of exposure of biota higher up in the food web (e.g. Li et al. 2004; Sipia et al. 2001a; Negri and Jones1995). Xie et al. (2005) present data that demonstrate that MC has a general tendency to accumulate up the food chain, with concentrations being highest in carnivorous and lowest in herbivorous fish species. PST concentrations in Daphnia magna grazing on Aphanizomenon exceeded those in the cyanobacterium (bioaccumulation factor > 1). Bioaccumulation in most papers however use a loose definition an d usually it just means that toxins are present in biota. When a more fo rmal – and informative – definition of bioaccumulation, and the related proc esses of bioconcentration and biomagnification are used (see Havens et al, this issue for definitions) there is very little evidence to support the notion of bioaccumulation of MC and NODLN in aquatic food webs. Rather the opposite, i.e. biodilution of hepatotoxins in the food web, is suppor ted by the data (Karjalainen et al. 2005). Data in Table A.1 do show how ever that bioconcentration of NODLN may take place. In an experi mental setting two copepods and a ciliate took up dissolved NODLN and accu mulated this to concentrations far higher than in the water (BCF ranged from 12–22). Predators of these zooplankters and protozoa would be e xposed to substantial concentrationsof toxin in their food and may suffer consequences like decreased ingestion rates, as was shown for pike larvae and mysid shrimps feeding on the zooplankton (Karjalainen et al. 2005). Biomagnification factors express the concentration of a toxin in biota as a percentage of that in their diet. BMF for the Baltic Sea and IJsselmeer

are shown in Table 1.1. BMF in the Baltic biota and in Daphnia and Dreissena from the IJsselmeer are well below 100 %, indicating that the concentration was much below the concentration in the seston. Biomagnification obviously is absent in these cases. BM F in the Baltic of Copepods and clams are exceptionally low; only a very small part of NODLN which is present in the cyanobacteria is take n up by these grazers. The calculation in Table 1 of BMF for grazers of the phytoplankton is very sensitive to the toxin content of the seston used in the calculation, and this content may be highly variable. Values of BMF in the table should be taken as indicative

rather than absolute. The BMF of ruffe and especially smelt in the IJsselmeer are > 100 and seem to indicate that biomagnification of MC has Chapter 32: Cyanobacterial Toxins

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# taken place. In this case however concen trations of MC in a whole organism (like Daphnia ) are compared to values fo r a selected organ where thetoxin specifically accumulates (the liver), and this gives a skewed representation of biomagnification (Gray 2002). The difference in BMF between freshwater mussels in the IJsselmeer and their marine counterparts in the Baltic is striking. The very low concentration of MC in Dreissena led Ibelings et al. (2005) to the conclu sion that the food web linked to filter feeding mussels is hardly exposed to toxins. In contrast Kankaanpaa et al. (2005a) concluded that in the Baltic food webs involving mussels are especially exposed to hepatotoxins. Th e tenfold difference in BMF supports these apparently opposing conclusions.

Table 1. Biomagnification factors Baltic Sea and IJsselmeer (The Netherlands). BMF were calculated as NODLN (Baltic) or MC (IJsselmeer) content in biota as apercentage of toxin in their diet (e.g. in eiders as a % of that in mussels). For comparison BMF is also calculated as percentage of the concentration in the seston (although this would only qualify as biomagnification for organisms that actually feed on seston, like Daphnia and the mussels). Data compiled from (Engstrom– Ost et al. 2002; Kankaanpaa et al. 2005a; Karj alainen et al in press; Lehtonen et al. 2003; Sipia et al. 2001b; Sipia et al. 2002; Sipia et al. 2004) for the Baltic sea.

BMF for the IJsselmeer have been modified from Ibelings et al. (2005). Data on

which calculation of BMF are based are taken from Table A.1.

Baltic biota BMF

seston BMF diet IJsselmeer biota BMF

seston BMF

diet Copepods 0.3 0.3 Daphnia galeata 20 20 Blue mussel (Mytilus edulis) 8.9 8.9 Zebra mussel(Dreissena polymorpha) 0.9 0.9

Baltic clam (Macoma baltica) 0.6 0.6 Perch (Percia fluviatilis) 5.9 11

Mysid shrimp (Mysis relicta) 0.3 100 Ruffe (Gymnocephalus cernuus) 13.2 120

Pike larvae (Esox lucieus) 0.2 59 Smelt (Osmerus eperlanus) 53.5 286

Sticklebacks (Gasterosteus aculeatus) 0.05 24

Flounder (Platichthys flesus) 1.6 19

Eider (Somateria mollissima) 0.8 8 6 B.W. Ibelings and K.E. Havens

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# Exposure routes

**Laboratory studies**



# Early studies on fish (Tencalla et al. 1994; Kotak et al. 1996b) primarily used the method which is preferred fo r exposure of mammals in laboratory studies – intra–peritoneal injection. Direct injection of toxins like MC proved to be highly toxic to fish. The effects are comparable to those seen

in mammals, but differences are seen as well. Whereas mammals die from haemorrhagic shock following hepatocyte insult, fish die from direct liver failure, necrosis (e.g. Malbrouck et al. 2003, Li et al 2004). The LC 50 

# for MC–LR in perch (1500 μ g g DW -1 

# , Ibelings et al unpublished data) is well

above the LC 50 

# for mice, indicating that these fish species are less sensitive

to the toxin than warm blooded animal s. Nevertheless Sipia et al. (2001a) notes that salmon hepatocytes seem more sensitive to algal toxins than rat hepatocytes. When MC was administered orally (up to 1150 μ g MC kg -1



# bw given by gavage 8 times over 96 h (a total dose of 9200 μ g MC kg -1



# bw) to perch from the IJsselmeer no mortality was seen, although histopathology of the livers showed that MC were having severely detrimental effects. Similar differences between IP injection and gavage can be found in Table A.2 (e.g. Tencalla et al. 1994).

**Directly from the water**



# In ecotoxicology there is a general assumption that uptake from the wateris a common route for aquatic vertebrates to accumulate xenobiotic substances (Karjalainen et al. 2003). Indeed some of the studies have demonstrated direct uptake of cyanobacterial toxins from the water, even to the

extent that the concentration in biota exceeds those in the water. However, most studies where biota are exposed to dissolved toxins have been in the laboratory (see Table A.2) using purif ied toxins. These studies have proven valuable in finding the mechanisms through which biota are affected bycyanobacterial toxins but are less informative about the importance of uptake of dissolved toxins in the ecosy stem. A specific effect of dissolved MC is inhibition of ATP–ase activity of Na + 

# K + 

# pumps in the gills of fish and crabs, resulting in ion imbalance (Best et al. 2003; Vinagre et al. 2003; Zambrano and Canelo 1996). Concentra tions of dissolved MC are much increased when surface blooms of floating cyanobacteria lyse. When cyanobacteria float to the surface they are exposed to extreme conditions,in particular an increase in irradiance, potentially to damaging levels. Photoprotective mechanisms that may protect the cells from photooxidation are hampered by the co–occurrence of light stress and other stress factors, Chapter 32: Cyanobacterial Toxins

# 6



# notably an increase in temperature, desiccation and depletion of inorganic carbon (Ibelings and Maberly 1998). Lehtonen et al. (2003) suggested that the major fate of cyanobacterial blooms in the Baltic is to disintegrate in the water column so that very little reaches the bottom. If this were the case exposure to dissolved toxins would be a major event. Lysis of surface blooms is not unlikely, but we maintain that exposure of biota to high

concentrations of dissolved toxin are the exception rather than the rule because processes like mixing, adsorption to clay particles, photolysis and bacterial degradation rapidly reduce the availability of dissolved toxins (Ozawa et al. 2003). Moreover it has been shown by several authors that some aquatic biota are not sensitive to dissolved cyanob acterial toxins – e.g., brown trout (Best et al. 2001), pike–larvae (Karjalainen et al. 2005) and Daphnia

magna (Lurling and van der Grinten 2003). Microcystins tend to be quite water soluble and polar, and do not readily pass the lipid bilayer of membranes. It is important to note however that whenever effects of purified dissolved toxins are compared to whol e cell extracts biological effects tend to be much enhanced for the latter (Palikova et al. 1998; Oberemm et al. 1999), a possible indication of synergistic effects between MC and other bioactive compounds in cyanobacteria l cells. There are exceptions, however. For example, in A. salina purified CYN showed a lower LC 50 

# than

crude Cylindrospermopsis extracts, and this may indicate that unidentified compounds in the cyanobacterial cell extracts lowered the bioavailability of the toxin (Metcalf et al. 2002).

**Via food (vectorial transport)**



# Feeding seems to be the most important route for exposure of aquatic biota to cyanobacterial toxins. This seems natural for organisms that directlyfeed on seston that includes cyanobacteria. Zooplankton, filter feeding bivalves and phytoplanktivorous fish w ould be among the organisms that are directly exposed to toxins in their f ood (unless they manage to avoid toxic cyanobacteria – see below ‘protective mechanisms’). For those biota that do not feed directly on cyanobacteria, toxins must reach them via the foodweb. The risk of being exposed to toxins via the food web is much increased if biomagnification takes place. This is commonly found for lipophilic toxicants like PCB, but is less likely for hydrophilic compounds like MC–LR. This congener has a very lo w octanol to water partition coefficient, but as demonstrated by Ward and Codd (1999) other variants may have higher coefficients, and toxicity to Tetrahymena has been shown to vary accordingly. As discussed above biomagnification of MC and NODLN is unlikely and not substantiated by data from the field. Of the

88 B.W. Ibelings and K.E. Havens

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# amount of toxin ingested with the food very little is actually taken up into the body (e.g. 2.7 % in Daphnia in Rohrlack et al. 2005). And even the little toxin that is actually taken up into the blood of Daphnia and transported to its organs is subject to de toxication (see later in this paper) and excretion. These processes that dilute toxin concentration act at every step in a food chain. Rather than biomagnification MC and other toxins may be subject to biodilution in the foodweb . Nevertheless toxins are found at higher trophic levels, so there must be some vectorial transport, and as will be discussed below in sufficient quan tities to have harmful effects. Thepresence of toxins in grazers like the zooplankton indicate that cyanobacteria are indeed ingested, despite their reputation of being hard to handle because of their large size. Indeed Work and Havens (2003) found cyanobacteria in the gut of all crustacean zooplankton in a large subtropical lake, including taxa known to produce toxins such as Anabaena

. A special case of exposure via food is coprophagy, described in a studyon blue mussels (Svensen et al. 2005). A fair number of studies have analyzed toxins in the faeces of various species. Concentrations may be relatively high compared to concentrations in organs and tissues, and the faeces laden with toxins provide a medium for further transport of toxins in aquatic systems, especially towards the benthic community. Examples in Tables 1 and 2 include the faeces of silver carp and C. gibelio (Jang et al. 2004), M. galloprovincialis (Amorim and Vasconcelos 1999) as well as M.

edulis (Svensen et al. 2005), and faecal pellets of calanoid copepods (Lehtiniemi et al. 2002).

**Effects on biota**

**Acute vs. chronic effects**



# Acute effects are those that result from a single exposure to a toxin. This is conceivable under laboratory settings or after large scale lysis of a surface

bloom. Biota in the field however will mainly be exposed repeatedly to toxins over a long period of time. Th is is sub–chronic and chronic exposure (definitions in Havens et al., this issue). An example of a study of acute exposure is that by Kankaanpaa et al. (2002) on sea trout. The fish were exposed to a single bolus of toxic Nodularia and time dependent accumulation / depuration of NODLN was coupled to the analysis of damage and recovery of the liver. An example of a sub–chronic exposure study (i.e. on a time scale intermediate be tween acute and chronic) is that by Pinho et al. (2003) where estuarine crabs were exposed daily for 4–7 d to

cell extracts from toxic Microcystis or the exposure of carp to Microcystis Chapter 32: Cyanobacterial Toxins

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# during 28d (Li et al. 2004). Experimental chronic exposure studies where biota are exposed to toxins for the great er part of their lifespan have – necessarily – been restricted to organisms with short generation times, especially zooplankton. There are a fair number of studies in which the effects of cyanobacterial toxins on the life–history of Daphnia have been studied. Examples are the studies by (Lurling 2003) and (Hietala et al. 1997). Some of the bivalve studies (accumulation / depuration) lasted for several weeks (e.g., Pires et al. 2004 and Bury et al. 1996) exposed brown trout to MC– LR for a period of 63d, but the great majority of data on chronic exposure to toxic cyanobacteria come from fi eld studies where animals are exposed to toxic cyanobacteria via natural routes (many examples in Tables A.1 and A.2) during extended periods of time.Overall Table A.2 indicates wide ranging effects of different cyanobacterial toxins on various aquatic organisms. Effects vary from mortality to subtle changes in behavior. Effects in fish include changes in liver enzymology, liver damage and ionic imbalance. Effects of cyanobacterial toxins on the embryonic development of fishes have been studied in two species: zebra fish and loach. Whereas immersion of zebra fish embryos in a solution of purified MC did not result in morphological changes except atthe very highest concentration (Oberemm et al 1999), embryonic development of loach was affected by exposure to MC (Liu et al. 2002). In the study on zebra fish it was seen – like in studies on other biota – that crude cell extracts had much stronger effects, resulting in malformations of the

fishes. Effect studies are especially rich in the zooplankton literature. In Table A.2 it can be seen that effects on zooplankton vary from feeding inhibition to reduced reproduction, growth and mortality. Feeding inhibition may actually serve as a protective mechanism, and there is some evidence that especially species that are highly sus ceptible to MC may protect themselves by strong inhibition of the intake of cyanobacteria (Demott 1999). Studies also have shown that zoopla nkton is relatively insensitive to dissolved toxins (Demott et al. 1991; Lurling and van der Grinten 2003) so that feeding inhibition may indeed be very effective in preventing harmful exposure to the toxins. A complicating factor in zooplankton studies is that also ‘non–toxic’ cyanobacteria induce effects like reduced growth and reproduction. Cyanobacteria are generally believed to be food of low quality to zooplankton, especially Daphnia , so that direct toxic effects can not always be separated from the effect of insufficient food of good quality (LaurenMaatta et al. 1997). Experimental tests in which toxicity effects were separated from food effects – by adding a sufficient amount of high quality food like the green alga Scenedesmus – clearly demonstrate however that nutritional insufficiency of Microcystis cannot be solely respon90 B.W. Ibelings and K.E. Havens

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# sible for the effects on Daphnia. Negative effects on survival, growth and population development persisted even when green algae were added.

Moreover this was true even when a Microcystis mutant was used that no longer produces MC (Lurling 2003). He nce the author concluded that harmful effects by Microcystis cannot be the result of MC only. Feeding inhibition (starvation) and unknown bioactive compounds must also play a

part.The studies by Rohrlack and co–workers (Rohrlack et al. 1999b; Rohrlack et al. 2004; Rohrlack et al. 2005) enable direct insight into the effects

of MC on Daphnia because the wild type Microcystis and its mutant only differ in their capacity to produce MC. Several interesting observations were made. Although MC was not respons ible for feeding inhibition – the mutant had an equally strong effect – clearly MC had direct toxic effects.Visible first symptoms of MC poisoning included an inhibition of movements of thoracic legs, mandibles, foregut, second antennae, as well asstimulation of gut muscles leading to a permanent contraction of the midgut. These effects became apparent as soon as MC was taken up into theblood. Contraction of the midgut interferes with digestion, nutrient assimilation and uptake of ions. Eventually MC resulted in a breakdown of Daphnia metabolism, exhaustion and eventual ly death. In nature intake of microcystins will be modulated by various factors that were not considered in the study by Rohrlack et al. (2005) like Microcystis colony size, presence of alternative food, temperature or condition of the animal (see section below on ‘modulating factors’).

**Dose effect relationships**



# In studies on Baltic Sea flounder as well as on fish from the IJsselmeer no relationship could be detected between liver histopathology and toxin concentrations (Ibelings et al 2005.; Kankaanpaa et al. 2005a). The lesions that are seen in fish livers caught fro m systems supporting dense blooms of cyanobacteria may be attributed to hepatotoxin exposure, but other factors like liver parasites and anthropogeni c pollutants will also play a part. Kankaanpaa et al. (2005a) concluded that liver histopathology can not be used as a reliable bioindicator of exposure to cyanobacterial toxins. A complicating factor is the dynamic na ture of liver damage and recovery. The acute exposure study mentioned ea rlier where flounders were given a single dose of toxin (Kankaanpaa et al. 2005a) demonstrated that damage is transient, and recovery from liver da mage is rapid (on the order of days). Many studies have failed to relate e ffects to concentrations of toxin. Egg production of Daphnia in the IJsselmeer had no relationship with toxin content of the cyanobacteria in the lake (Ibelings et al. 2005). According to Chapter 32: Cyanobacterial Toxins

# 6



# Rohrlack et al. (1999) it may not be the presence of toxins in the seston but

the actual intake of toxins that matters. Daphnia species that were presumed to differ in susceptibility to MC may actually be equally susceptible – where they actually differ may be in their ingestion rate of toxic Microcystis cells. Rohrlack et al. (2004) established a clear relationship between MC ingestion rate and LT 50 

# (survival time) of Daphnia . Despite all the complicating factors, significant dose–ef fect relationships have been foundand are included in Table A.2. Another example is the dose (and time) dependent mortality in brine shrimp expo sed to CYN and MC (Metcalf et al. 2002).

**Protective mechanisms**



# Aquatic biota are by no means defenseless against toxic cyanobacteria.

Blooms of toxic Nodularia have been around for at least 7000 years in theBaltic, giving other biota sufficient time to adapt to these nuisance cyanobacteria (Bianchi et al. 2000). Several studies indicate that species which are most frequently exposed to the toxins have the highest physiological tolerance. Baltic shrimp are less sensitiv e than fish larvae, but these larvaeonly feed on phytoplankton during the first stages of their life. Baltic copepods feed upon and ingest toxic Nodularia and they survive and reproduce without apparent harmful effects of the toxins (Engstrom et al. 2000).

Where Thamnocephalus exhibited reduced survival after grazing upon

Planktothrix filaments, other zooplankton – naturally co–existing with toxic cyanobacteria– were unaffected (Kurmayer and Juttner 1999). Sessile organisms like mussels cannot move away from cyanobacteria,but zooplankton and fish may migrate to parts of the system where concentrations of cyanobacteria are low, as has been suggested for fish in the Baltic (Karjalainen et al. 2005). More over zooplankton, mussels and fish may temporarily stop feeding when toxic cyanobacteria are present and avoid ingestion in this way. If toxic cyanob acteria can not be avoided and cells are indeed ingested, very little of the toxin present may actually be taken up into the body. Muco id cyanobacteria like Microcystis are resistant to digestion, and there are barriers for the uptake of MC across the gut epithelium into the blood (Fischer and Dietrich 2000). Rohrlack et al. (2005) however showed that presence of Microcystis in the midgut of Daphnia caused the epithelium to loose cohesi on. Cells loose contact with each other and this may facilitate the uptake of MC into the blood. Microcystin was transported by the blood to various organs, where beat rates were slowed down, until finally Daphnia died. Although both the MC producing wild type and the mutant (that no longe r is capable of MC production) affected cohesion of the epithelium, beat rates were only affected by the MC 92 B.W. Ibelings and K.E. Havens

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# producing strain, a clear demonstration that MC – if taken up into the blood – is indeed highly toxic to Daphnia

. Feeding inhibition in the presence of t oxic cyanobacteria is an efficient means to prevent ingestion of the toxi ns. However the avoidance of toxins must be balanced with the risk of starvation. Demott (1999) found that in Dahnia magna exposure to toxic Microcystis in a mixture with Scenedesmus resulted in a rapid feeding inhibition, but feeding recovered when exposure was continued. DeMott concluded that this pattern of inhibition and recovery may balance the benefits of reduced ingestion of toxin with the disadvantage of a reduced food intake . In an environment with a patchy occurrence of toxic cyanobacteria feed ing inhibition would be adaptive if the environment could be sensed correctly (chemical cues) and animals are able to recover quickly from inhibiti on in the absence of toxic strains.

**Detoxication and oxidative stress**



# Another important process is detoxication of MC, which now has beendocumented in many aquatic biota, including several animals and macrophytes (Pflugmacher 2004). Metabolic breakdown of MC results in conjugate formation, amongst others with GSH. The formation of these conjugates is catalyzed by the enzyme GST, which has a microsomal and a cytosolic fraction. The activity of cG ST has been demonstrated to increase after exposure to MC in zebra fish (Wiegand et al. 1999) and brine shrimp (Beattie et al. 2003), although exceptions have also been described, where activity of the enzyme remained unchanged, e.g. in goldfish (Malbrouck et al. 2004) and carp (Li et al. 2003). GST activity (cGST and mGST) also

increased in Daphnia after exposure to CYN a nd an unidentified hepatotoxin (Nogueira et al. 2004). The MC–GSH conjugates have a much reduced toxicity and may be subject to enhanced excretion. Detoxication thus is a significant mechanism that protects biota to acute toxic effects of MC, as long as the capacity for detoxication is not exceeded. As a result of detoxication the cellular GSH pool is depleted and this exposes cells tooxidative stress through the formation of ROS like hydrogen peroxide. Organisms have a wide range of protec tive mechanisms to oxidative stress, including enzymes like SOD, CAT and GSH–reductase. The latter enzyme needs GSH as a co–substrate and its activity may be reduced when GSH is depleted through its conjugation with MC. Thus exposure to MC may have damaging effects in direct (inhibition of PP) and indirect ways (disbalance in ROS). Jos et al. (2005) showed that crushed cyanobacterial cells (MC released) resulted in enhanced oxidativ e stress resulting in lipid peroxidation, despite the fact that also levels of defensive enzymes were enhanced. Chapter 32: Cyanobacterial Toxins

# 69



# Studies by several authors in Table A.2 (i.e. Best et al. 2002) indicate that LPS (which are present on the cell–surface of cyanobacteria and onbacteria associated with cyanobacterial blooms) interfere with the detoxication process. In the study by Best and others GST activity in zebra fish

was reduced when MC and LPS were offered in combination. Since LPSfrom different bacterial sources are always present in the aquatic environment (although not all LPS from differen t sources equally disturbed detoxication when tested) it would be rewarding to study the process of detoxication in the field. Another study (Best et al. 2003) demonstrated that LPS stimulate drinking in fish, the increased volume of water in the gut potentially increases the opportunity for uptake of toxins (including MC) from the water and promotes osmoregulatory imbalance.

**Modulating factors**



# A common theme in much of what has been discussed in this paper is that of modulating factors. The effects that these cyanobacterial toxins have on aquatic biota are seldom straightforward but are modulated by factors in

the environment or the status of biot a themselves. Examples of modulating factors include condition of the animals, temperature and pre–acclimation /adaptation to cyanobacterial toxins. Hepatotoxicity of MC–LR has been shown to increase in fasted compared to fed animals (Malbrouck et al. 2004). Fasted goldfish showed a more severe and rapid inhibition of PP, and this may be related to differences in the glycogen content of the livers and the rate of MC removal from the body via the bilary excretion system.

The tolerance of Daphnia pulex to toxic Microcystis was shown to be temperature dependent (Hietala et al. 1997) and decreased with higher temperatures. Adaptation to toxic cyanobacteria may play an important role

too. Daphnia from locations where it is repeatedly exposed to toxic blooms would develop a higher tolerance to the toxins (Gustafson and Hansson 2004). Whether this is truly an adaptive evolutionary response remains to be tested since adaptation during 4–6 generations in the experiments by Gustafson and Hansson (2004) seem insufficient (although adaptation in Daphnia has indeed been shown to be a rapid process (e.g. Ebert et al. 2000). The essential message from their work is clear however: whenever the ecological effects of cyanobacterial toxins on biota are considered it is important to understand the history of the species involved. Modulating factors are an important reason why it is so hard to generalize the effects toxic cyanobacteria have on the biota in their environment. 4 B.W. Ibelings and K.E. Havens

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# Knowledge gaps

Throughout this paper remarks have be en made about knowledge gaps that limit our understanding of the ‘true’ ecological effects of cyanobacterialtoxins. At this point there is no need however to list those gaps extensively, since this is the subject of the paper by Havens et al, this issue. To summarize their main findings, Havens et al recommend further study on the following subjects:

• Studies at the whole community level in the presence vs. absence of cyanobacterial toxins;

• Studies that mitigate the bias toward s microcystin, especially MC–LR, i.e. more knowledge is needed about ecological effects of toxins like

CYN;

• Studies into synergistic effects of combinations of cyanobacterial toxins and of cyanobacterial toxins and other bioactive compounds from cyanobacterial cells;

• Studies in which biota are exposed to toxins under environmentally relevant conditions (synergistic effects with other stressors like temperature, low oxygen etc);

• More emphasis should be placed into (sub)chronic studies having sub– lethal effects, including those on be havior or genotoxicty; these may be

more relevant than acute lethal e ffects, and more knowledge is needed

here;

• What is the fate of toxins produced by cyanobacteria in the ecosystem?;

what is for instance the role of det oxication and covalent binding of MC

on transfer of toxins in the foodweb?

• Effects of toxins on benthic communities are not well understood;

• In which way and to what extend does toxicity of cyanobacteria

interfere with lake restoration?



# Conclusions

The qualitative meta–analysis identifies the following general patterns for major groups of aquatic biota (birds , fish, macroinvertebrates and zooplankton). Chapter 32: Cyanobacterial Toxins

# 6



# Birds. On basis of the limited number of studies on the role of toxic cyanobacteria on waterfowl we conclude that aquatic birds are at risk of cyanobacterial toxicosis. Anatoxins s eem to play a relatively large role, they are often present when dead bird s are found and the symptoms in diseased birds indicate a neurotoxin. Birds may be at high risk because they may directly feed on floating scum of cyanobacteria (personal observation)

and are warm blooded animals, like th e mammals which have been shown to be sensitive to cyanobacterial toxi ns in laboratory studies. A disease that

must be mentioned here is avian vac uolar myelinopathy (AVM), which is a neurologic disorder primarily affecting bald eagles (Haliaeetus leucocephalus ) and American coots ( Fulica americana ). The agent of this disease is an uncharacterized neurotoxin produced by a novel cyanobacterial epiphyte of the order Stigonematales (Wilde et al. 2005).

Fish . On basis of their study of common carp exposed to Microcystis , Li et al. (2005) conclude that fish kills during blooms of cyanobacteria can be assumed to result from extensive liver damage. Zambrano and Canelo (1996) on the other hand state that bloc kage of the gill activity could be the cause of mass mortalities during blooms of Microcystis . Both papers have in common that they put forward that cyanobacterial toxicosis can directly be responsible for the death of fish. We maintain that this is unlikely. Studying the collected data in Table 2 it seems doubtful whether naturally occurring concentrations of cyanobacter ial toxins (either dissolved in the water or contained in the cell) are suff iciently high to be directly lethal. Again, generalizations are difficult because there appear to be important differences between fish species. Fischer and Dietrich (2000) related thecapacity for uptake of toxins to the morphology of the GI. Perhaps combinations of stress factors that co–occur during blooms of toxic cyanobacteria (high temperature and pH, enhanced levels of ammonia, low oxygen in addition to cyanobacterial toxins) are more likely to cause fish mortality. Important sub–lethal effects of cyanobacterial toxins in fish are more than probable, however. Harmful effects have been seen on embryonic development, on growth of juvenile fish and on adult species. Several organs may be affected (e.g., kidney, heart, go nads), but the liver is the main target. Several studies have shown that around 50 % of fish caught from lakes or estuaries that support cyanobacteria l blooms show hepatic lesions that could – partially – be the result of exposure to cyanobacterial toxins.

Macroinvertebrates. Most of the studies on bivalves agree that these animals are quite resistant to different cyanobacterial toxins. This has been

shown for freshwater and marine muss els and clams and has been found for hepatotoxins, neurotoxins and CYN. More attention is given to the potential accumulation of toxins in mussels, and especially the risk of vecto96 B.W. Ibelings and K.E. Havens

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# rial transport to predators (incl uding man). However depuration studieshave shown that mussels clear toxins fairly rapidly, so that there is little retention. Nevertheless depuration is seldom complete, and low concentrations may even be carried through to the start of the next cyanobacterial growing season.

Zooplankton . Whenever the ecological significance of cyanobacterial toxins is discussed the primary suggestion is often that they deter grazing by zooplankton. Highly selective grazers like copepods would exert a stronger selection pressure than less selective grazers like Daphnhia, but the study by Kurmayer and Juttner (1999) shows that Daphnia may play a persisting role in the evolution of MC production (see also studies by Jang et al. 2004) who demonstrated that MC concentrations increased up to five–fold

when Microcystis was exposed to filtered zooplankton growth medium ( Daphnia and Moina spp). The literature concerning the effects of toxic cyanobacteria on zooplankton is exte nsive (Table A.2 only shows a selection) but there appear to be many contradictions. This is not surprising since there are numerous complicating factors. Furthermore it is now well

established that not all toxic effects can be traced back to the well known cyanobacterial toxins like MC. Although work by Rohrlack et al. (2005) has proven decisively that microcystins are toxic, the same work has shown that also the mutant incapable of producing MC has negative effectslike inhibition of feeding. Effects of cyanobacterial blooms also exert effects at the community level. Zo oplankton community composition may change towards dominance of smaller cladocerans which have a lowergrazing pressure. In this way cyanobacterial blooms may stabilize the turbid state on which some of the cyanobacteria like Planktothrix agardhii depend (Scheffer et al. 1997) and interfe re with lake restoration. The specific contribution of toxins – as opposed to general negative effects of cyanobacteria – at this high level of integration is unclear however.



# Acknowledgments

The authors wish to acknowledge Calvin Walker and Mike Coveney for their comments on the paper. Chapter 32: Cyanobacterial Toxins

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# 70 6 B.W. Ibelings and K.E. HavensEider (Somateria mollissima)NODLN Feeding on musselsin Baltic sea(i) 0.003-0.18(ii) 0.1-5.8μg per liverELISA andLC-MSSipiä et al, 2004FISHRound goby (Neogobiusmelanostomus)MCNatural routes inlakeGoby liver: 0.6-3.0 91-820 Babcock-Jackson et al,2002Sticklebacks(Gasterosteus aculeatus)NODLN Feeding on copepods pre-exposed tocyanobacteria(i) 0.15 (ELISA)(ii) 0.8 (PPase)ELISA andPPase assayEnsgtröm-Öst et al,2002(i) perch (Percia fluviatilis)(ii) ruffe (Gymnocephalus cernuus)(iii) smelt (Osmerus eperlanus)MCNatural routes inlake foodweb(i) perch liver: 17-51(ii) ruffe liver: 9-194(iii) smelt liver: 59-8747-3912 HPLC Apparent biomagnification based upon skewedrepresentationIbelings et al, 2005(i) silver carp (Hypophthalmichthysmolitrix),(ii) Carassius gibelioMC-LR; MC-RR Feeding on Microcystis cells(i) homogenized tis-sues:< 5 (H. molitrix)< 0.8 (C. gibelio)(ii) faeces:11 - 46 (H. molitrix)21 (C. gibelio)MC-RR levels in tissues>>MC-LRJang et al, 2004Sea trout (Salmo trutta)NODLN Gavage with singledose of Nodularia(i) liver: 0.019-1.2;max 1.6(ii) muscle: max. 0.125ELISA andHPLCSingle oral dose: loss liverarchitecture 1-2d, partialrecovery 4-8 d, completeafter 8 d: damage reversibleKankaapää et al, 2002Flounder (Platichthysflesus)NODLN, MC Feeding on musselsin Baltic sea(i) 0.02-0.1 TEH prebloom exposure(ii) 0.02-2.23 TEH(ELISA)(iii) nd-0.47 (LC-MS)ELISA andLC-MS(i) NODLN variable between individuals: peakconc. in sub-populations(ii) 50 % livers smallscale necrosisKankaapää et al, 2005a 

# 707 Chapter 32: Cyanobacterial ToxinsFlounder (Platichthysflesus)NODLNNatural routes in seaLiver: 0.82-6.37 LC-MSNo biotransformationproducts (e.g. glutathioneadduct) foundKarlsson et al, 2003(i) northern pike (Esoxlucieus(ii) white sucker (Catostomus comersonii)MC Feeding on MCcontaining prey(i.a. gastropods)Liver: not detected 1.2–6.1μg L-1HPLC Kotak et al, 1996aCarp (Cyprinus carpio) MC Feeding on Microcystis scum in tanks(i) hepatopancreas: 2.6(ii) muscle: 0.4Li et al, 2004Tilapia rendalli MCNatural routes inlake(i) viscera: 0 – 67.8(ii) liver: 0 – 31.1(iii) muscle: 0.003 –0.026max. 980μg L-1HPLC,ELISAMC found in 75 % fishsamplesMagalhães et al, 2001Fish (unspecified) MCNatural routes inlakeMuscle: 0.01 – 0.4 0.12–0.78μg L-1ELISA Magalhães et al, 2003Goldfish (Carassius auratus)MC-LR IP injection Liver: 0.5 - 3PPase Time course accumulation(first 48 h), depuration(48-96 h); patterns not affected by fasteningMalbrouck et al, 2003,2004Tilpapia (Oreochromisniloticus)MC On fish-farm to Microcystis bloom(i) gut: 1.8 – 8.3(ii) liver: 4.1 – 5.3(iii) kidney: 3.6 - 4(iv) muscle: 0.4 – 1.01120 ELISA At times MC in liver > gutMohamedet al, 2003Flounder (Platichthysflesus)NODLN and MCNatural routes inBaltic seaMax 0.4 (TEH) 150-8700 THE;< 2400 NODLNELISA andMALDITOF-MSSipiä et al, 2001, 2002Tilapia rendalli MC (i) Feeding on Microcystis cells + fishfood; one exp cellsdisrupted prior feeding(i) liver: 2.8(ii) muscle: 0.08(iii) faeces: 0.07(all max values)14.6 ELISA (i) 15 or 42 d exposure(ii) accumulation MC lessin presence alternativefood(iii) depuration phase:only small percentage MCremovedSoares et al, 2004 

# 7 08 B.W. Ibelings and K.E. HavensSilver carp (Hypophthalmichthysmolitrix)MC-LR and RR Feeding on Microcystis bloom intanks(i) faeces: 44.5;MC-LR:RR = 0.75(ii) intestines: 49 – 115;MC-LR:RR = 0.17(iii) blood (MC-RR):0.4 – 50(iv) liver (MC-RR): 8 -18(v) muscle (MC-RR):0.5 – 1.4MC: 286–866;MC-LR:RR =0.57PPase (i) in the various tissuesand organs always MCRR found, rarely MC-LR(ii) active degradation ofMC-LR during digestion(?)Xie et al, 2004Various fish species:phytoplanktivorous(Hypophthalalmichthysmolitrix), herbivorous(Parabramis pekinensis), omnivorous(Carassius auratus), carnivorous (Culter ilishaeformis)MC-LR and RRNatural routes inlake(i) intestines: 22 (26 %LR)(ii) blood:14.5 (45 %LR)(iii) liver: 7.8(iv) bile 6.3 (48 % LR)(v) kidney: 5.8 (30 %LR)(vi) muscle: 1.8 (18 %LR)240 HPLC (i) MC in carnivorous >omnivorous > phytoplanktivorous fish(ii) fish at top of foodwebmost at risk; general tendency MC to accumulateup the foodchainXie et al, 2005Salmon (Salmo salar) MC-LR IP injection (i) 2.6 – 263 (MeOH)(ii) 138 – 1181 (Lemieux)(i) PPase afterMeoH extract. (ii) Lemieux (GCMS)Covalently bound MCmade up ~ 74 % of totalWilliamset al, 1997aMACROINVERTEBRATESMytilus galloprovincialisMC Grazing on Micro-cystis cells(i) mussel: 10.7 duringaccumulation rising tomax of 16 on day 2 depuration(ii) faeces: 1403.4μg / 107cellsELISA (i) no mussel mortality(ii) during depuration initial increase, followed decrease MCAmorim & Vasconcelos(1999)Zebra mussels(Dreissena polymorpha)MCNatural routes inlake0.2 91-820 Babcock-Jackson et al,2002 

# 709 Chapter 32: Cyanobacterial Toxins(i) freshwater shrimps(Palemon modestus;Macrobrachium nipponensis)(ii) red swamp crayfish(Procambarus clarkii)MC-LR and RRNatural routes inlake(i) P.modestus -stomach: 4.53hepatopancreas: 4.29gonads: 1.17eggs: 2.34muscle: 0.13gills: 0.51(ii) M. nipponensis -stomach: 2.92hepatopancreas: 0.53gonads: 0.48eggs: 0.27muscle: 0.04gills: 0.05(iii) P. clarkia -stomach: 9.97hepatopancreas: 0.08gonads: 0.93muscle: 0.05gills: 0.27LC-MS (i) Proportion of MC-LRof total MC varied withtype of tissue and the species; ratio of MCLR:MC:P. modestus decreasedgonad (94 %) > stomach(61 %) > eggs (56 %) >HP (30 %) > muscle (5 %)> gills (0 %)M. nipponensis: stomach(71 %) > muscle (48 %) >HP (39 %) > eggs (39%)> gonads (33 %) > gills(0 %)P. clarkia: muscle (100 %)> intestine (68 %) > stomach (58 %) > gonad (57%) > gills(52 %)(ii) considerable part toxinburden in eggs (29 % in P.modestus), i.e.MC transferred to offspringChen & Xie, 2005Freshwater snail(Bellamya aeruginosa)MC-LR and RRNatural routes inlake(i) digestive track: 0.8-4.54 –MC LR:RR=0.44(ii) HP: 1.06-7.42 -LR:RR = 0.63(iii)) gonad: 0-2.62 -LR:RR=0.96(iv) foot: 0.01240 HPLC MC hepatopancreas > digestive track: selectivebioaccumulation?Chen & Xie, 2005 

# 7 10 B.W. Ibelings and K.E. HavensZebra mussels(Dreissena polymorpha)MC-LR Feeding on toxicMicrocystis cells(i) 11 – feeding solelyMicrocystis(ii) 3.9 feeding on mixture Microcystis andgreen algae3.1 LC-MS;MMPB(i) maximum share covalently bound MC 38 % total(ii) only 0.5 % offeredMC found in mussels(iii) rapid depuration, after3 wk nearly completeDionisio Pires et al,2004Mysid shrimp (Mysisrelicta)NODLN Feeding on copepods pre-exposed tocyanobacteria(i) 0.74 (ELISA)(ii) 0.52 (PPase)ELISA andPPase(i) time dependent accumulation in shrimps, notin fish (results ELISA)(ii) copepods as vectors tohigher trophic levelsEnsgtröm-Öst et al,2002Zebra mussels(Dreissena polymorpha)MCNatural routes inlake foodweb1-30 7-3912 HPLC Ibelings et al, 2005Black tiger prawns(Penaeus monodon)MC, NODLN (i) natural routes inponds(ii) oral uptakeNODLN via food inexperiments(iii) injection MCLR(i) ponds -HP (TEH) 0.006-0.08(ii) experiment(NODLN) -brain and heart: 0.36HP: 0.25 (0.83 peaklevel)gut: 0.1gills: 0.014muscle: 0.01(iii) experiment (MC) -HP: 0.130 (peak)(i) ponds:TEH: 1.2(ii) exp: 10μgkg bw-1ELISA;HPLCRapid depuration fromprawnsKankaanpää et al,2005bBaltic Sea zooplankton:Acartia tonsa,Eurytemora affinis,Strombium sulcatumNODLN Exposure to dissolved NODLN(3Hdihydronodularin)A. tonsa: 0.37μg g-1CE.affinis: 0.60S. aulcatum: 1.555μg L-1(i) minimum BCF 12 – 18for copepods;(ii) max BCF ciliate 22(iii) possible vectorialtransport with significantsublethal effectsKarjalainen et al, 2003 

# 711 Chapter 32: Cyanobacterial Toxins(i) pike larvae (Esoxlucius)(ii) mysid shrimps(Neomysis integer)NODLN Fed with zooplankton pre-exposed to(i) Nodularia extract(ii) purifiedNODLN(20μg L-1)(i) pike larvae (12h):0.47(ii) Neomysis: (12h):0.31Only 0.12 (pike) and 0.03% (shrimps) of ingestedtoxin was detected inanimalsKarjalainen et al, 2005Gastropods (Lymneastagnalis, Helisomatrivolis, Physa gyrina)MC Grazing on (settled)lake phytoplankton11 – 121 1.2 – 6.1μg L-1~1220μg g-1DWHPLC MC in seston not expressed on DW basisKotak et al, 1996aBaltic clam (Macomabalthica)NODLN Exposure to dissolved toxin andNodularia cells intanks0.16 - 16.6 (24 h) –30.3 (96 h) (TEH); < 5% of this NODLN10-50μg L-1(cells)4-20μg L-1(dissolved)ELISA,HPLC,MALDITOF-MSNo NODLN-GSH conjugates detectedLehtonen et al, 2003Signal crayfish (Pacifastacus leniusculus)MC Exposure to toxicand non toxicPlanktothrixagardhiiMC present but notquantified3610 (i) crayfish ingested 430μg MC – no effects(ii) possible vectorialtransport to fish, birds,minkLiras et al, 1998Crab (unspecified) MCNatural routes inlakeMuscle: 0.02 – 1.0 0.12 – 0.78μgL-1ELISA Magalhães et al, 2003Freswhater mussel(Alathyria condola)PST feeding on toxicAnabaena5.7 1580Negri & Jones, 1995Pearl oyster (Pinctadamaxima)SAX (i) natural routes insea;(ii) exposure juvenile oyster toTrichodesmium0.73 in viscera diseasedoysterNot detectable HPLC; mouseassay; sodiumchannel andsaxiphilinbinding assayno mortality juvenile oysters in experimentNegri et al, 2004 

# 7 12 B.W. Ibelings and K.E. Havens(i) freshwater snail(Sinotaia histrica)(ii) freshwater clam(Corbicula sandai)MC-LR and RR (i) natural routes inlake(ii) feeding on toxicMicrocystis cells inexp.(i) S. histrica (lake)intestine 2.7 – 19.5(MC-LR + RR)HP 0 – 3.2(MC-RR only)(ii) C. sinai: nd(iii) S. histrica (exp.)HP max. 43651.8 – 284(lake)20.1μg L-1(exp.)HPLC (i) lag phase in depurationfrom snail tissue; biological half life 8.4 d(ii) MC in lake snail stillpresent next springOzawa et al, 2003Freshwater mussel(Anodonta cygnea)PST Feeding on toxicAphanizomenon -accumulation 14d,depuration 14d0.26 1.9-2.6HPLC (i) Anodonta exposed to1.4e9cells L-1d-1, removed 65 % these; clearance rate negatively related to PST content(ii) slow-fast-slow depuration; s-shaped kinetics(8.2 % d-1)Pereira et al, 2004Freshwater clam (Anodonta grandis)(i) exposure dis-solved MC(ii) natural routes inlakeViscera 0.59(i) gills 0.31(ii) muscle 0.36MC 51-55μgL-1dissolved upto 8.3μg L-1inlake(i) toxin burden evenlydistributed over threebody parts(ii) rapid depuration first6d (~70% gone), stablefor 15d afterwards(iii) suggestion of bioconcentrationPrepas et al, 1997 

# 713 Chapter 32: Cyanobacterial ToxinsFreshwater mussel(Anodonta cygnea)CYN Feeding on Cylindrospermopsis culture(i) haemolymph 61.5(= 408μg L-1)(ii) viscera: 5.9(iii) mantle: 0.13(iv) foot + gonads: 0.75(v) whole body extract:2.9(all maximum conc. after 10-16 d accumulation)14-90μg L-1HPLC (i) no adverse effects onmussels despite bioaccumulation CYN inhaemolymph to conc.higher than in water(ii) bi-phasic depuration,increase in CYN contentfrom day 22-28Saker et al, 2004Blue mussel (Mytilusedulis)NODLNNatural routes in theseaMax 2.15 Max 2400 Sipiä et al, 2001(i) balthic clam(Macoma balthica)(ii) blue mussel (Mytilus edulis)NODLN and MCNatural routes in thesea(i) mussels 1.5 (max ofTEH)(ii) clams 0.1-0.13150-8700 ELISA andMALDITOF-MS(i) NODLN-GSH conjugates confirmed with MS(ii) mussels MC 30 foldincrease summer(iii) time lagged accumulation deep sitesSipiä et al, 2002Blue mussels (Mytilusedulis)NODLN (i) grazing on Nodularia(ii) coprophagy (exposure to feces ofmussels grazing onNodularia)(i) Mussels preexposure to Nodularia:0.05-0.1(ii) Mussels grazing onNodularia -digestive track 245body 80gills 2(iiii) feces when feeding on Nodularia 95(iv) Body after coprophagy0.065(v) Feces coprophagy 1(vi) PF from (ii) 714Nodularia culture 16μg L-1LC-MS (i) high NODLN in PF in-dicative of selective feeding; (ii) cells Nodularia inPF may survive butgrowth inhibitedSvensen et al, 2005 

# 7 14 B.W. Ibelings and K.E. HavensMytilus galloprovincialisMC-LR Grazing on Micro-cystis cells10.5 (of which 96 % indigestive gland + stomach)28μg 108cells-1Depuration bi-phasic andfairly rapid (13d)Vasconcelos 1995Crayfish (Procamabarusclarkia)MC Feeding on toxic(and non toxic) Microcystis cells2.9 (max after 11 d),of which 53 % in intestine, 38 % HP and 9 %rest body2300 ELISA Juvenile crayfish enhanced mortality on nontoxic Microcystis: compounds other than MCmore relevant to invertebrates (?)Vasconcelos et al, 2001Dungeness crab larvae(Cancer magister)MC-LRNatural routes infoodweb(i) 0.006 (after MeOHextraction)(ii) 84.4 (after Lemieux)(i) PPase afterMeoH extract. (ii) Lemieux (GCMS)10.000 fold greater MCconcentration usingMMPB after Lemieux extractionWilliams et al, 1997aBlue mussel (Mytilusedulis)MC-LR (i) feeding on toxicMicrocystis cells(ii) natural routes infoodweb(i) 3369 decreasing to113 after 4 d depuration (Lemieux)(ii) 2 - dropping to0.14 over 53 d depuration (MeOH)(i) PPase afterMeoH extract. (ii) Lemieux (GCMS)Less than 0.1 % of MCextractable with MeOHWilliams et al, 1997bFreshwater snail(Bellamya aeruginosa)MC-LR, RR (i) HP 1.06-7.42(ii) digestive track 0.8-4.54(iii) gonad 0-2.62(iv) foot 0-0.06 (MCLR only)HPLC,LC-MS(i) bioaccumulation in HP(ii) ratio LP:RR increasedfrom digestive track to HPto gonad (iii) suggestionLR more resistant to degradation?Xie et al, 2005Freshwater mussel(Unio douglasiae)MC-LR Exposure to Microcystis cellsHP:(i) at 15 oC – 130(ii) at 25oC - 25027 - 50μg L-1(i) accumulation temperature dependent(ii) depuration relativelyfast; but slower at 15 than25oC; halted in winter(iii) no adverse effects onmusselsYokoyama & Park(2003) 

# 715 Chapter 32: Cyanobacterial ToxinsPulmonate snails(Lymnea stagnalis,Helisoma trivolis, Physagyrina)MCNatural exposure inlakes of varyingtrophic statusUp to 144 0 – 1526 HPLC Possible uptake routes viafood and directly fromwaterZurawell et al, 1999ZOOPLANKTONEchinogammarusischnusMCNatural routes inlake2 91-820 Exposure of this detrivorevia pseudofaeces Dreissena?Babcock-Jackson et al,2002Community of differentspp, i.a. Daphnia galeataMCNatural routes inlake0-1352 HPLC Ibelings et al, 2005(i) calenoid copepods(Eurytemora affinis,Acartia tonsa)(ii) ciliate (Strombidiumsulcatum)NODLN Uptake of dissolvedNODLN from waterduring 15 min - 6 d(i) E.affinis 1.14(ii) A. tonsa 0.66(iii) S. sulcatum 4.8(maxima)5μg L-1(i) BCF copepods 12-18(ii) ciliate 22(iii) vectorial transport toshrimps and planktiv. fishKarjalainen et al, 2005Community of differentspp., a/o Daphnia pulexMC Grazing on lakephytoplanktonUp to 67 1.2 – 6.1μg L-1HPLC Kotak et al, 1996(i) Thamnocephalus,platyurus(ii) Eudiaptomus gracili,(iii) Daphnia hyaline(iv) Cyclops abyssorumMC (i) grazing on mixtures of Cryptomonas and Planktothrix(ii) Planktothrix at0.05 or 0.1 mg C L-1- MC(+) or MC(-)4.7 ng MC-LRperμg C-1HPLC Kurmayer & Jüttner,1999 

# 71 6 B.W. Ibelings and K.E. HavensCalanoid copepod (Eurytemora affinis)NODLN (i) feeding on Nodularia and non-toxicflagellates(ii) natural seston(i) fed with Nodularia:0.032 ng copepod-1(background) rising to0.007 (ELISA)(ii) 0.0095 to 0.101(PPase)(iii) feacal pellets -0.0067ng pellet-1(ELISA)(iv) idem 0.0050(PPase)ELISA andPPase(i) lower conc. when fedwith natural seston(dominated by non-toxAphanizomenon)(ii) ELISA and PPase different but not significant(iii) no acute effects oncopepods despite accumulation(iv) vectorial transport tofoodweb(v) toxin content feacalpellets low: no vector tocoprophagous animalsLehtiniemi et al, 2002Daphnia magna PST (i) grazing onAphanizomenon(1.2e6cells mL-1)(ii) lyophilized material (1 mg mL-1)(i) exposure Apha:0.065-0.378 pmol PSTanimal-1(ii) exposure lyophilized material: 0.007(i) cells: 643-1170 pmol mL-1(ii) lyoph.: 2745pmol mL-1HPLC-FLD Bioaccumulation factor >1 after 12h feedingNogueira et al, 2004Daphnia magna CYN Grazing on Cylindrospermopsisstrains (+/- CYN)(i) 0.025 ng animal-1(24 h)(ii) 0.020 ng animal-1(48 h)(i) intracellulartoxin 219.6 –236.4 ng mL-1(ii) extracellulartoxin 12.4 –47.4 ng mL-1(increase intime)(iii) total CYN234.2 – 278.4ng L-1HPLC-MS Bioaccumulation factorssmaller than unity: 0.71(24 h) and 0.46 (48 h);levels not high enough toindicate bioaccumulationNogueira et al, 2004(b) 

# 717 Chapter 32: Cyanobacterial ToxinsDaphnia magna MC (i) grazing on Microcystis alone ormixed withScenedesmus (3.2mg C L-1)(ii) exposure to lakewater enriched withMicrocystis fromenclosures (8.9 mgC L-1)Daphnia 0.2-24.5μg g-1DW (highest when exposed solely to Microcystis)(i) Microcystis:2000μg g DW-1(ii) MC variedbetween 5 – 156μg L-1in enclosuresELISA (i) calculation shows MCreally accumulated inbody, not just gut content(ii) calculation showstransport to roach resultsonly in sublethal effectsThostrup & Christoffersen, 1999 

# 718 B.W. Ibelings and K.E. HavensTable A.2.Effects of various cyanobacterialtoxins on aquatic biota. Where possible dose and effect are shown.ExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceBIRDSFISHZebra fish(Danio rerio)MC–LR Purified MC, dissolved inwater(i) 0.5, 5μg L-1: increased motility(ii) 15, 50μg L-1: decreasedmotility(iii) 50μg L-1:reduced spawning(i) effects on behaviour dose dependent (reversalat higher conc.)(ii) LOEC < 0.5μg L-1Baganzet al, 1998Brown trout(Salmo trutta)MC i) purified MCii) aqueous extracts of Microcystis5, 50, 500μg L-1MC: cardiovascular effectsPurified MC no effect on (some)cardiac responses, in contrast tocell extracts – always effects;cardiac output only affected byMC at higher dosesBestet al, 2001Zebra fish (Daniorerio)MC–LR;LPSIn vivo exposure fish embryos to dissolved toxins;LPS different bacterial origins (including cyanobacterial)LPS (up to 29.8 106EU mg-1DW) + MC (up to 12.6 ng mgDW-1): reduced mGST +sGSTLPS from axenic bacteria or lakeblooms reduced GST activity, interfered detoxicationBest et al, 2002Rainbow trout (Oncorhynchus mykiss)MC Aqueous suspension wholeor broken Microcystis cells100μg L-1MC–LR: osmo-regulatory imbalance, increased liver massInteraction MC with LPS Best et al, 2003 

# 719 Chapter 32: Cyanobacterial ToxinsExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceBrown trout(Salmo trutta)MC–LR (i) purified MC(ii) lysed Microcystis cellsi) lysed tox Microcystis, containing 41 – 68μg L-1MC:ionic imbalance, reducedgrowthii) purified toxin: effects ongrowth (less than toxic lysate)iii) lysed non–tox Microcystis:effects on growth (less thantoxic lysate)Ammonia similar effects to MC Bury et al, 1995Carp (Cyprinus carpio)MC Fish exposed to toxins vianatural routes in the lake,including Microcystis scumsUp to 4000μg g-1DW: hepatic lesionsLesions in > 50 % fish examined Carbis et al, 1997Carp (Cyprinus carpio)MC–LR Gavage freeze dried Micro-cystis cells (single sublethalbolus)400μg kg-1bw: pathologicalchanges in hepatopancreas,kidney, gastrointestinal tract(i.a. apoptosis)In carp – cyprinid – compared tosalmonids liver pathology develops faster and at lower MC concentrationsFischer & Dietrich,2000Rainbow trout (Oncorhynchus mykiss)MC Gavage freeze dried Micro-cystis cells5700μg kg-1bw: liver necro-sis, hepatocyte apoptosisUnbound MC results in fast ef-fects, including PP inhibition andliver necrosis; covalently bindingto MC results slower effects, i.a.apoptosisFischer et al, 2000(i) Hypophthalmichthys molitrix(ii) Carassius gibelioMC–LR;MC–RRFeeding on cultured Microcystis cellsunspecified dose: reducedgrowthJang et al, 2004 

# 7 20 B.W. Ibelings and K.E. HavensExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceTilapia (Oreochromis sp)MC–LR Feeding on lyophilized Mi-crocystis bloom (crushedand non–crushed cells)60μg fish-1day-1MC–LR:oxidative stress in liver, kidney and gills(i) crushed cells more effectnon–crushed on LPO and levelsantioxidant enzymes(ii) time dependent developmentoxidative stress: not yet presentafter 14d; 21 dlipid peroxidation+ enhanced antioxidant enzymeactivitiesJos et al, 2005Sea trout (Salmotrutta)NODLN Gavage single dose Nodularia210–620μg kg-1bw: rapid (1–2 d) but reversible (4–8 d)liver damage(i) rapid detoxication, cross reaction conjugates in ELISA(ii) no effects on swimmingKankaanpää et al,2002Rainbow trout (Oncorhynchus mykiss)MC–LR IP injection of purified MC (i) 1000μg kg-1: 100 % mor-tality(ii) 400μg kg-1: no mortality;increased ratio liver to bodymass, liver necrosis; kidneylesionsFish less sensitive than mice:LD50= 400 – 1000μg kg-1Kotak et al, 1996bCarp (Cyprinus carpio)MC–LR In vitro exposure hepato-cytes to 10μgL-1MC–LRIncrease ROS, depletion GSH,increase SOD, CAT and GS–Px activity, GST unchanged:oxidative shock by exposureMCAntioxoidant enxymes did notprevent oxidative shock: apoptosis and necrosis hepatocytesLi et al, 2003Carp (Cyprinus carpio)MC Feeding on Microcystisscum in tanks50μg kg-1bw for 28 days: re-duced growth, increased liverenzyme activity, damagedhepatocytes,Long term subchronic effects Li et al, 2004 

# 721 Chapter 32: Cyanobacterial ToxinsExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceLoach (Misgurunsmizolepis)MC–LR Dissolved MC–LR MC–LR at 1, 3, 10, 100, 1000μg L-1:(i) mortalilty, delayed hatching, liver (necrosis) and cardiotoxicity(ii) (late stage) embryos andlarvae more sensitive than juvenile fish: LC50larvae = 164;juveniles = 593μg L-1(i) embryonic development loachaffected by MC in contrast to zebrafish(ii) results dose dependent, i.e.increasing effects with increasingconc. MC. No mortality at thelower doses.Liu et al, 2002Goldfish (Carassiusauratus)MC–LR IP injection 125μg kg-1bw: no changeionic homeostasis, liver lesions, changes liver enzymeactivityLiver damage reversible Malbrouck et al,2003Goldfish (Carassiusauratus)MC–LR IP injection of fed and fastedjuvenile goldfish(i) 125μg kg-1bw: inhibitionPPase, complete after 6 h infasted, less inhibition in fedfish(ii) recovery after 96 h(iii) GSH levels and GST unaffectedFeeding status has effect on toxicity MC, possibly by acting onbile formation and secretionMalbrouck et al,2004a(i) silver carp (Hypophthalmichthysmolitrix)(ii) carp (Cyprinuscarpio)MC–LR (i) ip injection purified MC(ii) cyanobacterial biomassper os / anus(i) 400μg kg-1bw (purifiedMC): no – minor effects(ii) 3 – 1,200μg kg-1bw(biomass): changes blood indices and immunologicalchanges(i) effects crude biomass muchstronger purified toxin(ii) oxidative stressPalikovaet al, 1998 

# 7 22 B.W. Ibelings and K.E. HavensExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceSeven fish (cyprinids) and three amphibian speciesMC–LR, RRand YR,SAX,ANA(a)(i) in vivo emersion(ii) exposure to dissolvedpurified toxin(iii) exposurecrude cyanobacterial cell extracts(i) MC 0 – 50μg L-1: no acuteeffects on embryonic development(ii) MC–RR > 0.5μg L-1; YR> 5μg L-1; LR > 50μg L-1:timing hatching affected(ii) morphological effects onlyat highest conc. MC–LR 10mg L-1(iii) SAX > 10μg L-1delayedhatching malformations at 500μg L-1(iv) ANA(a) 400μg L-1hearthrate affected, no chronic effects(v) far more pronounced effects with crude extracts: malformations and mortalityCrude cell extracts of severalcyanobacteria gave severe effects(more so than purified toxin):cannot be attributed MC aloneOberemm et al,1999 

# 723 Chapter 32: Cyanobacterial ToxinsExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceRainbow trout(Oncorhynchusmykiss)MC–LR (i) IP injection or oral dosage purified MC(ii) IP injection or oral dosage freeze dried Microcystiscells(i) 550μg kg-1bw MC (ip):severe liver damage, death(ii) 550μg kg-1bw Microcystis (ip): severe liver damage,death(iii) 1200μg kg-1bw MC(oral): no effects(iv)1700μg kg-1bw Microcystis (oral): no effects(v) 6600μg kg-1bw Microcystis (oral): severe liver damage,death(vi) 550μg kg-1bw Microcystis (8 times oral dosage):modest – severe liver damageTreatments included repeated(oral) exposure to MC, i.e. closeto natural exposure but during‘limited’ period of time (8 timesat 12 h intervals)Tencallaet al, 1994Zebra fish (Daniorerio) embryosMC–LR Purified dissolved MC (i) > 0.1μg L-1: increasedsGST and GPx activity(ii) > 2.0μg L-1: effects ongrowth and survivalDose dependent relationshipMC~GSTWiegand et al, 1999Carp (Cyprinus carpio)MC–LR Purified dissolved MCinhibition ATP–ase activity ofNa+K+pump in gills, disruption ion homeostasisHarmful effects of dissolved MC(without ingestion of cells)Zambrano & Ca-nelo, 1996MACRO–INVERTEBRATESBrine shrimp (Artemia salina)MC–LR,MCHtyR andNODLNPurified dissolved toxin 0.5μg L-1: elevation GST activityBeattie et al, 2003 

# 7 24 B.W. Ibelings and K.E. HavensExposedorganismToxin Exposure route Concentration (dose)and effectRemark Reference(i) pike larvae (Esoxlucius)(ii) mysid shrimps(Neomysis integer)NODLN (i) purified NODLN (20μgL-1)(ii) crude extract Nodularia(i) crude extract:pike larvae: decreased ingestion and faeces productionratesshrimps: no effects on molting, faeces production, C:N orgrowth(ii) purified NODLN:no effectsPurified NODLN no effects onpike larvae; crude cell extractsstronger effectsKarjalainen et al,2005Baltic clam(Macoma balthica)NODLN (i) exposure to dissolvedtoxin(ii) exposure to toxic andnon toxic Nodularia cells intanks4–20μg NODLN per day:(i) conc. dependent neurotoxiceffects (increase / decreaseAChE activity when exposedto low and high NODLN respectively)(ii) some treatments low siphon activityAbundant unidentified compound with NODLN like spectralcharacteristics (found in bothtoxic and non–toxic Nodulariatreatments)Lehtonen et al, 2003Brine shrimp (Artemia salina)CYN andMC(i) purified dissolved CYNand MC(ii) extracts Cylindrospermposis, Microcystis(i) CYN LC50decreased from4.48 to 0.71μg mL-1between24 and 72 h(ii) likewise MC LC50from4.58 to 0.85μg mL-1(i) dose and time dependent mortality(ii) LC50cell extracts typically >than purified CYN (reducedbioavailabilty?)Metcalf et al, 2002Estuarine crab(Chasmagnathusgranulatus)MC Cell extracts Microcystis Injected daily for 4–7 d with17.6 ng MC: (i) increased enzyme activity (GST, CAT) inhepatopancreas(ii) no change LPO – no oxidative damage (?)(iii) yet histological damagePinho et al, 2003 

# 725 Chapter 32: Cyanobacterial ToxinsExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceEstuarine crab(Chasmagnathusgranulatus)MC Crabs injected twice withcell extracts Microcystis(i) decreased Na+, K+–ATPaseactivity anterior gills(ii) increased GST posteriorgills(iii) increased TOSCIncreased TOSC in response MCas protective mechanism againstLPO (level unchanged by exposure to MC)Vinagre et al, 2003ZOOPLANKTONMoina macrocopa Unidentifiedmetabolites(possiblycyanopeptolinsA–D)Freeze dried Microcystis Inhibitionof proteases Agrawal et al, 2005Thamnocephalusplatyurus[D–Asp3,(E)–Dhb7]MC–RR,MC–LR,MC–YR,MC–RR,NODLNPurified toxins LC50NODLN < [D–Asp3,(E)–Dhb7]MC–RR < otherMCLC50insufficient to study response of organisms to exposure;LC10+ LC90required to get slopeBlom et al, 2001 

# 72 6 B.W. Ibelings and K.E. HavensExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceDaphnia pulicaria,D. Pulex, D. hyaline,Diaptomus birgeiMC–LR,NODLN,ANA(a)Exposure to:(i) purified toxins(ii) cell extracts(iii) toxic or non–toxiccyanobacterial strains(i) species specific responsesto toxin exposure; not alldaphnids equally sensitive interms feeding inhibition(ii) LC50for MC after 24 h(varied from > 50μg L-1in D.pulicaria to < 1.0 in D.birgei);for NODLN > 20 and < 0.6respectively(iii) LC50decreased withlonger exposure(i) feeding inhibition protectsagainst toxic effects: less sensitive Daphnia stronger inhibiton(ii) zooplankters insensitive dissolved MC, more so dissolvedANA(a)DeMott et al, 1991Daphnia spp MC Feeding on mixtures Microcystis and Scenedesmus; 0,50 or 80 % Microcystis intotal food conc. of 0.5 mg CL-1(i) rapid feeding inhibition(but recovery after continuedexposure to same mixture)(ii) reduced growth and reproduction(iii) reduction ingrowth/ingestion: direct toxiceffects and feeding inhibitionClear differences between Daphnia sppDeMott, 1999Temperate andtropical cladocerans(Ceriodaphniacornuta, Daphniapulex; D. pulicaria,D similes, Moinamicrura,Moinodaphniamacleayi)MC (i) grazing on toxic Microcystis strains mixed withAnkistrodesmus, total conc.1.0 mg C L-1;(ii) acute and chronic exposureMC contents 2810–4080μg g1DW:(i) decreased survival in presence tox Microcystis(ii) toxic Microcystis inhibitedfeeding rate, even when just 5% in mixture with greens(ii) non–toxic cyanobacteriumas sole food:poor growth(i) species from low productivitysites – adapted to starvation –showed lowest sensitivity totoxic Microcystis(ii) small and large bodied fastgrowing spp prone to starvationand most sensitive to MC; smallbodied slow growing spp mostresistant and least sensitiveFerrão–Filho et al,2000 

# 727 Chapter 32: Cyanobacterial ToxinsExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceTropical cladocerans(Moina micrura,Ceriodaphniacornuta)MC Grazing on:(i) large Microcystis colonies(ii) single cells and (small)colonies MicrocystisMC 2.0 – 16.0μg L-1:(i) toxic Microcystis inhibitedgrowth reduced reproductioncladocerans; partly throughfeeding inhibition(ii) effects unicellular lab cultures stronger than colonialcultures or natural seston, especially with large colonies(although very toxic 3.9 mgMC g DW-1)Colony forming cyanobacteria:little effect despite high toxicity;possible explanation why fieldstudies fail to demonstrate effects of toxic bloomsFerrão–Filho &Azevedo 2003Daphnia magna MC Grazing on mixtures of Microcystis and Scenedesmus;max of 140,000 Microcystiscells mL-1(1.0 mg C L-1)(i) reduced growth, reproduction and survival, increasingeffects with increased proportion Microcystis cells(0; 50 or 100 %)0(ii) pre–exposure (acclimation) reduced harmful effects:development of toleranceGustafsson & Hansson, 2004Daphnia pulex MC–LR (i) acute exposure to Microcystis cells0–2.43 mg C L-1= 0–360,000 cells mL-1(ii) chronic exposure 30,000cells mL-1MC–LR 7.6 10-5ng cell-1:(i) variation in acute tolerance(EC50) to toxic Microcystis(ii) increase temperature: decrease EC50(iii) chronic exposure, reducedsurvival and reproduction,clonal differences reversedcompared to acute exposureSuggestion made that more resistant clones show stronger feeding inhibitionHietala et al, 1997 

# 7 28 B.W. Ibelings and K.E. HavensExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceCalanoid copepods(Acartia bifilosa,Eurytemora affinis)NODLN Cultures of Nodularia addedto enclosuresNODLN ~ 11μg mg DW-1(3–4μg mL-1):Baltic copepds feed, ingest,reproduce and survive in presence toxic Nodularia, no negative effectsPerhaps hatching success moresensitive to toxins (not measured)Koski et al, 2002Eudiaptomusgracilis, Thamnocephalus platyurus,Daphnia hyaline,Cyclops abyssorumMC Grazing on (artificiallyshortened) filaments Planktothrix(i) reduced survival Thamnocephalus(ii) survival naturally co–existing zooplankton unaffected;(iii) Eudiaptomus high sensitivity but also strict foodavoidance(iv) Daphnia and Cyclopsgreater physiological resistance to MC, less avoidance –ingested filaments(i) Daphnia feeding rates increased (not so for copepods)when prior toexposure MC wereextracted from filaments (MCacts as feeding deterrent)(ii) unidentified lipophilic toxinpresent(iii) high avoidance linked tohigh sensitivityKurmayer & Jüttner(1999)Daphnia pulex MC–LR Grazing on toxic Microcystis cells(0 – 320,000 cells mL-1) +low or high densityScenedesmus (20,000 or80,000 cells mL-1)MC–LR 8.9 10-5ng cell-1:(i) decreased population density(ii) delayed maturity(iii) increased number ephippia ind-1Effects toxic Microcystis comparable to food of low quality orlack of foodLaurén Määttä et al,1997 

# 729 Chapter 32: Cyanobacterial ToxinsExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceDaphnia magna MC Grazing on MC(+) wt andMC(–) mutant + Scenedesmus; Microcystis 0 –100 %of 5 mg C L-1(i) severe disturbance Daphniapopulation development (increased mortality, decreasedreproduction)(ii) MC(+) killed Daphniafaster MC(–)(iii) also MC(–) had negativeeffects on survival andgrowth, even in presenceScenedesmus(i) nutritional insufficiency Microcystis not responsible effectson Daphnia(ii) feeding inhibition and/orcompounds other MC(cyanopeptolines?) may play role(iii) MC alone cannot explainharmful effects on DaphniaLürling, 2003Daphnia magna MC–LR (i) grazing on MC(+) andMC(–) strains of Microcystis in max conc. of 5 mgC L-1in mixtures withScenedesmus + in sometreatments in addition:(ii) exposure to purified, dissolved MC, max 3.5μg L-1(i) no effects dissolved MC(ii) exposure cellbound MC:reduced feeding and growth(iii) reductions also in treatment 50 % Microcystis ofMC(–) strain + 50 %ScenedesmusInhibition of Daphnia feedingand growth in presence ofScenedesmus and MC(–) strainMicrocystis: unknown toxiccompoundsLürling & van derGrinten, 2003Daphnia magna PST Grazing on Aphanizomenon 1.2e6cells mL-1containing643–1170 pmol mL-1PST:(i) reduced fitness, growth andsurvival(ii) reduced activity cGSTNogueira et al, 2004 

# 7 30 B.W. Ibelings and K.E. HavensExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceDaphnia magna CYN, unidentifiedhepatotoxinGrazing on 2 differentstrains Cylindrospermopsis(+/– CYN);1.8 – 3.6 106cells mL-1;CYN 4.78 ng mg cells DW-1:(i) reduced growth and increased mortality, also incomparison to starvationtreatment,(ii) above also true for CYN(-)strain(iii) sGST + mGST activityincreasedUnknown toxic compounds present in strain that does not produce CYNNogueira et al,2004(b)Daphnia pulex; D.longispinaMC, unidentified toxinsGrazing on mixturesScenedesmus and Microcystis cells (10,000 or 40,000cells mL-1~ 0.076 – 0.304mg C L-1)(i) increased allocation resources to reproduction(ii) lower dose resultedsmaller clutch size D. Pulex;D. longispina no effect(iiii) higher dose virtual inhibition reproduction D. Pulex;D. longispina reduced sizeneonates(i) severe and dose dependent effects Microcystis on reproduction in Daphnia(ii) toxicity and food quality playa roleReinikainen et al,1999Estuarine calanoidcopepods(Eurytemora affinis;Acartia bifilosa)MC–LR,ANA(a),NODLNPurified dissolved toxins,single and in combination(i) 1μg mL-1MC or ANA noeffect on egg hatching(ii) 0, 0.25, 0.5 and 1μg mL1l: reduced survival for MC >0.1μg mL-1(iii) ANAa and NODLN onlyweak effectsReinikainen et al,2002Daphnia galeata MC Daphnia feeding on toxic(wt) and non toxic (mutantdoes not produce MC) Microcystis strains(i) wt toxic to Daphnia: decreased swimming + death,mutant not toxic(ii) both wt and mutant inhibitingestion rateDose response relationship notbetween MC content of the foodand effects in Daphnia but between ingestion rate and effectsRohrlack et al, 1999 

# 731 Chapter 32: Cyanobacterial ToxinsExposedorganismToxin Exposure route Concentration (dose)and effectRemark ReferenceDaphnia pulicaria MicroviridinJ(i) Grazing on Microcystisstrains containing 1.05 –1.57μg mm-3microviridin(ii) Purified, dissolved microviridin (0–12 mg L-1)Lethal molting disruption Microviridin is a protease inhibitorRohrlack et al, 2004Daphnia spp. MC Daphnia feeding on toxicMC(+) wt and non–toxic(MC(-) mutant Microcystisstrains(i) inhibition of feeding rate,equal inhibition for tox andnon–tox Microcystis(ii) reduced survival time(LT50) Daphnia when feedingon wt(iii) Daphnia feeding on mutant signs starvation(i) MC major source of acuteDaphnia poisoning(ii) clear relationship – dose–response – between LT50andMC ingestion rateRohrlack et al, 2001Daphnia galeata MC Daphnia feeding on MC(+)and MC(–) strainToxic strain 0.87 mg L-1MC:(i) both tox and non–tox Microcystis negatively affect thecohesion of midgut epithelium(within 9 h)(ii) tox strain: uptake MC inblood, increase 0.25 to 1 ng L1(6–9 h)(iii) tox strain: decreased beatrates (5–9 h), constant contraction midgut, finally complete loss beat rates, death(32–41 h)(i) midgut disrupting factor is notMC, disruption stimulates uptakebioactive compounds fromcyanobacteria in blood Daphnia(ii) results not easily translated tofieldRohrlack et al, 2005Daphnia magna MC (i) Daphnia exposed to suspensions of toxic Microcystis cells(ii) Daphnia exposed to lakewater (filtered or not)MC=2000μg g DW-1:reduced survival, fecundityand growthNegative correlation betweentoxin content and growth(strong) and between toxin concentration and fecundity (weak)Thostrup &Christoffersen, 1999 

# 7 32 B.W. Ibelings and K.E. Havens