Chapter 32: Cyanobacterial toxins: a qualitative 

meta–analysis of concentrations, dosage and 

effects in freshwater, estuarine and marine biota

Bas W Ibelings1, Karl E Havens2 

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 expanding 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 patterns from the literature or signals 

contradictions. The meta–analysis is set up by putting together two large 

tables – embodying a large and representative part of the literature (see 

Appendix 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 lab 

or 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 partly 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 at 

times 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 assess 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 occur in the foodweb with toxins being subject to degradation and excretion at every level. Nevertheless toxins 

were present at all tropic levels, indicating that some vectorial transport 

must take place, and in sufficient quantities 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 different 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, detoxication of MC and NODLN by conjugation 

with glutathione, and fairly rapid depuration and excretion. A common 

theme 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 adaptation to toxic cyanobacteria). This 

makes it harder to generalize on what is known about ecological effects of 

cyanobacterial toxins. The paper concludes 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 appear 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 (concentration 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) 

H2O2 Hydrogen peroxide (one of the ROS formed during oxidative 

stress) 

HP Hepatopancreas 

IP Intraperitoneal injection 

LC50 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 

NODLN Nodularin 

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

PST Paralytic Shellfish Toxin 

ROS Reactive Oxygen Species (formed 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 research involving aquatic organisms which may 

be exposed to toxin producing cyanobacterial blooms in their natural environment” (Zurawell et al. 1999) were fully justified. In contrast, there now 

exists a large literature from observational and experimental studies dealing with cyanobacterial toxins in aquatic systems. Yet major information 

gaps remain, and our ability to understand 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 the 

protease 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, chronic 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 supporting 

cyanobacterial 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–analysis 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 clarify 

the 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 
 679

of toxins on biota. While Tables A.1 and A.2 do not include 100% of the 

published papers on these subjects, they 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 toxin 

studied; (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 quantify 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 cyanobacterial toxins is often speculated to 

increase the risk of exposure for aquatic biota (especially at higher trophic 

levels), but bioaccumulation has seldom 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 consequences 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 relationships 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 μg kg bw-1). It is more common however to find toxin contents of 

the source expressed in μg L-1. Differences in units hamper interpretation 

across different studies. What will also emerge from Table A.2 is that the 

exposure route has a strong influence on biological effects. In fish IP injecChapter 32: Cyanobacterial Toxins
6

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 toxic 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 papers 

investigating 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 understanding of the effects these toxins 

have on aquatic biota) is another matter. Below we formulate an answer to 

this question – which is central to this 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 perspective 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 hypertrophic, 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 
 681

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 LD50 in mice is 50 μg kg-1, considerably lower than for instance the LD50 of MC–RR (600 μg kg-1) (Spoof 2005). Yet as the data in 

Table 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 suggested 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 LC50 of MC–RR (and of MC–YR) was actually very close to the LC50

of MC–LR. The real difference in this invertebrate was a much gentler 

slope of MC–RR compared to MC–LR (and YR) when LC10 and LC90 values were included (Blom et al. 2001). 

A further relevant distinction within the microcystins is that between the 

methyldehydroalanine–containing microcystins 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 is 

facilitated (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. More 

research 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 of 

NODLN 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 (anatoxin–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 and 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 cyanobactertial toxins (neurotoxins 

and hepatoxins) may play a simultaneous role and have synergistic effects 

(a more accurate description of the interaction between different 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 toxic 

cyanobacteria. 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 been exposed to toxins via natural routes 

(exposure through the food web or to dissolved toxins after lysis of 

blooms). In the ecosystem the feeding guild of a fish species appears to be 

a primary determinant of exposure to toxins. Phytoplanktivorous fish like 

silver 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 cyanobacteria. Likewise a species like 

flounder predates on filter feeding blue–mussels. Piscivorous fish that prey 

on zooplanktivorous fish are one step further removed from the toxin producing cyanobacteria. It may be expected 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 (benthic) 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 
 683

Carbis et al. (1997) explain that the neutral 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 feeding 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 much 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 cyanobacterial toxins produced in the pelagic zone. 

There is a surprisingly large number of studies on the presence of toxins 

in 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 generally 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 time 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 much 

less 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 ducks) still contain traces of toxins. 

Chapter 32: Cyanobacterial Toxins
68

Thus the mussels may be considered a vector that prolongs the time when 

toxins 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 zooplankton, 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 detection (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 – total extractable hepatotoxins, which includes the biotransformation products. TEH almost invariably exceeds the 

concentrations of untransformed hepatotoxins 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 – those that contain methyldehydroalanine 

– covalently bind to PP in plant and 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 
 685

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 cyanobacterial 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 Jones 

1995). 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 and usually it just means that toxins 

are present in biota. When a more formal – and informative – definition of 

bioaccumulation, and the related processes 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 supported by the data (Karjalainen et al. 

2005). Data in Table A.1 do show however that bioconcentration of 

NODLN may take place. In an experimental setting two copepods and a 

ciliate took up dissolved NODLN and accumulated this to concentrations 

far higher than in the water (BCF ranged from 12–22). Predators of these 

zooplankters and protozoa would be exposed to substantial concentrations 

of 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. BMF in the Baltic of Copepods and 

clams are exceptionally low; only a very small part of NODLN which is 

present in the cyanobacteria is taken 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
68

taken place. In this case however concentrations of MC in a whole organism (like Daphnia) are compared to values for a selected organ where the 

toxin 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 conclusion 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. The 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 a 

percentage 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; Karjalainen 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 
 687

Exposure routes 

Laboratory studies 

Early studies on fish (Tencalla et al. 1994; Kotak et al. 1996b) primarily 

used the method which is preferred for 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 LC50 for 

MC–LR in perch (1500 μg g DW-1, Ibelings et al unpublished data) is well 

above the LC50 for mice, indicating that these fish species are less sensitive 

to the toxin than warm blooded animals. 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 μg MC kg-1

bw given by gavage 8 times over 96 h (a total dose of 9200 μg 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 water 

is 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 purified toxins. These studies have proven 

valuable in finding the mechanisms through which biota are affected by 

cyanobacterial toxins but are less informative about the importance of uptake of dissolved toxins in the ecosystem. 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). Concentrations 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 cyanobacterial 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 whole 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 cyanobacterial cells. There are exceptions, however. For example, in A. salina purified CYN showed a lower LC50 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 directly 

feed on seston that includes cyanobacteria. Zooplankton, filter feeding bivalves and phytoplanktivorous fish would be among the organisms that are 

directly exposed to toxins in their food (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 food 

web. 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 low 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 
 689

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 detoxication (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 quantities to have harmful effects. The 

presence 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 study 

on 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. This 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 between 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
6

during 28d (Li et al. 2004). Experimental chronic exposure studies where 

biota are exposed to toxins for the greater 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 field 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 at 

the 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 susceptible to MC may protect themselves by strong inhibition of the intake of cyanobacteria (Demott 1999). 

Studies also have shown that zooplankton 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 
 691

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). Hence 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 responsible 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 as 

stimulation of gut muscles leading to a permanent contraction of the midgut. These effects became apparent as soon as MC was taken up into the 

blood. Contraction of the midgut interferes with digestion, nutrient assimilation and uptake of ions. Eventually MC resulted in a breakdown of 

Daphnia metabolism, exhaustion and eventually 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 from systems supporting dense blooms of 

cyanobacteria may be attributed to hepatotoxin exposure, but other factors 

like liver parasites and anthropogenic 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 nature of liver damage and recovery. 

The acute exposure study mentioned earlier where flounders were given a 

single dose of toxin (Kankaanpaa et al. 2005a) demonstrated that damage 

is transient, and recovery from liver damage is rapid (on the order of days). 

Many studies have failed to relate effects 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 LT50 (survival time) of Daphnia. Despite all the 

complicating factors, significant dose–effect relationships have been found 

and are included in Table A.2. Another example is the dose (and time) dependent mortality in brine shrimp exposed 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 the 

Baltic, 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 sensitive than fish larvae, but these larvae 

only 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). Moreover zooplankton, mussels and fish may 

temporarily stop feeding when toxic cyanobacteria are present and avoid 

ingestion in this way. If toxic cyanobacteria can not be avoided and cells 

are indeed ingested, very little of the toxin present may actually be taken 

up into the body. Mucoid 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 cohesion. 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 longer 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 
 693

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 toxic cyanobacteria is an efficient 

means to prevent ingestion of the toxins. 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 feeding inhibition would be adaptive if 

the environment could be sensed correctly (chemical cues) and animals are 

able to recover quickly from inhibition in the absence of toxic strains. 

Detoxication and oxidative stress 

Another important process is detoxication of MC, which now has been 

documented 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 cGST 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 and 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 to 

oxidative stress through the formation of ROS like hydrogen peroxide. Organisms have a wide range of protective 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 oxidative 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 on 

bacteria 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 LPS 

from different bacterial sources are always present in the aquatic environment (although not all LPS from different 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 biota 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 
 695

Knowledge gaps 

Throughout this paper remarks have been made about knowledge gaps that 

limit our understanding of the ‘true’ ecological effects of cyanobacterial 

toxins. 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 towards 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 behavior or genotoxicty; these may be 

more relevant than acute lethal effects, 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 detoxication 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 seem to play a relatively large role, 

they are often present when dead birds 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 the mammals which have been shown 

to be sensitive to cyanobacterial toxins in laboratory studies. A disease that 

must be mentioned here is avian vacuolar 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 blockage 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 cyanobacterial toxins (either dissolved in the 

water or contained in the cell) are sufficiently high to be directly lethal. 

Again, generalizations are difficult because there appear to be important 

differences between fish species. Fischer and Dietrich (2000) related the 

capacity 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, gonads), but the liver is the main target. Several studies have shown that around 50 % of fish caught from lakes 

or estuaries that support cyanobacterial 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 mussels 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 
 697

rial transport to predators (including man). However depuration studies 

have 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 extensive (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 effects 

like inhibition of feeding. Effects of cyanobacterial blooms also exert effects at the community level. Zooplankton community composition may 

change towards dominance of smaller cladocerans which have a lower 

grazing 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 interfere 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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Chapter 32: Cyanobacterial Toxins
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Chapter 32: Cyanobacterial Toxins
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Chapter 32: Cyanobacterial Toxins
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Chapter 32: Cyanobacterial Toxins
Appendix A 

Table A.1. Concentrations of various cyanobacterial toxins in a range of aquatic biota. Where possible concentrations are expressed as 

μg g DW-1. Conversion of wet weight to dry weight in animals using a factor of 0.1 and conversion of C to DW requires multiplication by 1/0.526 {Winberg, 1971). Ash content of DW is neglected. MC is commonly expressed as MC–LR eq, since the majority of studies only used MC–LR as a standard in HPLC. Concentrations are usually presented as a range (lowest – highest values 

found). Concentration in the ‘source’ applies to either purified toxin (experiments), cultured cells of cyanobacteria, dissolved toxins in natural waters or cyanobacteria in the seston. For abbreviatons see ‘list of abbreviations’. Organism Toxin Exposure routeConc. in 

organism (μg g-1 DW) 

Conc. in source (μg g-1 DW)

Analysis method 

Remark Reference 

BIRDS (i) coots (Fulica atra) (ii) grebes ( Podiceps nigricollis, P. cristatus) 

ANA(a-s), MC 

Natural routes: 

feeding in foodweb from lake with Anabaena bloom 

(i) ANA - Coot: 0.021 Grebe: 0.90 (ii) MC - 0.0005-0.001

ANA: 4-3300 MC: 0.1-0.9 

ELISA (MC) AChE assay (ANA) 

(i) Anabaena and its toxin found in stomach birds: cyanobacterial toxicosis? (ii) ratio ANAeq : MCeq 40 to > 1000 

Henriksen et al, 1997 

Lesser flamingo (Phoeniconaias minor) MC-LR, RR, LF and YR, ANA(a) Feeding on mats of cyanobacteria, as 

well as on planktonic Arthrospira 

(i) stomach: MC 1.96 (all 4 variants); ANA 43.4 (ii) intestines: MC 0.36 (MC-LR only); ANA 7.62 (iii) faeces: MC 0.48 (all but MCLF); ANA 2.45 

(i) mats - MC: 221-845 ANA: 10-18 (ii) plankton – MC: max 4600 ANA: max 223 

HPLC, MALDITOFMS 

(i) intoxication birds likely but lack of relevant data on susceptibility (ii) multiple stress factors

Krienitz et al, 2003; Ballot et al, 2004 

706 B.W. Ibelings and K.E. Havens 
Eider (Somateria mollissima) 

NODLN Feeding on mussels 

in Baltic sea 

(i) 0.003-0.18 (ii) 0.1-5.8 μg per liver

 ELISA and 

LC-MS 

 Sipiä et al, 2004 

FISH 

Round goby (Neogobius 

melanostomus) 

MC 

Natural routes in 

lake 

Goby liver: 0.6-3.0 91-820 Babcock-Jackson et al, 

2002 

Sticklebacks (Gasterosteus aculeatus) 

NODLN Feeding on copepods pre-exposed to 

cyanobacteria 

(i) 0.15 (ELISA) (ii) 0.8 (PPase) 

 ELISA and 

PPase assay 

 Ensgtröm-Öst et al, 

2002 

(i) perch (Percia fluviatilis) (ii) ruffe (Gymnocephalus cernuus) (iii) smelt (Osmerus eperlanus) 

MC 

Natural routes in 

lake foodweb 

(i) perch liver: 17-51 (ii) ruffe liver: 9-194 (iii) smelt liver: 59-874

7-3912 HPLC Apparent biomagnification based upon skewed 

representation 

Ibelings et al, 2005 

(i) silver carp (Hypophthalmichthys 

molitrix), 

(ii) Carassius gibelio 

MC-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-LR 

Jang et al, 2004 

Sea trout (Salmo trutta) 

NODLN Gavage with single 

dose of Nodularia 

(i) liver: 0.019-1.2; max 1.6 (ii) muscle: max. 0.125

 ELISA and 

HPLC 

Single oral dose: loss liver architecture 1-2d, partial recovery 4-8 d, complete after 8 d: damage reversible 

Kankaapää et al, 2002 

Flounder (Platichthys flesus) 

NODLN, MC Feeding on mussels 

in Baltic sea 

(i) 0.02-0.1 TEH prebloom exposure (ii) 0.02-2.23 TEH (ELISA) (iii) nd-0.47 (LC-MS) 

 ELISA and 

LC-MS 

(i) NODLN variable between individuals: peak conc. in sub-populations (ii) 50 % livers small scale necrosis 

Kankaapää et al, 2005a

Chapter 32: Cyanobacterial Toxins 707
Flounder (Platichthys 

flesus) 

NODLN 

Natural routes in seaLiver: 0.82-6.37 LC-MS 

No biotransformation products (e.g. glutathione 

adduct) found 

Karlsson et al, 2003 

(i) northern pike (Esox lucieus (ii) white sucker (Catostomus comersonii) 

MC Feeding on MC 

containing prey (i.a. gastropods) 

Liver: not detected 1.2–6.1 μg L-1 HPLC Kotak et al, 1996a 

Carp (Cyprinus carpio) MC Feeding on Microcystis scum in tanks (i) hepatopancreas: 2.6 (ii) muscle: 0.4 Li et al, 2004 Tilapia rendalli MC 

Natural routes in 

lake 

(i) viscera: 0 – 67.8 (ii) liver: 0 – 31.1 (iii) muscle: 0.003 – 0.026 

max. 980 μg L-1 HPLC, 

ELISA 

MC found in 75 % fish samples 

Magalhães et al, 2001 

Fish (unspecified) MC 

Natural routes in 

lake 

Muscle: 0.01 – 0.4 0.12–0.78 μg L-1ELISA Magalhães et al, 2003 

Goldfish (Carassius auratus) MC-LR IP injection Liver: 0.5 - 3 PPase Time course accumulation (first 48 h), depuration 

(48-96 h); patterns not affected by fastening 

Malbrouck et al, 2003, 2004 

Tilpapia (Oreochromis niloticus) 

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 – 1.0 

1120 ELISA At times MC in liver > gutMohamed 

et al, 2003 

Flounder (Platichthys flesus) 

NODLN and MC

Natural routes in 

Baltic sea 

Max 0.4 (TEH) 150-8700 THE; 

< 2400 NODLN

ELISA and MALDITOF-MS 

 Sipiä et al, 2001, 2002 

Tilapia rendalli MC (i) Feeding on Microcystis cells + fish 

food; one exp cells disrupted 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 less in presence alternative food (iii) depuration phase: only small percentage MC removed 

Soares et al, 2004 

708 B.W. Ibelings and K.E. Havens 
Silver carp (Hypophthalmichthys 

molitrix) 

MC-LR and RR Feeding on Microcystis bloom in 

tanks 

(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 – 1.4 

MC: 286–866; MC-LR:RR = 0.57 

PPase (i) in the various tissues 

and organs always MCRR found, rarely MC-LR (ii) active degradation of MC-LR during digestion (?) 

Xie et al, 2004 

Various fish species: phytoplanktivorous 

(Hypophthalalmichthys molitrix), herbivorous (Parabramis pekinensis), omnivorous (Carassius auratus), carnivorous (Culter ilishaeformis) 

MC-LR and RR 

Natural routes in 

lake 

(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 foodweb most at risk; general tendency MC to accumulate up the foodchain 

Xie et al, 2005 

Salmon (Salmo salar) MC-LR IP injection (i) 2.6 – 263 (MeOH) 

(ii) 138 – 1181 (Lemieux) 

 (i) PPase after 

MeoH extract. (ii) Lemieux (GCMS) 

Covalently bound MC made up ~ 74 % of total 

Williams et al, 1997a 

MACROINVERTEBRATES Mytilus galloprovincialis MC Grazing on Micro- cystis cells (i) mussel: 10.7 during accumulation rising to 

max of 16 on day 2 depuration 

(ii) faeces: 140 

3.4 μg / 107 cellsELISA (i) no mussel mortality 

(ii) during depuration initial increase, followed decrease MC 

Amorim & Vasconcelos (1999) 

Zebra mussels (Dreissena polymorpha) 

MC 

Natural routes in 

lake 

0.2 91-820 Babcock-Jackson et al, 

2002 

Chapter 32: Cyanobacterial Toxins 709
(i) freshwater shrimps 

(Palemon modestus; Macrobrachium nipponensis) (ii) red swamp crayfish (Procambarus clarkii) 

MC-LR and RR 

Natural routes in 

lake 

(i) P.modestus - stomach: 4.53 hepatopancreas: 4.29 gonads: 1.17 eggs: 2.34 muscle: 0.13 gills: 0.51 (ii) M. nipponensis - stomach: 2.92 hepatopancreas: 0.53 gonads: 0.48 eggs: 0.27 muscle: 0.04 gills: 0.05 (iii) P. clarkia - stomach: 9.97 hepatopancreas: 0.08 gonads: 0.93 muscle: 0.05 gills: 0.27 

 LC-MS (i) Proportion of MC-LR 

of total MC varied with type of tissue and the species; ratio of MCLR:MC: P. modestus decreased gonad (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 toxin 

burden in eggs (29 % in P. 

modestus), i.e.MC transferred to offspring 

Chen & Xie, 2005 

Freshwater snail (Bellamya aeruginosa) 

MC-LR and RR 

Natural routes in 

lake 

(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: 0.01 

240 HPLC MC hepatopancreas > digestive track: selective 

bioaccumulation? 

Chen & Xie, 2005 

710 B.W. Ibelings and K.E. Havens 
Zebra mussels 

(Dreissena polymorpha) 

MC-LR Feeding on toxic 

Microcystis cells 

(i) 11 – feeding solely Microcystis (ii) 3.9 feeding on mixture Microcystis and green algae 

3.1 LC-MS; 

MMPB 

(i) maximum share covalently bound MC 38 % total (ii) only 0.5 % offered MC found in mussels (iii) rapid depuration, after 3 wk nearly complete 

Dionisio Pires et al, 2004 

Mysid shrimp (Mysis relicta) 

NODLN Feeding on copepods pre-exposed to 

cyanobacteria 

(i) 0.74 (ELISA) (ii) 0.52 (PPase) 

 ELISA and 

PPase 

(i) time dependent accumulation in shrimps, not in fish (results ELISA) (ii) copepods as vectors to higher trophic levels 

Ensgtröm-Öst et al, 2002 

Zebra mussels (Dreissena polymorpha) 

MC 

Natural routes in 

lake foodweb 

1-30 7-3912 HPLC Ibelings et al, 2005 

Black tiger prawns (Penaeus monodon) 

MC, NODLN (i) natural routes in 

ponds 

(ii) oral uptake 

NODLN via food in 

experiments (iii) injection MCLR 

(i) ponds - HP (TEH) 0.006-0.08 (ii) experiment (NODLN) - brain and heart: 0.36 HP: 0.25 (0.83 peak level) gut: 0.1 gills: 0.014 muscle: 0.01 (iii) experiment (MC) -HP: 0.130 (peak) 

(i) ponds: TEH: 1.2 (ii) exp: 10 μg 

kg bw-1 

ELISA; HPLC 

Rapid depuration from 

prawns 

Kankaanpää et al, 2005b 

Baltic Sea zooplankton: Acartia tonsa, Eurytemora affinis, Strombium sulcatum 

NODLN Exposure to dissolved NODLN 

(

3Hdihydronodularin) 

A. tonsa: 0.37 μg g-1 C 

E.affinis: 0.60 S. aulcatum: 1.55 

5 μg L-1 (i) minimum BCF 12 – 18 

for copepods; (ii) max BCF ciliate 22 (iii) possible vectorial transport with significant sublethal effects 

Karjalainen et al, 2003 

Chapter 32: Cyanobacterial Toxins 711
(i) pike larvae (Esox 

lucius) (ii) mysid shrimps (Neomysis integer) 

NODLN Fed with zooplankton pre-exposed to 

(i) Nodularia extract (ii) purified 

NODLN 

(20 μg L-1) 

(i) pike larvae (12h): 0.47 (ii) Neomysis: (12h): 0.31 

 Only 0.12 (pike) and 0.03 

% (shrimps) of ingested toxin was detected in animals 

Karjalainen et al, 2005 

Gastropods (Lymnea stagnalis, Helisoma trivolis, Physa gyrina) 

MC Grazing on (settled) 

lake phytoplankton 

 11 – 121 1.2 – 6.1 μg L-1 

~1220 μg g-1

DW 

HPLC MC in seston not expressed on DW basis Kotak et al, 1996a 

Baltic clam (Macoma balthica) 

NODLN Exposure to dissolved toxin and 

Nodularia cells in 

tanks 

0.16 - 16.6 (24 h) – 30.3 (96 h) (TEH); < 5 % of this NODLN 

10-50 μg L-1

(cells) 4-20 μg L -1 (dissolved) 

ELISA, HPLC, MALDITOF-MS 

No NODLN-GSH conjugates detected Lehtonen et al, 2003 

Signal crayfish (Pacifastacus leniusculus) MC Exposure to toxic and non toxic 

Planktothrix agardhii 

MC present but not quantified 

3610 (i) crayfish ingested 430 

μg MC – no effects (ii) possible vectorial transport to fish, birds, mink 

Liras et al, 1998 

Crab (unspecified) MC 

Natural routes in 

lake 

Muscle: 0.02 – 1.0 0.12 – 0.78 μg 

L-1 

ELISA Magalhães et al, 2003 

Freswhater mussel (Alathyria condola) 

PST feeding on toxic 

Anabaena 

5.7 1580 

Negri & Jones, 1995 

Pearl oyster (Pinctada maxima) 

SAX (i) natural routes in 

sea; (ii) exposure juvenile oyster to Trichodesmium 

0.73 in viscera diseased oyster 

Not detectable HPLC; mouse 

assay; sodium channel and saxiphilin binding assay

no mortality juvenile oysters in experiment 

Negri et al, 2004 

712 B.W. Ibelings and K.E. Havens 
(i) freshwater snail 

(Sinotaia histrica) (ii) freshwater clam (Corbicula sandai) 

MC-LR and RR (i) natural routes in 

lake (ii) feeding on toxic Microcystis cells in exp. 

(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. 436 

51.8 – 284 (lake) 20.1 μg L-1 

(exp.) 

HPLC (i) lag phase in depuration 

from snail tissue; biological half life 8.4 d (ii) MC in lake snail still 

present next spring 

Ozawa et al, 2003 

Freshwater mussel (Anodonta cygnea) 

PST Feeding on toxic 

Aphanizomenon - accumulation 14d, depuration 14d 

0.26 1.9-2.6 HPLC (i) Anodonta exposed to 

1.4e9 cells L-1 d-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, 2004 

Freshwater clam (Anodonta grandis) (i) exposure dis- solved MC 

(ii) natural routes in lake 

Viscera 0.59 (i) gills 0.31 (ii) muscle 0.36 

MC 51-55 μg 

L-1 dissolved up 

to 8.3 μg L-1 in 

lake 

 (i) toxin burden evenly 

distributed over three body parts (ii) rapid depuration first 6d (~70% gone), stable for 15d afterwards (iii) suggestion of bioconcentration 

Prepas et al, 1997 

Chapter 32: Cyanobacterial Toxins 713
Freshwater mussel 

(Anodonta cygnea) 

CYN Feeding on Cylindrospermopsis culture (i) haemolymph 61.5 (= 408 μg 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) 

14-90 μg L-1 HPLC (i) no adverse effects on 

mussels despite bioaccumulation CYN in haemolymph to conc. higher than in water (ii) bi-phasic depuration, increase in CYN content from day 22-28 

Saker et al, 2004 

Blue mussel (Mytilus edulis) 

NODLN 

Natural routes in the 

sea 

Max 2.15 Max 2400 Sipiä et al, 2001 

(i) balthic clam (Macoma balthica) (ii) blue mussel (Mytilus edulis) 

NODLN and MC

Natural routes in the 

sea 

(i) mussels 1.5 (max of TEH) (ii) clams 0.1-0.13 

150-8700 ELISA and 

MALDITOF-MS 

(i) NODLN-GSH conjugates confirmed with MS (ii) mussels MC 30 fold increase summer (iii) time lagged accumulation deep sites 

Sipiä et al, 2002 

Blue mussels (Mytilus edulis) 

NODLN (i) grazing on Nodularia 

(ii) coprophagy (exposure to feces of 

mussels grazing on 

Nodularia) 

(i) Mussels preexposure to Nodularia: 0.05-0.1 (ii) Mussels grazing on 

Nodularia - digestive track 245 body 80 gills 2 (iiii) feces when feeding on Nodularia 95 (iv) Body after coprophagy 

0.065 (v) Feces coprophagy 1(vi) PF from (ii) 714 

Nodularia culture 16 μg L-1 LC-MS (i) high NODLN in PF in- dicative of selective feeding; (ii) cells Nodularia in 

PF may survive but growth inhibited 

Svensen et al, 2005 

714 B.W. Ibelings and K.E. Havens 
Mytilus galloprovincialis MC-LR Grazing on Micro- cystis cells 10.5 (of which 96 % in digestive gland + stomach) 28 μg 108 cells-1 Depuration bi-phasic and fairly rapid (13d) Vasconcelos 1995 

Crayfish (Procamabarus clarkia) 

MC Feeding on toxic 

(and non toxic) Microcystis cells 

2.9 (max after 11 d), of which 53 % in intestine, 38 % HP and 9 % rest body 

2300 ELISA Juvenile crayfish enhanced mortality on nontoxic Microcystis: compounds other than MC 

more relevant to invertebrates (?) 

Vasconcelos et al, 2001

Dungeness crab larvae (Cancer magister) 

MC-LR 

Natural routes in 

foodweb 

(i) 0.006 (after MeOH extraction) (ii) 84.4 (after Lemieux) 

 (i) PPase after 

MeoH extract. (ii) Lemieux (GCMS) 

10.000 fold greater MC concentration using MMPB after Lemieux extraction 

Williams et al, 1997a 

Blue mussel (Mytilus edulis) 

MC-LR (i) feeding on toxic 

Microcystis cells (ii) natural routes in foodweb 

(i) 3369 decreasing to 113 after 4 d depuration (Lemieux) (ii) 2 - dropping to 0.14 over 53 d depuration (MeOH) 

 (i) PPase after 

MeoH extract. (ii) Lemieux (GCMS) 

Less than 0.1 % of MC extractable with MeOH 

Williams et al, 1997b 

Freshwater 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 increased from digestive track to HP to gonad (iii) suggestion LR more resistant to degradation? 

Xie et al, 2005 

Freshwater mussel (Unio douglasiae) 

MC-LR Exposure to Microcystis cells HP: (i) at 15 oC – 130 

(ii) at 25 o C - 250 

27 - 50 μg L-1 (i) accumulation temperature dependent 

(ii) depuration relatively fast; but slower at 15 than 25 oC; halted in winter 

(iii) no adverse effects on mussels 

Yokoyama & Park (2003) 

Chapter 32: Cyanobacterial Toxins 715
Pulmonate snails 

(Lymnea stagnalis, Helisoma trivolis, Physa gyrina) 

MC 

Natural exposure in 

lakes of varying trophic status 

 Up to 144 0 – 1526 HPLC Possible uptake routes via 

food and directly from water 

Zurawell et al, 1999 

ZOOPLANKTON 

Echinogammarus ischnus 

MC 

Natural routes in 

lake 

2 91-820 Exposure of this detrivore 

via pseudofaeces Dreissena? 

Babcock-Jackson et al, 2002 

Community of different spp, i.a. Daphnia galeata 

MC 

Natural routes in 

lake 

0-1352 HPLC Ibelings et al, 2005 

(i) calenoid copepods (Eurytemora affinis, Acartia tonsa) (ii) ciliate (Strombidium sulcatum) 

NODLN Uptake of dissolved 

NODLN from water 

during 15 min - 6 d 

(i) E.affinis 1.14 (ii) A. tonsa 0.66

(iii) S. sulcatum 4.8 (maxima) 

5 μg L-1 (i) BCF copepods 12-18 

(ii) ciliate 22 (iii) vectorial transport to shrimps and planktiv. fish 

Karjalainen et al, 2005 

Community of different spp., a/o Daphnia pulex 

MC Grazing on lake 

phytoplankton 

 Up to 67 1.2 – 6.1 μg L-1 HPLC Kotak et al, 1996 

(i) Thamnocephalus, 

platyurus 

(ii) Eudiaptomus gracili, (iii) Daphnia hyaline (iv) Cyclops abyssorum 

MC (i) grazing on mixtures of Cryptomonas and Planktothrix 

(ii) Planktothrix at 0.05 or 0.1 mg C L-1

- MC(+) or MC(-) 

 4.7 ng MC-LR 

per μg C-1 

HPLC Kurmayer & Jüttner, 

1999 

716 B.W. Ibelings and K.E. Havens 
Calanoid copepod (Eurytemora affinis) 

NODLN (i) feeding on Nodularia and non-toxic 

flagellates (ii) natural seston 

(i) fed with Nodularia: 0.032 ng copepod-1 

(background) rising to 0.007 (ELISA) (ii) 0.0095 to 0.101 (PPase) (iii) feacal pellets - 0.0067 ng pellet-1 (ELISA) (iv) idem 0.0050 (PPase) 

 ELISA and 

PPase 

(i) lower conc. when fed with natural seston (dominated by non-tox Aphanizomenon) (ii) ELISA and PPase different but not significant (iii) no acute effects on copepods despite accumulation (iv) vectorial transport to foodweb (v) toxin content feacal 

pellets low: no vector to 

coprophagous animals 

Lehtiniemi et al, 2002 

Daphnia magna PST (i) grazing on 

Aphanizomenon (1.2e6 cells mL-1) 

(ii) lyophilized material (1 mg mL-1) 

(i) exposure Apha: 0.065-0.378 pmol PST animal-1 

(ii) exposure lyophilized material: 0.007 

(i) cells: 643-1170 pmol mL-1 

(ii) lyoph.: 2745 pmol mL-1

HPLC-FLD Bioaccumulation factor > 

1 after 12h feeding 

Nogueira et al, 2004 

Daphnia magna CYN Grazing on Cylindrospermopsis 

strains (+/- CYN) 

(i) 0.025 ng animal-1

(24 h) (ii) 0.020 ng animal-1

(48 h) 

(i) intracellular toxin 219.6 – 236.4 ng mL-1

(ii) extracellular toxin 12.4 – 47.4 ng mL-1

(increase in time) (iii) total CYN 234.2 – 278.4 ng L-1 

HPLC-MS Bioaccumulation factors 

smaller than unity: 0.71 (24 h) and 0.46 (48 h); levels not high enough to indicate bioaccumulation 

Nogueira et al, 2004(b)

Chapter 32: Cyanobacterial Toxins 717
Daphnia magna MC (i) grazing on Microcystis alone or 

mixed with Scenedesmus (3.2 mg C L-1) 

(ii) exposure to lake water enriched with Microcystis from enclosures (8.9 mg C L-1) 

Daphnia 0.2-24.5 μg g-1

DW (highest when exposed solely to Microcystis) 

(i) Microcystis: 2000 μg g DW-1

(ii) MC varied between 5 – 156 μg L-1 in enclosures 

ELISA (i) calculation shows MC 

really accumulated in body, not just gut content (ii) calculation shows transport to roach results only in sublethal effects 

Thostrup & Christoffersen, 1999 

718 B.W. Ibelings and K.E. Havens 
Table A.2. Effects of various cyanobacterial toxins on aquatic biota. Where possible dose and effect are shown. 

Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

BIRDSFISHZebra fish (Danio rerio) 

MC–LR Purified MC, dissolved in 

water 

(i) 0.5, 5 μg L-1: increased motility 

(ii) 15, 50 μg L-1: decreased 

motility (iii) 50 μg L-1: reduced spawning 

(i) effects on behaviour dose dependent (reversal at higher conc.) (ii) LOEC < 0.5 μg L-1 

Baganz et al, 1998 

Brown trout (Salmo trutta) 

MC i) purified MC 

ii) aqueous extracts of Microcystis 

5, 50, 500 μg L-1 MC: cardiovascular effects Purified MC no effect on (some) cardiac responses, in contrast to 

cell extracts – always effects; cardiac output only affected by MC at higher doses 

Best et al, 2001 

Zebra fish (Danio rerio) 

MC–LR; LPS 

In vivo exposure fish embryos to dissolved toxins; LPS different bacterial origins (including cyanobacterial) 

LPS (up to 29.8 106 EU mg-1

DW) + MC (up to 12.6 ng mg DW-1): reduced mGST + 

sGST 

LPS from axenic bacteria or lake blooms reduced GST activity, interfered detoxication 

Best et al, 2002 

Rainbow trout (Oncorhynchus mykiss) MC Aqueous suspension whole or broken Microcystis cells 100 μg L-1 MC–LR: osmo- regulatory imbalance, increased liver mass Interaction MC with LPS Best et al, 2003 

Chapter 32: Cyanobacterial Toxins 719
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Brown trout (Salmo trutta) 

MC–LR (i) purified MC 

(ii) lysed Microcystis cells 

i) lysed tox Microcystis, containing 41 – 68 μg L-1 MC: 

ionic imbalance, reduced growth ii) purified toxin: effects on growth (less than toxic lysate) iii) lysed non–tox Microcystis: effects on growth (less than toxic lysate) 

Ammonia similar effects to MC Bury et al, 1995 

Carp (Cyprinus carpio) MC Fish exposed to toxins via natural routes in the lake, 

including Microcystis scums 

Up to 4000 μg g-1 DW: hepatic lesions Lesions in > 50 % fish examined Carbis et al, 1997 

Carp (Cyprinus carpio) MC–LR Gavage freeze dried Micro- cystis cells (single sublethal 

bolus) 

400 μg kg-1 bw: pathological 

changes in hepatopancreas, kidney, gastrointestinal tract (i.a. apoptosis) 

In carp – cyprinid – compared to salmonids liver pathology develops faster and at lower MC concentrations 

Fischer & Dietrich, 2000 

Rainbow trout (Oncorhynchus mykiss) MC Gavage freeze dried Micro- cystis cells 5700 μg kg-1 bw: liver necro- sis, hepatocyte apoptosis Unbound MC results in fast ef- fects, including PP inhibition and 

liver necrosis; covalently binding to MC results slower effects, i.a. apoptosis 

Fischer et al, 2000 

(i) Hypophthalmichthys molitrix (ii) Carassius gibelio 

MC–LR; MC–RR 

Feeding on cultured Microcystis cells unspecified dose: reduced growth Jang et al, 2004 

720 B.W. Ibelings and K.E. Havens 
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Tilapia (Oreochromis sp) MC–LR Feeding on lyophilized Mi- crocystis bloom (crushed 

and non–crushed cells) 

60 μg fish-1 day-1 MC–LR: 

oxidative stress in liver, kidney and gills 

(i) crushed cells more effect non–crushed on LPO and levels antioxidant enzymes (ii) time dependent development oxidative stress: not yet present after 14d; 21 dlipid peroxidation + enhanced antioxidant enzyme activities 

Jos et al, 2005 

Sea trout (Salmo trutta) 

NODLN Gavage single dose Nodularia 210–620 μg kg-1 bw: rapid (1– 2 d) but reversible (4–8 d) 

liver damage 

(i) rapid detoxication, cross reaction conjugates in ELISA (ii) no effects on swimming 

Kankaanpää et al, 2002 

Rainbow trout (Oncorhynchus mykiss) MC–LR IP injection of purified MC (i) 1000 μg kg-1: 100 % mor- tality 

(ii) 400 μg kg-1: no mortality; 

increased ratio liver to body mass, liver necrosis; kidney lesions 

Fish less sensitive than mice: LD50 = 400 – 1000 μg kg-1

Kotak et al, 1996b 

Carp (Cyprinus carpio) MC–LR In vitro exposure hepato- cytes to 10 μg 

L-1 MC–LR 

Increase ROS, depletion GSH, increase SOD, CAT and GS–Px activity, GST unchanged: oxidative shock by exposure MC 

Antioxoidant enxymes did not prevent oxidative shock: apoptosis and necrosis hepatocytes 

Li et al, 2003 

Carp (Cyprinus carpio) MC Feeding on Microcystis scum in tanks 50 μg kg-1 bw for 28 days: re- duced growth, increased liver 

enzyme activity, damaged hepatocytes, 

Long term subchronic effects Li et al, 2004 

Chapter 32: Cyanobacterial Toxins 721
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Loach (Misguruns mizolepis) 

MC–LR Dissolved MC–LR MC–LR at 1, 3, 10, 100, 1000 

μg L-1: 

(i) mortalilty, delayed hatching, liver (necrosis) and cardiotoxicity (ii) (late stage) embryos and larvae more sensitive than juvenile fish: LC50 larvae = 164; 

juveniles = 593 μg L-1

(i) embryonic development loach affected by MC in contrast to zebrafish (ii) results dose dependent, i.e. increasing effects with increasing conc. MC. No mortality at the lower doses. 

Liu et al, 2002 

Goldfish (Carassius auratus) 

MC–LR IP injection 125 μg kg-1 bw: no change 

ionic homeostasis, liver lesions, changes liver enzyme activity 

Liver damage reversible Malbrouck et al, 

2003 

Goldfish (Carassius auratus) 

MC–LR IP injection of fed and fasted 

juvenile goldfish 

(i) 125 μg kg-1 bw: inhibition 

PPase, complete after 6 h in fasted, less inhibition in fed fish (ii) recovery after 96 h (iii) GSH levels and GST unaffected 

Feeding status has effect on toxicity MC, possibly by acting on bile formation and secretion 

Malbrouck et al, 2004a 

(i) silver carp (Hypophthalmichthys molitrix) (ii) carp (Cyprinus carpio) 

MC–LR (i) ip injection purified MC 

(ii) cyanobacterial biomass per os / anus 

(i) 400 μg kg-1 bw (purified MC): no – minor effects (ii) 3 – 1,200 μg kg-1 bw 

(biomass): changes blood indices and immunological changes 

(i) effects crude biomass much stronger purified toxin (ii) oxidative stress 

Palikova et al, 1998 

722 B.W. Ibelings and K.E. Havens 
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Seven fish (cyprinids) and three amphibian species MC–LR, RR and YR, SAX, 

ANA(a) 

(i) in vivo emersion (ii) exposure to dissolved purified toxin (iii) exposure crude cyanobacterial cell extracts 

(i) MC 0 – 50 μg L-1: no acute 

effects on embryonic development (ii) MC–RR > 0.5 μg L-1; YR 

> 5 μg L-1; LR > 50 μg L-1: 

timing hatching affected (ii) morphological effects only at highest conc. MC–LR 10 mg L-1(iii) SAX > 10 μg L-1 delayed 

hatching malformations at 500 μg L-1 

(iv) ANA(a) 400 μg L-1 hearth 

rate affected, no chronic effects (v) far more pronounced effects with crude extracts: malformations and mortality 

Crude cell extracts of several cyanobacteria gave severe effects (more so than purified toxin): cannot be attributed MC alone 

Oberemm et al, 1999 

Chapter 32: Cyanobacterial Toxins 723
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Rainbow trout (Oncorhynchus mykiss) 

MC–LR (i) IP injection or oral dosage purified MC 

(ii) IP injection or oral dosage freeze dried Microcystis cells 

(i) 550 μg kg-1 bw MC (ip): severe liver damage, death (ii) 550 μg kg-1 bw Microcystis (ip): severe liver damage, 

death (iii) 1200 μg kg-1 bw MC 

(oral): no effects (iv)1700 μg kg-1 bw Microcystis (oral): no effects 

(v) 6600 μg kg-1 bw Microcystis (oral): severe liver damage, 

death (vi) 550 μg kg-1 bw Microcystis (8 times oral dosage): 

modest – severe liver damage 

Treatments included repeated (oral) exposure to MC, i.e. close to natural exposure but during ‘limited’ period of time (8 times at 12 h intervals) 

Tencalla et al, 1994 

Zebra fish (Danio rerio) embryos 

MC–LR Purified dissolved MC (i) > 0.1 μg L-1: increased 

sGST and GPx activity (ii) > 2.0 μg L-1: effects on 

growth and survival 

Dose dependent relationship MC~GST 

Wiegand et al, 1999 

Carp (Cyprinus carpio) MC–LR Purified dissolved MC inhibition ATP–ase activity of Na+ K+ pump in gills, disruption ion homeostasis Harmful effects of dissolved MC (without ingestion of cells) Zambrano & Ca- nelo, 1996 

MACRO–INVERTEBRATES Brine shrimp (Artemia salina) MC–LR, MCHtyR and 

NODLN 

Purified dissolved toxin 0.5 μg L-1: elevation GST activity Beattie et al, 2003 

724 B.W. Ibelings and K.E. Havens 
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

(i) pike larvae (Esox lucius) (ii) mysid shrimps (Neomysis integer) 

NODLN (i) purified NODLN (20 μg 

L-1) 

(ii) crude extract Nodularia 

(i) crude extract: pike larvae: decreased ingestion and faeces production rates shrimps: no effects on molting, faeces production, C:N or growth (ii) purified NODLN: no effects 

Purified NODLN no effects on pike larvae; crude cell extracts stronger effects 

Karjalainen et al, 2005 

Baltic clam (Macoma balthica) 

NODLN (i) exposure to dissolved 

toxin (ii) exposure to toxic and non toxic Nodularia cells in tanks 

4–20 μg NODLN per day: (i) conc. dependent neurotoxic effects (increase / decrease AChE activity when exposed to low and high NODLN respectively) (ii) some treatments low siphon activity 

Abundant unidentified compound with NODLN like spectral characteristics (found in both toxic and non–toxic Nodularia treatments) 

Lehtonen et al, 2003 

Brine shrimp (Artemia salina) CYN and MC (i) purified dissolved CYN and MC 

(ii) extracts Cylindrospermposis, Microcystis 

(i) CYN LC50 decreased from 

4.48 to 0.71 μg mL-1 between 

24 and 72 h (ii) likewise MC LC50 from 

4.58 to 0.85 μg mL-1

(i) dose and time dependent mortality (ii) LC50 cell extracts typically > 

than purified CYN (reduced bioavailabilty?) 

Metcalf et al, 2002 

Estuarine crab (Chasmagnathus granulatus) 

MC Cell extracts Microcystis Injected daily for 4–7 d with 

17.6 ng MC: (i) increased enzyme activity (GST, CAT) in hepatopancreas (ii) no change LPO – no oxidative damage (?) (iii) yet histological damage 

 Pinho et al, 2003 

Chapter 32: Cyanobacterial Toxins 725
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Estuarine crab (Chasmagnathus granulatus) 

MC Crabs injected twice with 

cell extracts Microcystis 

(i) decreased Na+, K+–ATPase 

activity anterior gills (ii) increased GST posterior gills (iii) increased TOSC 

Increased TOSC in response MC as protective mechanism against LPO (level unchanged by exposure to MC) 

Vinagre et al, 2003 

ZOOPLANKTON

Moina macrocopa Unidentified 

metabolites (possibly cyanopeptolins A–D) 

Freeze dried Microcystis Inhibition of proteases Agrawal et al, 2005 

Thamnocephalus platyurus 

[D–Asp3, 

(E)–Dhb7]MC–

RR, MC–LR, MC–YR, MC–RR, NODLN 

Purified toxins LC50 NODLN < [D–Asp3, 

(E)–Dhb7]MC–RR < other 

MC 

LC50 insufficient to study response of organisms to exposure; 

LC10 + LC90 required to get slope

Blom et al, 2001 

726 B.W. Ibelings and K.E. Havens 
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Daphnia pulicaria, D. Pulex, D. hyaline, Diaptomus birgei 

MC–LR, NODLN, ANA(a) 

Exposure to: (i) purified toxins (ii) cell extracts (iii) toxic or non–toxic cyanobacterial strains 

(i) species specific responses to toxin exposure; not all daphnids equally sensitive in terms feeding inhibition (ii) LC50 for MC after 24 h 

(varied from > 50 μg L-1 in D. 

pulicaria to < 1.0 in D.birgei); for NODLN > 20 and < 0.6 respectively (iii) LC50 decreased with longer exposure

(i) feeding inhibition protects against toxic effects: less sensitive Daphnia stronger inhibiton (ii) zooplankters insensitive dissolved MC, more so dissolved ANA(a) 

DeMott et al, 1991 

Daphnia spp MC Feeding on mixtures Microcystis and Scenedesmus; 0, 

50 or 80 % Microcystis in total food conc. of 0.5 mg C L-1

(i) rapid feeding inhibition (but recovery after continued exposure to same mixture) (ii) reduced growth and reproduction (iii) reduction in growth/ingestion: direct toxic effects and feeding inhibition 

Clear differences between Daphnia spp DeMott, 1999 

Temperate and tropical cladocerans (Ceriodaphnia cornuta, Daphnia pulex; D. pulicaria, D similes, Moina micrura, Moinodaphnia macleayi) 

MC (i) grazing on toxic Microcystis strains mixed with 

Ankistrodesmus, total conc. 1.0 mg C L-1; 

(ii) acute and chronic exposure 

MC contents 2810–4080 μg g1 DW: 

(i) decreased survival in presence tox Microcystis (ii) toxic Microcystis inhibited feeding rate, even when just 5 % in mixture with greens (ii) non–toxic cyanobacterium as sole food:poor growth 

(i) species from low productivity sites – adapted to starvation – showed lowest sensitivity to toxic Microcystis (ii) small and large bodied fast growing spp prone to starvation and most sensitive to MC; small bodied slow growing spp most resistant and least sensitive 

Ferrão–Filho et al, 2000 

Chapter 32: Cyanobacterial Toxins 727
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Tropical cladocerans (Moina micrura, Ceriodaphnia cornuta) 

MC Grazing on: 

(i) large Microcystis colonies (ii) single cells and (small) colonies Microcystis 

MC 2.0 – 16.0 μg L-1: (i) toxic Microcystis inhibited growth reduced reproduction cladocerans; partly through feeding inhibition (ii) effects unicellular lab cultures stronger than colonial cultures or natural seston, especially with large colonies (although very toxic 3.9 mg MC g DW-1) 

Colony forming cyanobacteria: little effect despite high toxicity; possible explanation why field studies fail to demonstrate effects of toxic blooms 

Ferrão–Filho & Azevedo 2003 

Daphnia magna MC Grazing on mixtures of Microcystis and Scenedesmus; 

max of 140,000 Microcystis cells mL-1 (1.0 mg C L-1) 

(i) reduced growth, reproduction and survival, increasing effects with increased proportion Microcystis cells (0; 50 or 100 %)0 (ii) pre–exposure (acclimation) reduced harmful effects: development of tolerance 

 Gustafsson & Hansson, 2004 

Daphnia pulex MC–LR (i) acute exposure to Microcystis cells 

0–2.43 mg C L-1 = 0–

360,000 cells mL-1

(ii) chronic exposure 30,000 cells mL-1 

MC–LR 7.6 10-5 ng cell-1: 

(i) variation in acute tolerance (EC50) to toxic Microcystis 

(ii) increase temperature: decrease EC50

(iii) chronic exposure, reduced survival and reproduction, clonal differences reversed compared to acute exposure 

Suggestion made that more resistant clones show stronger feeding inhibition Hietala et al, 1997 

728 B.W. Ibelings and K.E. Havens 
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Calanoid copepods (Acartia bifilosa, Eurytemora affinis) 

NODLN Cultures of Nodularia added 

to enclosures 

NODLN ~ 11μg mg DW-1 (3–

4 μg mL-1): 

Baltic copepds feed, ingest, reproduce and survive in presence toxic Nodularia, no negative effects 

Perhaps hatching success more sensitive to toxins (not measured) 

Koski et al, 2002 

Eudiaptomus gracilis, Thamnocephalus platyurus, Daphnia hyaline, Cyclops abyssorum 

MC Grazing on (artificially 

shortened) filaments Planktothrix 

(i) reduced survival Thamnocephalus (ii) survival naturally co–existing zooplankton unaffected; (iii) Eudiaptomus high sensitivity but also strict food avoidance (iv) Daphnia and Cyclops greater physiological resistance to MC, less avoidance – ingested filaments 

(i) Daphnia feeding rates increased (not so for copepods) when prior toexposure MC were extracted from filaments (MC acts as feeding deterrent) (ii) unidentified lipophilic toxin present (iii) high avoidance linked to high sensitivity 

Kurmayer & Jüttner (1999) 

Daphnia pulex MC–LR Grazing on toxic Microcystis cells 

(0 – 320,000 cells mL-1) + 

low or high density Scenedesmus (20,000 or 80,000 cells mL-1) 

MC–LR 8.9 10-5 ng cell-1: 

(i) decreased population density (ii) delayed maturity (iii) increased number ephippia ind-1

Effects toxic Microcystis comparable to food of low quality or lack of food 

Laurén Määttä et al, 1997 

Chapter 32: Cyanobacterial Toxins 729
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Daphnia magna MC Grazing on MC(+) wt and 

MC(–) mutant + Scenedesmus; Microcystis 0 –100 % of 5 mg C L-1 

(i) severe disturbance Daphnia population development (increased mortality, decreased reproduction) (ii) MC(+) killed Daphnia faster MC(–) (iii) also MC(–) had negative effects on survival and growth, even in presence Scenedesmus 

(i) nutritional insufficiency Microcystis not responsible effects on Daphnia (ii) feeding inhibition and/or compounds other MC (cyanopeptolines?) may play role (iii) MC alone cannot explain harmful effects on Daphnia 

Lürling, 2003 

Daphnia magna MC–LR (i) grazing on MC(+) and 

MC(–) strains of Microcystis in max conc. of 5 mg C L-1 in mixtures with 

Scenedesmus + in some treatments in addition: (ii) exposure to purified, dissolved MC, max 3.5 μg L-1

(i) no effects dissolved MC (ii) exposure cellbound MC: reduced feeding and growth (iii) reductions also in treatment 50 % Microcystis of MC(–) strain + 50 % Scenedesmus 

Inhibition of Daphnia feeding and growth in presence of Scenedesmus and MC(–) strain Microcystis: unknown toxic compounds 

Lürling & van der Grinten, 2003 

Daphnia magna PST Grazing on Aphanizomenon 1.2e6 cells mL-1 containing 

643–1170 pmol mL-1 PST: 

(i) reduced fitness, growth and survival (ii) reduced activity cGST 

 Nogueira et al, 2004 

730 B.W. Ibelings and K.E. Havens 
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Daphnia magna CYN, unidentified 

hepatotoxin 

Grazing on 2 different strains Cylindrospermopsis (+/– CYN); 1.8 – 3.6 106 cells mL-1; 

CYN 4.78 ng mg cells DW-1: 

(i) reduced growth and increased mortality, also in comparison to starvation treatment, (ii) above also true for CYN(-) strain (iii) sGST + mGST activity increased 

Unknown toxic compounds present in strain that does not produce CYN Nogueira et al, 2004(b) 

Daphnia pulex; D. longispina 

MC, unidentified toxins Grazing on mixtures Scenedesmus and Microcystis cells (10,000 or 40,000 

cells mL-1 ~ 0.076 – 0.304 

mg C L-1) 

(i) increased allocation resources to reproduction (ii) lower dose resulted smaller clutch size D. Pulex; D. longispina no effect (iiii) higher dose virtual inhibition reproduction D. Pulex; D. longispina reduced size neonates 

(i) severe and dose dependent effects Microcystis on reproduction in Daphnia (ii) toxicity and food quality play a role 

Reinikainen et al, 1999 

Estuarine calanoid copepods (Eurytemora affinis; Acartia bifilosa) 

MC–LR, ANA(a), NODLN 

Purified dissolved toxins, single and in combination 

(i) 1 μg mL-1 MC or ANA no 

effect on egg hatching (ii) 0, 0.25, 0.5 and 1 μg mL1l: reduced survival for MC > 

0.1 μg mL-1

(iii) ANAa and NODLN only weak effects 

 Reinikainen et al, 

2002 

Daphnia galeata MC Daphnia feeding on toxic 

(wt) and non toxic (mutant does not produce MC) Microcystis strains 

(i) wt toxic to Daphnia: decreased swimming + death, mutant not toxic (ii) both wt and mutant inhibit ingestion rate 

Dose response relationship not between MC content of the food and effects in Daphnia but between ingestion rate and effects 

Rohrlack et al, 1999 

Chapter 32: Cyanobacterial Toxins 731
Exposed 

 organism 

Toxin Exposure route Concentration (dose) 

and effect 

Remark Reference 

Daphnia pulicaria Microviridin 

J 

(i) Grazing on Microcystis strains containing 1.05 – 1.57 μg mm-3 microviridin 

(ii) Purified, dissolved microviridin (0–12 mg L-1) 

Lethal molting disruption Microviridin is a protease inhibitor Rohrlack et al, 2004 

Daphnia spp. MC Daphnia feeding on toxic 

MC(+) wt and non–toxic (MC(-) mutant Microcystis strains 

(i) inhibition of feeding rate, equal inhibition for tox and non–tox Microcystis (ii) reduced survival time (LT50) Daphnia when feeding 

on wt (iii) Daphnia feeding on mutant signs starvation 

(i) MC major source of acute Daphnia poisoning (ii) clear relationship – dose–response – between LT50 and 

MC ingestion rate 

Rohrlack et al, 2001 

Daphnia galeata MC Daphnia feeding on MC(+) 

and MC(–) strain 

Toxic strain 0.87 mg L-1 MC: 

(i) both tox and non–tox Microcystis negatively affect the cohesion of midgut epithelium (within 9 h) (ii) tox strain: uptake MC in blood, increase 0.25 to 1 ng L1 (6–9 h) 

(iii) tox strain: decreased beat rates (5–9 h), constant contraction midgut, finally complete loss beat rates, death (32–41 h) 

(i) midgut disrupting factor is not MC, disruption stimulates uptake bioactive compounds from cyanobacteria in blood Daphnia (ii) results not easily translated to field 

Rohrlack et al, 2005 

Daphnia magna MC (i) Daphnia exposed to suspensions of toxic Microcystis cells 

(ii) Daphnia exposed to lake water (filtered or not) 

MC=2000 μg g DW-1: 

reduced survival, fecundity and growth 

Negative correlation between toxin content and growth (strong) and between toxin concentration and fecundity (weak) 

Thostrup & Christoffersen, 1999 

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