Measured concentrations of cyanobacterial toxins in freshwater fish and aquatic invertebrates eaten by humans. Adapted, in part, from Ibelings and Chorus (2007).
Cyanobacteria (or “blue-green algae”) are among the oldest known groups of organism on Earth, with fossil records spanning approximately 3.5 billion years, and inhabit nearly every ecological niche on the planet. In addition to their conspicuous occurrence as aquatic “blooms” (e.g. visible scums on ponds and lakes; Fig. 1), as well as various colonial or macrophytic forms, in marine and freshwater habitats, these photosynthetic prokaryotes are widely found in terrestrial soils, and as part of numerous symbiotic relationships with a range of organisms including animals, plants and fungi. To illustrate the extent of their global abundance and importance, approximately 50% of all primary productivity, and related oxygen production, occurs in the ocean, and the majority of this is derived from two genera of cyanobacteria,
Alongside their global biological importance, the cyanobacteria are widely recognized as producers of a chemically diverse array of biologically active secondary metabolites (see, for example, reviews by Gerwick et al., 2001; Tan, 2007; Tan, 2010). Considerable work over the past four decades (see Tan, 2007) has, in particular, focused on exploring this chemical diversity as a source of bioactive compounds with possible relevance to biomedicine, and specifically development of potential chemotherapeutics. Furthermore, a rather convincing body of evidence (Proksh et al., 2002; Simmons et al., 2008) also suggests that many of the pharmacologically active marine natural products, originally isolated from various marine animal sources – and including several currently either in clinical trials, or being commercially developed as drugs - may, in fact, originate from cyanobacterial (or other microbial) sources as a result of trophic transfer (e.g. herbivory, filter-feeding), and symbiotic or commensal relationships.
In addition to their potential in biomedicine, however, a number of these bioactive metabolites have been identified as naturally occurring toxins, and have been associated – as so-called “cyanotoxins” - with various human and environmental health concerns. Perhaps most notably, in freshwater systems, cyanobacterial populations can proliferate to the extent that they form large “blooms,” typically manifesting as “films” or “scums” on lakes, ponds and other freshwater systems (see Fig. 1). When comprised of toxin-producing representatives, their occurrence is generally categorized as “harmful algal blooms” (HABs), or frequently as “cyanoHABs” (to distinguish from similar blooms of several, unrelated, but likewise toxigenic, marine microalgae). In particular, toxins from cyanoHABs - or even simply high abundance of cyanobacterial cells – can contaminate water, and exposure to toxins via drinking water, recreational exposure and related routes has been linked to various cases of human and animal intoxication, as well as possible sub-acute and/or chronic health effects (e.g. increased rates of certain cancers, effects on fetal development). As these direct routes of exposure are beyond the scope of this chapter, and have been thoroughly covered by many previous authors, the reader is direct to several good reviews on the topic (e.g. Rao et al., 2002; Stewart et al., 2006; Funari and Testai, 2008).
Although the vast majority of studies related to the health impacts of cyanobacteria have focused on direct exposure to cyanotoxins via drinking water and related routes, there is a growing body of evidence to suggest that toxic cyanobacterial metabolites can bioaccumulate in aquatic food-webs, and may consequently pose additional health concerns as food-borne contaminants. The relatively limited number of studies on food-borne cyanobacterial contaminants may be attributed, in part, to the perceived lack of a mechanism for their bioaccumulation. Unlike more lipophilic contaminants, many of the recognized, water-soluble cyanotoxins would, as such, not be expected to biomagnify by otherwise well-documented mechanisms (i.e. storage, and subsequent transfer, in fatty tissues of animals) to higher trophic levels most frequently consumed by humans. However, despite the lack of a clear means transfer of these hydrophilic toxins in food webs, numerous studies have, indeed, demonstrated presence and apparent bioaccumulation in a range of trophic levels. Also likely limiting the attention paid to cyanobacterial toxins in food-webs is the fact that best documented cases of intoxication have been generally limited to direct exposure to these toxins, and specifically acute human or animal poisonings with clear links to consumption of contaminated water, or various related route, whereas there are - at present - few, if any, clear cases of recognized human intoxication by food-borne cyanotoxin. That said, growing recognition that cyanobacterial toxins may contribute to a sub-acute and/or chronic health effects – ranging from increased rates of cancers, neurodegeneration and development toxicity - which are considerably more difficult to identify, would suggest that, despite the lack of currently documented toxicoses, health threats posed by diet-derived toxins remains a very real concern.
The following chapter will present the current state of knowledge regarding the bioaccumulation of cyanobacterial toxins in the food web, and the possible role of these food-borne toxins as it relates to human and environmental health. To begin, the chapter will present a brief summary of the recognized cyanobacterial toxins, and their known toxicology and health effects. Subsequently, the current evidence related to the bioaccumulation, trophic transfer and bioavailability of these cyanobacterial toxins in food webs will be reviewed, along with related methodologies (including methodological limitations and innovations) for investigating these aspects. In addition to the widely recognized water-soluble cyanotoxins, cyanobacteria produce a host of bioactive metabolites, including a number of lipophilic representatives. Accordingly, the discussion will include a consideration of the less characterized bioactive metabolites that, despite relatively unknown health effects, may represent – due to their potential for biomagnification –relevant food-web contaminants. Finally, the chapter will summarize the current state of knowledge regarding the impacts of cyanobacterial toxins as it relates to human and environmental (i.e. ecosystem, animal) health.
2. Recognized cyanobacterial toxins: Chemistry and toxicology
Detoxification of a wide range of toxic metabolites occurs - via multiphasic enzymes, and associated processes (e.g cellular transporter and “pumps”) - in the liver or equivalent organ systems in animals. Accordingly, many toxic metabolites are actively transported (for subsequent detoxification) to, and thus accumulate primarily in, hepatocytes. Not surprising, therefore, two of the most commonly recognized cyanobacterial toxins are, in fact, associated with hepatotoxicity. However, aside from active transport of these toxins to – and consequent toxicity in - hepatocytes, it is becoming increasingly clear that the same metabolites may accumulate in a range of tissues (even if not associated with acute toxicity in these cells), and may – along with their generally uncharacterized toxicology in these tissues - be thusly transferred to higher trophic levels.
2.1.1. Microcystins (MCs) and nodularin (NOD)
Perhaps the most widespread, and consequently well studied, of the cyanobacterial toxins, microcystins (MCs) and nodularin (NOD) are, respectively, hepatotoxic hepta- and pentapeptide toxins. Both share structural similarity (Fig. 2), specifically characterized by a peptide macrocyle incorporating common and unusual amino acids. However, the former (i.e. MCs) represents a chemically diverse group of toxins, comprised of more than ninety variants (Welker and Van Dohran, 2006). Although structural variation throughout the macrocycle of the MCs has been reported, the primary differences occur in “X” and “Y” positions (Fig. 2), as per the accepted nomenclature for the group. As an example, the most common, and generally considered the most toxic, of these variants is MC-LR in which the X and Y positions, respectively, are occupied by leucine (L) and arginine (R) residues. Although chemical variations exist, both NOD and most MC variants are characterized by a relatively well-conserved unusual β-amino acid, 3-amino-9-methoxy-2,6,8-trimethyl-10-phenyldeca-4,6-dienoic acid (Adda), which is involved (see below) in the toxicology of these metabolites (Gulledge et al., 2003). Finally, whereas NOD is generally limited to a single late summer blooming species,
Toxiologically, both NOD and the MCs are inhibitors of Ser/Thr type 1 and 2A protein phosphatases (PPases). Data generally suggest that Adda of NOD and MCs are involved in the binding of the metabolite to the active site of PPases (Nishiwaki-Matsumishi et al., 1991; Gulledge et al., 2003). To demonstrate this, several analogues of the MCs, specifically comprised of only a single amino acid (i.e. Gly, or L- or D-Ala) coupled via peptidyl linkage to the carboxylic acid of the
Given the importance of PPases in a wide range of cellular functions, inhibition of these enzymes, following exposure to NOD/MCs, can result in a range of acute toxicoses. Accumulating primarily in hepatocytes (i.e. liver) and associated organ systems, inhibition of PPase by MCs and NOD most typically manifests as acute failure and hemorrhaging in these systems. However, recent studies, specifically pointing to the presence of similar active transporters in mammalian (e.g. rat) brains, have proposed a possible connection between MC uptake and apparent inhibitory effects on short- and long-term memory (Maidana et al., 2006). Moreover, emerging evidence supports an additional role of MCs in various chronic health effects, and particularly, as recognized tumor promoters, increased rates of certain cancers. Most notably, studies in China (Yu, 1995; Ueno et al., 1996; Yu et al., 2001) have linked chronic exposure to MC through ingestion of contaminated surface (i.e. ditch) waters to endemically high rates of primary liver cancers. These studies suggest, in particular, that health concerns associated with exposure to even quite low (e.g. sub-picogram per day) doses of cyanobacterial toxins, such as MC/NOD, may promote negative health effects that might not clearly manifest as acute intoxication. This may be particularly germane to discussion of the bioaccumulation of these toxins since there have been to-date no known cases of overt intoxication from food-borne MCs or NOD, whereas the possible long-term effects associated with chronic exposure to these toxins (and others, e.g. BMAA) in the diet continues to present a possible concern.
2.1.2. Cylindrospermopsin (CYN)
Cylindrospermopsin (CYN) is a zwitterionic tricyclic alkaloid, specifically containing a unique hydroxyuracil (Fig. 3). It was first identified following a relatively large intoxication event (the so-called “Palm Island Mystery”) in Queensland, Australia. In this original case, children from more than one hundred families on Palm Island, and nearby mainland community of Townsville, were stricken with severe gastroenteritis. Subsequent studies (Bourke et al., 1983; Hawkins et al., 1985) linked the illness to the Solomon Dam – the primary water reservoir for the community – and identified several bloom-forming species of cyanobacteria. Among these, a toxic (in mouse bioassay) strain of the species,
Toxicological studies primarily suggest that CYN is an inhibitor of protein synthesis (Terao et al., 1994; Froscio et al., 2001). Specifically, Terao et al. (1994) demonstrated ribosomal detachment (from endoplasmic reticula) in hepatocytes treated with CYN, and
Several of the prominent cyanobacterial toxins are known to presumably cross the blood-brain barrier, and have been consequently associated with neurotoxicity. Neurotoxic cyanotoxins have been particularly identified based on observation of acute toxicity following exposure to these toxins (see below). However, in at least one case (i.e. BMAA; see below), toxicity has been associated with possible chronic neurodegeneration.
2.2.1. Anatoxin-a (ATX-a) and anatoxin-a(s)
Although chemically unrelated, two of the most active neurotoxins produced by cyanobacteria are related both in name and mode of action. First identified from species of
2.2.2. Saxitoxin (STX) and “Paralytic Shellfish Toxins” (PSTs)
Potent inhibitors of voltage-gated sodium channels, saxitoxin (STX) and several chemically related metabolites have been frequently associated with contamination of shellfish, and consequent toxicity (i.e. “paralytic shellfish poisoning” [PSP]), as the so-called “paralytic shellfish toxins” (PSTs). Specifically, in marine systems, the origins of STX/PSTs have been identified as
STX/PSTs are potent inhibitors of voltage-gated sodium channels in neuronal cells, specifically acting on (via binding to, and consequent blockage of) sodium passage through channel pores (Aráoz et al., 2010). Inhibition of sodium channels, by blocking sodium influx involved in the propogation of action potentials in neurons, leads to the aforementioned PSP syndrome which manifests in a range of neurotoxic symptoms including numbness, tingling, weakness and difficulty breathing as a result of the neuromuscular paralysis (Etheridge, 2010). Interestingly, STX/PST binds to the nearly identical location as the equally potent neurotoxin, tetrodotoxin (TTX), associated with poisoning by consumption of several species of pufferfish (Stevens et al., 2011), and STX has, in fact, been identified alongside TTX in pufferfish (Nakamura et al., 1984; discussed below). Notably, despite the identification of STX/PSTs from numerous cyanobacteria species found in freshwater sytems, as well as apparent bioaccumulation of presumptively cyanobacteria-derived toxin in fish and shellfish consumed by humans, reported poisoning by the PSTs has been generally limited to ingestion of shellfish contaminated by apparent marine dinoflagellate sources.
2.2.3. β-Methylamino-L-alanine (BMAA)
As perhaps the best studied case of apparent long-term toxicity resulting from a cyanobacterial toxin bioaccumulated within food webs, the non-essential, non-protein amino acid, β-methylamino-L-alanine (BMAA; Fig. 6) has been linked to high rates of the otherwise rare amyotrophic lateral sclerosis (ALS), and possibly other related neurodegenerative diseases (e.g. Parkinson’s Disease, Alzheimer’s Disease). Indeed, the first reports of BMAA as a neurotoxic cyanobacterial metabolite specifically stemmed from studies of extraordinarily high rates of ALS amont the indigenous Chamorro populations on the island of Guam. BMAA was originally identified as a plant-derived natural product, and specifically found in non-flowering plants of the genus,
Toxicologically, BMAA is a recognized agonist of glutamate receptors. Staton and Bristow (1997) found that BMAA excited glutamate receptors, leading to apoptotic and necrotic cell death, in cerebellar granule cells. Subsequent studies (e.g. Rao et al., 2006; Lobner et al., 2007; Cucchiaroni et al., 2010) have confirmed a similar effect in a range of relevant neurons (e.g. spinal motor neurons, cortical neurons, dopaminergic substantia nigra pars compacta cells). More recently, however, it has been proposed that BMAA – as an amino acid – may become erroneously incorporated via translation into proteins. One of the hallmarks of several related neurodegenerative diseases (including ALS, Alzheimer’s Disease, Parkinson Disease/Dementia) is the formation of misfolded protein aggregates, and it has consequently been proposed that the possible mis-incorporation of BMAA may represent an alternative mechanism of action for this putative toxin.
Although first identified in the Chamorro/ALS case, potential health concerns associated with BMAA have continued to grow with recent reports of a widespread occurrence of the metabolite among cyanobacteria, and its bioaccumulation in a wide range of systems, as well as additional epidemiological findings that link the compound to a complex of related neurodegenerative diseases. In a study by Cox et al. (2005) chemical analysis of a wide range of cyanobacteria, including marine, freshwater and terrestrial representatives, indicated that as many as 95% of the genera produce BMAA, and pointed to a potentially widespread occurrence of the metabolite. Since this study, the analytical techniques used with respect to BMAA have rapidly evolved (see
3. Evidence for bioaccumulation of cyanobacterial toxins in aquatic food-webs
Toxins from a number of marine HAB species – particularly including diverse eukaryotic taxa within the dinoflagellates and diatoms (Bacilliariophyta) – bioaccumulate and/or biomagnify in marine food-webs, and have been clearly linked to contamination of fish and seafood, and consequent intoxication of humans and wildlife (Van Dolah, 2000). Notably, marine algal toxins are (1) frequently associated with commercially important seafood species, including filter-feeding/grazing shellfish (e.g. clams, mussels) and several plantivorous fish species, and/or (2) alternatively characterized by relatively high lipophilicity enabling uptake and storage in fat tissues as a means of biomagnifications to higher trophic levels, including marine fish species eaten by humans (Van Dolah, 2000). Relevant examples of the former include dinophysotoxins and various other metabolites associated with “diarrhetic shellfish poisoning” (DSP), domoic acid associated with “amnesic shellfish poisoning” (ASP) and contamination of shellfish by so-called PSP toxins (i.e. STX and other PSTs, see above) derived from dinoflagellates (e.g.
Compared to marine HAB toxins, bioaccumulation of cyanobacterial toxins in food webs, and its consequent relevance to human and environmental health, has been relatively much less studied. There are likely several reasons for this. The most obvious is that, in contrast to the well-documented contamination of fish and other seafood by marine algal toxins, there are very few recognized cases of acute human or animal intoxication via consumption of bioaccumulated cyanobacterial toxins. It is further proposed that this may be due, in part, to less commercial fishing in freshwater water habitats - and thus consumption of freshwater fish and shellfish - compared to marine fisheries. Indeed, a recent report by the United Nations’ Food and Agriculture Organization (FAO) estimated the 2008 global fisheries catch as approximately 90 million tonnes, but it was comprised of only a “record 10 million tonnes from inland waters,” compared to more than 80 million tonnes from marine sources (FAO, 2010).
Regardless of the relatively limited focus on cyanobacterial toxins in freshwater fish and shellfish, a growing body of knowledge – summarized in Table 1 – does, in fact, support the occurrence of cyanobacterial toxins in a range of trophic levels, including species with direct potential for human exposure, as well as possible implications for ecosystem health (see
From inspection of the available data on the accumulation of cyanobacterial toxins in fish and shellfish (Table 1), it is clear that MCs are, by far, the most commonly reported. Indeed, MCs are generally considered the most widespread of the freshwater cyanobacterial toxins. Frequently, these data are reported in terms of “MC-LR equivalents,” as typical quantitative analyses (e.g. ELISA, LC-MS) use this common variant as a reference standard, despite the fact that as many as ninety variants have been reported (see Welker and von Döhren, 2006). In addition to being among the most commonly detected of the microcystins, MC-LR is also one of the most toxic variants (Zurawell et al., 2005). That said, studies suggest variability in the uptake and detoxification of the variants. Xie et al. (2004), for example, studied MC-LR and MC-RR distribution and depuration in phytoplanktivorous carp, and proposed, based on these findings, a possible preferential uptake of MC-RR, or inhibited uptake of and/or active mechanism to “degrade” MC-LR.
|Cazenave et al., 2005|
|Silver Carp (
|Chen et al., 2005
Shen et al., 2005
|Li et al., 2004
Berry et al., 2011a
Osswald et al., 2007
|Berry et al., 2011a
||MC-LR (eq)||Wholec||0.0185||Berry et al., 2011a|
|Redbreast Tilapia (
||MC-LR (eq)||Muscle||0.002-0.337||Magalhaes et al., 2001|
||MC-LR (eq)||Muscle||0.102||Mohammed et al., 2003|
|Blue Tilapia (
|Berry et al., 2011b|
|Berry et al., 2011b|
||NOD||Muscle||0.0005-0.1||Sipia et al., 2006|
||NOD||Muscle||0.0004-0.2||Sipia et al., 2006|
|Wood et al., 2006
Osswald et al., 2011 (experimental studies)
|Yellow Perch (
||MC-LR (eq)||Muscle||0.0008 (max)d||Wilson et al., 2008|
|Unidentified species||MC-LR (eq)||Muscle||0.04||Magalhaes et al., 2003|
|NOD||Muscle||0.0007-0.025||Van Buynder et al, 2001|
|Chen & Xie, 2005a|
||STX/PSTs||Wholec||2.6||Pereira et al., 2004
||STX/PSTs||Wholec||57||Negri & Jones, 1995
|Chen & Xie, 2005a|
|Chen & Xie, 2005a|
|Chen & Xie, 2005a|
Osswald et al., 2008 (experimental study)
|Unidentified mussel species||CYN||Wholec||0.247||Saker et al., 2004|
|Unidentified mussel species||NOD||Wholec||2.5||Van Buynder et al., 2001|
|Apple Snails (
|Berry and Lind, 2010|
|Crustaceans: Shrimp, Crab
||MC-LR (eq)||Muscle||0.005 (mean)
|Chen & Xie, 2005b|
|Red Claw Crayfish (
|Saker & Eaglesham, 1999|
|Freshwater Shrimp (
|Chen & Xie, 2005b|
|Freshwater Shrimp (
|Chen & Xie, 2005b|
|Unidentified crab species||MC-LR (eq)||Muscle||0.103||Magalhaes et al., 2003|
|Unidentified prawn species||CYN
|Saker et al., 2004
Van Buynder et al., 2001
As shown in Table 1, concentrations of MC in these tissues are generally quite low, and might imply a consequently low concern with respect to human exposure. However, there is evidence – as discussed above - to suggest that chronic expoure to low levels of these toxins may pose concern for long-term health (e.g. increased rates of cancer). Moreover, not shown in this table is the generally higher accumulation of MCs by liver and associated organ systems due to active transport of these toxins to hepatocytes and related cells (as discussed above). Although, in the case of fish, in particular, muscle tissues (i.e. “flesh,” e.g. filets, etc.) are most typically eaten, there are exceptions. Berry et al. (2011a), for example, evaluated the MC content (see Table 1) of fish caught from a persistent cyanobacterial bloom in Lake Patzcuaro (Mexico), and specifically reported considerable levels for those fish (i.e. “charales” and
Bioaccumulation, however, is not limited to the MCs, and a growing number of studies (see recent review by Kinnear, 2010) have, for example, also reported variable levels of the hepatotoxic CYN in relevant fish and shellfish species (Table 1). In fact, soon after the identification of CYN as the toxin responsible for the Palm Island Mystery (Ohtani et al., 1992), Saker and Eaglesham (1999) reported quite high levels of the toxin in both fish (“Rainbow Fish,”
Although not as well recognized (nor investigated), emerging evidence suggests that cyanobacterial neurotoxins may also accumulate in relevant freshwater species (Table 1). With regards to human health, STX and related “paralytic shellfish toxins” (PSTs) are – based on their well-described association to intoxication via seafood – perhaps of most obvious concern. STX/PSTs have widely documented as contaminants of marine shellfish, and particularly bivalves, representing a recognized concern for public health (Van Dolah, 2000). More recently, there have been increasing reports of STX/PSTs in fish, and particularly species of “pufferfish” (Family Tetraodontidae), alongside the toxicologically related (i.e. voltage-gated sodium channel blocking) tetrodotoxins that have been well described from these species. Similar to contamination of shellfish, however, it has been recently shown (Landsberg et al., 2006) that marine dinoflagellates (e.g.
In addition to STX/PSTs, cyanobacteria are known to produce several other neurotoxic metabolites, including (as discussed above) the toxicologically related ATX-a and anatoxin-a(s). Compared to other cyanotoxins, the neurotoxic ATX-a is generally considered chemically quite labile, and it is generally anticipated that the potential for bioaccumulation of this unstable toxin would be, accordingly, rather low. That said, in experimental studies, it has been shown that both fish – including trout (
Similarly, despite the emerging picture of its biomagnification in terrestrial species (e.g. fruit bats feeding on cycads; see above) over the past several decades, as well as the particularly conspicuous abundance of cyanobacteria in aquatic systems, relatively limited attention has been paid to the possible bioaccumulation of neurotoxic BMAA in aquatic food-webs. Several recent studies (Jonasson et al., 2010; Brand et al., 2010; Mondo et al., 2012), however, have suggested both accumulation, and possible biomagnifications of BMAA in marine systems, including those species (i.e. fish, seafood) directly related to human health. In one very recent case, the fins of several species of sharks, as “apex” marine predators, were examined, and found to be laden with BMAA (Mondo et al, 2012), and consequently proposed to present – via widespread consumption in the form of “sharkfin soup” – a potentially important route of exposure to this toxin, and thus a public health concern, in Asian countries where shark fins are considered a delicacy. Likewise, one of these studies, specifically evaluating BMAA in South Florida waters, and more specifically including Caloosahatchee River, did, in fact, detect this putatively toxic amino acid in both invertebrate (i.e. mussel) and fish species, including those consumed – at least occasionally - by humans (e.g. bass, bowfin, alligator gar) in this freshwater system. Most interestingly perhaps, it was found in this, as well as concurrent studies of marine food-webs, that measured BMAA levels were, in fact, higher in higher trophic levels suggesting the possibility of biomagnifications of this metabolite. As a highly water-soluble amino acid - with a low octanol/water-partitioning coefficient - it is not expected that BMAA would biomagnify by conventional means (i.e. via deposition in fat bodies, etc.); however, alternative mechanisms to this end are proposed (discussed below).
Finally, it bears mention that a growing number of studies have documented apparent uptake of cyanobacterial toxins by various plant species via toxin-contaminated irrigation water. Uptake of cyanobacterial toxins by plants was first suggested in a study by Pflugmacher et al. (2001) that reported both uptake - and associated metabolism - of MC-LR by the water reed (
4. Trophic transfer and bioavailability of cyanobacterial toxins
Despite emerging evidence to suggest the bioaccumulation of cyanobacterial toxins within food webs (as summarized above; Table 1), relatively little is known regarding the process of trophic transfer, and the subsequent bioavailability of “food-derived” cyanotoxins. For lipophilic contaminants, including recognized anthropogenic pollutants (e.g. PCBs, DDT) and even some HAB toxins (e.g. ciguatoxins), uptake and storage in fat tissues have been largely implicated as a means of trophic transfer. However, there is no clear mechanism for bioaccumulation and/or biomagnification of the most widely recognized and, moreover, typically water-soluble cyanobacterial toxins. Likewise, although growing evidence suggests that cyanobacterial toxins are, in fact, present in relevant components of freshwater food webs (see section
4.1. Trophic transfer
As discussed in the previous section, a growing number of studies do, indeed, suggest that cyanobacterial toxins are transferred via dietary/trophic transfer within aquatic food-webs. It has been shown, in particular, that a range of phytoplanktivorous species, including zooplankton, fish, benthic grazers and filter-feeders consume toxin-laden algal cells, and directly accumulate these toxins. Alternatively, it has been shown (Karjalainen et al., 2003) that certain species, specifically including zooplankton, can accumulate, i.e.
The potential for trophic transfer of cyanobacterial toxins is, generally speaking, controlled by three interrelated factors:
A preponderance of evidence, in fact, supports a possible avoidance of toxigenic cyanobacteria by phytoplanktivores. In particular, selectivity with regards to trophic transfer is perhaps best demonstrated by numerous studies that have investigated feeding by zooplankton, and particularly
Even if toxin-containing items are selected for ingestion, however, several lines of evidence suggest both rather limited uptake of diet-derived toxin, as well as active and passive mechanisms for detoxification and/or elimination of these toxins, which together would be expected to limit the potential for trophic transfer. Understanding the contribution of these factors to trophic transfer, requires knowledge of – and/or means to investigate - the physiological, cellular and possible molecular processes involved in both uptake and potential detoxification/elimination. In the cases of MCs, for example, it has been suggested by multiple studies that the gastrointestinal tract, and particularly mid-gut wall, of fish may be an important site for toxin absorption (Chen et al., 2007; Dyble et al., 2011). Given the assumption that bioaccumulation is generally limited to lower trophic levels, it is not perhaps surprising, however, that most insight in this regard has been, likewise, largely limited to phytoplanktivore models. In order, for example, to evaluate the uptake (and subsequent elimination) of MC-LR by fish, in relation to possible human exposure, specifically using the juvenile yellow perch (
As pointed-out, the initial step for trophic transfer (from cyanobacterial cell to grazer) might be expected to represent, in terms of selectivity, uptake and detoxification/elimination, a rather distinct process compared to subsequent uptake by higher trophic levels. To understand this higher-level transfer, therefore, it is necessary to evaluate the role of toxin derived from primary consumer with respect to secondary (and subsequent) consumers. In a particularly elegant example, Karjalainen et al. (2005) experimentally demonstrated uptake of NOD by planktivorous fish larvae (i.e. Northern Pike) and invertebrate (i.e. mysid shrimp,
In addition to understanding physiological, cellular and molecular aspects of potential grazers/predators, uptake and detoxification can also be closely tied to the chemistry of the toxin. Certainly, among the cyanobacterial toxins, the potential for uptake, and subsequent detoxification/elimination, might be expected – due the chemically diverse nature of these compounds – to vary considerably with this chemical variability. This is most obviously exemplified by the distinction of so-called “hepatotoxins” (e.g. CYN, MCs) that, as implied by this classification, and unlike other cyanotoxins (e.g. PSTs, ATX-a, BMAA), are actively transported via characterized organic anion transporter (OAT) proteins to hepatocytes for subsequent detoxification/elimination. However, even with toxin families, variability in uptake and elimination has been reported. For example, in studies on the uptake of MCs, Xie et al. (2004) compared relative distribution following dosing with two common variants, namely MC-LR and MC-RR, in phytoplanktivorous silver carp. Interestingly, dietary exposure to MC-LR and MC-RR (in algal cells) resulted in considerably higher levels of the latter distributed in various tissues, but rather limited tissue concentration/distribution of the former, and more toxic, variant (Xie et al., 2004). Moreover, detection of relative amounts of the two variants in gut and feces specifically supported an apparent barrier to uptake of MC-LR, compared to the less toxic MC-RR (Xie et al, 2004). It should be pointed-out, however, that results in this phytoplanktivorous model differ substantially from similar studies in a generally carnivorous fish model, namely rainbow trout, particularly with respect to apparently rapid uptake of MC-LR by the latter species, and consequently suggest a role of both consumer species, and differential species physiology, relative to the potential for uptake (and subsequent trophic transfer). Finally, uptake (and subsequent trophic transfer) may even be determined, in part, by chemical presentation of the toxin. For example, it has been shown that in a benthic grazer model, namely the snail,
The potential for trophic transfer is not likely limited to the most studied cyanobacterial hepatotoxins (i.e. MCs). Although relatively few studies have evaluated their bioaccumulation, the transfer of the neurotoxic STX/PSTs, for example, has been studied with respect to its accumulation in marine animals, and specifically in relation to non-cyanobacterially (i.e. marine dinoflagellate) derived toxin. As a particularly important vector for PSTs, several filter-feeding mollusks are recognized to accumulate toxin-containing algal cells, and represent a possible route for both direct exposure (i.e. “shellfish” consumption), as well as possible indirect exposure to these toxins (i.e. trophic transfer to, and consumption of, secondary consumers/vectors, e.g. fish). There is, in fact, emerging (albeit currently limited) evidence to suggest that these neurotoxins may be transferred, to some extent, from filter-feeding invertebrates to higher trophic levels. In particular, however, these studies suggest that biotransformation via metabolism of PSTs – represented, as a group, by as many as fifty variants (Wiese et al., 2010) - may be a critical consideration. In studies by Kwong et al. (2006), black sea bream were exposed to green-lipped mussels, previously exposed to the PST-producing dinoflagellate,
Before moving on, to consider bioavailability, perhaps the one exception to the observed pattern of biodilution, which consequently bears discussion, appears to be the trophic transfer of BMAA. In the limited studies that have investigated BMAA in marine and freshwater food-webs, it was shown, in fact, that levels of the toxic amino acid were higher for top trophic levels (e.g. predatory fish) compared to lower trophic levels (e.g. Brand et al., 2010; Jonasson et al., 2010). In a recent study, for example, Jonasson et al. (2010) examined BMAA within food webs of the Baltic Sea, and reported a discernible positive correlation between levels of the toxin and trophic level. It seems likely, though, that the pattern is not quite as simple as classic biomagnification. For example, the Jonasson et al. (2010) and other studies (e.g. Brand et al., 2010) also suggest particularly high levels for benthic versus pelagic species (of both vertebrates and invertebrates). These studies also suggest differences in tissue distribution of the toxin with highest levels of BMAA observed in brain compared to, for example, muscle, and, therefore, underscore the importance of feeding ecology within food-webs, as well as the likely important role of subsequent bioavailability (including uptake and metabolism) and tissue distribution of toxins, in regards to trophic transfer.
Just as selection, chemical availability, uptake and detoxification/elimination would determine trophic transfer of cyanobacterial toxins through food-webs, these factors are, likewise, expected to primarily determine the bioavailability of these toxins to human as would be ostensibly considered – with respect to the current discussion of emerging public health concerns – the “top predator” in this regard. Although not perhaps, strictly speaking, a “bioavailability factor” selectivity with respect to human consumption can certainly contribute to the potential for exposure to food-borne cyanotoxins. In a general sense, several Tiers of selectivity can dictate the likelihood of exposure (in concert with other bioavailability factors) to cyanobacterial toxins in food. As mentioned previously, the generally limited consumption of freshwater fish and shellfish, relative to much more common consumption of fish and other seafood from marine sources, would be expected – given the recognized abundance of cyanobacterial toxins in freshwater systems – to, likewise, generally limit the possible exposure to these toxins. Even within freshwater systems, the relative consumption of fish and shellfish species from lower trophic (i.e. phytoplanktivorous) levels of food webs would similarly contribute to the possible exposure. However, as detailed above (
Cyanobacterial toxins, as discussed previously (see
Although the potential for bioavailability of cyanobacterial toxins to target organs is implied by their patterns of bioaccumulation, and observed toxic effects on certain organ systems, as well as limited number of
Rather, as with other aspects of health concerns regarding cyanobacterial toxins, the very few studies that have considered bioavailablity of these toxins – and implicitly uptake and detoxification/elimination as key factors - have generally relied on data, and subsequent inferences, extrapolated from water-borne cyanotoxins, including dissolved or algal cell-derived toxins. Most notably, several authors have considered World Health Organization (WHO) guidelines regarding acceptable concentrations of MCs - as the clearly most widespread cyanobacterial toxin family - in water, and subsequently derived guideline values for total daily intake (TDI) of this toxin. Values of TDI are generally based on observed
Aside from considerations of uptake and subsequent detoxification/elimination, as it relates to bioavailablity, it has become clear that, in certain cases, the potential (or lack thereof) for human bioavailability may be considerably affected by
As discussed earlier in the chapter (see
A preponderance of evidence continues to suggest that bound MCs do, indeed, represent a considerable pool of the toxin, however, very few studies have investigated whether these bound MCs are, in fact, biologically available. To address this question, Smith et al. (2010) recently investigated the potential for digestive enzymes to release covalently bound MC from PPases. Whereas digestive proteases (e.g. trypsin, chymotrypsin, pepsin) were found, as expected, to effectively hydrolyze a control protein (i.e. angiotensin), they had no effect on the cyclic peptides (i.e. MC-LR and MC-LY). Furthermore, based on the assumption that protein-bound MCs could be partially released by these peptidolytic enzymes, the investigators synthesized four Cys-containing MC-oligopeptide adducts, specifically predicted for hydrolytic digestion of the PPase active site by these enzymes, and subsequently evaluated them for toxicity (i.e. inhibition of protein phosphatase). Although inhibition was reduced (compared to MC-LR alone) to approximately 58% for MC-peptide adducts - composed of the cyclic MC-LR covalently bound, via cysteine, to predicted tetra- and nonomeric peptide fragments - this residual biological activity supports the possible bioavailability of potentially toxic bound MCs following protein hydrolysis in the digestive system. Interestingly, concurrent studies (Zhang et al., 2010) evaluated the effects of cooking as an alternative mechanism for release of covalently bound MCs with respect to potential availability of the toxin. In these studies, it was specifically found that levels of MC-LR in carp (injected intraperitoneally with the toxin) were significantly higher (approximately 4-fold) in both muscle tissue and water following boiling, compared to lyophilization and subsequent solvent extraction only, and it was suggested that elevated levels were due to release of covalently bound toxin from these tissues. Although such studies do point to the possible chemical availability of covalent bound toxins, clearly further studies are needed to fully elucidate the possible bioavailability of these in relation to human exposure.
5. Methodologies for evaluating cyanobacterial toxins in the food-web
Techniques for chemical detection and quantitative analysis of cyanobacterial toxins have evolved alongside recognition of their potential health impacts. The majority of the previously established analytical methods (e.g. HPLC-UV, LC-MS, ELISA) have, therefore, primarily focused on the identification of toxin in algal cells and/or dissolved in water, with water (i.e. contamination of drinking water, recreational exposure) being an established direct route of exposure. As for these applications, analytical techniques applied to measurement of toxins in food webs have, likewise, generally focused on two approaches (Sivonen, 2008; Humpage et al., 2010). The first, and arguably most common - given the complex nature of these biological matrices (discussed further below) - has included a number of so-called “hyphenated methods” in which analytical separation, including liquid chromatography (LC) and capillary electrophoresis (CE), in particular, are coupled to one or more suitable detection/measurement technique, including UV absorbance, fluorescent derivatization/detection (FL), mass spectrometry (MS) and electrochemical detection. Alternatively, with the relatively recent commercial availability of enzyme-linked immunosorbent assay (ELISA) kits for several cyanobacterial toxins, as well as growing understanding of the toxicology of these metabolites - and thus development of several biochemical techniques (e.g. protein phosphatase inhibition assays for MCs) - these bioanalytical techniques have been also applied to the evaluation of cyanobacterial toxin in relation to food-webs and bioaccumulation (e.g. Lance et al., 2006; Berry and Lind, 2010; Berry et al., 2011; Berry et al., 2012). However, unlike detection of several non-cyanobacterial, marine algal toxins as contaminants of fish and seafood for which there are validated analytical techniques, there are presently no validated methods for evaluating cyanobacterial toxins in biological matrices.
Both of the aforementioned approaches present potential limitations, but generally speaking, it would be argued that the two consequently complement one another. In particular, ELISA-based methods have been somewhat criticized (Metcalf et al., 2000; Mountfort et al., 2005) as being potentially susceptible to non-targeted molecules (i.e. matrix components) in samples that may immunologically cross-react with antibodies, or alternatively not being able to distinguish more toxic variants among co-occurring congeners within toxin groups. As antibodies used in ELISAs are typically generated in relation to a particular representative variant of these toxins, relying on chemical similarity and cross-reactivity to detect other variants, they do not enable co-occurring congeners to be distinguished from total toxin concentrations. This latter limitation is perhaps best exemplified by the ELISA-based analysis of MCs. Although commercially available ELISAs for MCs exploit the generally conserved Adda moiety found in most variants (Fig. 2), MC variants lacking (or containing modified versions of) Adda have been reported (Namikoshi et al., 1990 and 1992; Sivonen et al., 1992; Oksanen et al., 2004), and would be missed in these analyses leading to some degree of “false negatives,” or at least possible underestimation of MC content. Alternatively, studies have shown that improper use of ELISA kits – as well as factors such as organic solvents, salinity and pH - can contribute to the potential for false positives (Metcalf et al., 2005). Additionally, it is recognized that toxicity of MCs varies with congener, and therefore the inability to distinguish particular variants, with respect to this relative toxicity, does not enable what is essentially a proxy of “total MC” to be evaluated in terms of actual relevance to toxicity (Mountfort et al., 2005). As an alternative, enzyme assays - and particularly the various PPase inhibition assays developed for MCs (Mountfort et al., 2005) – are, in fact, capable of assessing cyanobacterial toxins based on relevant biological activity. However, such assays are typically not as sensitive as, for example, ELISA, and likewise may be susceptible to matrix components, as well as being unable to chemically distinguish particular toxin variants. That said, sensitivity of methods such as ELISA are generally higher than for most other methods, and moreover, although coupling analytical separation to detection (e.g. LC-MS) may enable identification of particular chemical variants, this is only applicable to those variant which are specifically targeted. In other words, even though LC coupled to tandem mass spectrometry (MS/MS) can, for example, selectively detect several common MC variants based on characteristic molecular ions, and subsequent “daughter ions,” without prior knowledge of the optimal parameters (i.e. parent/daughter ions, ionization energy, etc.) - and/or availability of suitable analytical standards - to use for other for other less common (or perhaps yet uncharacterized) variants, and their metabolic products, these would be generally missed by such an approach (Mountfort et al., 2005). Accordingly, a strategy which incorporates both approaches and their relative benefits (i.e. highly sensitive detection of “total” MC by ELISA, target-based bioassay and selective analytical separation and detection of individual MC variants) might be expect to provide the most comprehensive toxin profile.
In general, the obvious challenge posed in adapting analytical techniques, originally developed for water (and, to a lesser extent, algal) samples, to bioaccumulation in food webs is the relatively more complex matrix of biological specimens (i.e. animal tissues). Other components of these biological matrices can interfere with analyses by specifically requiring a higher degree of selectivity (to discern the analyte from other components of the matrix), and as well as leading to suppression of the detection response (e.g. suppression of ionization in MS). To some extent, these challenges are inherently addressed when coupling detection to optimized analytical separation (e.g. LC-MS) that essentially isolates components (e.g. as chromatographic
One of the particular challenges of a biological matrix is due to the lack of specificity of certain detection methods. For example, although none of the recognized cyanobacterial toxins discussed have a particularly specific chromophore, as to enable unambiguous detection, HPLC coupled to UV spectrophotometric detection has been successfully used – specifically based on shortwave UV detection and established chromatographic retention time, and in conjunction with analytical standards – to detect and measure several cyanobacterial toxins in water samples (e.g. Harada et al., 1994; Gugger et al., 2005; McElhiney and Lawton, 2005; Berry and Lind, 2010). However, in more complex bio-matrices, this lack of a distinguishing UV chromophore, and potential for co-eluting non-targeted components of the matrix generally limit this approach. Similarly, although fluorescence derivatization – and subsequent fluorescence detection coupled to chromatography or other analytical separation techniques (e.g. CE) - has been used as a highly sensitive means of detection/measurement of cyanobacterial toxins (e.g. Harada et al., 1997; James et al., 1998), the derivatization chemistry frequently employed in these approaches exploit common functional groups (e.g. amines, carboxylic acids, dienes). As such, non-toxin analytes (e.g. peptides) present in biological matrices can be coincidentally derivatized and, by co-elution and/or simply poor resolution, interfere with identification of the analyte of interest. Possible overlap in analytical response is not limited to these analytical separation/detection techniques, and indeed, it has been suggested that for ELISA, antibody cross-reactivity with chemically related components of the matrix might, likewise, lead to non-selective detection, and erroneous results in analyses.
Even in the case of highly selective detection techniques, interference due to the bio-matrix can arise. This is particularly seen with quantitative analyses based on mass spectrometry, including LC-MS, and particularly the most commonly used (at present) method of electrospray ionization (LC-ESI-MS). Components of the sample matrix, including inorganic (e.g. pH, salts/ions) and organic (e.g. proteins and other biomolecules) components, can both interfere with ionization; the former can directly interfere with ionization, whereas the latter can indirectly effect ionization of the analyte, particularly through competitive ionization. In the case of MCs, for example, it has been shown that dissolved organic carbon, pH and ionic strength can suppress signal in LC-MS with the latter being the most significant (Li et al., 2010). More notably in relation to the present discussion, Karlsson et al. (2005) investigated the effect of a biological matrix with respect to the LC-MS detection of MCs and NOD in biological tissues, including aquatic invertebrates (i.e. Blue Mussels,
No chemical information, e.g. identification of chemical variants
No toxicity information
|Highly sensitive (< ppb)
Rapid/analyze multiple samples at once
Easy/little training required
Relatively inexpensive instrumentation (i.e plate readers)
e.g. PPase inhibition
|No chemical information, e.g. identification of chemical variants
Somewhat lower sensitivity
Low selectivity, e.g. false positives
|Provides toxicity information
Relatively inexpensive (reagents)
Rapid/analyze multiple samples at once
Easy/little training required
Relatively inexpensive instrumentation (i.e plate readers)
(i.e. analytical separation/detection)
Low selectivity, e.g. false positives
Limited chemical information, e.g. can’t identify unknown variants
Moderately expensive instrumentation
Relatively slow/can’t analyze multiple samples at once
Requires sample clean-up
|Provides some chemical information, e.g. can identify chemical variants (if standards available)
Analysis can be automated
Somewhat low selectivity, e.g. false positives
Limited chemical information, e.g. can’t identify unknowns
Moderately expensive instrumentation
Relatively slow/can’t analyze multiple samples at once
Requires sample clean-up
Provides some chemical information, e.g. can identify chemical variants (if standards available)
Analysis can be automated
Relatively slow/can’t analyze multiple samples at once
Requires (some) sample clean-up
|Provides chemical information, e.g. can identify chemical variants (if standards available), possible information regarding unknowns
Analysis can be automated
Aside from issues specifically related to the complex biological matrices encountered in food webs, some of the same general challenges associated with quantitative analysis of cyanobacterial toxins in water are, likewise, associated with analysis of bioaccumulation. A common consideration, in this regard, includes a requirement of selectively to detect, and discern, isomers and chemically related congeners. Indeed, this is exemplified by all of the recognized cyanobacterial toxins. For example, both PSTs and the MCs belong to rather large families of chemically related, but structurally distinct, variants – specifically represented, at present, by as many as fifty, and more than ninety, reported variants, respectively - with equally variable toxicity for each. In both cases, especially toxic and common variants (i.e. STX and MC-LR) are most frequently considered as a proxies for these toxin families, however, it is known that several other congeners from these groups can, in fact, potentially contribute to toxicity, and moreover, due to differential uptake and bioavailability hold, likewise, variable potential with respect to bioaccumulation. This presents, as discussed above, a clear limitation to ELISA-based analyses of the MCs that is unable to distinguish specific contributions of individual congeners. Likewise, although LC-MS and related methods, which employ analytical separation prior to detection, are able (if sufficiently optimized) to chromatographically resolve/separate congeners, the requirement for established molecular ionization and fragmentation parameters to detect these variants (by mass spectrometry) limits which variants will, and will not, be detected. Although considerably less complex, on the other hand, both CYN and ATX-a, likewise, have structurally related congeners and/or structural isomers, including for homoATX-a and 7-epiCYN, respectively, that are found alongside the “primary” toxins. In both of these cases, the congeners are also associated with toxicity, and evaluation of their contribution with respect to food-borne toxins needs to be included. On the other hand, although only BMAA (among several structural isomers) has been reported as potentially toxic, it has been suggested that natural occurrence of numerous possible isomeric congeners, all sharing the same molecular mass (Fig. 9), may greatly confound mass spectrometric analysis of this toxic amino acid (Banack et al., 2010), and specifically it has been suggested that, due to this, the once considered widespread occurrence of this neurotoxin may, in fact, be considerably over-estimated (Jiang et al., 2012; Li et al., 2012).
Although many of the cyanotoxins are generally considered chemically quite stable, and thus persistent in the environment, stability has, likewise, been suggested to limit quantitative analysis in some cases - and particularly ATX-a as perhaps the most chemically labile of the known toxins from cyanobacteria. Indeed, ATX-a has an estimated (Stevens and Krieger, 1991) half-life of only about 1-2 days – or even as low as 4-10 hours - in solutions emulating relevant biological conditions, including sunlight (i.e. photolysis) and pH (i.e. acidification). Degradation products, moreover, are generally not considered toxic (Stevens and Krieger, 1991). This instability poses obvious challenges, and particularly the potential for this toxin being underestimated or even missed in chemical analyses, and may, in fact, contribute to the absence of any reports on its bioaccumulation in food webs. Accordingly, it is generally advised that appropriate precautions regarding sample collection, transport and storage be taken to minimize recognized factors (i.e. light exposure, pH), and methods have been developed to increase speed of analysis (e.g. Smith and Lewis, 1987), to minimize this factor in assessing the possible role of ATX-a in relation to food-borne health concerns.
As we learn more about the fate of cyanobacterial toxins in food webs, it is becoming increasingly recognized that at least two of the recognized cyanotoxins, namely MCs and BMAA, may present specific analytical challenges due to their specific chemical toxicology. As previously discussed (see
Perhaps the most elegant of these has been the development of the so-called “MMPB method,” based on detection of the Adda oxidative cleavage product, 3-methyl-4-methoxy-phenylbutanoic acid (MMPB), as a proxy for total (i.e. bound and unbound) microcystins. As stated previously, Adda is both a key component of the MCs with respect to their toxicity, and is, indeed, conserved in the vast majority of variants (Fig. 2). Taking advantage of this conserved moiety, Sano et al. (1992) developed an analytical technique, specifically based on
Similar to the covalent binding of MCs to PPase targets, the possible incorporation of BMAA into proteins, and consequent potential for underestimation, has been addressed in recent – albeit much fewer - studies. Incorporation of BMAA into proteins, with respect to both analytical challenges, and its possible role in bioavailability and trophic transfer of the toxin, was first reported by Murch et al. (2004). In this study, acid hydrolysis (i.e. 24 h boiling in 6 M HCl) – to digest proteins, and release BMAA - coupled to subsequent HPLC-FL analysis measured levels of the toxin in cycad flours exceeding those previously measured, following simple solvent extraction (i.e. presumptively unbound BMAA), by as much as 90-fold. In the same study, BMAA in brain tissues from patients who had died of Alzheimer’s Disease were, likewise, analyzed following acid digestion, and similarly showed levels of the toxin that were 60- to 130-fold great than those measured following solvent extraction (i.e. “free BMAA” only). The presence of this bound pool of BMAA in proteins was accordingly suggested to represent a previously unrecognized reservoir of the toxin available for both trophic transfer (i.e. following proteolytic enzyme digestion), and as a slowly released form of the toxin with respect to the recurrent damage and “latency period” observed in the onset of these neurodegenerative diseases (Murch et al., 2004). Moreover, these results point to a limitation in the prior analytical approach, and specifically a likely underestimation of the toxin in food webs. Accordingly, subsequent studies (Rosen and Hellenäs, 2008; Baptista et al., 2011; Cervantes Cianca et al., 2012) have undertaken development and validation – as well as application - of various analytical techniques (e.g. HPLC-FL, LC-MS/MS and CE-UV) that incorporate acid hydrolysis a means of characterizing “total BMAA” content. As these methods are validated, they will no doubt provide a key tool for evaluating both the ecology (e.g. trophic transfer within food-webs) and toxicology of this cyanobacterial metabolite, and clarify its possible role in human health.
6. Other cyanobacterial toxins and their bioaccumulation
A rather limited number of secondary metabolites produced by cyanobacteria – as discussed throughout this chapter (i.e. MCs, CYN, STX/PSTs, ATX-a and BMAA) – are generally considered, specifically based on association with documented intoxication events, as “toxins.” However, the blue-green algae are, in fact, widely recognized to produce a myriad of biologically active metabolites (Gerwick et al., 2001; Tan, 2007; Tan, 2010). Rather than toxins though, the majority of these chemically diverse, bioactive compounds have been identified as part of “drug discovery” efforts (Tan, 2007). That said, many of the biological systems used to prospect for potential pharmaceuticals – most notably perhaps including cytotoxicity as a means of identifying anticancer drugs – could clearly be extended to include those compounds (i.e. “cytotoxins”) which may have negative impacts on human health, as toxins, rather than, or in addition to, their intended therapeutic targets. Indeed, it has been argued (e.g. Berry et al., 2008) that many (or perhaps even most) of the bioactive secondary metabolites from cyanobacteria – including several investigated as drug leads - are likely produced as
Although limited to no literature exists, due to the nature of their discovery (as drug candidates), on either the potential for either toxicity or bioaccumulation of these many previously identified bioactive metabolites, indirect evidence suggests that as-of-yet unidentified metabolites do, indeed, contribute to toxicity of the cyanobacteria. Specifically, several studies (e.g. Pietsch et al., 2001; Kurmayer and Jüttner, 2009) have compared relative toxicity of pure cyanobacterial toxins to crude extracts, and consistently found higher degrees of biological activity for the latter suggesting an additive or even synergistic role of congeners in these mixtures, including both chemically related (i.e. variants within a toxin family) and/or potentially unrelated, and likely uncharacterized, metabolites. These finding, therefore, generally point to the higher toxicity of metabolic mixtures, and specifically suggest the possibility of unknown toxins which might be relevant to bioaccumulation – and, thus, consequent health concerns - within food-webs. Identification of these additional toxic metabolites will, therefore, be essential to a holistic understanding of the health effects of cyanobacterial toxins, including their potential contribution to food-borne toxicity.
Due, perhaps in part, to the nature of the biological assays used (e.g. cytotoxicity assays, enzyme inhibition assays) that rely on water-solubility of test compounds, the majority of the bioactive metabolites identified from blue-green algae as potential drug candidates have typically focused on rather water-soluble or polar compounds, and particularly the diverse non-ribosomal peptides (NRPs) characteristic of the cyanobacteria (Welker and Von Döhren, 2006). Indeed, a few such NRPs have, in fact, been identified alongside recognized toxins, specifically based on potential relevance to chemical ecology. A particularly notable example is the identification of several chemically related, and apparently quite widespread, peptides – isolated, alongside MCs, from
In marine environments, a particularly salient example is lyngbyatoxin, and severally chemically related cyclic peptides (e.g. aplysiatoxin, debromoaplysiatoxin), isolated from marine species of the genus,
Although the vast majority of bioactive secondary metabolites identified have, for reasons stated above, particularly included a number of peptides and other hydrophilic compounds, the cyanobacteria are known to produce a variety of lipophilic metabolites as well. In support of a possible role of these metabolites in relation to food webs, when evaluating the effects of toxic metabolites from the recognized toxigenic cyanobacterial species,
Although much of the chemistry behind these observations remains to be characterized, one particularly well-studied group of lipophilic metabolites is a diverse family of indole alkaloids produced by members of the relatively widespread, but otherwise understudied, Stigonemataceae (e.g. Raveh and Carmeli, 2007; Mo et al., 2010; Kim et al., 2012; and many more). Although this chemically diverse, and taxonomically restricted, group of metabolites have been generally identified based on antimicrobial activity (e.g. Raveh and Carmeli, 2007; Mo et al., 2012), and in fact, have been linked to possible allelopathy (i.e. inhibition of photosynthetic microbial competitors) in their natural environment, several biological activities supporting animal toxicity have also been reported (e.g. inhibition of RNA polymerase; Doan et al., 2001). This finding, and their seemingly widespread occurrence, raises the question as to whether these quite lipophilic metabolites may contribution to toxicity within food webs, including perhaps human health concerns.
Similarly, considerable work over the past 15 years or more has identified a diversity of lipopeptides, particularly isolated from the widespread marine cyanobacterial species,
More recently – following-up on prior studies (Berry et al., 2009; as discussed briefly above) – an apparently widespread group of toxic, lipophilic metabolites have been identified from several species of otherwise toxigenic cyanobacteria. Utilizing the zebrafish (
7. Implications for ecosystem health
Finally, although beyond the scope of this chapter (and volume), it is worthwhile to consider – before concluding our discussion with respect to public health implications of cyanobacterial toxins in food webs – the apparent contribution of cyanobacterial toxins in food webs with respect to animal and ecosystem health. Obviously, understanding the impacts of these toxins on animal health with respect to ecosystems is essential to understanding the potential for trophic transfer of these toxins. Moreover, insights regarding the potential for human health concerns of food-borne toxins – especially as contaminants of shared food-webs - can be often studied indirectly, and to some extent extrapolated, by understanding the role of these toxins in non-human animals. Indeed, with regards to the toxicity of cyanobacterial metabolites, the first reported case of intoxication by a cyanobacterial bloom was made based on animal health, and specifically reported livestock poisonings associated with pond scums (later identified as cyanobacteria) in a so-called “poisonous lake,” famously detailed in the pioneering works by George Francis in the late 19th century (Francis, 1878).
Given that current evidence generally suggests limited trophic transfer of most of the known cyanobacterial toxins, and consequent restriction of foodborne cyanobacterial toxins to lower trophic levels in food-webs, it is not surprising that toxicity to animals has particularly focused on those vertebrate (i.e. fish) and invertebrates which are exposed either directly (through phytoplanktivory), or through “single vector” (e.g. zooplanktivory), transfer. For example, the previously discussed studies of Karjalainen et al. (2005), not only demonstrated that pre-exposure of zooplankton to cell-free, NOD-containing extracts of
Numerous studies, encompassing essentially all of the other known cyanotoxins, as well as yet uncharacterized, but apparently toxic, metabolites, likewise, have demonstrated the potential toxic effects to various aquatic species exposed at this lower end of the trophic scale. At the level of zooplanktivorous grazers, numerous studies have documented the apparent toxic effects of, not only MCs (as discussed above), but also other cyanobacterial toxins, including CYN (Nogueira et al., 2004), STX/PSTs (Filho et al., 2008) and ATX-a (Sieroslawska et al., 2010) - as well as perhaps other unidentified metabolites (present in extracts) - on
On the other hand, relatively fewer studies to-date have clearly documented toxic effects of subsequent predation on toxin-laden invertebrates (e.g. zooplankton, benthic invertebrates), or phytoplanktivorous fish. However, the limited studies that have do generally point to the potential for toxicity – along with toxin transfer – to higher trophic levels. In one particularly notable study, Qiu et al. (2007) examined – using biochemical and histopathological approaches - four trophic levels, comprised of a phytoplanktivorous, omnivorous and carnivorous fish species in a Chinese lake in relation to a MC-producing bloom. Surprisingly in this case, the most pronounced histopathological signs of toxicity were observed for carnivorous fish, whereas the largest biochemical response (i.e. particularly the production of several antioxidant enzymes/pathway, e.g. superoxide disumutase, catalase, glutathione, glutathione peroxidase), were measured for phytoplanktivorous species. These results suggest not only that carnivores can be exposed to cyanobacterial metabolites, and toxic effects via food webs, but that grazers of cyanobacteria (i.e. phytoplanktivores) may, as such, be specifically adapted to the direct exposure to these toxins.
The toxic effects of cyanobacterial metabolites on higher trophic levels, aside from fish, in aquatic food webs remain quite scarce, despite the fact that numerous taxa, including bird and mammalian species, are recognized as frequent “top consumers” in these systems. However, examples are beginning to emerge in the literature. In an especially insightful example, Miller et al. (2010) recently reported on the apparent toxic effects - including several animal deaths among - among southern sea otters, along the Pacific coast of the U.S., exposed to MCs bioaccumulated by bivalves (i.e. clams, mussels and oysters) consumed by these carnivorous predators. Following an unusually high number of sea otters deaths in the Monterey Bay, and surrounding coastal areas, particularly during the period of 2005-2008, necropsy on these stranded animals was performed. Based on the detection of relatively high levels (up to 348 ppb) of several variants of microcystins, including MC-RR, -LR and -desmethyl LR, in livers of sea otters, along with gross and microscopic pathological indications - particularly in livers of the animals - consistent with MC intoxication, it was concluded that animals had, indeed, died from exposure to this cyanobacterial toxin. The source of the toxin was ultimately traced to outflow from the nearby Pinto Lake into the marine waters of Monterey Bay, and subsequent bioaccumulation of the toxin by mollusks, consumed by sea otters, in the Bay. Characterized as a “super-bloom” of cyanobacteria, levels of MC in Pinto Lake, during this time, were measured as high as 2,100 ppm (more than six orders of magnitude higher than the WHO limit of 1 ppb), and use of Solid Phase Adsorption Toxin Tracking (SPATT) samples enabled tracking of the toxin from the lake toward the Bay. Subsequent laboratory studies confirmed the bioaccumulation of MC - at levels as high 1,324 ppb - by various bivalve (i.e. clam, mussel, oyster) species that make up a primary component of the sea otters’ diet. Indeed, this case is particularly revealing as it not only supports hypothesis that higher trophic levels – including mammalian carnivores - can, in fact, be exposed to toxic (and even lethal) levels of cyanobacterial toxins through food-webs, but also that toxins can transfer not only within ecocystems, but between (in this case, freshwater and marine) systems. Moroever, although levels of MCs, in this case, were exceptionally high, this study further provides – through quantificationof the toxin in (livers of) exposed animals – a first estimation of relevant (i.e. lethal) exposure doses for mammalian consumers in relation to environmental concentration both in water, and vectors (i.e. bioaccumulation in bivalves) of the toxin.
Cyanobacteria are prolific producers of toxic, and otherwise biologically (i.e. pharmacologically) active, metabolites. Although ubiquitous in the environment, the cyanobacteria are arguably most conspicuous, and generally more abundant, in aquatic systems, particularly in association with so-called “harmful algal blooms.” As such, exposure to several cyanobacterial toxins through contamination of drinking water, and related routes, has been clearly linked to both acute toxicoses, including human and animal mortalities, as well as being increasingly tied to several long-term health effects (e.g. cancer, neurodegenerative disease). Consequently, waterborne cyanotoxins are widely acknowledged as a global health concern.
Given this particularly widespread occurrence of the “blue-green algae” in marine and freshwater ecosystems, it is not, therefore, surprising, that bioaccumulation of nearly all of the “recognized” cyanobacterial toxins within relevant species of aquatic food webs has been reported. In light of the enormous human reliance on the world’s oceans and freshwater systems, particularly as a source of food (i.e. fish, seafood), as well as culminating evidence to suggest that global climate change is fueling an apparently rapid increase in cyanobacterial abundance, bloom frequency and perhaps even toxigenicity in aquatic systems, cyanobacterial toxins in food webs, likewise, represent a clearly important human health issue.
As an emerging concern, however, clearly more questions than answers remain, at present, regarding the potential role of cyanobacterial toxins in food webs, particularly in terms of possible human and environmental health concerns. Although a preponderance of evidence indicates that diverse taxa of aquatic animal species do, in fact, accumulate cyanotoxins, the biochemical, physiological and ecological processes that control trophic transfer within food-webs remains to be clarified. Likewise, although trophic transfer of the largely water-soluble “known” cyanobacterial toxins appears to follow a pattern of biodilution, rather than biomagnification to top consumers, a growing number cases indicate that species, known to be consumed as part of human diets, do bioaccumulate significant quantities of these toxins, yet the potential implications for human health remains to be elucidated. This, it is argued is due, in part, to the lack of information regarding bioavailability, as well as limitations and challenges in current analytical methodologies used to assess this contribution.
Although the scientific evidence that does exist strongly suggests a potential (and, in fact, high probability) for human exposure to, and consequent health concerns associated with, food-borne cyanobacterial toxins, the implications with respect to public health policy remains almost entirely to be addressed. Indeed, while our scientific understanding of the bioaccumulation of cyanobacterial toxins in relation to human health currently remains rather limited - particularly relative to many other environmental health concerns – the public health implications, including relevance to policy makers and stakeholders, lags even more so. And, in fact, it is asserted that it is many of the same gaps in knowledge that limit our scientific understanding which, likewise, limit public health policy with regards to food-borne cyanotoxins, and thus addressing these gaps will be critical in this regard. First, and perhaps foremost, is the preeminent need to understand the toxicology of food-borne cyanotoxins, including both laboratory and (currently non-existent) epidemiological studies to elucidate what (if any) health effects exist, and which groups are most likely affected. Similarly, clarifying the actual health effects, including relevant doses, mechanisms of action, bioavailability, etc., will be fudnamental to developing effective regulatory guidelines. Although, as described above, attempts have been made (e.g. Ibelings and Chorus, 2007) to extrapolate current (albeit limited) toxicological knowledge in this regard to possible acceptable levels for food-borne cyanotoxins, these are based entirely on data from water-borne toxins, and are not likely to be accurate in terms of exposure through food. Furthermore, even the provisional guidelines that exist for cyanobacterial toxins in water are only recommendations, and policy will not only need to clarify acceptable levels, but also address monitoring and enforcement of these guidelines. As such, improvements, validation and standardization of methods for chemical analysis of cyanobacterial toxins – toward effective monitoring and enforcement - in food will be key. Continued investigations in these areas will, therefore be of the crucial toward developing a comprehensive picture of this emerging public health concern.
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