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Human Health Risk Assessment Related to CyanotoxinsExposureEnzo Funari a; Emanuela Testai aa Environment and Primary Prevention Department, Istituto Superiore di Sanità,Rome, Italy

Online Publication Date: 01 February 2008To cite this Article: Funari, Enzo and Testai, Emanuela (2008) 'Human Health RiskAssessment Related to Cyanotoxins Exposure', Critical Reviews in Toxicology, 38:2,97 - 125To link to this article: DOI: 10.1080/10408440701749454URL: http://dx.doi.org/10.1080/10408440701749454

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Critical Reviews in Toxicology, 38:97–125, 2008Copyright c© 2008 Informa Healthcare USA, Inc.ISSN: 1040-8444 print / 1547-6898 onlineDOI: 10.1080/10408440701749454

Human Health Risk Assessment Relatedto Cyanotoxins Exposure

Enzo Funari and Emanuela TestaiEnvironment and Primary Prevention Department, Istituto Superiore di Sanita, Rome, Italy

This review focuses on the risk assessment associated with human exposure to cyanotoxins, sec-ondary metabolites of an ubiquitous group of photosynthetic procariota. Cyanobacteria occurrespecially in eutrophic inland and coastal surface waters, where under favorable conditions theyattain high densities and may form blooms and scums. Cyanotoxins can be grouped accordingto their biological effects into hepatotoxins, neurotoxins, cytotoxins, and toxins with irritatingpotential, also acting on the gastrointestinal system. The chemical and toxicological propertiesof the main cyanotoxins, relevant for the evaluation of possible risks for human health, are pre-sented. Humans may be exposed to cyanotoxins via several routes, with the oral one being by farthe most important, occurring by ingesting contaminated drinking water, food, some dietarysupplements, or water during recreational activities. Acute and short-term toxic effects havebeen associated in humans with exposure to high levels of cyanotoxins in drinking and bathingwaters. However, the chronic exposure to low cyanotoxin levels remains a critical issue. Thisarticle identifies the actual risky exposure scenarios, provides toxicologically derived referencevalues, and discusses open issues and research needs.

Keywords Cyanobacteria, Cyanotoxins, Human Health, Toxicological Risk Assessment

1. INTRODUCTIONCyanobacteria are a group of ubiquitous photosynthetic pro-

cariota. They occur especially in surface waters, where theycan tolerate remarkable changes of salinity and temperatureand photosynthesize under conditions of low light intensity, thatis, high turbidity (Rai, 1990). In favorable conditions for theirgrowth (i.e., nutrient availability, temperature, light), cyanobac-teria form blooms, giving rise to biomass accumulation and scum(Ressom et al., 1994). Planktonic cyanobacteria produce, as sec-ondary metabolites, a high variety of toxins known as cyanotox-ins that give rise to some concern for human health. Indeed,cyanobacteria have been included among emerging pathogenicmicroorganisms, even though they do not colonize and invadethe host (OECD, 2005).

Cyanobacteria are characterized by a wide morphologicalvariability (Chorus and Bartram, 1999; Sivonen and Jones,1999). In most cases their cells are surrounded by a gelatinousstratum, which increases their chance to survive even in harshenvironmental conditions (Whitton, 1992).

In the last years many papers have been published, report-ing the occurrence of cyanobacteria in surface waters and envi-

Address correspondence to E. Funari/E. Testai, Istituto Superioredi Sanita, Environment and Primary Prevention Department, VialeRegina Elena, 299, 00161 Roma Italy. E-mail: [email protected];[email protected]

ronmental factors influencing their production, reviewing tox-icological and ecotoxicological properties of selected toxins,and reporting methods of detection (van Apeldoorn et al., 2007;Codd et al., 2005; Dittmann and Wiegand, 2006; Zurawell et al.,2005; Duy et al., 2000). The readers are invited to consult thecited papers for a detailed description of those issues, which areonly mentioned or not treated here. Indeed, in this article atten-tion is focused on the evaluation of the risk for human healthassociated with the different sources and routes of exposures tothose cyanotoxins known so far.

Besides this introductory section, the review consists of fourparts: Section 2 very briefly introduces the different cyanobac-teria producing toxins and their occurrence in the environment;section 3 describes the toxicological profile of the main “known”cyanotoxins; from this, in section 4 the risks for human healthderiving from different exposure scenarios are presented; andfinally, in section 5, based on the analysis of the available data,some research needs and open issues are highlighted, althoughthe list is far from exhaustive.

2. TOXIC CYANOBACTERIA AND KNOWNCYANOTOXINS2.1. Cyanotoxins Classification

Cyanotoxins can be classified into categories that reflect theirbiological effects on the systems and organs that they affect moststrongly (Codd et al., 2005).

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98 E. FUNARI AND E. TESTAI

TABLE 1Toxigenic cyanobacteria from marine, brackish and freshwaters

Cyanotoxin Main producing cyanobacteria Bibliographic source

Microcystins Most of Microcystis spp and Planktothrix spp, someAnabaena, Nostoc and Synechocystis andCyanobium bacillare, Arthrospira fusiformis,Limnothrix redekei, Phormidium formosum,Hapalosiphon hibernicus

Sivonen and Jones, 1999; Cronberg et al.,2003; Odebrecht et al., 2002; Ballotet al, 2004; Gkelis et al., 2001;Steffensen et al., 2001; Prinsep et al.,1992

Nodularins Nodularia spumigena (in transitional waters) Rinheart et al., 1988Cylindrospermopsin Cylindrospermopsis raciborskii, Umezakia natans,

Aphanizomenon ovalisporum, Aphanizomenonflos-aquae, Rhaphidiopsis curvata, Anabaenalapponica, Anabaena bergii

Ohtani et al., 1992; Harada et al., 1994;Banker et al., 1997; Schembri et al.,2001; Li et al., 2001; Fastner et al.,2007; Spoof et al., 2006

Anatoxin-a Most of Anabaena spp., some Aphanizomenon (A.flos-aquae, A. issatschenkoi), Cylindrospermum,Microcystis and Planktothrix spp. and Raphidiopsismediterranea

Edwards et al., 1992; Sivonen et al., 1989;Park et al., 1993; Namikoshi et al.,2003; Wood et al., 2007

Homoanatoxin-a Oscillatoria formosa, Raphidiopsis mediterranea Skulberg et al., 1992; Steffensen et al.,2001; Namikoshi et al., 2003

Anatoxin a-(s) Anabaena flos-aquae and A. lemmermannii Carmichael and Gorham, 1978;Henriksen et al., 1997

Saxitoxins (PSP) Aphanizomenon, Anabaena, Lyngbya andCylindrospermopsis spp.

Humpage et al., 1994

LPS endotoxins All cyanobacteria McElhiney and Lawton, 2005Aplysiatoxin,

LyngbyatoxinDebromoaplysiatoxin

Lyngbya majuscula (marine waters), Oscillatorianigro-vridis

Serdula et al., 1982; Mynderse et al., 1997

Microviridin J Microcystis spp Rohrlack et al., 2003β-N-methylamino-L-alanine

Microcystis, Anabaena, Nostoc and Planktothrix sppand most of cyanobacteria symbionts tested

Cox et al., 2005

In this sense, they include:

• Hepatotoxins (more than 70 microcystin variants,6 known nodularine variants).

• Neurotoxins (anatoxin-a, homoanatoxin-a, anatoxina-(s), 21 known saxitoxin variants, known also as par-alytic shellfish poisoning toxins).

• Cytotoxins: cilyndrospermopsin.• Irritants and gastrointestinal toxins: aplysiatoxin, de-

bromoaplysiatoxin, lyngbyatoxin (produced by marinecyanobacteria); lipopolysaccharidic (LPS) endotoxins.

• Other cyanotoxins whose toxicological or ecotoxi-cological profile is still only partially known, suchas microviridin J and β-N-methylamino-L-alanine(BMAA).

At present it is not clear which is the proportion of known versusunknown cyanotoxins.

2.2. Cyanotoxin ProductionEach cyanotoxin can be produced by more than one

cyanobacterial species; likewise, the same species is able to pro-

duce more than one toxin (Table 1). Moreover, within a singlespecies, different genotypes occur, some of which possess thegene for a given toxin and others that do not. This was firstdemonstrated for microcystins (MCs) (Kurmayer et al., 2002).In 50–75% of cyanobacterial blooms, the toxicity is associatedwith a simultaneous production of diverse cyanotoxins (An andCarmichael, 1994), whose relative importance and spatial distri-bution are subjected to a wide variability. The toxicity of a givenbloom is determined by its strain composition, i.e., the relativeshare of toxic versus nontoxic genotypes.

The dynamics of toxigenic cyanobacteria in surface waterbodies can be studied by means of DNA-based tests; availabledata on a German lake indicate that remarkable variations of thegenotype ratio can be found, even on a weekly or biweekly scale(Kurmayer and Kutzenberger, 2003). Therefore, the variations ofthe MC concentrations detected in the lakes might be a result ofpopulation dynamics, altering the proportion of toxic genotypeswithin the population of cyanobacteria (Dittmann and Borner,2005).

The amount of MC production by a cyanobacterial popula-tion in culture appears to be directly proportional to its growth

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HUMAN HEALTH RISK OF CYANOTOXINS EXPOSURE 99

TABLE 2Maximum levels of different cyanotoxins detected in blooms material found in the literature

Cyanotoxins Maximum levels Location References

Microcystins 7,300 µg/g15,600–25,000 µg /l

China and PortugalJapan, Germany

Nagata et al., 1997; Fastner et al., 1999;Chorus et al., 1998

Nodularins 18,000 µg/g Baltic Sea Kononen et al., 1993Cilyndrospermopsin 5,500 µg/g Australia Sivonen and Jones, 1999Anatoxin-a 2,600–4,400 µg/g Finland, Japan Sivonen et al., 1989; Harada et al., 1993Saxitoxins 2,040–3,400 µg/g Australia Negri et al., 1997; Humpage et al., 1994Anatoxin-a (s) 3,300 µg/g Denmark Henriksen et al., 1997

rate, with the highest amount being produced during the late log-arithmic phase; the amount of MCs contained in a single cell ofMicrocystis aeruginosa is constant within a narrow range (two-to threefold) (van Apeldoorn et al., 2007). Beside the populationdynamics, the remarkable variability in MC concentrations inwater bodies has been attributed to environmental condition vari-ation, which can influence cyanotoxin production rate (Sivonenand Jones, 1999).

Yet the role of environmental factors in cyanotoxin produc-tion is not sufficiently understood. Some studies showed thatthe variations of parameters such as light, culture age, temper-ature, pH, and nutrients give rise to differences in the cellularMC content not exceeding a factor of five (Sivonen and Jones,1999).

The production of cyanotoxins by cyanobacteria hasbeen confirmed by laboratory tests using mono-cyanobacteriaisolated cultures; however, the possibility that associatedeterothropic bacteria may have a role in cyanotoxin productionor in its modulation cannot be definitely ruled out (Codd et al.,2005).

2.3. Cyanotoxins in Surface WatersA high number of studies have been published on the oc-

currence of cyanotoxins, particularly MCs, in surface waters,which have been extensively reviewed in dedicated publications(Chorus and Bartram, 1999; van Apeldoorn et al., 2007). Thissubsection only represents a brief synthesis of the available data,relevant to the assessment of the risks to human health associatedwith exposure to cyanotoxins. Specific data on different sourcesof human exposure to cyanotoxins are reported into each specificparagraph.

Cyanotoxins may be localized both within the cyanobacterialcells and dissolved in the water, depending on both the nature ofthe toxin and the growth stage (Chorus and Bartram, 1999; vanApeldoorn et al., 2007).

The highest total (intracellular plus dissolved) cyanotoxinlevels have been found in blooms and scums (Table 2). TotalMC concentrations in surface waters vary in a very wide range ofvalues (from trace to several milligrams per liter), being stronglyinfluenced by the occurrence of these forms of biomass. Total

MC concentrations of 10–50 µg/L, up to 350 µg/L, have beenreported in surface waters in Germany (Fastner et al., 1999), butmuch higher levels (up to 25,000 µg/L) elsewhere (Sivonen andJones, 1999). Total cylindrospermopsin (CYN) and anatoxin-a(s) concentrations up to 12.1 and 3300 µg/L have been deter-mined in surface waters (Rucker et al., 2007; Sivonen and Jones,1999).

The sum of intracellular plus dissolved cyanotoxin level isgenerally the most relevant index of exposure to be considered inrisk evaluation; nevertheless, in some cases it may be necessaryto differentiate between particulate and dissolved form. As anexample, when for drinkin-water purposes simple treatments arein place to remove cells, the levels of dissolved cyanotoxins areof interest in order to evaluate human exposure.

Intracellular MC content is generally higher than that dis-solved in the surrounding water (van Apeldoorn et al., 2007;Ibelings and Chorus, 2007), where they are partially released,probably due to active transport (Rapala et al., 1997). On thecontrary, CYN may often be found at higher levels in dissolvedform than within cells (Rucker et al., 2007); limited or no infor-mation is available about the proportion of dissolved form withrespect to the total level for the other cyanotoxins.

After a collapse of aging, declining blooms or their treat-ment with algaecides, high concentrations of dissolved cyan-otoxins can be found in the surrounding water (van Apeldoornet al., 2007; Jones and Orr, 1994). However, these high levels areusually not long-lasting, due to strong dilution in the water body,wind mixing, adsorption to the sediment, and (bio)degradation.

Indeed, once released in the water, cyanotoxins persist inthe environment, depending on the efficiency of degradation(i.e., photolysis, hydrolysis and bacterial degradation). MCsand nodularins (NODs) can persist in water for relatively longtimes, ranging between 21 days and 2–3 months (Ressom et al.,1994; Jones and Orr, 1994), and up to 6 months in dry scum(Jones et al., 1995). A half-life of 11–15 days has been reportedfor CYN in surface waters (Chiswell et al., 1999). Similarlyanatoxin-a showed a half-life of about 14 days in normal lightconditions with basic pH and low initial concentrations (Smithand Sutton, 1993), whereas a much shorter half-life (1–2 hours)has been shown in the presence of high light intensity (Stevens

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and Krieger, 1991), indicating a high efficiency of the photoly-sis reaction. Anatoxin-a(s) is relatively stable under neutral andacidic conditions (Matsunaga et al., 1989). A persistence of 1–2 months has been reported for saxitoxin (STX) in surface water(Batoreu et al., 2005).

3. CHEMICAL AND TOXICOLOGICAL PROFILESOF CYANOTOXINS

A basic understanding of cyanotoxin chemical properties andtoxicological potential is crucial for the human health risk as-sessment associated with their exposure. Therefore, a brief de-scription of the most relevant features of the different cyanotox-ins is in the following, summarizing general chemical structureinformation, mechanism of action, toxicokinetics, acute and re-peated toxicity data, and, when available, data on genotoxicity,reproductive toxicity and carcinogenicity potential. A synopticview of the major features of main cyanotoxins is presented inTable 3.

3.1. Mycrocistins and NodularinsMicrocystins and NODs are cyclic peptides consisting of

seven (MC) or five (NOD) amino acids. A common charac-teristic of both hepatotoxins is the amino acid Adda, which isunique for cyanobacteria. Structural variations occur by chang-ing of two (MC) or one (NOD) amino acid(s), and several otherchanges in small side groups: these differences give rise to morethan 70 MC variants and about 6 NOD variants (Sivonen andJones, 1999). The chemical structure of MC-LR is shown inFigure 1; this is the most common MC congener, characterizedby the presence of leucin (L) and arginin (R) as L-amino acids inpositions 2 and 4 (Carmichael, 1988). On the basis of acute tox-icity, MC-LR is considered among the most potent hepatotoxinswithin the different variants and is by far the most studied.

Microcystin mechanism of action is associated with specificinhibition of protein serine/threonine phosphatases (PP1 andPP2A), altering phosphorilation of cellular proteins involvedin signal trasduction (Gehringer, 2004). At high levels of ex-posure (representative of acute intoxication), MC-LR producesa cascade of events (cytoskeleton alterations, lipid peroxida-tion, oxidative stress, apoptosis) leading to centrolobular toxic-ity with intrahepatic hemorrhagic areas due to damage of sinu-soidal capillaries. At low doses (typical of long-term exposure),phosphatases inhibition induces cellular proliferation and hep-atic hypertrophy (Gehringer, 2004). MC-LR is able to induceoxidative stress and apoptosis in human cell lines (Botha et al.,2004). The binding of the hepatotoxin to the human hepaticmitochondrial aldehyde dehydrogenases (ALDH2) has recentlybeen shown: Since the alteration of this enzymatic activity isinvolved in oxidative stress induction, it has been hypothesizedthat the enzyme could be one of the main targets of MC-LRmechanism of toxicity in humans (Chen et al., 2006).

The degree of severity of MC-induced toxicity depends onthe levels and duration of exposure, determined by the balancebetween MC absorption, detoxification, and excretion. MC-LR

is highly hydrophilic and can not enter cell membranes by pas-sive transport. It is actively absorbed by the intestinal mucosa,thanks to the organic anion transport system (OATP) and then en-ters hepatocytes through to the activity of bile acid transportersand OATP (Fisher et al., 2005). These active transporters areexpressed also in the kidney and in the blood–brain barrier, par-tially explaining some neurological disorders observed in hu-mans during a fatal incident in Brazil (Azevedo et al., 2002).MC are conjugated with reduced glutathione in the liver ofboth aquatic organisms (Pflugmacher et al., 1998) and mam-mals (Kondo et al., 1992, 1996). The reaction, catalyzed by glu-tathione S-transferases, involves the methyl group of N-methyl-dihydroalanine (opposed to Adda): Conjugates retain only aminimal residual inhibitory activity with respect to the parentcompound and are mainly excreted in the urine (Dittmann andWiegand, 2006).

The MC-LR acute toxicity after intraperitoneal (ip) admin-istration to mice results in a LD50 = 50 µg/kg bw; when givenby the oral route, MC-LR is less toxic (LD50 = 5000 µg/kg bwand even higher in the rat) (Fawell et al., 1994, 1999a). Thelower acute toxicity (30- to 100-fold) shown by the oral route islikely due to toxicokinetic differences: The active transport sys-tem used for the absorption through the gastrointestinal (GI) mu-cosa is bypassed by ip injection and MC-LR is directly availablefor internalization into hepatocytes. This observation is partiallysupported by studies on organ distribution after oral and ip ad-ministration in mice, indicating a 80-fold difference in hepaticcontent of radiolabeled 3H-dihydro-MC-LR (Nishiwaki et al.,1994; Robinson et al., 1989; Robinson et al., 1991). Therefore,ip administration is not fully representative of the actual condi-tions of human exposure, mainly associated with consumptionof possibly contaminated drinking water and food, and has onlya limited value for risk assessment.

Acute toxicity is highly variable among MC variants: Someof them, such as MC-LA, -YR, and -YM, show ip LD50 similarto MC-LR; for the other congeners, LD50s are spread in a widerange of values (from 50 up to 1200 µg/kg) (Table 4), due to thepresence of different substituents. As an example, for MC-RR,containing polar amino acids in positions 2 and 4, the LD50 is 10-fold higher than for MC-LR (Kotak et al., 1993; Wolf and Frank,2002), whereas the presence of hydrophobic amino acids, suchas alanine or phenylalanine, does not affect the acute toxicity(Zurawell et al., 2005).

Adda group stereochemistry and its double bonds configura-tion are crucial for MC-induced PP1 and PP2A inhibition, due toa covalent binding between a protein cysteine and the Adda-glugroup (Harada et al., 1990). However, Adda per se is not ableto inhibit PP1 and PP2A, and it is not toxic when ip injectedin mice even at very high doses (10 mg/kg bw) (Harada et al.,2004).

Freeze-dried algal aqueous suspensions from both Microcys-tis and Anabaena blooms showed very low potential for skinirritation, and gave contrasting results for eye irritation, whileclear positive results were obtained in the skin sensitization test.

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TAB

LE

3M

ain

toxi

colo

gica

ldat

aof

som

ecy

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s

i.p.L

D50

Ora

lLD

50Ta

rget

orga

nan

dm

echa

nism

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OE

LA

DI/

TD

IC

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n(µ

g/kg

b.w

.)(µ

g/kg

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tion

NO

EL

(µg/

kg/d

)(µ

g/kg

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(µg/

kg/d

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Mic

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and

PP2A

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ases

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bitio

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tivity

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(MC

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;mic

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wee

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gava

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330

(MC

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inB

GA

Sex

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mic

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200

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(UF

=10

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ND

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(ref

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MC

-LR

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D—

Cyl

indr

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2100

(24

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Lim

ited

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nic

risk

ND

0.51

(UF

=10

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Hom

oana

toxi

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330

ND

Sim

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toan

atox

in-a

ND

Lim

ited

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nic

risk

——

Ana

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)10

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263

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Acu

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LPS

End

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40–1

90m

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Not

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FIG. 1. Chemical structure of microcystin-LR, characterized by the presence of leucin (L) and arginin (R) as L-amino acids inpositions 2 and 4.

However, components other than cyanotoxins are likely to bepresent in the algal extract, with possible irritation and sensi-tizing potential. Indeed, no correlation was found between thetoxin content and the allergenic character; pure MC-LR showedonly slight skin sensitizing potential, even when tested at highconcentrations (1.5 mg/ml) (Torokne et al., 2001). In addition,the axenic strains were not allergenic at all (Torokne et al., 2001).

Among the available studies on repeated toxicity, the mostrelevant one has been carried out with mice, shown to bemore susceptible to MC-induced acute effects than other ro-dent species. Microcystin-LR was administered orally (by gav-age) for 90 days at 3 different doses (Fawell et al., 1994). The

TABLE 4Differences in acute toxicity of MC variants

Toxin i.p. LD50 (µg/kg) M.W. Structure∗

MC-LR 50 994 cyclo-(D-Ala-L-Leu-D-MeAsp-L-Arg-Adda-D-Glu-Mdha-)[D-Asp3]MC-LR 50 970 cyclo-(D-Ala-L-Leu-D-Asp-L-Arg-Adda-D-Glu-Mdha-)[L-MeLan7]MC-LR 1000 1115 cyclo-(D-Ala-L-Leu-D-MeAsp-L-Arg-Adda-D-Glu-L-MeLan-)[6(Z)-Adda5]MC-LR >1200 994 cyclo-(D-Ala-L-Leu-D-MeAsp-L-Arg-6(Z)Adda-D-Glu-Mdha)MC-LA 50 909 cyclo-(D-Ala-L-Leu-D-MeAsp-L-Ala-Adda-D-Glu-Mdha-)MC-RR 500 1037 cyclo-(D-Ala-L-Arg-D-MeAsp-L-Arg-Adda-D-Glu-Mdha-)[Dha7]MC-RR 180 980 cyclo-(D-Ala-L-Arg-D-MeAsp-L-Arg-Adda-D-Glu-Dha-)[6(Z)-Adda5]MC-RR >1200 1037 cyclo-(D-Ala-L-Arg-D-MeAsp-L-Arg-6(Z)Adda-D-Glu-Mdha)MC-YR 150–200 1044 cyclo-(D-Ala-L-Tyr-D-MeAsp-L-Arg-Adda-D-Glu-Mdha-)MC-YA 60–70 959 cyclo-(D-Ala-L-Tyr-D-MeAsp-L-Ala-Adda-D-Glu-Mdha-)MC-AR 250 952 cyclo-(D-Ala-L-Ala-D-MeAsp-L-Arg-Adda-D-Glu-Mdha-)MC-M(O)R 700–800 1028 cyclo-(D-Ala-L-Met(O)-D-MeAsp-L-Arg-Adda-D-Glu-Mdha-)

∗Aminoacidic differences with respect to MC-LR are indicated in bold.i.p. = intraperitoneal; M.W. = Molecular Weight.Data from Zurawell et al. (2005).

study allowed the identification of a no-observed-effect level(NOEL) of 40 µg/kg bw/day (Fawell et al., 1994). Slight hep-atic damages were observed at the lowest-observed-effect level(LOEL) of 200 µg/kg bw/day in a limited number of treated an-imals, whereas at the highest dose tested (1 mg/kg bw/day) allthe animals show hepatic lesions, consistent with the known ac-tion of MC-LR. When mice were subchronically administeredwith MC-LR-containing extracts through the diet, a regimenmore similar to human exposure, the NOAEL value was higher(333 µg/kg bw/day) (Schaeffer et al., 1999), due to toxicokineticdifferences. Indeed, gavage corresponds to a bolus dose, result-ing in tissue concentrations higher than those attained after the

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FIG. 2. Chemical structure of the most common congener of nodularin.

more gradual introduction of a dietary treatment, giving time tothe detoxification/excretion systems to be efficient.

As for the other toxicological properties relevant to the riskassessment, pure MC-LR was not teratogenic in mice, in culturedmouse embryos and in frog embryos (Chernoff et al., 2002).Other studies found MC-producing cyanobacteria extracts to beembryotoxic and teratogenic in frog (Dvorakova et al., 2002;Buryskova et al., 2006). The evidence was strong, but unrelatedto MC content, suggesting that the extracts may contain bioactivecompounds other than MCs.

Furthermore, a direct interaction with DNA, responsible forgenotoxic activity, can be reasonably excluded (Runnegar andFalconer, 1982; Repavich et al., 1990). Indeed, contrasting re-sults have been reported and positive results have been obtainedboth in vivo and in vitro only at highly cytotoxic doses, suggest-ing the involvement of DNA endonucleases (Ding et al., 1999;Rao et al., 1998; Zhan et al., 2004). The possibility exists thatMC-LR induces oxidative stress, resulting in indirect oxidativeDNA damage (Lankoff et al., 2004).

The tumor-promoting activity of MC-LR was described al-ready early in cyanotoxin research (Falconer, 1991; Nishiwaki-Matsushima et al., 1992) and more recently confirmed by cyan-otoxin administration with known carcinogenic compounds,such as aflatoxin B1 and diethyl-nitrosamine (Wanght andZhuth, 1996; Sekijima et al., 1999). On the contrary, MC-LR didnot show any tumor induction when the cyanotoxin was given tomice by gavage for 28 weeks (80 µg/kg bw/day) (Ito et al., 1997).

The International Agency for Research on Cancer (IARC)recently reviewed available data on MC-LR carcinogenity, con-cluding that there is inadequate evidence in both humans and ex-perimental animals for the carcinogenicity of MC-LR; however,on the basis of data on tumor promoting mechanisms, IARC hasclassified MC-LR as possibly carcinogenic to humans (Group2B) (IARC, 2006).

Nodularins share with MCs not only a similar chemical struc-ture, which is depicted in Figure 2, but also the mechanism ofaction, that is, phosphates inhibition (Yoshizawa et al., 1990).However, they have not been studied as extensively as MCs.

Nodularins display cumulative toxicity and are tumor pro-moters without any initiation capability (Ohta et al., 1994; Songet al., 1999; Sueoka et al., 1997). However, according to theIARC evaluation, NODs are not classifiable as to their carcino-genicity to humans (Group 3), due to the scant amount of dataavailable (IARC, 2006).

The ip LD50 in mice is similar to the one calculated for MC-LR (50–70 µg/kg bw), but no data on repeated toxicity relevantfor risk assessment are available. However, it is reasonable toconsider that, at least as a worst case, the situation can be com-pared with MC-LR and therefore to adopt the NOEL value iden-tified for MC-LR, expressing NOD concentrations as MC-LRequivalents.

3.2. CylindrospermopsinThe cylindrospermopsin (CYN) molecule consists of a tri-

cyclic guanidine group combined with a hydroximethyl uracil(Figure 3).

It is considered a cytotoxin, since it produces both hepato-toxic and nephrotoxic effects, although other organs may also be

FIG. 3. Chemical structure of cylindrospermopsin.

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damaged following exposure to the toxin (Falconer et al., 1999;Ohtani et al., 1992; Seawright et al., 1999; Terao et al., 1994).

Cylindrospermopsin is highly hydrophilic; its intestinal ab-sorption needs active transport systems as well as the entranceinto hepatocytes, making use of the bile acid transport system(Chong et al., 2002). However, due to the small size of themolecule, a limited passive diffusion through biological mem-branes occurs, as indicated by in vitro studies, showing cytotoxiceffects in a cell line, not expressing bile acid transport system(Chong et al., 2002). When [14C]-CYN (0.2 mg/kg bw) was ipinjected to mice, radioactivity was recovered mainly in the liverand, to a lesser extent, in the kidney (20.6% and 4.3% of theadministered dose 6 hours after the treatment, respectively). Inthe liver about 2% of the injected dose was still detectable aweek after dosing. Urinary excretion is the major route of elimi-nation: 50–70% of administered radioactivity was present in theurine within 6–12 hours. Most of the urinary radiolabel (about72%, corresponding to around 50% of the administered dose)was associated with the parent compound (Norris et al., 2001),indirectly indicating that about one-half of the parent compoundundergoes biotrasformation.

Experimental evidences, produced both in vivo on rodentspecies (Norris et al., 2001, 2002) and in vitro in primary hep-atocytes (Runnegar et al., 1994, 1995), indicated that CYN isbioactivated by cytochrome P-450 (P450). The substantial GSHdepletion, observed after CYN oral administration to rats, leadsto the hypothesis that CYN can be further metabolized by GSHconjugation (Runnegar et al., 1994). Nevertheless, the reductionin GSH content may be also attributed to the inhibition of GSHsynthesis (Runnegar et al., 1995). However, metabolic informa-tion about CYN is very limited and based mainly on indirectobservations.

Cylindrospermopsin has a late and progressive acute toxi-city: After treatment with a lethal dose, death usually occurs24–120 hours after treatment. Indeed, LD50 in mice after ipinjection of pure CYN is 2.1 mg/kg bw after 24 h, but it is10-fold lower (LD50 = 0.2 mg/kg bw) when the observationperiod is prolonged to 120–144 hours (Ohtani et al., 1992). Asimilar trend was seen when Cylindrospermopsis raciborskii ex-tracts containing CYN were injected to mice: The LD50 valueat 7 days corresponded approximately to 0.18 mg CYN equiv-alent/kg bw. When mice were dosed with freeze-dried extractsfrom C. raciborskii via the oral route, the acute toxicity waslower (oral LD50 = 4.4-6.9 mg CYN equivalent/kg bw after 2–6 days), likely due to toxicokinetics differences (Seawright et al.,1999; Chorus and Bartram, 1999).

Acute hepatic damage is localized in the centrilobular areas,being characterized by hepatocyte vacuolization and increasedpigmentation of nuclei and cytoplasm. Necrosis and increasedlumen of proximal tubules and alterations in the glomerulus arethe main features of renal toxicity (Falconer et al., 1999).

The mode of action of CYN as such has been associatedwith inhibition of protein synthesis (Terao et al., 1994), whereasmetabolites very likely act with a different mechanism (Froscio

et al., 2001). Indeed, in mice hepatocytes, CYN concentrations≤0.5 µM were able to inhibit protein synthesis in 4 hours; necro-sis occurred only at 10-fold higher concentrations (in 18 hours).The presence of P450 inhibitors prevented cell death (observedin 18 hours) but not protein synthesis inhibition. This suggestthat at low CYN concentrations toxicity is mainly due to theparent compound through protein synthesis inhibition, whereasat higher concentrations the toxicologically relevant compoundsare very likely represented by CYN metabolites (Froscio et al.,2003).

CYN irritation potential was tested in rabbits in an intra-dermal test, by injecting 0.2 ml lyophilized extract from Aph-anizomenon ovalisporum. Results showed a moderate skin ir-ritation response (Torokne et al., 2001). When lyophilized C.raciborskii extract containing 0.015 mg CYN/g and a nontoxicAphanizomenon strain extract were tested in a maximizationtest for skin sensitization, a clear positive result was obtainedin both cases (Torokne et al., 2001). Therefore, the high sensi-tizing action displayed by the cyanobacterial extracts cannot beassociated with the presence of CYN, and may be attributed toother cellular constituents, such as LPS endotoxins.

Among the available repeated toxicity studies, two have beenconsidered relevant for the risk assessment: (1) mice treated for10 weeks with 3 doses of CYN-containing C. raciborskii ex-tracts dissolved in drinking water (corresponding to 0, 216, 432,and 657 µg toxin equivalents/kg/day bw), and (2) mice treatedfor 11 weeks by gavage with purified CYN (0, 30, 60, 120,and 240 µg/kg/day bw) (Humpage and Falconer, 2003). Bothtreatments resulted in a dose-dependent increase in liver andkidney weight, alteration in plasma enzymes (used as markersfor hepatic and renal toxicity), and consistent hystopathologicalchanges at the high doses. No NOEL could be derived from thestudy with extract, since effects were evidenced in all treatedanimals at each dose. In the study with pure CYN, renal effectsoccurred at lower doses: The no-observed-adverse-effect level(NOAEL) of 30 µg/kg bw per day has been identified on thebasis of increased kidney weight observed at the immediatelyhigher dose tested (i.e., 60 µg/kg bw per day). When compar-ing the two studies, at similar levels of toxin equivalents (i.e.,administration of 216 and 240 µg toxin equivalents/kg/day bw),the degree of severity of the effects was higher following ad-ministration of the extract rather than the pure toxin. The resultsuggests that cyanobacterial constituents other than CYN maycontribute to toxicity.

An additional subchronic study is available, describing oraltoxicity in mice exposed to CYN in drinking water for 42 weeks(Sukenik et al., 2006): Results qualitatively support the alreadymentioned findings, confirming the liver and the kidney as themajor targets for CYN-induced toxicity. In addition, increase incholesterol levels in the plasma and liver and variations in bloodparameters (e.g., elevated hematocrit and deformation of redblood cells) were also reported. In this study, the animals in theexperimental group (one for gender) received CYN at graduallyincreasing daily doses, ranging from 10 to 55 µg/kg bw, with

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changes occurring every 8 weeks (Sukenik et al., 2006). Dueto the nonstandard experimental design for a toxicity test, it isnot possible to derive and adequate NOAEL or lowest-observed-adverse-effect level (LOAEL).

Regarding the other toxicological endpoints relevant for therisk assessment, a possible interaction with DNA, leading even-tually to genotoxic activity, may be suggested by the chemicalstructure of the toxin (presence of guanidine-like groups and po-tentially reactive sulphates). In a limphoblastoid cell line, lack-ing metabolic competence, CYN induced cytogenetic damage(micronuclei formation and chromosomal loss) but only at con-centrations (1–6 µM) at which protein synthesis inhibition andovert cytotoxicity were already evident (Humpage et al., 2000).Similar results were obtained also in CHO K1 cells, lacking anyP450-mediated metabolic activity (Fessard and Bernard, 2003).On the contrary, a possible genotoxic activity, not secondaryto cytotoxic effects, was suggested by positive comet assay re-sults obtained with murine hepatocytes treated with nontoxiclow CYN concentrations (≤0.05 µM) (Humpage et al., 2005).Hepatocytes differ from the previously used cell lines in thatthey are metabolically competent; indeed, positivity in comet as-say was prevented by treating hepatocytes with P450 inhibitors(Humpage et al., 2005), suggesting that the genotoxic potentialis very likely dependent on CYN bioactivation. In vivo DNA-adducts formation (Shaw et al., 2000) and DNA fragmentationin the liver (Shen et al., 2002) have also been reported, but sincethe doses tested were similar to or slightly less than the LD50

value, they can be very likely a consequence of cytotoxicity anddo not allow any final conclusion on CYN genotoxicity.

A study has been conducted in order to evidence a pos-sible carcinogenic activity induced by CYN (Falconer andHumpage, 2001): C. raciborskii extracts containing CYN (500and 1500 µg/kg bw) have been orally administered to mice (1dose every 2 weeks for 3 times) followed by administration of10 µg O-tetradecanoyl-forbole 13-acetate (TPA), a known tumorpromoter, twice a week for 30 weeks (Falconer and Humpage,2001). The results suggest some tumorigenic activity, althoughthe unusual study design (very high doses and frequency ofadministration of an extract instead of the toxin), the limitednumber of animal tested, and the lack of both dose depen-dence and statistical significance of results do not allow us,at present, to draw any conclusions. Although the study hasbeen considered a preliminary one, no other report on CYNcarcinogenicity has been published in the mean time. No re-productive effects induced by CYN have been described up tonow.

3.3. NeurotoxinsAlthough with different mechanisms, all the known neuro-

toxins act on the neuromuscolar system, by blocking skeletaland respiratory muscles, leading to death for respiratory failure.The major groups are anatoxins and saxitoxins, whose majortoxicological features are described in the following.

FIG. 4. Chemical structure of (A) anatoxin-a and (B) homoana-toxin.

3.3.1. Anatoxins-a and Homoanatoxin-aAnatoxin-a is a bicyclic alkaloid; the presence of an addi-

tional methyl group is the only difference with homoanatoxin-a(Figure 4); the toxicological properties of the two structurallyrelated molecules are almost identical.

Anatoxin-a is a potent pre- and postsynaptic depolarizingagent, acting by binding to nicotinic receptors for acetylcholinein the central and peripheral nervous system, and in neuromus-cular junctions (Carmichael, 1998). The toxin has a high acutetoxicity: The ip LD50 in mice is 375 µg/kg bw; death is due tomuscular paralysis and respiratory failure in a very short time(Fawell and James, 1994), whereas after oral administration theLD50 is >5000 µg/kg bw and death occurrs after a latency period(Fitzgeorge et al., 1994). The direct intravenous (iv) injection re-sults in a higher toxicity (iv LD50 < 100 µg/kg bw) and morerapid death (Astrachan et al., 1980). The toxin is rapidly ab-sorbed after ingestion and it is readily degraded, and therefore alow bioaccumulating potential can be anticipated.

The acute toxicity of homoanatoxin-a is similar to its analo-gous toxin (ip LD50 = 330 µg/kg bw in mice), with overlappingsymptoms and death within 5–10 minutes (Namikoshi et al.,2003).

Some repeated toxicity studies (treating periods <2 months)are available on anatoxin-a (Fawell et al., 1999b; Astrachan et al.,1980); in all cases no effect was observed even at the highestdoses tested (120 and 510 µg/kg bw per day in the 2 studies,respectively). Therefore a NOEL value could not be derived. Itcan be concluded that acute effects seem to represent the majorconcern for human health.

When anatoxin-a was administered to pregnant rodents inthe appropriate gestational days (i.e., the susceptibility windowsfor neurological development), no fetal abnormalities neitherneurobehavioral late effects have been evidenced after in uteroexposure, neither any effects of maternal toxicity (Fawell et al.,1999b; Rogers et al., 2005; MacPhail et al., 2005).

3.3.2. Anatoxins-a(s)Anatoxin-a(s) is the phosphoric ester of N-hydroxyguanidine

(Figure 5); similarly to organophosphorous insecticides, towhich is structurally related, it irreversibly inhibits acethyl-cholinesterase (AChE) activity in the neuromuscular junctions(Carmichael and Falconer, 1993), blocking hydrolysis of the

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FIG. 5. Chemical structure of anatoxin-a(s).

neurotransmitter. As a consequence, acethylcholine accumu-lates, leading to nerve hyperexcitability. Anticholinesterase ac-tivity of anatoxin-a(s) is limited to peripheral nervous system;indeed, the brain AChE activity is unaffected also at neurotoxinlethal doses (Cook et al., 1988). The affinity for the human AChEin red blood cells is relatively high, and therefore the risk of anacute intoxication for the human population is not negligible,higher than the one shown by some aquatic species (Carmichael,1990).

The LD50 in mice after ip injection is 20–40 µg/kg bw(surviving time 5–30 min) (Mahmood and Carmichael, 1987;Matsunaga et al., 1989; Cook et al., 1988). Data on oral admin-istration as well as on subchronic and/or chronic toxicity are notavailable.

3.3.3. SaxitoxinsParalytic shelfish poisoning (PSP) toxins are a family of more

than 20 congeners of the same molecule, consisting of a tetrahy-dropurine group and 2 guanidine subunits; saxitoxin (STX) andneosaxitoxin structure is similar to that of carbamates (Figure 6).

The great majority of reported toxicological data have beenobtained with STX produced by marine organisms and only lim-ited information is available for STX produced by freshwatercyanobacteria. However, the chemical structure and the toxico-logical profile of the toxins are the same, independently on theirsource.

The various PSP toxins significantly differ in toxicity, withSTX being the most toxic; they prevent electrical transmission(within the peripheral nerves and skeletal or cardiac muscles),followed by muscle and respiratory paralysis (Kao, 1993; Suet al., 2004; Wang et al., 2003). The mechanism of action is

FIG. 6. General structure of PSP toxin. R4-1: carbamate toxins,including STX and neo-saxitoxin; R4-2: N-sulfocarbamoyl (orsulfamate) toxins, including GTX5 and GTX6; R4-2 and R4-3decarbamoyl toxins, including dcSTX.

based on the blocking of Na channels in neuronal cells (Kao,1993) and on Ca2+ and K+ channels blocking in cardiac cells,thus preventing the propagation of the action potential (Su et al.,2004; Wang et al., 2003). The cause of death is asphyxiation dueto progressive respiratory muscle paralysis.

The 7,8,9-guanidine function has been identified as the oneinvolved in the channel blockade, whereas the removal of thecarbamoyl group side-chain gives rise to a molecule with about60% of the original toxic activity. The biological mechanismof action has been clarified for 50% of the natural analogues,suggesting that it could be basically the same for all the toxinswithin the PSP family.

Saxitoxins are readily absorbed by GI tract; they diffusethrough the blood–brain barrier and are excreted mainly inthe urine. Studies on their metabolism are very scant. How-ever, clinical observations in patients surviving PSP intoxicationfor 24 hours suggest that PSPs undergo either rapid excretion,metabolism, or both. They can bioaccumulate in crustaceansand mollusks, which seem to be resistant to their toxic effects(Llewellyn et al., 1997); this feature determines the possibility ofhigh levels of exposure for predators, including humans (Negriand Jones, 1995). In mice the ip LD50 for STX is 10 µg/kg bwand the oral LD50 is 263 µg/kg bw (Mons et al., 1998).

The level at which PSP intoxications occur in humans variesconsiderably; indeed, an oral consumption of about 300 µgPSP toxin per person in one case was reported as nontoxic,whereas in another one it was fatal (FAO, 2004). Accordingto the FAO report (2004), a total acute ingestion in the range0.144–1.66 mg STX-eq (STX equivalants) per person mainlyinduces mild symptoms, whereas for consumption in a simi-lar overlapping range of doses (0.456–12.4 mg STX-eq/person)a broad spectrum of effects has been described, ranging frommoderate symptoms up to paralysis and death (Shumway, 1995;FAO, 2004). This high variability can be attributed to the dif-ferent approaches and methods applied to quantify the actuallevel of exposure as well as to differences in individual suscep-tibility. The determinants of the severity of the effects are thespecific toxicity of the PSP toxin(s) in the ingested food, theamount of food ingested, and the rate of elimination of the PSPtoxin(s) from the body. When symptoms are mild to moderate,recovery from an STX intoxication is usually complete (Orret al., 2004). In fatal cases, respiratory paralysis occurs within2 to 12 hours after consumption of the PSP-contaminated food(FAO, 2004).

The N-sulfocarbamoyl compounds are appreciably less toxicthan the corresponding carbamoyl toxins. However, under acidicconditions, such as those in the gastric environment, the SO3

−

group is lost, converting the toxins in the carbamoyl analogue(Aune, 2001), with increases in toxicity of up to 40-fold. Thisconversion may therefore have relevant health consequences,since fish containing weakly toxic N-sulfocarbamoyl toxinscould result in severe poisoning episodes after ingestion andacidic hydrolysis in the stomach. No repeated toxicity data areavailable at present, nor are data on the genotoxic potential.

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FIG. 7. General structure of LPS endotoxins.

Some evidences of teratogenic activity have been providedin fish and amphibian larvae, in which STX concentrations≥10 µg/L caused growth retardation; malformation and mor-tality occur at 500 µg/L (IPCS, 1984; Oberemm et al., 1999).However, no data are available on mammals.

3.4. LPS EndotoxinsLPS endoxins are external components of cell membranes

of most cyanobacteria as well as gram-negative bacteria. Themolecule consists of three regions: an internal acylated glycol-ipid (termed lipid A), and a central area of liposaccharides, link-ing the internal subunit with the external specific carbohydratepolymer (O-specific chain) (Jann and Jann, 1984) (Figure 7).Among bacteria this external subunit shows the most diversityand is the basis for serological specificity, but also the lipid Amoiety is variable. Cyanobacterial LPS endotoxins slightly dif-fer from those typical of other bacteria in the three components,including the presence of small quantities of phosphates (Mayerand Weckesser, 1984; Kaya, 1996).

LPS endotoxins exposure has been associated with local ef-fects due to direct contact, such as skin or eye irritation, gastroin-

testinal problems or allergic reactions. However, these types ofeffects have never been experimentally reproduced and the LPSendotoxin mechanism of action is unknown. Indeed, the po-tential induction of gastroenteritis and inflammation problemshas been often assumed in analogy with known effects of LPSfrom gram-negative bacteria, which have been extensively stud-ied. It has been shown that they bind to transmembrane recep-tors within the Toll-like receptor family, initiating a cascade ofhost-mediated responses, among which the release of cytokinesand other inflammatory mediators, stimulation of monocytesand macrophages, and congregation of neutrophils and plateletsmicrocapillaries, followed by vascular injury (Heumann et al.,2002). Therefore, LPS endotoxins are not directly toxic, but theirtoxicity is associated with the interaction with host-mediatedfactors.

Among different gram-negative bacteria, the lipid A moietyis considered the LPS component responsible for toxic effects,which can be extremely variable, up to totally inactive (Stewartet al., 2006a). Since the structure of the lipid A subunit in thecyanobacterial LPS molecule has not been clearly identified sofar, no definite conclusion should be drawn on the degree of their

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toxic potential, although some indirect data seem to suggest arole for LPS in cyanobacterial intoxication.

As already mentioned, results from maximization testsclearly indicated that different cyanobacterial extracts inducehigh degree of skin sensitization, independent from the produc-tion of intracellular cyanotoxins (Torokne et al., 2001), suggest-ing a possible role for other cyanobacterial constituents, amongwhich LPS endotoxins. However, during cyanobacterial bloomsmany other organic compounds (such as aldehydes, terpenoidsand ketons), some of which are endowed with irritating and sen-sitizing properties, are dissolved in water. Therefore, it is pos-sible that irritating and sensitizing effects observed so far weredue to the concurrent presence of different etiologic agents.

Very little is known on LPS endotoxins systemic effects; somedata on lethality in mice after injection of LPS extracts from dif-ferent cyanobacteria indicate that LD50 values range between 40and 190 mg/kg bw, although some exctracts caused no death at250 mg/kg bw; in addition, qualitative dermonecrotic lesions inrabbit skin were described, following sequential subcutaneousand iv injections of the same LPS extracts (Stewart et al., 2006a).Due to the fact that cell wall fragments are readily aerosolized,inhalation of LPS might contribute in explaining cyanobacterial-related adverse effects known as “flue-like symptoms,” charac-terized by cough, chills, sore throat, and fever. However, no clearassociation have been found and therefore it can be concludedthat the health implications of cyanobacterial LPS are poorlyunderstood and this topic requires more research.

3.5. Other ToxinsBeside the already mentioned toxins, which have been de-

tected in freshwater and brackish waters worldwide, other cyan-otoxins have been identified, mainly in marine coastal areas inHawaii and the Indo-Pacific, among which are aplysiatoxin, de-bromoaplysiatoxin, and lyngbyatoxin. These toxins have beenindicated as the causative agents of contact dermatitis (swim-mer’s itch) in Hawaii (Serdula et al., 1982) and of intoxicationsdue to the ingestion of contaminated meat from Chelonia mi-das, a marine turtle (Yasumoto, 1998). In the years 1993–1998in Madagascar, consumption of meat from marine turtles ledto some poisoning episodes; 414 people were intoxicated, 29 ofwhom died. Described symptoms included acute gastritis, mouthulcers, burning of buccal mucosa and tongue with appearance ofpapule, salivation, headache, weakness, and fever (Champetieret al., 1998).

Aplysiatoxin is a phenolic bislactone (Figure 8). Whenip injected to mice, it showed an LD50 value around 100–120 µg/kg bw; the cause of death was bleeding from the smallintestine preceded by dilation of the lymphatic vessel and con-gestion of capillaries in the lamina propria (Ito and Nagai, 1998,2000). At sublethal doses the major effect was diarrhea, inducedby hypersecretion from edema in the caecum (Ito and Nagai,1998). After oral administration, toxicity was reduced and sub-lethal effects on the small intestine were observed at much higherdoses (3 mg/kg bw). When the toxin was iv injected (100 mg/kg),

FIG. 8. Chemical structure of aplysiatoxin.

the target vessels were in the lung: Fibrin deposition in the di-lated pulmonary artery caused the appearance of a gap in theartery wall and, consequently, bleeding. In addition, bleedingeffects on the small intestine were also present, due to fibrindeposition in the lumen via distension of the capillary wall (Itoand Nagai, 2000).

Lyngbyatoxin has many similarities with aplysiatoxin in itsmechanism of toxicity; indeed, ip injection of lethal doses inmice (250 µg/kg bw) induced severe damages in the villi capil-laries, leading to bleeding in the small intestine; sublethal dosescaused erosion in the stomach, small and large intestine and in-flammation in the lung (Ito et al., 2002). After oral administra-tion, the pathological outcome at sublethal doses was almost thesame, but effects occured at higher doses (600–1000 µg/kg bw).

No data on aplysiatoxin- or lyngbyatoxin-induced repeatedtoxicity are available at the moment, but it is known thatthey are potent tumor promoters, as well as debromoaplysia-toxin, acting through potentiation of protein kinase C as 12-O-tetradecanoylphorbol-13- acetate (TPA) does (Fujiki et al., 1981,1982).

Cyanobacteria may also produce a number of other bioactivepeptides, including microviridins, microginins, and cyanopep-tolides. Their function, actual presence in the environment, andimpact on human and environmental health are poorly known(Welker and von Doeren, 2006). Among the already mentionedpeptides, microviridin J and BMAA have been the subject ofsome recent publications.

Microviridin J, another metabolite of Microcystis spp., hasbeen indicated as the cause for a lethal molting disruption inDaphnia spp., upon ingestion of living cyanobacterial cells. Thetoxin consists of an acetylated chain of 13 amino acids arrangedin three rings and two side chains. Proximal hydrophobic inter-actions between Arg and other regions of the molecule result inthe formation and stabilization of an additional ring system. Thepresence of Arg and its distinctive conformational interactionshas been associated with microviridin J inhibition of trypsin,chymotrypsin, and trypsin-like proteases in the daphnid, pre-sumably linked to the molting disruption. No data are availableon the toxicological profile of this toxin in mammals (Rohrlacket al., 2003, 2004)

It has been recently demonstrated that a wide variety of bothfree-living and symbiont cyanobacteria are able to produce β-N-methylamino-L-alanine (BMAA), a nonessential amino acid, atsignificant levels (Cox et al., 2005). Although with contrastingopinions, neurotoxic effects have been attributed to BMAA

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(Miller, 2006; Lobner, 2007). Its dietary consumption has beenassociated with an atypical neurodegenerative disease knownas amyotrophic lateral sclerosis–Parkinsonism–dementia com-plex (ALS/PDC) of the Western Pacific. The hypothesis for theassociation relies on the biomagnification from cyanobacteriasymbionts on cycads root, to fruit bats foraging on cycad seeds,to humans that usually eat the bats in the area (Cox et al., 2003).The fact that BMAA has been detected in the brain tissue ofpatients from Guam affected by ALS/PDC supports the hypoth-esis. In addition, the presence of BMAA up to 6 µg/g in brainsamples of 8 out of 9 Canadians who died of Alzheimer’s disease(Murch et al., 2004) suggested that cyanobacteria could repre-sent the ultimate source of BMAA for populations other than theones on Western Pacific islands. BMAA can be associated withendogenous proteins, giving rise to high bound concentrationsin the human tissues (“sink”); the subsequent slow release couldexplain the lag phase between exposure and late neurodegener-ative disease.

The major criticisms to the role of BMAA in ALS/PDC andAlzheimer’s disease are linked to the high doses that have tobe administered to primates to induce some degenerative ef-fects; these doses correspond to unrealistic consumption of con-taminated food in humans. In addition, the administration ofBMAA has not been associated with in vivo neurotoxic ef-fects in mice (Cruz-Aguado, 2006); furthermore high BMAAconcentrations (1–3 mM) are required to induce in vitro neu-ronal death (Ross et al., 1987). Interestingly, recent findingshave suggested that although BMAA alone is unlikely to causeneurological deficits unless very high levels are dosed, concen-trations as low as 10–30 µM can potentiate the action of otherneurotoxic agents (Lobner et al., 2007), leading to the develop-ment of neurological diseases after long-term chronic exposure.However, more research is warranted in order to ascertain therole of BMAA produced by cyanobacteria in neurodegenerativediseases.

4. RISK ASSESSMENT ASSOCIATED WITHCYANOTOXIN EXPOSURE

Humans may be exposed to cyanotoxins through severalroutes: The oral one is by far the most important, occurring byconsumption of contaminated drinking water or food (includingsupplements) or by ingesting water during recreational activi-ties. However, dermal and inhalation exposure may also occur.These routes of exposure are associated with recreational, sport,and professional activities (i.e., fishing) in infested waters, orto the domestic use of cyanotoxin-containing water, as in thecase of showering. The possibility of the parenteral route of ex-posure has also been described, when water from contaminatedsuperficial water bodies has been used for hemodialysis.

The identification of the appropriate scenario and the routeof exposure may strongly affect the bioavailability determin-ing the amount of internal dose of cyanotoxins in humans. Forthis reason in the following, the possible risks for human health

are presented by grouping the different exposure scenarios andcyanotoxin sources.

4.1. HemodialysisWhen surface waters infested by cyanobacteria are used for

hemodialysis, they can represent a remarkable risk for the pa-tients; indeed, the paternal route of exposure considerably in-creases the internal dose of toxins, directly entering the blood-stream. Therefore, although this is a quite low-frequency event,it represents an extremely relevant route of exposure with respectto the risk evaluation for human health.

The most important episode of human health consequencesassociated with exposure to cyanotoxins has been reported inBrazil, where 56 patients out of 130 in hemodialysis treatmentdied after receiving water that subsequently turned out to becontaminated by MCs (Jochimsen et al., 1998; Azevedo et al.,2002).

The quality of water used for dialysis is regulated in somecountries worldwide, but it is not subjected to any mandatory reg-ulations in most European countries. The recommendations ofthe European Pharmacopoeia (1997) concerning inorganic ionsand few organic pollutants are generally met, but cyanotoxin de-tection is not performed as a routine quality control for dialysiswater. However, for therapeutical use, in face of the particularexposure together with the pathological conditions of patients,the water used for hemodialysis should be free of cyanotox-ins. Therefore, if there are no alternatives to the use of surfacewaters infested by cyanobacteria in hemodialysis therapy, it isextremely important to guarantee the maximum effectiveness inthe preparation of water and to plan adequate quality controls.

4.2. Drinking WaterCyanotoxins can occur in drinking water, dependening on

their level in raw surface water and the effectiveness of treatmentmethods for removing cyanobacteria and cyanotoxins. World-wide, particularly in developing countries, many people need touse unfiltered/untreated surface waters, being therefore exposedat the same time to both cell-bound and dissolved cyanotoxins.

Depending on cyanotoxin levels in drinking water, bothacute/short-term and chronic effects in humans may occur(Chorus and Bartram, 1999).

Acute/short-term effects are associated either with the con-sumption of raw waters infested by cyanobacteria or with highcyanotoxin dissolved concentrations in drinking water as a con-sequence of either the breakdown of a natural cyanobacterialbloom or its artificial lysis followed by the failure of watertreatments.

Among the different episodes described so far, an outbreakin Brazil has been attributed to a bloom of Anabaena and Mi-crocystis genera in raw waters; 2000 cases of gastroenteritis and88 deaths in a period of 42 days were reported (Teixera et al.,1993). In Australia, due to the treatment with copper sulfateof a cyanobacterial bloom, causing cell lysis, 140 children and10 adults required hospitalization for liver and kidney damage

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within a week (Byth, 1980). All the patients recovered after ap-propriate therapeutic treatment. Cylindrospermopis raciborskiiwas successively identified as the etiologycal agent (Hawkinsand Griffiths, 1993). Several other episodes of gastroenteritisand liver damage potentially associated with drinking water con-taminated by cyanotoxins have been reported (Botes et al., 1985;Fawell et al., 1993; Falconer, 1989, 1994; Zilberg, 1996; El Saadiet al., 1995).

Acute/short-term effects can be prevented with adequatetreatments that allow a strong reduction of both cell num-ber (>99%) and dissolved cyanotoxins (Jones and Orr, 1994;Dietrich and Hoeger, 2005). At present, little is known aboutthe possible risk posed by ozonation treatment via the formationof byproducts, which have been frequently detected, especiallywhen an insufficient ozone dose is used (Dietrich and Hoeger,2005). For those countries where efficient treatments are notapplied, the acute risk associated with consumption of contami-nated water is higher and tailored “practical” management mea-sures should be defined. As examples, recommendations canbe made not to use surface waters infested by cyanobacterialblooms without filtering to remove cells (i.e., simple sand filters),and to avoid the use of water when the bloom is senescent andextracellular cyanotoxin concentration is expected to be higher.

Chronic effects are difficult to identify and demonstrate, andindeed, information from epidemiological studies is scarce andinconclusive, even because cyanotoxins were not proven to bethe actual cause of the observed effects, but merely the mostlikely one. In China the high incidence of hepatic tumors in apopulation has been associated with consumption of raw waterscontaining MC (Ueno et al., 1996). However, the extremely poorwater quality (contaminated also with other chemicals and mi-crobiological agents) and the concomitant consumption of foodcontaminated with aflatoxin B1, a known hepatic carcinogen,represent such relevant confounding factors that the validity ofthese results is low. Similarly, no conclusions could be drawn intwo ecological studies in Florida about a possible association be-tween hepatic and colorectal tumors and consumption of drink-ing water potentially contaminated with cyanotoxins (Fleminget al., 2001, 2002). In light of the quality of the available epi-demiological data, the IARC (2006) concluded that it was notpossible to associate the excess risk of hepatocellular carcinomaand of colorectal cancer specifically with exposure to MC.

Although the epidemiological data are not conclusive, theavailable toxicological data can help, at least for some cyanotox-ins, to evaluate the risk associated with contaminated drinkingwater consumption, making use of consolidated, internationallyaccepted risk assessment procedures, recommending the adop-tion of a conservative approach for human health protection.

In the case of MC-LR, WHO (2004) selected the alreadymentioned subchronic NO(A)EL of 40 µg/kg bw/day (Fawellet al., 1994). The choice of this NO(A)EL represents an exam-ple of the application of a conservative approach, indeed: It hasbeen obtained in a study on mice, which are more sensitive toacute effects of MC-LR than rats; furthermore, in the study the

space among doses is large, the effects at LO(A)EL are slightand involve a limited number of animals, and finally the routeof exposure is gavage rather than dietary (Fawell et al., 1994).By dividing the NO(A)EL value by an uncertainty factor (UF)of 1000, inter- and intraspecies variability (100) and the lackof chronic toxicity data (the subchronic toxicity study corre-sponds to about one-sixth of mouse life span) have been takeninto account. A provisional tolerable daily intake (TDI) value of0.04 µg MC-LR/kg bw/day is obtained, meaning that an adultwith a body weight of 60 kg could be orally exposed to 2.4 µgper day all life long, without experiencing any toxicological ef-fect. In light of the approach used, this value is conservativeenough to consider that the exposure for a limited period of timeto MC-LR values similar or slightly exceeding the TDI valuedo not represent a real risk for the human population. On thisbasis, the Wold Health Organization (WHO, 2004) has calcu-lated a provisional guideline value (GV) of 1 µg/L for MC-LRin drinking water. For this purpose, a daily consumption of 2 Ldrinking water and an allocation factor (AF) of 0.8 (meaningthat drinking water was assumed to contribute for the 80% ofthe total intake of MC-LR) have been assumed. The followingequation summarizes this procedure:

GV = TDI × bodyweight × AF

daily consumption (C)

= 0.04 µg/kg × 60 kg × 0.8

2L∼= 1 µg/L

A recommendation to use concentration equivalents as defaultvalue for the total concentration of all MC variants has beenmade (Chorus and Bartram, 1999), to overcome the problem ofnot having a specific GV for each congener. Since MC-LR is oneof the most acutely toxic MC variants, the expression of total MCconcentrations as MC-LR equivalents could represent a furtherconservative element in this approach, although the estrapolationfrom the acute toxicity ranking among MC congeners to thechronic toxicity remains to be demonstrated.

WHO has not derived GV for any other cyanotoxins, due tothe lack of adequate toxicological data.

Regarding anatoxin-a, WHO considered the repeatedtoxicity database not adequate to derive a TDI. As mentionedearlier, an NO(A)EL has not been identified, since no effectswere observed at the highest tested doses in the 2-monthstudy (120–510 µg/kg bw/day) (Astrachan et al., 1980; Fawellet al., 1999). However, if the highest value is considered asthe NO(A)EL (by being extremely conservative), a provisionalTDI of 0.51 µg/kg bw/day can be derived by applying a UF of1000 (considering the same uncertainties evidenced in the caseof MC-LR), leading to GV ∼= 12 µg/L (Duy et al., 2000). Basedon these considerations, it has been proposed that GV = 1 µg/Lfor the total concentrations of anatoxins in drinking watercould provide an adequate margin of safety (about 3 ordersof magnitude) to protect human health of potentially exposedpopulations (Fawell et al., 1999). On this basis, New Zealandhas established a limit of 6 µg/L for total anatoxins content in

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drinking water and Australia has adopted a limit of 3 µg/L foranatoxin-a (Chorus, 2005).

Concerning CYN, in order to evaluate the chronic risk as-sociated with its oral exposure, the most appropriate startingpoint is the subchronic NOEL = 30 µg/kg bw/day (Humpageand Falconer, 2003). Dividing the NOEL for a UF of 1000(100 for inter- and intraspecies variability, and 10 for the lackof chronic data), a TDI value of 0.03 µg/kg bw/day is derived.Therefore, no risk is expected to be associated with CYN inges-tion up to 1.8 µg per person (weighing 60 kg) every day duringthe life span. This TDI should be updated, once genotoxic and/orcarcinogenic properties of this cyanotoxin are demonstrated.

Considering an AF of 90% (due to lack of information aboutother possible sources of CYN exposure, including contami-nated food) and 2 L drinking water consumed every day, a GVof 0.81 µg/L (rounded to 1 µg/L) can be derived (Humpage andFalconer, 2003; Codd et al., 2005).

Toxicological data can be also used for defining safe con-centrations with regard to the acute risk. This evaluation can becarried out starting from the identification of an acute NOEL,which is the highest dose at which no effects are observed. Un-fortunately, it is not possible to derive an acute NO(A)EL fromthe oral studies, since signs of hepatic toxicity were presenteven at the lowest dose tested (LOAEL = 500 µg/kg bw). Onthe other hand, some ip acute studies allow the derivation ofa range of doses (25–50 µg/kg bw) producing no effects in themouse liver, the target organ (Fromme et al., 2000). In thesestudies, the ratio between the concentration with no effects andthe one inducing severe adverse effects, up to mortality, is verysmall (≤2), suggesting steep dose-response curves for MC-LR(Fromme et al., 2000). Considering the 30- to 100-fold differ-ence evidenced between the oral and ip routes of exposure, theapplication of a correction factor (CF) of 10 would representa conservative approach. In addition, similarly to the chronicrisk, a UF of 100 can be applied for inter- and intraspeciesvariability. An acute no-effect dose of 2.5 µg/kg bw is then ob-tained. In agreement with the steepness of the dose/responsecurve, this value is very close to the LOAEL for hepatic ef-fects after oral administration (5 µg/kg bw, obtained by divid-ing the LOAEL in mice for UF = 100, accounting for inter- andintraspecies extrapolation). Hence, special attention should begiven when the exposure to MC-LR is close to the acute no-effectdose.

For an adult with a 60-kg bw, the acute no-effect total doseis 150 µg/person, according to the equation:

Acute no-effect dose = acute NOELx bw × CF

UF

= 25 µg/kg × 60 kg × 10

100∼= 150 µg/person

At this level of exposure to total MCs, no acute effect is expected,also considering that the evaluation has been based on data onMC-LR, which is among the most toxic variants.

Also, STX could represent a source of concern for acute ef-fects, when infested waters are used for drinking, due to theiroccurrence in fresh waters within cyanobacteria cells and dis-solved even at quite high levels (up to 2700 µg/L)(Batoreu et al.,2005). However, no evidence of human intoxication has been re-ported so far.

In order to manage this possible health problem, some coun-tries proposed GVs or adopted mandatory regulatory require-ments. As examples, a guideline concentration of 3 µg/L STXequivalents in drinking water has been adopted in Australia(NHMRC, 2001) and of 1 µg/L in New Zealand (Orr et al.,2004). The daily intake corresponding to these regulatory lim-its (2–6 µg/person, considering an intake of 2 L drinking water)represents only a small fraction of the limit established by the Eu-ropean Union (EU) for bivalve mollusks (80 µg STX-eq/100 gof meat, see next sibsection), in order to protect consumers fromacute effects (EU Directive 91/492/EEC).

4.3. FoodIn the following paragraphs the levels of contamination in

different food items are addressed at first, and then the relatedrisk for human health is assessed.

4.3.1. Fish, Shellfish, and MollusksAquatic organisms (fish, bivalves, snails, and other macroin-

vertebrates) may accumulate cyanotoxins via ingestion ofcyanobacteria cells/contaminated food or via the transdermalroute as dissolved toxins. The latter route may be especiallyrelevant for CYN, since it may often be found in dissolvedform (Preussel et al., 2006; Rucker et al., 2007). In the caseof MCs, although high concentrations may be measured inwater due to lysis of surface bloom, it can be reasonably as-sumed that exposure of biota to high levels of dissolved toxinis an unfrequent event, and, when occurring, it is likely to beshort-lived (Ozawa et al., 2003). Therefore, ingestion of MC-contaminated food seems to be the major route for toxin up-take in freshwater biota. A review of accumulation pattern ofcyanobacterial toxins in freshwater edible organisms and of thepossibility of biomagnification versus biodilution processes inthe food web has been recently produced (Ibelings and Chorus,2007).

In many edible shellfish and molluscs, MCs efficiently accu-mulate in the hepatopancreas, where they persist for many days(Vasconcelos, 1995); similarly, in fish the highest levels weredetected in viscera (liver and gut), with an average value of 13.4mg/kg (Magalhaes et al., 2001) and maximum values up to 478mg/kg (Chen et al., 2007). From a comprehensive analysis ofavailable data on MCs, maximum concentrations of 300, 2700,and 16,000 µg/kg have been reported in the edible parts of fish,crustaceans, and mussels, respectively (Dietrich and Hoeger,2005). Recently, an even higher MC concentration (370 µg/kg)has been reported in bighead carp in a Chinese lake (Chen et al.,2007).

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Data available on other cyanotoxins are limited and obtainedwith laboratory tests rather than coming from field monitoring.After exposure of Anodanta cygnea to cultures of CYN pro-ducing cyanobacteria, levels of 1099 and 247 µg/kg wet weightwere reported in viscera and whole body, respectively (Sakeret al., 2004). Concentrations up to 4.3 mg and 0.9 mg of CYN/kgwere found in freeze-dried hepatopancreas and muscle tissue incrayfish (Cherax quadricarinatus) and up to 1.2 mg/kg in freezedried viscera tissue in rainbow fish (Melanotaenia eachamensis)harvested in an aquaculture pond (Saker and Eaglesham, 1999).In laboratory conditions, cyanobacterial STXs administered tofreshwater mussels may accumulate up to >80 µg/100 g musselflesh (Negri and Jones, 1995). Other laboratory studies showedthat STXs accumulate also in other freshwater bivalves and inDaphnia magna (van Apeldoorn et al., 2007; Nogueira et al.,2004).

4.3.2. Meats, Milk, VegetablesCyanotoxins uptake can be assumed in animal husbandry

through consumption of contaminated waters. Results fromstudies where MCs or toxic M. aeruginosa were administeredto reared cows did not show any presence in the meat (Orr et al.,2003) or transport to the milk (Orr et al., 2001; Feitz et al., 2002),consistent with their nature of hydrophilic compounds. No dataare available on other cyanotoxins.

Cyanotoxins could also be transmitted to plants from sur-face irrigating waters. In a study, lettuce plants were grown withspray irrigation with water containing M. aeruginosa. Analysisof an extract of liophylized blooms and scums material from theirrigation water supply revealed the presence of MC-LR equiva-lents up to 3.23 µg/mg dry weight (dw) (Codd et al., 1999); after10 days from the last irrigation, MC-LR equivalent levels rang-ing from 0.094 to 2.487 µg/g dw were detected in lettuce leavesextracts (Codd et al., 1999), very likely derived from the intra-cellular content of cyanobacteria deposed on the lettuce leaves,observed before the extraction procedure. However, it can be an-ticipated that cyanobacteria cells could be readily removed by theusual washing and rinsing procedure before eating. In relation toabsorption of dissolved cyanotoxins present in irrigation waterby the plant tissues, extracts from rape and rice seedlings ex-posed to water with MC-LR equivalents up to 3 mg/L containedMC levels correlated with the applied toxin concentrations withmaximum values of 651 and 5.4 ng/g fresh weight of rape andrice, respectively (Chen et al., 2004). The study showed that, al-though with pronounced differences, some vegetables can retainMCs present in irrigation water. Similar results were obtainedin another study with different plants (McElhiney et al., 2001).However, the already described situations correspond to the useof water infested by blooms or scums, at such high cyanotoxinslevels able to inhibit the plant growth, to induce visible toxiceffects (such as the appearance of brown leaves) (Chen et al.,2004), thus posing a threat to the yield and quality of crops thatvery likely will not be used for eating purposes. This strongly

limits the possibility for human exposure. These considerationsabout the low level of concern posed by cyanotoxin exposure viavegetable consumption are supported by results from a recentlypublished paper (Jarvenpaa et al., 2007). It showed that whenwatering broccoli and mustard seedlings with water containingMC concentrations typically found in natural surface waters (1–10 µg/L), the toxins were found only in the roots, at levels of nohuman health concern (Jarvenpaa et al., 2007).

Nevertheless, it can be easily recommended not to use bloom-and scum-infested waters for irrigation, even though this maybe not so easily implemented in hot climate areas, where scumsand lack of water for irrigation tend to co-occur. However, intemperate climates such events are often only episodic, lastingfor a limited period of time.

An additional consideration should be made of the use ofcyanobacterial bloom and scums as organic fertilizers in agri-culture, as occurs in China (Chen et al., 2006). Indeed, in favor-able conditions and specific soil, MCs can leach and possibly beresponsible for groundwater contaminations (Chen et al., 2006),being possibly transferred to drinking water.

The overall available data on food contamination are still notsufficient for an adequate risk assessment. However, they seemto indicate that fish and other freshwater organisms consumptionis the major food source representing a risk to human health.

The risk associated with chronic exposure to contaminatedfood can be assessed by comparing the TDI value with the esti-mated (or measured) cyanotoxin levels daily ingested. The ex-posure assessment is a crucial step in the process. In the case offreshwater organisms, in order to obtain a reliable estimate of theaverage ingestion, it is necessary to measure cyanotoxin contentin the edible parts of representative species, over an adequateperiod of time, to take into account seasonal variation in cyan-otoxins production and possible differences in fish species intheir capacity of retention of cyanotoxin residues. The monitor-ing program should be carried out locally and tailor made, basedon the feature of the water body and on the dietary habits of thepotentially exposed population. When such an estimation for theactual exposure is not accurate or not available, the reliability ofthe risk assessment could be considered only approximate andprovisional. In addition, another factor of uncertainty in evalu-ating human exposure derives from the fact that it is not clearwhether the levels measured in freshwater organisms correspondto doses bioavailable for absorption or to total cyanotoxin levels,including free as well as protein bound or conjugated forms.

As described earlier, most of the available data have beenproduced in laboratory conditions; when field data are available,they are often obtained without adequate monitoring plans, orrefer to organisms which are not commonly eaten by humans, sothat they are poorly representative of the actual human exposure.

On the basis of the available data, the following risk assess-ment is proposed.

Referring to MC-associated acute risk, the worst case cor-responds to a peak level of 370 µg/kg in edible parts of carps(Chen et al., 2007). With this value, considering an episodic

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consumption of 300 g of flesh, a dose of 100 µg/person wouldbe ingested. This intake is below the already calculated acuteno-effect level dose of 150 µg/person, but sufficiently close torecommend some caution, due to the steepness of the dose-response curve. On the other hand, the available data indicatethat doses corresponding to (or exceeding) the acute risk can bereached ingesting even small quantities of fish viscera, whereMC levels of more than 10 mg/kg have been detected.

In terms of the chronic risk, the peak values that are usually re-ported are not very useful, since to evaluate chronic human expo-sure, average MC concentrations should be considered togetherwith local dietary habits. However, by using the available liter-ature data, 0.25–100 µg/kg can be assumed as the range of MCconcentrations in edible parts of fish, and considering an averagedaily consumption of 200 g, intakes of 0.05–20 µg/person/daycan be calculated. The low-end value of the range is well belowthe TDI value, whereas the higher end value poses concern forhuman health following chronic fish consumption, being 10-foldhigher than the TDI value for an adult of 60 kg bw.

It could be inferred that daily consumption of cyanotoxin-contaminated freshwater organisms every day for life-span du-ration could represent a quite unreasonable exposure scenario. Itappears that the hypothesis of a daily consumption for a limitedperiod of time (i.e., blooming season) could be more represen-tative of actual exposure conditions for the human population.This scenario for highly acutely toxic compounds, such as cyan-otoxins, is usually covered by the acute reference dose (ARfD),which is derived from the subacute/subchronic NO(A)EL, towhich UFs have to be applied. For MC-LR an UF of 100,covering inter- and intraspecies variability, can be consideredappropriate; the result is an ARfD of 0.4 µg/kg bw/day, corre-sponding to 24 µg/person/day for a 60 kg bw adult. The com-parison of the ARfD with the already calculated intakes (0.05–20 µg/person/day), shows that while the low end of the rangeis well below the ARfD, the higher end value is very close tothe limit value and some caution should be taken, also takinginto account cumulative intake possibly derived by exposure todifferent sources (i.e., food, drinking water, bathing activity).

A similar approach comparing exposure to TDI and ARfDvalues could be applied to evaluate the chronic risks associ-ated with consumption of freshwater organisms contaminatedby CYN or anatoxins; however, the available data set is too lim-ited for such general considerations.

Referring to STX, EU Directive 91/492/EEC has set a limitfor PSP toxins in bivalve mollusks at 80 µg STX-eq/100 g ofmeat. This limit is intended to protect consumers from acuterisks; indeed, it is derived exclusively from data on acute intox-ication in humans indicating 144–304 µg STX/person for mildsymptoms and >450 µg STX/person for severe intoxications.These data cannot be used to evaluate chronic risks, associatedwith repeated exposure to STX, for which no information areavailable.

STX levels above 80 µg/100 g (Negri and Jones, 1995) havebeen reported in freshwater mussels, suggesting a possible risky

human acute exposure. In fact, by assuming an average con-sumption of the edible part between 100 and 300 g, the marginof safety toward mild symptoms is around 1. Since the data seton PSP intoxication derives from effects seen in many individu-als within a quite large exposed population, displaying possibledifferences in susceptibility, the UF of 10 for intraspecies differ-ences may not be necessary. Nevertheless, the margin of safetybetween the regulatory limit and the dose causing severe intox-ications or death appears to be quite small (<10). On the otherhand, according to an FAO report (2004), it is neither practicalnor realistic to establish a lower tolerance level, because the of-ficial method of analysis, the mouse bioassay, has a detectionlimit of about 40 µg PSP (STX-eq)/100 g shellfish. It is stated inthe legislation that when needed, a chemical detection methodis associated, but if the results are challenged, the referencemethod remains the bioassay (EU, 1991). Therefore, once moresensitive methods are available, the toxicity figures of STX andderivatives after acute exposure should be reevaluated.

As a general consideration, it can be recommended that when-ever ingestion of cyanotoxins is around or above the thresholdfor concern, it is necessary to carefully monitor the contamina-tion of the water body and of fishery products, increasing thenumber of both sampling sites and analyzed samples. In thepresence of a potential risk for the population it is necessaryto promote adequate management measures to be decided at thelocal level, including information to the population. As an exam-ple, for water bodies affected by blooms of toxic cyanobacteria,common advice is not to eat the viscera of fish. In industrial-ized countries, on the basis of the outcome of risk assessment,a decision like fishery temporary banning could be taken with-out expecting significant side effects on nutrition, because ofeconomically affordable available alternatives to fish in the diet.On the other hand, in developing countries, banning fish con-sumption may have unacceptable consequences, like shortage ofprotein intake; therefore, this issue should be faced by carryingout a proper risk/benefit analisys.

4.4. Dietary Supplements ConsumptionIn recent times the use of extracts from cyanobacteria as di-

etary supplements has become popular in Western countries,similarly to what has been done in the past in China and Africa(Carmichael et al., 2000). These supplements, known as blue-green algae supplements (BGAS), are mainly obtained fromSpirulina spp. and Aphanizomenon flos-aquae, grown in artifi-cial ponds or collected directly from the natural environment.They are presumed to have beneficial effects on human health,including support in loosing weight during hypocaloric diets,and increasing alertness and energy and elevated mood for peo-ple suffering depression (Jensen et al., 2001). In addition, BGASare administered to children as an alternative, natural therapy totreat attention deficit hyperactivity disorders (ADHD).

Since BGAS are not considered a drug, there is no prescrip-tion nor indication for a specific daily dosage: consequentelythey are consumed following individual programs. It is therefore

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extremely difficult to appropriately evaluate the actual exposure.However, a consumption of high doses, up to 20 g per day, hasbeen reported (Dietrich and Hoeger, 2005). Some recent studiesevidenced that BGAS could be contaminated by variable MC-LR levels, up to 35 µg/g, with a substantial percentage of sam-ples characterized by MC-LR levels ≥1 µg/g dry weight (Diet-rich and Hoeger, 2005; Saker et al., 2005). When two genes ofthe microcystin synthetase gene cluster were sequenced in somefood supplement samples, results showed that M. aeruginosawas the source of contamination (Saker et al., 2005), suggest-ing that toxin-producing cyanobacteria other than Spirulina spp.and Aphanizomenon flos-aquae are included during collectionprocesses from the natural environment.

The risk assessment associated with BGAS use depends onthe duration of the exposure period; that is, it is a chronic risk ifthe consumption is prolonged for very long time or comparablewith a subchronic risk, when ingestion is limited to a relativelyshort period.

Considering the former scenario, it is possible to comparethe MC-LR level daily ingested with the provisional WHO TDIvalue of 0.04 µg/kg bw per day: When the exposure levels arelower than or similar to the TDI, it can be anticipated that there isno human health concern associated with the chronic consump-tion of BGAS in the condition of intended use.

In the event that exposure is around the TDI value, it would beappropriate to consider also other sources of exposure. Indeed,contaminated drinking water or fish products consumption couldsignificantly contribute to exceed the threshold of concern es-tablished by the TDI value and it could be useful to estimate thepercentage of total ingestion attributable to BGAS.

The possibility exists that an adult consumes daily 20 g BGAScontaminated by the highest MC-LR level measured (35 µg/gdry weight); this scenario can be considered as the “worst case.”It would result in the ingestion of 700 µg/person per day, whichcorresponds to about 250-fold the TDI.

The described worst case would imply such a high dailydose to represent also an acute risk. Indeed, by referring tothe already mentioned acute no-effect dose (2.5 µg/kg bw), thestandard adult weighing 60 kg could ingest up to 150 µg with-out experiencing any appreciable health effects. Therefore, the“worst case” ingested dose (700 µg/person per day) would behigh enough to produce significant hepatic acute effects, consid-ering the steepness of the dose-response curve for acute toxicity.In such a case, the contaminated BGAS batch should be imme-diately withdrawn from the market, as it is impossible to controlits consumption individually or to give indications for dose re-duction.

The second scenario can be considered representative of acontinuative use for limited times, such as 2–3 months (i.e.,during a hypocaloric dietary regimen to loose weight). In thiscase it could be useful to consider the acute reference dose(ARfD), which is 0.4 µg/kg bw/day for MC-LR, correspondingto 24 µg/person/day for a 60-kg bw adult.

In the earlier described worst case, the intake is 25-foldgreater than the threshold for an adult, and much higher for chil-dren administered BGAS for curing ADHD. Indeed, consideringthey have lower body weight, even with a reduced consumption,the intake of MCs could result in a severe risk for their health.

In each of the described “worst case” scenarios for an adult,the related reference value is exceeded 250-, 25-, and 5-foldfor chronic, subchronic, and acute risk, respectively, giving aclear idea of the necessary reduction of BGAS contaminationlevels, maintaining the assumption for the maximal ingesteddose unaltered.

By applying an approach similar to the one used by WHOfor drinking water GV derivation, a provisional limit value forBGAS contamination by MC-LR of 1 µg/g dry weight has beenproposed (Gilroy et al., 2000). However, this limit has no legalvalue, and higher levels of contamination have been demon-strated in a substantial number of samples on the market (Gilroyet al., 2000; Dietrich and Hoeger, 2005). A case report de-scribing the death due to liver failure of a 34-year-old womanhas been very recently published (Dietrich et al., 2007). Al-though the cause–effect evidence has not been definitely es-tablished, available results suggest that chronic consumption ofMC-contaminated A. flos-aquae algae products should be sus-pected as responsible for the patient’s death. Indeed, BGAS usedby the patient contained 2.62–4.06 µg MC-LR equivalents/g dw,and MC-positive immunostaining was observed in the patientliver section (Dietrich et al., 2007).

The presence of cyanotoxins as contaminants of BGAS isa real cause for concern, particularly for children, and cannotbe disregarded or simply underestimated: The Health Authorityshould monitor the situation and, if necessary, establish specificlegal limits and quality controls before BGAS marketing.

4.5. Recreational ActivitiesRecreational activities can represent a source of exposure to

cyanobacteria and their products through direct contact, inhala-tion, and/or ingestion. In settings with transient cyanobacterialblooms, these exposures are unlikely to be associated with achronic risk, but in regions with persistent cyanobacterial bloomsand intensive recreational activities, subacute/sub-chronic expo-sure may be a public health issue.

Stewart et al. (2006b) thoroughly reviewed the available dataand information on the possible relationships between exposureto cyanobacteria during recreational activities and effects onhuman health. The results from the analysis of anecdotal andcase reports show that a range of diverse symptoms is associatedwith exposure to cyanobacteria in recreational settings.

A unique case of fatality has been associated a posteriori bya coroner in the United States to anatoxin-a exposure: The deathof a teenage boy has been attributed to accidental ingestion ofanatoxin-a from a golf-course pond, since the toxin was detectedin his blood and stool postmortem samples. Some uncertain-ties are however related to this association, mainly concerning

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the long latency between exposure and death (some 48 hours),inconsistent with the much shorter time observed in experimen-tal animals (Stewart et al., 2006b).

Some serious acute illnesses have also been described, withsymptoms like severe headache, pneumonia, fever, myalgia, ver-tigo, and blistering in the mouth (Dillenberg and Dehnel, 1960;Carmichael et al., 1985; Stewart et al., 2006c; Turner et al.,1990). An episode has been reported where exposure throughinhalation, after prolonged activities of swimming or canoeingin a surface water affected by a strong Microcystis bloom, ledto 10 out of 20 soldiers suffering respiratory symptoms and 2 ofthem having pneumonia (Turner et al.,1990).

Some descriptions of allergic responses to cyanobacteria havebeen also published (Stewart et al., 2006b). Cutaneous effectsare strongly suggestive of allergic reactions, and symptoms suchas rhinitis, conjunctivitis, asthma, and urticaria also indicateimmediate hypersensitivity responses, which are probably ex-plained by a cascade action of proinflammatory cytokines. Thispicture is consistent with inflammation reactions attributed tocyanobacterial LPS endotoxins, which have been assumed onthe basis of LPS from gram-negative bacteria actions, as alreadymentioned. Indeed, the binding of bacterial LPS to transmem-brane receptors initiate a cascade of host-mediated responses(release of cytokines and other inflammatory mediators, stimula-tion of monocytes and macrophages, congregation of neutrophilsand platelets microcapillaries, vascular injury) (Heumann et al.,2002). However, at present no direct information is available forcyanobacterial LPS endotoxins, and hence no definite conclu-sion should be drawn on the degree of their toxic potential. Asmentioned before, LPS endotoxins could be a good candidatefor a possible role in skin sensitization induction by extractsfrom nonaxenic cyanobacteria, which is independent on othercyanotoxin presence and may indeed be caused by accompa-nying bacteria rather than by cyanobacteria themselves. On theother hand, it should be considered, in relation to effects as-sociated with recreational activities, that during cyanobacterialblooms many other organic compounds (such as aldehydes, ter-penoids, and ketons), are dissolved in water and some of themare endowed with irritating and sensitizing properties. There-fore, it cannot be excluded that in these studies the irritatingand/or sensitizing effects of skin and mucosa were due to theconcomitant presence of different etiologic agents.

The WHO has examined the possible human health implica-tions associated with this exposure and has provided guidelinesfor preventing the risk from irritating and more severe effects(WHO, 2003). These guidelines are based on the outcome of aprospectic epidemiological study (Pilotto et al., 1997) and theknowledge of toxicological properties of MCs.

The guideline value protective from irritating effects hasbeen defined as a cyanobacterial density of 20,000 cells/ml,which corresponds to a low probability of adverse effects. Thisvalue derives from the earlier mentioned epidemiological study(Pilotto et al., 1997), where mild irritating effects were reportedat a density of 5000 cells/ml. Due to the nature of these topic

effects and to the small number of affected people with respectto the total exposed population, WHO has concluded that theydo not represent a sufficiently solid base to justify any action. Atdensities up to 20,000 cells/ml, if MC-producing cyanobacteriaare dominant, it is possible that 2–4 µg/L of MC might occur,and up to 10 µg/L in case of particularly toxic cyanobacteria.These levels are close to the guideline value defined by WHO fordrinking-water quality, which refers to a long-term exposure andto an ingestion of 2 L of water/day. Considering that daily swim-ming sessions might correspond to an involuntary/accidental in-gestion of about 200 ml per person, the corresponding intakeof 0.2–1 µg MC/day would be below the guideline value fordrinking-water quality. Therefore, these situations can be thenconsidered as having low risk, also taking into account that MCswould be orally assumed only for a limited period of time (i.e.,the bathing season).

At densities higher than 20,000 cells/ml the probability ofirritating effects increases; therefore, a second guideline valuehas been set at a density of 100,000 cells/ml, which correspondsto a moderate probability of adverse effects. In the presenceof 100,000 cells/ml when Microcystis spp. are dominant (0.2pg/cell), 20 µg/L MC might be reached (WHO, 2003). Ingestionof 100 ml water per bathing activity would correspond to MCintakes close to the TDI (2.4 µg/adult person and 0.4 µg/childof 10 kg bw) but lower than the acute risk (150 and 25 µg peradult and child, respectively).

When Planktothrix agardhii is dominant, even two-foldhigher concentrations can be attained (WHO, 2003).

Nevertheless, at densities of about 100,000 cells/ml, scumformation may readily cause very high cyanotoxin levels (on theorder of units to tens of milligrams per liter). In these conditions,an ingestion of even a small volume of water would correspondto an intake of fractions of a milligram of MCs, which couldbe dangerous for bathers, especially children. For these reasons,during scum episodes monitoring activities should be intensi-fied and management measures promoted in order to preventdangerous exposures.

These guidelines, defined in order to protect human health,particularly from dangerous exposures to MC, represent an im-portant tool for risk assessment and management of recreationalwaters. However, they have some limits. The guideline valueof 20,000 cells/mL is based on a unique epidemiological study(Pilotto et al., 1997), which shows some important shortcom-ings, discussed later in this article. Moreover, the other valuesrelated to systemic effects have been specifically tailored on thetoxicological profiles of MC-LR, hence excluding other cyan-otoxins. For example, STXs may represent an important po-tential sources of human exposure during recreational activities(Kaas and Henriksen, 2000). Finally, the WHO guidelines areexpressed as cell densities, whereas in some cases (i.e., senes-cent blooms for MCs) high levels of dissolved cyanotoxins occurin water. In these conditions, cell density can be misleading asan indicator for the absence of toxins and/or of risk for humanhealth.

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In the study by Pilotto et al. (1997), no significant differ-ences in symptom occurrence were observed between exposedand nonexposed recruited participants at the second day offollow-up; a significant increase in symptoms was observedonly 7 days after exposure to cyanobacteria densities higher than5000 cells/ml, and was attributed to delayed allergic responses.But, as underlined also by Stewart and colleagues (2006b), theso-called “late-phase” allergic and asthmatic responses tend tooccur earlier (4–24 hours post allergen exposure); furthermore,the size of the cohort was particularly small (93 exposed subjectsand 43 unexposed controls).

Beside the Pilotto study, five additional epidemiological stud-ies are available on recreational exposure to cyanobacteria: threeanalytical cross-sectional studies (Philipp, 1992; Philipp andBates, 1992; Philipp et al., 1992), a small case-control studyfrom Australia (El Saadi et al., 1995), and a prospective cohortstudy (Stewart et al., 2006c).

The three analytical cross-sectional studies were conductedin the United Kingdom on recreational users of six inland waterbodies, five of which were infested by cyanobacterial blooms.From these studies a common outcome was achieved: mostlyminor morbidity, with similar disease pattern across sites.

The case-control study was conducted after an extensive An-abaena circinalis bloom in an Australian river (El Saadi et al.,1995). The case group was represented by patients with GI ordermatological complaints. A significant increased risk of GIand cutaneous symptoms was found in patients using untreatedwater for domestic purposes. Only a small number of people(50 subjects) practiced recreational activities during the studyand for them a nonstatistically significant increase of GI andskin symptoms was observed.

In a prospective cohort study, 1331 subjects were recruitedfrom 19 recreational waterbodies in eastern Australia and cen-tral and northeast Florida (Stewart et al., 2006b). Exposure cate-gories were established according to low, intermediate, and highcyanobacteria levels, based on cell surface area, chosen as theexposure variable of interest. A statistically significant increaseof respiratory symptoms and a pooled “any symptom” category(rated as mild symptoms) was reported amongst subjects ex-posed to high cyanobacteria level (>12 mm2/ml). A similar butnonsignificant relationship was also found for skin, ear, and feversymptom groups. Cyanotoxins were very rarely detected in thewaters; when present the levels of MCs, CYN and anatoxin-awere generally low (around 1 µg/L), and far below the possibil-ity to cause any systemic acute effect.

Due to their nonspecific nature and their degree, whichis likely to be mild and self-limiting, acute cyanobacteria-associated symptoms from recreational exposure do not requiremedical assistance. Therefore, being not regularly registered,unless the problem exceeds the self-medication threshold, theycould be very likely underestimated. On the other hand, symp-toms such as dermatitis and GI effects can also be erroneously at-tributed to exposure to cyanobacterial metabolites. Indeed, thosekinds of disturbances can be caused by other chemicals dissolved

in water during cyanobacterial blooms, as well as by other riskfactors present in fresh waters, like avian cercariae, other bacte-ria and/or viruses, or physical stimuli that can induce nonallergicurticaria (Stewart et al., 2006c).

In conclusion, on the basis of anecdotal, epidemiological andtoxicological data, it appears that the risk of severe effects forbathers is posed by cyanobacteria only when they bloom or formscums.

This issue has been well addressed in an article of the newDirective of the European Union (EU Directive 2006/7/EC),concerning the management of bathing water quality. In relationto cyanobacterial risks, article 8 states the following:

Cyanobacterial risks

1. When the bathing water profile indicates a potential for cyanobacterial pro-liferation, appropriate monitoring shall be carried out to enable timely iden-tification of health risks.

2. When cyanobacterial proliferation occurs and a health risk has been identifiedor presumed, adequate management measures shall be taken immediately toprevent exposure, including information to the public.

With this article, the EU directive on one hand recognizes theimportance of cyanobacterial risk in bathing waters, and on theother, provides neither general nor specific limits, taking intoaccount the complexity of this issue and the still scant informa-tion available on known/unknown toxins and their toxicologicalprofile. Local authorities are asked to assess the risk in the spe-cific water body and to promote appropriate measures to pre-vent dangerous exposures, including information to the publicon the possible risks. Assessing the risk and managing it requireadequate specific skills and can be quite challenging for localauthorities. To this aim, central authorities should address thisissue, promoting also adequate training activities.

5. OPEN ISSUES AND RESEARCH NEEDSThe analysis of the available literature has evidenced a num-

ber of gaps, some of which are listed here, with no aim to beexhaustive. It appeared that additional efforts have to be put in(1) the description of patterns of cyanotoxins occurrence and re-lated levels of exposure for the population, (2) the identificationof possible “new” cyanotoxins, (3) the influence of environmen-tal factors in cyanotoxin production, and (4) the identificationof repeated-dose toxicity for cyanotoxins other than MC-LR,necessary for the derivation of guidance values and regulatorylimits. Some other specific issues have received very limitedattention up to now, and they are discussed next in order to stim-ulate some new research.

Usually the quality of the risk assessment is dependent onthe amount and reliability of available toxicological data, result-ing in the reduction of uncertainties. Differences among speciesare very often attibutable to differences in toxicokinetics, someof which have been shown, but only for few MC congeners(Meriluoto et al., 1990). Those results suggest that minimalstructural changes in the molecule could result in three- to four-fold differences in the uptake, tissue distribution, and excretion

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kinetics. The majority of these processes are regulated by theaffinity of single variants for both active transporters (uptake,distribution, and excretion) or for enzymes catalyzing conju-gation with GSH (metabolism and excretion), affecting plasmalevels, bioavailability, and clearance of specific MC congeners.These considerations imply that also species-specific differencesmay be expected. Therefore, the knowledge of the toxicokineticsproperties of different cyanotoxins and congeners may be ex-tremely relevant, when data should be extrapolated from exper-imental animals to humans during risk assessment processes.Unfortunately, such information is scant and limited to MCs.

Other possible susceptibility factors can modulate the out-come of cyanotoxin exposure in humans. They may be linkedboth to differential exposure (e.g., children drink a higher watervolume in proportion to their body weights) and/or to patholog-ical status (e.g., people with injury to liver and kidney, targetorgans of cyanotoxin action, could be more susceptible thanhealthy people).

In addition, the involvement in MC toxicokinetics of OATPand GSTs, as shown by animal testing, may suggest the pres-ence of groups of people differently susceptible to MC-inducedeffects. Indeed, both enzymatic families are characterized bygenetic polymorphism, responsible for different levels of ex-pression and enzymatic activity within the human population(Strange et al., 2000; Konig et al., 2006). As an example, indi-viduals characterized by deletion of GSTM1 and GSTT1 genes,endowed with a decreased level of glutathione transferases ac-tivity, could experience a lower formation of conjugates and alonger half life of the parent compound, resulting in a higherrisk of MC-induced toxic effects. GST genes deletions could beconsidered as a good candidate as biomarker of susceptibilityfor hepatotoxic effects associated with MC exposure.

These possibilities should be carefully taken into accountwhen intraspecies differences have to be considered. However,at present no data are available on the role of human enzymesin MC biotransformation reactions, and in the case of the othercyanotoxins, almost nothing is known about their toxicokineticseven in experimental animal models.

Cyanobacterial blooms are generally characterized by thepresence of several toxins. Therefore, the actual “toxic agent”is represented by a mixture of cyanotoxins. When only MCs arepresent, by adopting a conservative approach, toxicity may bereferred to MC-LR equivalents; since this congener is the mostacutely toxic (when administered ip), the toxicity of the mixtureis likely to be overestimated by this approach. MC and NODvariants share the same mechanisms of action and targets, al-though with different potency; on this basis, in analogy with themethod used for polychlorinated dibenzo[p]dioxins (PCDD), ithas been proposed that one derive a “toxicity equivalent factor”(TEF) from the available toxicological data for MCs and NODs(Wolf and Frank, 2002). Among the PCDD group, the referenceand most toxic congener is 2,3,7,8-TCDD, to which a defaultTEF value of 1 is attributed. The specific TEF for the other con-geners is established by comparison of their toxicity potency

TABLE 5Application of TEF method to a mixture of MCs and NOD∗

Concentration in the i.p. LD50 ToxicityToxin mixture (µg/L) (µg/kg) TEF Equivalent

MC-LR 30 50 1.0 30MC-RR 100 500 0.1 10MC-YR 15 150 0.33 5NOD 20 50 1.0 20MC-AR 60 250 0.2 12Total 225 77

TEF = Toxicity Equivalent Factor. i.p. = intraperitoneal.∗As described by Wolf and Frank (2002).

with TCDD as 1. The toxicity of the mixture is then obtained bysumming up the product of specific TEF with the concentrationof the related congeners. The same approach could be adoptedwhen the acute toxicity of MCs and NODs has to be evaluated.The approach should at present be limited to acute toxicity, sincerepeated toxicity data on different congeners are not available.

The reference cyanotoxin is MC-LR, with TEF = 1; the TEFof a specific toxin (X) is derived as the ratio between the LD50

values, according to the equation:

TEFX = LD50 MC-LR/LD50X

The total acute toxicity of the mixture is estimated by the sumof all the individual toxicity equivalents obtained as the productbetween the specific TEF and the toxin concentration. As shownin Table 5, by using this approach a more realistic assessmentis obtained for a hypothetic mixture when compared with the“worst case” approach, considering all the component as toxicas MC-LR, which is threefold higher.

As soon as data are available on variants of cyanotoxins char-acterized by different mechanisms and endpoints (i.e., neuro-toxins), the acute toxicity of their mixtures could be assessedsimilarly.

A model to predict the combined neurotoxic effects of binaryand ternary mixtures of STX derivatives that can accumulate infish has been recently proposed (Llewellyn, 2006). The behaviorof different mixtures of PSP indicates that the most potent toxinis by far the most relevant component, whereas the less toxicderivatives should be order of magnitude more concentrated tocontribute to the cumulative toxic potency (Llewellyn, 2006).

Since blooms are quite often characterized by the presenceof a mixture of cyanotoxins, combined exposure represents verylikely the rule rather than the exception. Nevertheless, this is-sue has not been sufficiently investigated so far. Interaction be-tween different cyanotoxins can occur, when they influence eachother toxicokinetics (i.e., by sharing common mechanisms ofabsorption or competing for the same detoxication enzymes),although their targets may be different. As an example, it hasbeen reported that LPS toxins may exert a synergistic effects

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on MC-LR-induced toxicity in fish (B. rerio), since LPS seemto decrease the GST activity, thus reducing MC detoxication(Best et al., 2002). At present, no information is available onmammals.

Another clear example is given by results showing thatcyanobacterial extracts containing known amounts of MC-LRare more toxic than the equimolar administration of the puretoxin. This issue is very elegantly presented and discussed in arecent letter to the editor (Falconer, 2007). The higher toxicityof the extracts suggests that studies on fractionated cyanobac-terial preparations are needed in order to identify the bioactivecompounds, and, at the same time, that availability of resultsobtained with pure cyanotoxins is crucial before any conclusionon their toxicological profile can be drawn.

Among the possible explanations for the higher toxicity of theextracts, which need to be experimentally confirmed, the possi-bility of interactions beetween different cyanobacterial compo-nents should be considered. Indeed, LPS may contribute to dis-rupting the gut mucosal barrier, increasing absorption and con-sequently the pathological effects of cyanobacterial exotoxins.The accumulation of different cyanotoxins up to high levels inaquatic organisms usually eaten by humans may contribute tofurther highlight the relevance of this issue. The possible mul-tiple exposure to these toxins should be considered at the locallevel, depending on both the presence of cyanotoxins in the wa-ter body, the features of the aquatic organisms, and the dietaryhabits of the popuylation, in order to understand which are themore plausible and frequent co-occurrence cases, to study theeventual implication for human health.

In addition, concomitant exposure with other chemicals, suchas products intended for human health, including drugs, shouldbe considered at least in those cases in which a prediction of in-teractions could be foreseen by comparing their toxicokinetics,as well as the toxicodynamics properties. Such interaction in-formation, once generated, should be progressively included inrisk assessment in order to improve risk characterization for spe-cial subgroups within the population. Indeed, all those chemicalsable to induce/inhibit transport systems or drug-metabolizing en-zymes could potentially interfere with cyanotoxin toxicity (andvice versa), as well as chemicals acting on the same targets. Agood example can be represented by organophosphorus pesti-cides, AChE inhibitors similarly to anatoxin-a(s), to which theyare structurally related. Exposure to the insecticides could po-tentiate the cyanotoxin-induced toxicity; indeed, they exert astrong inhibition of AChE activity also in the brain, whereasthe action of anatoxin-a(s) is limited to the peripheral nervoussystem (Cook et al., 1988). Therefore agriculture workers usingorganophosphorus pesticides may represent a group at higherrisk when exposed to anatoxin-a(s).

Analogously, people consuming antihypertensive and hy-polipidaemic drugs may be at higher risks of hepatotoxicity,according to a report describing the case of a 52-year-old maleregularly consuming Spirulina as a food supplement. He ex-perienced hepatic dysfunctions after 2 weeks of therapy with

antihypertensive and hypolipidemic drugs. Three weeks later hewas hospitalized, and liver degenerative changes were found.Since the withdrawal of dietary supplements led to a completerecovery, hepatotoxicity was attributed to consumption of Spir-ulina, causing interaction effects with the hypolipidemic drugsimvastatin, through potentiation of the LPS-induced inflamma-tory responses (Iwasa et al., 2002).

Another example is related to the postulated protective actionof alcohol consumption on PSP toxicity. In a case-control studycarried out in Alaska in order to study outbreaks associated withtoxin exposure, it has been shown that alcohol consumption andeating cooked rather than raw shellfish were associated with areduced risk of PSP-induced effects. The mechanism to explainthe risk reduction is unknown; it has been tentatively attributed tothe diuretic effect of alcohol, increasing PSP excretion throughthe urine, or to hepatic enzyme induction caused by ethanol(Mons et al., 1998). However, no data are available to supportthese considerations, and strong needs for future research in thefield are evident.

ACKNOWLEDGMENTSThe authors thank Drs. Annalaura Stammati and Daniela

Mattei for their support and advice.

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