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In: Dams: Strucuture, Performance and Safety Management Editor: Slaheddine Khilifi

ISBN: 978-1-62417-702-6 © 2013 Nova Science Publishers Inc.

Chapter 9 CYANOBACTERIAL BLOOMS IN DAMS: ENVIRONMENTAL FACTORS, TOXINS, PUBLIC HEALTH, AND REMEDIAL MEASURES Soumaya El Herry-Allani1 and Noureddine Bouaïcha2* 1 Laboratory of Microbial Ecology and Technology, Department of Biological and Chemical Engineering, National Institute of Applied Sciences and Technology, University of Carthage, Tunisia. 2 Laboratoire Ecologie, Systématique et Evolution, UMR 8079-Univ. Paris Sud/CNRS/AgroParisTech, Université Paris-Sud, France. * Email : [email protected]

ABSTRACT The eutrophication of dams becomes widely recognized as a sign of severe reduction in water quality arising as a consequence of the increase in levels of nutrients, through human activity, whether from intensification of agriculture or human waste disposal. These eutrophic waters provide growth opportunities for green algae or diatoms or they may be seasonally subject to substantial cyanobacterial blooms causing water quality problems for fisheries, aquaculture, farming, and sanitary hazard for humans and animals. This chapter describes the present knowledge of toxic species of cyanobacteria and how environmental factors may limit their development in dams, the chemistry and toxicology of the most relevant cyanotoxins, bioaccumulation of these toxins in aquatic organisms and its consequence on animal and human health, prevention and remediation of cyanobacterial blooms in dams, and water treatment technologies of cyanobacteria and their toxins. Key words: cyanobacteria, cyanotoxins, microcystins, public health, water treatment



The United Nations World Water Development Report affirmed that, by 2050, at least 25% of people are likely to live in a country affected by chronic or recurring shortages of freshwater. It seems, therefore, that the water is becoming one of the challenges of the future (such as demography, health and development) and that its conservation as its management requires an international mobilization. Consequently, a number of countries have been undertaking important efforts for the mobilization of surface water resources by the construction of more dams to face on one hand the limitation of groundwater reserves and the demographic growth and on the other hand to support their economic and social development. However, dams are the points of convergence of all the hydraulic streams of their ponds. These preferential zones of accumulation turn out to be the amplifier of the entire imbalance engendered upstream (Gonay and Lafforgue, 1997). The expansion of urban areas associated with the increasing agricultural pressure favor the progressive eutrophication of surface waters. Thus, this phenomenon has become inevitable in many water bodies in several regions. Eutrophication of aquatic ecosystems is the source of a natural phenomenon resulting from the enrichment of waters by nutrients. Indeed, the soil leaching and runoff and anthropogenic releases contribute to the increase of the stock of nutrients in surface


waters, generating a production increased of plant biomass (phytoplankton, macrophytes), an oxygenation decreased, and an accumulation of partially degraded material contributing to the thickening of the sediment layer. Besides the increase in plant biomass, eutrophication also implies a change in the composition and structure of phytoplankton communities. This often involves the massive development of a limited number of species or even a single species, which called algal blooms. In freshwater, blooms are often dominated by cyanobacteria, which can last from several days to several months. The problems associated with algal blooms are diverse, from the environment asphyxiation due to excessive consumption of oxygen, to purely aesthetic problems in recreational areas when the blooms are a colorful and often smelly scum on the water surface (Peter et al., 2009; Li et al., 2010). To these common problems, are added production by certain species of cyanobacteria of various toxins (hepatotoxic, neurotoxic, and dermatotoxins) that have adverse effects on aquatic fauna and flora, and animal and human health (Wood et al., 2010; Chen et al., 2009a). In recent years, many countries around the world have been facing serious problems due to toxicity of cyanobacterial blooms in dams used for the production of drinking water or recreation areas. Many cases of allergies or poisonings have been reported worldwide in humans after contact with toxic cyanobacteria or via drinking contaminated water. This chapter aims to provide a current description of present knowledge of the development of toxic cyanobacteria in dams and their consequences on animals and public health.



Cyanobacteria, also known as blue-green algae, are photosynthetic prokaryotes with no structured nucleus that form the most wide-spread group of photobacteria with 150 genera and more than 2000 species (Bourrelly, 1985; Holte et al., 1998). They have a cosmopolitan distribution and colonize a great variety of ecosystems, but especially associated with water bodies such as rivers, lakes and dams (Whitton and Potts, 2000; Kaebernick and Neilan, 2001). Favorable cyanobacterial growth conditions that often lead to blooms in dams (Figure 1) usually in summer to late autumn, in both subtropical and temperate latitudes are a combination of abiotic factors (stratified water bodies and light intensity, water temperature, alkaline pH, and eutrophic water conditions) and biotic factors (zooplankton grazing) (Park et al., 1993; Kaebernick and Neilan, 2001; Haider et al., 2003; Dörr et al., 2010).

Figure 1. A surface bloom of Microcystis (Dam Lalla Takerkoust, Marrakech, Morocco). Photograph by N. Bouaïcha.


Abiotic factors

Many planktonic cyanobacteria contain gas vesicles, which have a lower density than water, and therefore can provide cyanobacterial cells with buoyancy. In a stable water column, buoyant


cyanobacteria like Microcystis, Anabaena, Plankthotrix, Aphanizomenon, Nodularia, and Cylindrospermopsis may float towards a higher level, and thus can form scums on the water surface (Walsby, 1998; Stal et al., 2003; White et al., 2006). In fact, the genus Microcysti, which has higher light requirements, is known to engage in a vertical dial migration linked to changes in colony buoyancy caused by gas vesicules. This capacity confers a substantial ecological advantage to this genus, as it can congregate at favorable levels in the water of stratified lakes and also move up and down in the water column to maximize photosynthesis in the surface layers and nutrient uptake in the deep layers (Wallace and Hamilton, 1999; El Herry et al., 2008). However, nonbuoyant cyanobacteria like the genus Oscillatoria, which is known to tolerate very low light conditions, can accumulate at a depth where the combination of photon irradiance and nutrient availability is most favourable for its growth (El Herry et al., 2008). Therefore, the capacity to harvesting light and conducting photosynthesis of different cyanobacteria taxa in freshwaters is not homogeneous and also depends on the species. An average light intensity associated with good oxygenation of the water may be beneficial for a mass growth of cyanobacteria (Skulberg, 1984). Some species can tolerate low levels of light and can effectively dominate the other planktonic algae for the available light, mainly due to the presence of photosynthetic pigments including phycobiliproteins which gives them the ability to use more of the light spectrum and thus be developed at lower light intensities. Therefore, the increased turbidity with low light intensities also promotes adaptability and the dominance of cyanobacteria over other algal species. Indeed, in laboratory experimental conditions, it was shown that the ability of cyanobacteria to use a wide spectral band while depriving other species of light radiation, contributes to their competitiveness (Walsby et al., 1997; Huisman et al., 1999). Several other studies have been shown that the light intensity was also a major factor in the regulation of the production of toxins by cyanobacteria (Van der Westhuizen and Eloff, 1983; Watanabe and Oishi, 1985; Codd and Poon, 1988; Sivone, 1990; Rapala et al., 1997). In laboratory study, Deblois and Juneau (2010) have grown a strain of Microcystis aeruginosa (UTCC299) under a wide range of light conditions and they showed that microcystins (MCs) content decreased with increasing photo irradiation. However, Preubel et al. (2009) reported that temperature-light combinations seem to trigger cylindrospermopsin (CYN) production in two Aphanizomenon flos-aquae strains (10E9 and 22D11). Cyanobacteria blooms display a range of temporal dynamics. Some dams have seasonal blooms that start in summer and last into autumn, and some others have persistent blooms that encompass all seasons. Maximum growth rates are attained by most cyanobacteria at temperatures above 25°C. For example, within the North-African basin, in Egypt (Mohamed et al., 2003), in Morocco (Oudra et al., 2001), in Algeria (Nasri et al., 2007), and in Tunisia (El Herry et al., 2007; 2008), cyanobacterial blooms have been observed in dams during the warmest months particularly in summer and in early autumn. The temperature was also evaluated as a factor limiting the production of toxins by cyanobacteria. In fact, Runnegar et al. (1983) reported that toxin production was increased fourfold by varying the temperature of the culture media of the genus Microcystis from 18°C to 29°C. However; Watanabe and Oishi (1985) and Sivonen (1990) reported that toxin production by cyanobacteria was reduced at high temperatures. Several other studies have been shown that most strains of the species M. aeruginosa produce fewer toxins out of the temperature of 22°C considered optimum for this species (Van der Westhuizen and Eloff, 1983, Codd and Poon, 1988). The pH of the water also plays an important role in the development of cyanobacteria (Pick and Lean, 1987) and the production of cyanotoxins (Van der Westhuizen and Eloff, 1983). Thus, when algal blooms, the consumption of dissolved carbon dioxide becomes important during photosynthesis which increases the pH and subsequently promotes the growth of cyanobacteria. Contrary to algae, cyanobacteria can also use the bicarbonate as the source of inorganic carbon besides the CO2, what can be a competitive advantage when the pH is raised and when the system is concentrated in bicarbonate but weak in CO2 (Pick and Lean, 1987). Therefore, alkaline pH seems much more favorable to the growth of cyanobacteria (Shapiro, 1997). Indeed, strains of the genus Arthrospira prefer highly mineralized water, alkaline and hot and can grow up to a pH of 11 (Couté and Bernard, 2001). On the other hand an acid pH does not seem to much affect the proliferation of cyanobacteria, as they are able to proliferate in acidic media such as acid bogs (Couté and Bernard, 2001).


The availability of nutrients such as macro-elements, phosphorus (P) and to a lesser extent nitrogen (N) is essential for growth of cyanobacteria. For example, several studies have indicated that freshwater cyanobacteria blooms are typically associated with eutrophic (nutrient enrichment) and poorly flushed waters (Paerl, 1988; Carmichael, 1997; Lee et al., 2000; Albay et al., 2003). In fact, an increase in nutrient loads from either agricultural runoff or sewage treatment plants resulted in extensive populations of bloom-forming cyanobacteria (George and Bradby, 1993; Humphries and Robinson, 1995). Moreover, it has been indicated that eutrophication, especially by P, often leads to significant shifts in phytoplankton species composition towards bloomforming cyanobacteria (Steinberg and Gruhl, 1992; Steinberg and Hartmann, 1998). Similarly, Gobler et al. (2007) have been shown that the dominance of Microcystis sp. blooms during the summer is linked to nutrient loading, which can stimulate growth and cellular toxin synthesis, and in turn can suppress grazing by mesozooplankton and allowing for further accumulation of cyanobacterial cells. Furthermore, the low ratio of total N/total P is responsible for cyanobacteria dominance under eutrophic conditions (Smith, 1983; Paerl 1988; Havens and Walker, 2002; Scheffer et al., 1997). The predominance of high light requiring N2 fixers’ cyanobacteria, such as Anabaena circinalis and A. flos aquae, predictably is linked with stable water columns, depletion of dissolved inorganic N and high temperature. Both species are more common in relatively deep reservoirs, but can bloom in shallow reservoirs during calm summer periods. Phlips et al. (1997) indicate that changes in biomass of N2 fixing cyanobacteria and density of heterocysts were strongly coupled with depletion of dissolved inorganic N, and relatively high irradiance. Thus, we can generally predict that if an ecosystem is enriched with P, it is likely to have cyanobacteria dominance. Furthermore, some micronutrients such as iron and molybdenum have a rather special importance in the proliferation of cyanobacteria. For example, iron plays a direct role in N fixation and photosynthesis (Lukac and Aegerter, 1993). In dams, predicting cyanotoxin levels is even less certain than predicting cyanobacterial blooms occurrence. The production of cyanotoxins is linked with dominance by particular taxa of cyanobacteria, including species of Microcystis, Anabaena, Aphanizomenon, Oscillatoria, Nostoc, and Planktothrix. Some genera can fix N, such as Anabaena and Nodularia, and show high levels production of peptide MCs and nodularins (NODs), respectively even in the absence of N in the medium (Rapala et al., 1993). In contrast, Sivonen and Jones (1999) reported that species belonging to the genera Microcystis, Oscillotaria and Planktothrix (not molecular N -fixing) allow significant production of MCs in an environment rich in N. In addition, their research indicates that under limited conditions in P, low levels of cyanobacterial hepatotoxins have been recorded (Sivonen and Jones, 1999). Recently, Davis et al. (2009) assessed the effects of temperature and nutrients on the growth and dynamics of toxic and non-toxic strains of Microcystis by quantifying the microcystin synthetase gene (mcyD) and the small subunit ribosomal RNA gene (16S) as an indicator of total Microcystis biomass. Their results indicated that increases in temperature and P concentrations yielded the highest growth rates of toxic Microcystis cells suggesting that eutrophication and climatic warming may additively promote the growth of toxic, rather than nontoxic, populations of Microcystis, leading to blooms with higher microcystin content. Several earlier studies have suggested links between toxin concentrations and ratios of particulate to dissolved nutrients (Oh et al., 2001), concentrations of soluble P (Jacoby et al., 2000), total P levels (Rapala et al., 1997), total N levels, and irradiance (Rolland et al., 2005). Therefore, factors controlling the amount of toxin produced during a bloom are not well understood and additional research is required.


Biotic factors

Cyanobacteria are generally a poor source of nutrition for zooplankton and are often selectively avoided. Their populations can therefore increase relatively more than other algae which are more digestible. It has been well documented that many types of zooplankton are unable to graze in presence of high densities of cyanobacteria (Boon et al., 1994; Rohrlack et al., 1999; Paerl et al., 2001; Ghadouani et al., 2003; Soares et al., 2009). Consequently, during cyanobacterial blooms and following exhaustion of alternative sources of food, some populations of zooplankton disappear. Ger et al. (2009) investigated the toxicity and post-exposure effects of dissolved MCLR on the dominant copepods of the upper San Francisco Estuary, where blooms of the toxic cyanobacteria M. aeruginosa coincide with record low levels in the abundance of pelagic


organisms including phytoplankton, zooplankton, and fish. They observed that dissolved MC-LR above 0.14 mg/L proved likely to have chronic effects on the survival of copepods in this estuary. However, DeMott et al. (1991) reported that some other species such as Daphnia pulicaria and D. pulex may develop physiological and behavioral adaptations to survive in the presence of toxic species of cyanobacteria, which leading to changes in the population dynamics of zooplankton. Indeed, Ferrao Filho et al. (2008) assessed the effects of a saxitoxin-producer strain (T3) of the cyanobacterium species C. raciborskii on the swimming movements of three cladoceran species (D. gessneri, D. pulex, and Moina micrura), and they observed that the species D. pulex was extremely sensitive to the T3 strain and the species M. micrura was intermediate in sensitivity, however, the species D. gessneri was not sensitive. Therefore, the pressure of zooplankton on growing of cyanobacteria is reduced by the role of predator fish; they release nutrients such as P, which is later assimilated by cyanobacteria. Furthermore, the optimal growth of cyanobacteria may occur in the presence of other heterotrophic aquatic bacteria. These bacteria can also compete with cyanobacteria for nutrients (Zubkov et al., 2003). For example, Paerl (1996) has observed that the species Pseudomonas aeruginosa is localized around the heterocysts of the cyanobacterium species Anabaena sp. to take advantage of its activity of N fixation. Similarly, cells of M. aeruginosa can fix more carbon dioxide in the case of an association between them and other bacteria (Paerl, 1996). It was thus suggested that the production and excretion of metabolites such as extracellular cyanotoxins, may play a role in both the attraction of other hosts and the repulsion of antagonistic microorganisms and higher order grazers (Paerl, 1996).

3. TOXIC CYANOBACTERIA AND THEIR TOXINS Cyanobacteria biosynthesize a large number of secondary metabolites (Namikoshi and Rinehart, 1996; Van Wagoner et al., 2007; Rastogi and Sinha, 2009; Ducat et al., 2011) with various biological effects (for review see Rastogi and Sinha, 2009): antiviral and antifungal activities (Patterson et al., 1994), cytotoxic (Carmichael, 1994), inhibitors of protein phosphatases (Honkanen et al., 1991), antineoplastic (Sainis et al., 2010), and allelopathic (Pushparaj et al., 1999). However, some of these secondary metabolites, known collectively as cyanotoxins, are associated with a wide range of adverse human and animal health effects (Carmichael and Falconer, 1993; Falconer, 1996, 1999; Kuiper-Goodman et al., 1999, Hitzfeld et al. 2000), and some others known as taste and odorous compounds (especially geosmin and methylisoborneol) result in malodorous or unpalatable drinking water (Saadoun et al., 2001; Tsujimura and Okubo, 2003; Wang et al., 2005). Producing species of cyanotoxins that have been involved in such incidents belong mainly to the genera Microcystis, Anabaena, Aphanizomenon, Planktothrix, Oscillatoria, and Nodularia (Sivonen and Jones, 1999). Cyanotoxins can be divided into three families according to the organs on which they act: neurotoxins (nervous system), hepatotoxins (liver), and dermatotoxins (skin). The neurotoxins (for review see Aráoz et al., 2010) are classified in four groups (Figure 2): saxitoxins which block nerve cell voltage-gated sodium channels (Ikawa et al., 1982; Mahmood and Carmichael, 1986; Humpage et al., 1994; Haney et al., 1995; Pomati et al., 2001), anatoxins which are neuromuscular junction blocking agents (Aráoz et al., 2010), anatoxin-a(s) which resembles an organophosphate insecticides, with effects exerted through irreversible inhibition of acetylcholinesterase at the nerve synapse (Mahmood and Carmichael, 1986), and the unusual non protein neurotoxic amino acid L-beta-N-methylamino-L-alanine (BMAA) which has been associated to the neurological disorder amyaotrophic lateral sclerosis/Parkinsonium dementia complex (ALS/PDC) among the indigenous Chamorro people of Guam and other Marianas islands (Monson et al., 2003; Murch et al., 2004). Its neurotoxicity may be mediated via glutamate regulation (Papapetropoulos, 2007). Contrary to the other neurotoxins which their production depends on the phylogeny of the species, the BMAA can be produced by almost all groups of cyanobacteria from freshwater, brackish, and marine environments (Cox et al., 2005; Banack et al., 2007). The most ubiquitous cyanobacterial hepatotoxins can be divided into two groups (Figure 3): cyclic peptides MCs and NODs and alkaloid CYNs (Sivonen and Jones, 1999). MCs and NODs are respectively, cyclic heptapeptides with the general structure (-D-Ala-X-D-MeAsp-Z-Adda-DGlu-Mdha-) and pentapeptides with the general structure (-D-MeAsp-Z-Adda-D-Glu-Mdhb-),


where X and Z are variable L-amino amino acids, D D-MeAsp is D-erythro-β-methyl methyl aspartic acid, Mdha is N-methyldehydroalanine, Mdhb is N- methyldehydrobutyrine, and Adda is 3-amino-9-methoxy methoxy2,6,8-trimethyl-10-phenyldeca-4,6-dienoic dienoic acid. Multiple combinations of the variable amino acids (X and Z for MCs, and only Z for NODs) make the difference between more than 80 MC variants already reported (Sivonen and Jones, 1999; Codd et al., 2005; Del Campo and Ouahid, 2010)) and only 9 NODs (Codd et al., 2005; Mazur Mazur-Marzec et al., 2009a). Both MCs and NODs are waterwater soluble molecules and their cyclic structure provides them a high chemical stability (Sivonen and Jones, 1999).. Their toxicity resulted on a potent and specific inhibition of serine/threonine protein phosphatases (Mackintosh et al., 1990) 1990). They also have known to induce oxidative stress (Bouaïcha and Maatouk, 2004; Amado and Monserrat, 2010). Recently, Oziol and B Bouaïcha ouaïcha (2010) reported for the first time the estrogenic potential of the NOD-R (NOD-R) and the MC-LR LR in vitro in a stably transfect cell line with an estrogen estrogen-regulated luciferase gene. The CYN toxin is cyclic guanidine alkaloid which initially has been isolated from the filamentous C. raciborskii (Hawkins et al.,, 1985), and further from other species A. ovalisporum (Banker et al., 1997; 2000), A. bergii (Schembri et al., 2001), Umezakia natans (Terao et al., 1994) and Raphidiopsis curvata (Li et al., 2001). 1). It inhibits the synthesis of protein, resulting in wide spread necrosis of the tissues of many organs (Ohtani et al., 1992; Falconer et al., 1999; Runnegar et al., 2002). Two structural variants of CYN (7-epicylindrospermopsin epicylindrospermopsin and deoxycylindrospermopsin) have been characterized so far from bloom samples and isolated strains of cyanobacteria (Banker et al., 2000; Norris et al., 1999; Li et al., 2001).

Figure 2.. Chemical structures of cyanobacterial neurotoxins. ((A) Saxitoxin, (B) β-Nmethylamine-L-alanine, (C) Anatoxin-a, a, ((D) Homoanatoxin-a, and (E) Anatoxin-a(s). The freshwater cyanobacterial dermatotoxins such as lipopolysaccharides (LPS), which initially have been isolated from the cyanobacterium Anacystis nidulans (Weise et al.,, 1970), are external components of cell membranes of most cyanobacteria as well as Gram Gram-negative negative bacteria (Mayer and Weckesser, 1984; Kaya, 1996). They are frequently cited in the cyanobacteria literature as endotoxins associated with adverse human heath effects, such as skin rashes, gastrointestinal, and respiratory and allergic reactions (for review see Stewart et al., 2006). Nevertheless, the few results available indicate that cyanobacterial LPS is less toxic than that of Gram-negative enteric bacteria, acteria, such as Salmonella (Kuiper-Goodman et al., 1999).


Figure 3.. Chemical structures of cyanobacterial hepatotoxins. ((A)) General structure of microcystins, (B)) General structure of nodularins, ((C) Cylindrospermopsin.


Livestock and wildlife poisoning by cyanobacterial toxins

A significant number of cases of animal poisonings are reported in the literature. George Francis was the first, on 1878, to involve toxic cyanobacteria in the death of farm animals (cattle, cattle, sheep, horses and pigs) that had consumed water containing cyanotoxins from Alexandrina Lake, South Australia (Francis, 1878). Subsequently, aanimal poisonings by water contaminated with toxic cyanobacteriaa have been the subject of several studies in the world. Cyanotoxins can cause symptoms of disease or cause the death of mammals, birds or fish that ingest a sufficient quantity of toxic cells or extracellular toxins (Codd et al. 1992; Carmichael, 1994; Kuiper-Goodman et al., al 1999). Carmichael (1992) reported mortality of animals that have consumed water containing large numbers (> 106/mL)) of cyanobacterial cells. Animals unable to select their food such as birds and fish are both directly affected cted by hepatotoxins and neurotoxins (Kaebernick and Neilan, 2001). The major route of exposure of animals to cyanotoxins is ingestion ingestion. Recently,, there have been increasing studies to evaluate MCs contamination in aquatic vertebrates from surface waters with cyanobacterial blooms, but mainly focusing on fishes (Vasconcelos, 1999; Mohamed et al., ., 2003; Ibelings et al., 2005; Chen et al., 2007;; Wilson et al., ., 2008). Similar information is relatively rare for other aquatic vertebrates, although MCs MCs-producing roducing cyanobacteria have been associated with deaths of wild birds and turtles over recent years around the world, Kenya (Metcalf et al., ., 2006), Tanzania (Lugomela et al.,., 2006), Japan (Matsunaga et al., 1989), Canada (Murphy et al., ., 2003), Algeria (Nasri et al., 2008) and Belgium (Wirsing et al., 1998). All these disturbances could limit the introduction of herbivorous fish in the dams in order to fight against eutrophication.


Human poisoning by cyanobacterial toxins

Humans may be exposed to cyanotoxins via several routes, the oral one occurring by ingesting contaminated drinking water, food, some dietary supplements, or contact with water during recreational activities. Allergic llergic reactions after contact with cyanobacte cyanobacteria in surface waters are relatively frequent. Respiratory problems (asthma, hay fever), eye irritation and skin, dermatitis, fever, gastrointestinal disorders isorders and liver liver,, sore throat and signs of pneumonia were reported following contact with certain speci species of the genera Oscillatoria, Nodularia, Anabaena, Anabaena Aphanizomenon, or Microcystis (Carmichael and Falconer, 1993; Pilotto et al., 1997; Backer et al., al 2008; 2010).. Cases of irritation have also been reported after a shower with a poorly treated water


containing MCs (Falconer, 1999). The WHO guidelines, for example, are based on the following three levels of potential public health threat: (i) low probability of adverse health effects from water with 20,000 cells/mL or 10 g chlorophyll-a/L where cyanobacteria predominate, (ii) moderate probability of adverse health effects from waters with 100,000 cells/mL or 50 g chlorophyll-a/L, and (iii) high probability of adverse health effects from contact with, ingestion, or inhalation of cyanobacteria when algal scum appears on the water surface. The oral exposure to cyanotoxins, occurring by ingesting contaminated drinking water, is considered the most important route. Several human poisonings after consumption of water from different reservoirs, which had accumulated cyanobacteria, have been identified worldwide. The degree of poisoning depends on the person' s age and health status, with children being the most sensitive. Poisonings induced by cyanobacterial hepatotoxins are more common than those caused by neurotoxins (Hitzfeld et al., 2000). In 1983, at Palm Island in Australia, an outbreak of hepatitis and enteritis was caused by a bloom of toxic cyanobacteria in the dam of Solomon. One week after the treatment of the cyanobacterial bloom by copper sulfate, 148 people including 138 children who drank the water were severely ill. Symptoms were hepatitis, intestinal and kidney damage, vomiting, headache, abdominal pain, blood loss, glucose and protein in urine, constipation followed by profuse bloody diarrhea and severe electrolyte imbalances (Bourke et al., 1983). Two species of cyanobacteria have been incriminated in the poisoning: a non-toxic strain of A. circinalis and a highly toxic strain of C. raciborskii known to produce CYN (Hawkins et al. 1985; Falconer, 1991). During the same year (1983), in the small town of Armidale, Australia, liver damage characterized by an increase in the activity of certain liver enzymes were observed in people who drink water from a reservoir containing a bloom of M. aeruginosa. Peaks of γglutamyl transferase were found in samples of blood of intoxicated persons (Falconer et al., 1983). In 1993, a massive bloom of Anabaena sp. and Microcystis sp. in Itaparica dam in Brazil has caused poisoning of 2000 people showing signs of acute gastroenteritis, which has killed 88 patients mostly children (Teixeira et al., 1993). Another fatal poisoning occurred in 1996 and has led to the deaths of more than 60 people in a hemodialysis center in Brazil Caruaru. Involved poisoning was due to massive development of toxic cyanobacteria in a dam supplying water to the hemodialysis center (Jochimsen et al., 1998; Pouria et al., 1998; Azevedo et al., 2002). The incident in Brazil, and similar cases in Portugal, shows this issue as a priority theme in the strategies to protect water reservoirs. Thus, since the beginning of 1998, the WHO has established regulations for this type of hepatotoxins, including the maximum tolerated oral consumption of 1 µg microcystin-LR per liter of drinking water (WHO, 1998). Nevertheless, the damage caused by the repeated consumption of small amounts of toxins is probably more common than acute poisoning. In fact, MCs and NODs are considered as powerful tumor promoters (NishiwakiMatsushima et al., 1992; Ohta et al., 1994). However, the evidence for carcinogenecity of these hepatotoxins is today considered inadequate in humans and limited in animals (Bouaïcha et al., 2005). Nevertheless, it is necessary to indicate that an epidemiological study realized in China in the region of Jiangsu showed a strong correlation between the high proportion of primary cancers of the liver and the presence of high levels of MCs in the drinking water (Carmicheal et al., 1988; Yu, 1989; Harada et al., 1996). However, the presence of aflatoxin B1 and virus of hepatitis B in the drinking water could also explain the strong rate of hepatic cancers observed in this region (Ueno et al., 1996). Recently, MC-LR has been classified as “possibly carcinogenic to humans” (group 2B), and NOD-R as “not classifiable as to their carcinogenicity” (group 3) (Grosse et al., 2006).

5. BIOACCUMULATION OF CYANOTOXINS ON AQUATIC ORGANISMS AND ITS CONSEQUENCES ON PUBLIC HEALTH Cyanotoxins can be transferred through the food chain (Ibelings et al., 2005, Chen et al., 2009b; Ettoumi et al., 2011), suggesting a potential risk to high trophic level species and human consumption of contaminated aquatic products. Several studies have been reported the bioaccumulation of cyanotoxins in common aquatic vertebrates and invertebrates, including zooplankton, mollusks and crustaceans, and fish, which pose a risk to both animal and human heath if such aquatic animals are consumed (see review by Ibelings and Chorus, 2007). Thostrup


and Christoffersen (1999) and Mohamed (2001) reported that Daphnis can accumulate MCs at very high concentrations, up to 24.5 mg MCs/g DW. As primary consumers, freshwater mollusks and crustaceans constitute an important link between primary producers (potentially toxic cyanobacteria) and higher consumers. They can represent an intoxication route to predators in food webs. The bioaccumulation of cyanotoxins in some species of freshwater shellfish and mollusks has been reported by several studies (Saker and Eaglesham, 1999; Vasconcelos et al., 2001; Kankaanpaa et al., 2005; Chen and Xie, 2005a; b). These accumulations are due either to the consumption of toxic cyanobacteria only (Zurawal et al., 1999), or both a consumption of toxic cyanobacteria and ingestion of contaminated water (Saker and Eaglesham, 1999). Accumulation of cyanotoxins in fish is a potentially important route of exposure for humans. The presence and accumulation of these toxins in different fish tissues have been reported by experimental and field studies (Bury et al., 1998; Mohamed et al., 2003; Freitas de Magalhaes et al., 2001; Cazenave et al., 2005). Fishes may accumulate cyanotoxins via different routes: direct feeding on phytoplankton (phytoplanktivorous species like silver carp: Hypophthalmichthys molitrix), uptake of dissolved toxins via epithelium (gills, skin) or exposure via the foodweb. Generally it is believed that the oral route is the most important (Ernst et al., 2001). Several studies have proved that high concentration of cyanotoxins could be accumulated in different organs of fish, including the edible part, such as muscle (Bury et al., 1998; Xie et al., 2004). The use of water from sources containing cyanobacterial blooms and toxins for spray irrigation of crops presents potential health hazards through several exposure routes, including uptake into the food chain. There are several indications that terrestrial plants, including food crop plants, can take up MCs (Chen et al., 2012).

6. PREVENTION AND REMEDIATION OF CYANOBACTERIAL BLOOMS IN DAMS The treatment of water contaminated with toxic cyanobacteria should be achieved through prior actions of control and management of blooms that can include direct and/or indirect intervention. The indirect intervention is usually to manipulate the principal parameter that controls the growth of cyanobacteria such as nutrient availability. The direct actions correspond to techniques that minimize the access of masses of cyanobacteria supply systems. For example, the chemical control through the use of algaecides to destroy algal blooms upstream of a drinking water intake or a recreational area.


Nutrients reduction

Cyanobacteria are stimulated by excessive anthropogenic nutrient loading (Fogg, 1969; Reynolds, 1987; Paerl, 1988). In freshwater ecosystems, P availability is often the key factor limiting phytoplankton growth, including cyanobacteria species (Schindler et al., 2008). Accordingly, controlling P inputs has been the primary goal for resource managers. Nutrient loading dynamics have changed substantially over the past several decades. Practices such as bans on phosphate containing detergents, improved waste water treatment, and no till agriculture have been effective at reducing freshwater P loads but less for N, which is more mobile throughout the environment (Galloway and Cowling, 2002; Howarth, 2008). Overall, direct inputs from sewage represent the majority of P loads. The share controllable at the source is represented by the polyphosphates household detergents for clothes and dishes washing. Single nutrient input reductions, including a P-detergent ban and improved wastewater treatment for P during the 1980s in North Carolina' s (USA) Neuse River System, helped solve one problem (arrest freshwater blooms), but exacerbated blooms in downstream N-sensitive estuarine waters (Paerl, 2009). Techniques for trapping P by pre-selected upstream water bodies help protect the ultimate water body (Lürling and Faassen, 2012). An example given by Chorus and Bartram (1999) is the Lake Kis Balaton in Hungary, a reservoir with a surface area of 60 km2, which retains the P and protects Lake Balaton 10 times larger. It is shown that the impact of load reduction of phosphate on the development of algae is not directly visible because of a chemical balance that tends to recharge the water column by the P content in sediments. Trapped therein in certain forms of P or related to biogenic minerals such as


Fe, Mn, Ca, it becomes available for biomass because of anoxic conditions in the community. This new balance depends on time and depth of the water. In all cases, after reduction and/or removal of pollutant inputs, the return to a steady state can be time consuming (Lafforgue, 1998). It was estimated that in surface waters, the period required to achieve 90% reduction of the theoretical concentration is equal to 3 times the average residence time of water. A significant reduction in intake of N and P in lakes of Norway induced a transition period of 10 years during which eutrophication and the development of green algae have fallen sharply, while in parallel another species of cyanobacteria had colonized the resources using the internal load of P (Solheim, 2002). The same phenomenon was observed in Lake Bourget where a toxic cyanobacterium, Planktothrix rubescens, proliferated in recent years while P concentrations decreased from 120 to 30 g/L in the lake since the 1980s (Jacquet et al., 2005). Another technique for reducing P in situ is by precipitation of this element. The products most often mentioned are the salts of iron or aluminium that can precipitate the phosphates as phosphate of iron or aluminium insoluble (Cook, 1993). However, the addition of aluminium in waters with low buffering capacity can cause a decrease in the pH. Lam et al. (1995) showed that treatment with aluminium in the water caused an increase of MCs. Prepas et al. (1997) used iron salts and showed that they could be used as nutrients. In some cases, iron can stimulate the growth of cyanobacteria and algae. In addition, unlike aluminium, iron does not block in sustainable manner P. Under anoxic conditions that occur frequently at the bottom of the dam, the ferric iron is reduced to ferrous ion which is then accompanied by another release of P in the water column. The use of lime and calcium carbonate has shown its effectiveness on the control of phytoplankton blooms (Babin et al., 1994; Prepas et al., 2001b; Zhang et al., 2001), including limiting the available P. However, quite opposite results, were found by other authors (Reedyk et al., 2001). According to Prepas et al. (2001b) lime causes a saturation of the water column in Ca2+ which then produces a precipitation of P in the form of hydroxyapatite, which, once in the sediment continues to adsorb P. Prepas et al. (2001a) showed efficiency of P reduction but with a limited response over time: (i) the lime treatment must be repeated regularly, and (ii) the duration of the inactivation of P in the sediment decreases with pH and with the concentration of P in the water column. The experiences made by Reedyk et al. (2001) on various Canadian lakes have shown that lime was more effective than calcium carbonate. A lime treatment of 107 mg/L led to a P reduction of 85 to 92%. Despite, concerted efforts to control P cyanobacteria re-emerged. Large shallow lakes like Okeechobee, USA; Taihu, China, Kasumigaura, Japan; and the western basin of Lake Erie, USA tend to be co-limited by N and P (Havens et al., 2001; Guilford et al., 2005; North et al., 2007; Paerl and Scott, 2010; Xu et al., 2010); largely because previously-loaded P and N are effectively retained and recycled. While some N can be “lost” via denitrification, this process does not appear to keep up with “new” N inputs (Seitzinger, 1988). Overall, N and P co-limitation appears to be quite common in eutrophic systems (Elser et al., 2007) that are also most susceptible to cyanobacteria harmful algal bloom (CyanoHAB) outbreaks (Huisman et al., 2005).


Use of algicides

Copper is probably the most known algicide and it is usually applied as a copper sulfate (CuSO4 5H2O) (Murray-Gulde et al., 2002). Copper sulfate is used for a long time on the water bodies. Its use is reported for the first time in 1880 (Sawyer, 1962) in Europe, the United-States in 1904 (Moore and Kellerman, 1905) and in 1940 in Australia (Bruchet et al., 1998). Mainly due to its toxicity against algae (copper enters the cells causing the substitution of magnesium in the chlorophyll heme and also affects other molecules of biochemical processes in cells) and extremely low cost, it has been used since 1904 (Moore and Kellerman, 1905) for the control of noxious phytoplankton in surface waters. Toxicity and bioavailability of copper are influenced by a lot of factors, including parameters of the aquatic system (pH, DOC, alkalinity, ionic strength, conductivity), the form of copper used (hydrated Cu2+ or organo-copper complexes), dose and exposure time (Murray-Gulde et al., 2002; De Schamphelaere and Janssen, 2006). Copper toxicity against the phytoplankton is determined by the activity of the Cu2+ ion. It varies between 0.06 and 6.10-7 mg/L (McKnight et al., 1983). Its effect is reduced if the pH is high and in hard water. To facilitate dispersion, copper chelates have been used and in particular the copper citrate in hard water. Copper is commonly applied in low concentrations (from tens to hundreds of micrograms


per liter), which usually has to be repeated due to its rapid dynamics in natural waters. As early as within 2 h following the addition, more than 80% of the Cu can be found in a particulate form, while the rest can be associated with a dissolved organic or inorganic form (Haughey et al., 2002). It can be also exported from the reservoir or accumulated in sediments. Moreover, repeated dosing can lead to resistance of phytoplankton communities (Silver and Phung, 1996; Garcia-Villada et al., 2004). In the U.S., Raman (1995) recommends using a mixture of copper sulfate and acid. Brient et al. (2001) studied the application of copper sulfate and citric acid at 20% on three dams in Brittany (France) on long-term monitoring with up to 3 applications per year. Cyanobacteria are eliminated within 48 hours and environments are again colonized by diatoms. After 3-4 weeks, depending on the nutritional resources and weather, cyanobacteria re-appeared requiring further treatment with copper sulfate. As a short-term measure, copper-containing algicides such as copper sulfate have been applied successfully (McGuire et al., 1984; Murray-Gulde et al., 2002). However, while it is feasible to control algae in smaller reservoirs, it is difficult to control growth of microorganisms in larger reservoirs, where tons of algicides are needed to affect algal blooms. Furthermore, the application of copper sulfate often involves negative side-effects, including the potential release of intracellular taste and odor compounds (Fawell and James, 1994) and cyanotoxins (Jones and Orr, 1994) as well as negative effects on the aquatic life (for review see Jancula and Marsalek, 2011). The release of cyanotoxins occurs very quickly, within 3 days after treatment (Gupta et al., 2001). Serious accidents have shown the potential effect of such salting, such as Palm Island in Australia where many people were victims of a hepatoenteritis following treatment of a reservoir with copper sulfate (Bourke et al. 1983). In addition, the accumulation of copper in the sediments might become a long-term problem, with a possible release of copper due to changes in the water chemistry (Haughey et al., 2000) and can, therefore, induce the selection of resistant strains of algae and cyanobacteria. In fact, several studies have been reported that repeated treatment with copper sulfate led to a proliferation of more resistant species which needed, further, increasing doses of copper to remove them (Prepas and Murphy, 1998, Soldo and Behra, 2000). Recently, the use of a wide range of natural chemicals isolated (or synthesized) from both aquatic and terrestrial organisms seem to be an interesting and environment friendly substitute to traditional algicides such as copper sulfate (Schrader et al., 2003; Becher et al., 2005; Blom et al., 2006; Mizuno et al., 2008; Jancula et al., 2010). Unfortunately, the main limitation of such natural chemicals is price availability since the active compound must be intricately synthesized or extracted. This is perhaps why the studies are mainly focused in a laboratory, not in the field.


Biological remediation

The process of natural bioremediation for removal of cyanotoxins from water dams has been the subject of many studies. In order to control the algal concentration in some dams, some countries had introduced the algivorous Chinese silver carp fish, H. molitrix, in these reservoirs. For evaluating the control of toxic Microcystis blooms in Lake using phytoplanktivorous fish, three large fish pens (0.36 km2 of each) stocked with silver and bighead carp were set up in Meiliang Bay (China) (Ke et al., 2007). The results of this in situ study demonstrated that total phytoplankton biomass, Microcystis biomass and MCs concentration were lower in the fish pens than in the surrounding lake water, but the difference was not statistically significant. In addition, this study also reported that silver carp had a stronger ability of eliminating phytoplankton than bighead carp. Although the phytoplanktivorous silver carp directly fed on toxic cyanobacteria, and therefore ingested much MCs; they did not accumulate considerably more cyanotoxins in their organs or tissues, such as liver and muscle, than other fishes (Chen et al., 2007). Xie et al. (2005) reported that MC-LR was relatively low in tissues and organs of the phytoplanktivorous silver carp, despite the direct feeding on toxic cyanobacteria, and higher in predatory and omnivorous fishes. These results indicate that phytoplanktivorous fishes, such as silver carp, are probably more resistant to MCs exposure than other fishes, and therefore it is quite possible to use such fishes to counteract toxic cyanobacterial blooms. However, although the silver carpe bioaccumulate low levels of cyanotoxins, it remains a sanitary hazard to both animals and humans heath if consumed. Recently, Chen et al. (2009a) reported that MCs were identified for the first time in the serum of


chronically exposed fishermen at Lake Chaohu (China) together with indication of hepatocellular damage.


Conventional water treatment

The occurrence of toxic cyanobacteria and their toxins in dams used as source for drinking water production is recognized as being an increasing problem worldwide, presenting risks for humans and animals health and economic vitality. As mentioned in the section 2.1 many toxic species of cyaboacteria, especially those forming surface scums in dams, possess vesicles gas allowing them to float into the water column. Thus, the horizontal and vertical distribution of toxic cyanobacterial populations can vary throughout a water reservoir. Therefore, the choice of intake depth of raw water must take into account cell buoyancy of some cyanobactrial species in water body and to the possibility of some non-buoyant species belonging to the genus Oscillatoria which can accumulate in depth. Nevertheless, selection of the level draw water in depth in a dam can be important in reducing contamination by avoiding surface scums of buoyant cyanobacterial species which are more occurred worlwide. Since cyanotoxins are cell bound toxins, avoiding offtakes of intact cyanobacterial cells on water surface will decrease the amount of cyanotoxins to be dealt with further down the treatment train. Cyanobacterial toxins, particularly MCs, are chemically stable in natural water (Jones and Orr, 1994; Harada et al., 1996). Therefore, the management of toxic cyanobacteria in raw water supplies, and consequently their toxins in drinking water system is within the water treatment system from source to tap. Several studies have shown that conventional water treatment techniques such as combinations of coagulation, flocculation, clarification, and sand filtration are not effective for removing MCs (Himberg et al., 1989; Rositano and Nicholson, 1994; Lawton and Robertson, 1999). Coagulation and flocculation, using chemicals such as various aluminium and ferric iron salts as well as some synthetic polymers, involves the aggregation of smaller particles into larger flocs for further removal by sedimentation and filtration. Several studies have reported that coagulation can be an efficient method for removal of cyanobacterial cells from water, whereas soluble cyanotoxins are not very efficiently removed by this method (James and Fawell, 1991; Rositano and Nicholson, 1994). Mouchet and Bonnélye (1998) found that the efficiency of cyanobacterial removal is dependent on the optimization of chemical doses and coagulation pH, as inadequate coagulant dosage will remove other phytoplankton cells before cyanobacteria. However, James and Fawell (1991) reported that coagulation may cause additional problems such as lysis of cyanobacterial cells resulting to release of intracellular toxins. Others have found that aluminium or ferric chloride coagulation and filtration were effective to remove cyanobacterial cells without significant damage to the cell membrane integrity and release of intracellular toxins (Chow et al., 1999; Drikas et al., 2001). Once dissolved cyanotoxins are in the treatment plant, more advanced methods for their removal or transformation will be necessary. Lambert et al. (1996) have found low levels of microcystin removal (0 to 39%) when 60 mg/L aluminium was dosed in a small water treatment plant. Rapala et al. (2002) reported that the highest reductions in cyanobacterial endotoxin concentrations was occurred in the early stages of treatment, up to 83% through coagulation and sand filtration, suggesting that processes that reduce intact cyanobacterial cells remove endotoxins to a great extent. Sometimes flotation by saturating water under pressure and then releasing the pressure, called dissolved air floatation (DAF), is employed as a clarification step after coagulationfloculation that is generally more effective than sedimentation in clarifying algal-rich waters (Letterman, 1999). In several studies, Hrudey et al. (1999) reported that the DAF process achieved various levels of toxic cyanobacteria removal, up to 80% of Microcystis and 100% of Anabeana spp., but only 30% of Oscillatoria. The performance of rapid filtration, a method usually employed after coagulation to remove the flocs, does not effectively remove cyanobacterial cells (Lepisto et al., 1994; Steffensen and Nicholson, 1994). Mouchet and Bonnélye (1998) have indicated that direct filtration is generally not efficient for removing cyanobacterial cells. In fact,


Hoeger et al. (2001) found intact cyanobacterial cells after the filtration stage. Furthermore, sand filtration alone is unable to remove dissolved cyanotoxins (Falconer et al., 1983; Lambert et al., 1996). However, slow sand filters can be expected to remove 99% of algal cells (Mouchet and Bonnélye, 1998) and dissolved cyanotoxins (Keijola et al., 1988; Grützmacher et al., 2002). This can be explained by the formation of a biofilm on top of the filter that it allows for some biodegradation of cyanotoxins in slow sand filtration. Indeed, the biodegradation of hepatotoxic cyanotoxins (MCs, NOD and CYN) is well documented (Jones and Orr, 1994; Cousins et al., 1996; Christoffersen et al., 2002; Lemes et al., 2008; Mazur-Marzec et al., 2009b; Chen et al., 2010), with specific enzymatic pathways well characterized (Bourne et al., 1996; Okano et al., 2009; Zhang et al., 2010). It was recently reported that the data generated in laboratory and field studies strongly indicate that, in a large, shallow lake, low persistence and the natural elimination of MCs are due to biodegradation; suggesting that sediments play a crucial role in biodegradation by continuously supplying toxin-degrading bacteria to the water column (Chen et al., 2008; Mazur-Marzec et al., 2009b; Chen et al., 2010). Several previous studies have been indicated that MCs can be degraded by aquatic bacteria identified as pertaining especially to the genus Sphingomonas (Bourne et al., 1996; Harada et al., 2004; Ishii et al., 2004; Maruyama et al., 2006; Manage et al., 2009). Therefore, a microcystin-degrading gene cluster, mlrA, B, C and D was identified in these microorganisms, sequenced and the degradation process was proposed (Bourne et al., 2001; Saito et al., 2003; Imanishi et al., 2005). Subsequently the presence of these genes in some isolated bacteria was used in screening studies as indicator of microcystin-degrading species. Indeed, Ho et al. (2007b) have screened the presence of the 4 mlr genes in 32 taxa of bacteria isolated from a sand filter. The results of this study showed that the presence of these genes was occurred only in one species identified as a novel species Sphingopyxis witflariensis LH21which is able to efficiently degrade two congeners of microcystins, MC-LR and MC-LA at high concentrations 2 and 3 mg/L, respectively. In the last few years, several other species of bacteria capable of degrading peptidic cyanotoxins were identified, Sphingomonas sp. strain ACM-3962 (Jones et al., 1994), Paucibacter toxinivorans (Rapala et al., 2005), Sphingosinicella microcystinivorans (Maruyama et al., 2006), Burkholderia sp. (Lemes et al., 2008). The ability of these species to degrade other congeners of MCs and NODs was investigated and revealed that peptides with the Adda–Arginine bond were successfully degraded while MC-LF, with Adda– Phenyalanine bond and 6(z)-Adda-MC-LR and 6(z)-Adda-MC-RR were not significantly degraded (Imanishi et al., 2005). Another Japanese Sphingomonas isolate, 7CY, was shown to degrade a wider range of MCs, including MC-LR, -RR, -LY, -LW, and -LF but it was unable to degrade NOD-Har a NOD analogue where arginine is replaced by homoarginine (Ishii et al., 2004). The fact that the three products (linear MC-LR, a tetrapeptide, and the Adda group) obtained after the sequential enzymatic hydrolyses of the peptide bonds of the MC-LR are nontoxic compared with the parent toxin suggests that bacterial degradation is a safe and practical process for removing MCs from water (Bourne et al., 1996; Harada et al., 2004; Ho et al., 2007a). While biodegradation of MCs has been clearly demonstrated in some laboratory studies, only few studies have been extended to a pilot or a full scale treatment studies (Bourne et al., 2006). The alkaloid hepatotoxin CYN has also been shown to biodegrade (Senogles et al., 2002; Chiswell et al., 2004). In contrast, a laboratory study investigating biodegradation of CYN with bacterial communities from two water bodies in Spain, one having frequent exposure to CYN, the other rarely, has been shown that biodegradation of CYN by an active microbial community does not take place during a 40-day (Wormer et al., 2008). A recent study demonstrated that CYN was degraded by indigenous microbial flora in waters with a history of Cylindrospermopsis blooms (Smith et al., 2008). Despite isolation of many bacteria from CYN enriched cultures, only a single isolate Delftia sp. capable of degrading CYN has been obtained (Smith, 2005). Biologically active filters have been shown to remove both cells of Cylindrospermopsis and dissolved CYN once they have been ripened or conditioned (Garnett et al., 2003). However, for cyanobacterial neurotoxins there are few reports on their persistence and biodegradation compared to cyanobacterial heptotoxins. Kiviranta et al. (1991) reported the isolation of a Pseudomonas sp. capable of rapid degradation of anatoxin-a, with a rate of 6–30 mg/mL per 3 days. A second study reported by Rapala et al. (1994) has been shown the removal of anatoxin-a by microbial populations isolated from water and sediments of a eutrophic, oligotrophic, and humic lake.


A number of studies have investigated the suitability of two main types of activated carbon, powdered activated carbon (PAC) and granular activated carbon (GAC), to remove cyanobacterial toxins and taste and odor causing compounds, in concert with conventional treatment processes. In lab-scale experiments, Keijola et al. (1988) and Himberg et al. (1989) reported that simultaneous addition of PAC to coagulation processes at 5 mg/L removed up to 34% of MCs and more than 50% of the neurotoxin-a. Several other studies have shown that doses of PAC in excess of 20 mg/L are often necessary to achieve more than 90% removal of MCs (Bruchet et al., 1998; Hart et al., 1998; Mouchet and Bonnélye, 1998), CYN (Newcombe et al., 2002) and neurotoxins (Newcombe et al., 2002; Newcombe and Nicholson, 2002). Granular activated carbon (GAC) has also shown successful performance in removing cyanotoxins (Bruchet et al., 1998; Hart et al., 1998; Newcombe and Nicholson, 2002). Activated carbons, mostly in the powdered but also in the granular filter form, however, are not systematically added in a water treatment plant and may be cost-prohibitive since contamination with cyanobacterial toxins is seasonal and unpredictable. Therefore, to improve the elimination of dissolved cyanotoxins that are extremely stable across a wide range of pH and temperature and are generally not effectively removed by conventional treatment through coagulation/flocculation, sedimentation and filtration advanced oxidation processes will be necessary. The most efficient chemicals used for the oxidation of dissolved toxins are chlorine (for review see Merel et al., 2010), potassium permanganate and ozone. In general, chlorination is not an effective process in destroying cyanotoxins (Himberg et al., 1989; Rositano and Nicholson, 1994). The efficiency of chlorination seems to depend largely on the chloride compounds, the concentration used, and the pH (Nicholson et al., 1994; Rositano et al., 1998). Aqueous chlorine and calcium hypochlorite at 61 mg/L remove more than 95% of MCs or NOD, while sodium hypochlorite at the same dose or chloramine achieve 40–80% removal at the most (Nicholson et al., 1994; Rositano and Nicholson, 1994). Tsuji et al. (1997) have been shown that microcystin-LR could be easily decomposed by chlorination with sodium hypochlorite, and the decomposition depended on the free chlorine dose. It was reported that chlorination is not efficient to destroy the anatoxin-a (Nicholson et al., 1994; Rositano and Nicholson, 1994). However, a high removal of the saxitoxin group and the CYN was established with a free chlorine residual of 0.5 mg/L (Senogles et al., 2000; Newcombe and Nicholson, 2002). Several studies have shown that potassium permanganate was also an effective oxidant in reducing cyanobacterial MCs when applied to both raw and clarified water (Fawell et al., 1993; Rositano et al., 1998; Karner et al., 2001, Rodriguez et al., 2007a; b). However, two other studies have shown that permanganate induces some damage to cyanobacterial cell and increases levels of dissolved cyanotoxins but without further oxidation of those toxins (Lam et al., 1995; Schmidt et al., 2002). Ozone appears to be the most powerful oxidizing agents to achieve complete destruction of the intracellular and extracellular fractions of MCs, NOD, CYN, and anatoxin-a (Himberg et al., 1989; Rositano and Nicholson, 1994; Bruchet et al., 1998; Hart et al., 1998; Rositano et al., 1998; Brooke et al., 2006; Rodriguez et al., 2007b; Miao et al., 2010) but not the saxitoxin family (Rositano et al., 2001; Newcombe and Nicholson, 2002). However, several other studies showed the efficiency of the ozone in the elimination of the PSP toxins in a concentration up to 100µg/L (Chorus and Bartram, 1999). On the other hand, the action of the ozone on cyanobacterial toxins destruction is very fast. Bernazeau et al. (1995) and Rositano et al. (1998) found that the total elimination of MC and NOD in natural waters was removed in less than 1 min when treated with up to 0.2 mg/L ozone. For MCs, destruction was mainly involved in the attack of ozone on Adda side chain and the toxicity evaluation of these toxins ozonation products revealed that those endproducts had no adverse effects in vivo and in vitro, showing that ozonation could completely remove the MCs’ toxicity (Mia et al., 2010). The preoxidation by the ozone or the chlorine is widely used to improve the coagulation and especially for the elimination of algae and cyanobacteria. It allows the decomposition of cyanotoxins, cellular leases and liberation of toxin (Bonnélye et al., 1995; Lam et al., 1995; Mouchet and Bonnélye, 1998). On the other hand, Mouchet and Bonnélye (1998) have shown following a comparative study that the prechloration is slightly more effective than the preozonation in the improvement of the coagulation with respectively 97% and 94% of elimination of a bloom of algae in the same condition. Maatouk et al. (2002) found that total elimination of cyanobacterial cells and a low concentration of MCs (74 ng/L) was achieved through the combined action of preozonation at 0.07 mg/L and adsorption on powdered activated carbon at 20


mg/L. However, pre-chlorination at 0.42 mg/L followed by 20 mg/L of powdered activated carbon removed only 45% of toxins. Bernazeau et al. (1995) found that the oxidizing potential of the ozone is more effective if it is used before coagulation/floculation and after filtration.


Advanced water treatment

The use of alternative, more advanced methods for the removal or destruction of cyanobacteria and their toxins has been investigated in the last 10 years with varying degrees of success. These advanced treatment techniques include membrane filtration (Neumann and Weckesser, 1998; Mouchet and Bonnélye, 1998; Lee and Walker, 2008; Campinas and Rosa, 2010), ultraviolet photolysis (Welker and Steinberg, 2000; Hrudey et al., 1999; Afzal et al., 2010; Graham et al., 2010), nanoparticles adsorption (Liu et al., 2010), ultrasonic irradiation (Zhang et al., 2009), and bacterial biodegradation (Ho et al., 2012). All these advanced treatment options are a promising technology but are currently impractical for full-scale water treatment in different countries.

8. CONCLUSION The frequency of the appearance of cyanobacterial blooms in surface waters, especially in dams used for the purposes of producing drinking water, was reported world-wide. Although eutrophication in these dams has been recognized as a growing basis for cyanobacterial cells, many other information on the conditions leading to cyanobacterial bloom development require further research. The occurrence of these harmful cyanobacterial blooms in dams is often accompanied by a production of a variety of cyanotoxins and taste and odor compounds that can result in human and animal poisonings and in malodorous or unpalatable drinking water, respectively. Therefore, the management of toxic cyanobacteria in raw water supplies, and consequently their toxins in drinking water system is within the water treatment system from source to tap. Many factors can affect water treatment efficacy in reduction of cyanotoxins, the most important being the level of extracellular toxins. The mechanisms for effective removal of this part of toxins from influent water differ from those required for intracellular toxins. Therefore, operational strategies need to be developed to minimize cell lysis during the treatment processes and simple low-cost techniques for cyanobacterial cell removal should be, therefore, investigated and developed further. Although, the biological remediation by introducing in some countries algivorous fishes in order to control the algal concentration in dams gave some satisfying results; however, it is necessary to remain watchful because several studies have been shown that aquatic organisms could in a direct or indirect manner contribute to food chain cyanotoxins transfer, and by the way constitute a potential health risk source for animals and humans. Acknowledgements: We gratefully acknowledge the financial support of Agence Universitaire de la Francophonie (AUF) (Grants 6313PS002).

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