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wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 8

Available online at w

journal homepage: www.elsevier .com/locate /watres

Biological treatment options for cyanobacteria metaboliteremoval e A review

Lionel Ho a,b,*, Emma Sawade a, Gayle Newcombe a

aAustralian Water Quality Centre, South Australian Water Corporation, 250 Victoria Square, Adelaide, SA 5000, Australiab School of Earth and Environmental Sciences, University of Adelaide, Adelaide, SA 5005, Australia

a r t i c l e i n f o

Article history:

Received 16 August 2011

Received in revised form

25 October 2011

Accepted 4 November 2011

Available online 15 November 2011

Keywords:

Biodegradation

Biological filtration

Cyanotoxins

Tastes and odours

* Corresponding author. Australian Water QuAustralia. Tel.: þ61 8 7424 2119; fax: þ61 8 7

E-mail address: lionel.ho@sawater.com.a0043-1354/$ e see front matter ª 2011 Elsevdoi:10.1016/j.watres.2011.11.018

a b s t r a c t

The treatment of cyanobacterial metabolites can consume many resources for water

authorities which can be problematic especially with the recent shift away from chemical-

and energy-intensive processes towards carbon and climate neutrality. In recent times,

there has been a renaissance in biological treatment, in particular, biological filtration

processes, for cyanobacteria metabolite removal. This in part, is due to the advances in

molecular microbiology which has assisted in further understanding the biodegradation

processes of specific cyanobacteria metabolites. However, there is currently no concise

portfolio which captures all the pertinent information for the biological treatment of

a range of cyanobacterial metabolites. This review encapsulates all the relevant informa-

tion to date in one document and provides insights into how biological treatment options

can be implemented in treatment plants for optimum cyanobacterial metabolite removal.

ª 2011 Elsevier Ltd. All rights reserved.

1. Background production of secondary metabolites. In particular, metabo-

Cyanobacteria have evolved to adapt to almost every envi-

ronment on the planet, in particular marine and freshwater

sources. They continue to adapt to our changing environ-

mental conditions, and thrive in water sources impacted by

development and climate change (Paerl et al., 2011). A major

trigger for the proliferation of cyanobacteria in water is the

input of nutrients resulting from modern agricultural prac-

tices and treated sewage discharge. In addition, warmer global

temperatures, low rivers flows and reduced water quality

associated with drought conditions favour the growth of

cyanobacteria (Paerl and Huisman, 2009). As a result, issues

associated with cyanobacteria in drinking water are

increasing worldwide.

The major water quality implication from the proliferation

of cyanobacterial blooms for the drinkingwater industry is the

ality Centre, South Aust003 2119.u (L. Ho).ier Ltd. All rights reserve

lites which can cause aesthetic issues (eg. compounds which

impart tastes and odours, such as 2-methylisoborneol and

geosmin) and those which can severely impact human health

(eg. cyanobacterial toxins or cyanotoxins, such as micro-

cystins, cylindrospermopsin, saxitoxins and anatoxin-a).

Many of these metabolites are not well removed by conven-

tional water treatment practices and require more costly

treatments such as activated carbon and/or advanced oxida-

tion processes, which themselves may have limitations.

With the increasing frequency of cyanobacterial detection

in global water supplies, coupled with the changing climate

the world is facing, it is paramount that water authorities

implement and optimise successful treatment strategies for

the mitigation of cyanobacteria and their metabolites. Many

water safety plans such as those developed by the World

Health Organization (WHO) and the Australian Drinking

ralian Water Corporation, 250 Victoria Square, Adelaide, SA 5000,

d.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 8 1537

Water Guidelines (ADWG), stipulate that it is important to

utilise multi-barrier treatment options to ensure these

contaminants do not reach the customer tap. One such barrier

which has received widespread attention is the application of

biological treatment for the removal of these metabolites.

Biological processes have been used for centuries for the

treatment of drinking water and are considered integral

processes for the treatment of wastewater, including denitri-

fication processes. In general, they are low technology

processes that generally require little or no maintenance or

running costs, and have the advantage of being a “natural”

treatment, without the addition of chemicals that can produce

potentially harmful by-products.

Little information has been gathered with respect to the

effectiveness of biological treatment for a range of cyano-

bacterial metabolites under conditions that would be experi-

enced in water treatment plants (WTPs). This is particularly

relevant since multiple classes of these metabolites are now

being simultaneously detected in water bodies (Graham et al.,

2010; Paerl et al., 2011). This review encapsulates the relevant

information published to date on the biodegradation of cya-

nobacterial metabolites in water and lists some of the factors

which can affect biodegradation of such metabolites in water.

Furthermore, this review also provides insights into how

biological treatment options can be implemented in WTPs for

optimum cyanobacterial metabolite removal, in particular

through biological filtration processes.

Table 1 e Organisms implicated in the biodegradation ofMIB.

Organisms Reference(s)

2. Biodegradation of cyanobacterialmetabolites

2.1. MIB and geosmin

The metabolites 2-methylisoborneol (MIB) and geosmin are

cyclic aliphatic tertiary alcohols (see Fig. 1) that impart an

earthy/musty/camphorous odour. They are themost common

naturally occurring taste and odour (T&O) compounds

worldwide, and are particularly problematic as they can be

perceived by the consumer in drinking water at levels as low

as 5e10 ng L�1. MIB and geosmin can be produced by a range of

cyanobacteria including Anabaena, Aphanizomenon, Geitler-

inema, Symploca, Planktothrix (Oscillatoria), Phormidium, Nostoc,

Pseudanabaena and Lyngbya (Juttner and Watson, 2007). Even

OH

CH 3

CH 3

CH 3

CH 3

CH 3

CH 3OH

nimsoeGBIM

Fig. 1 e Molecular structures of 2-methylisoborneol (MIB)

and geosmin.

though their presence does not necessarily imply that the

quality of water is unsafe for drinking, it is an indicator that

the treatment processes are inadequate for the removal of

cyanobacterial metabolites.

Although MIB and geosmin pose no known human health

hazards, they have been shown to be toxic (at mg L�1

concentrations) to Salmonella typhimurium tester strains in the

Ames test (Dionigi et al., 1993). In addition, Nakajima et al.

(1996) documented that MIB and geosmin inhibit the early

development of sea urchin embryos with toxicity comparable

to that shown towards the Salmonella tester strains in the

Ames test (IC50 values of 69 and 17 mg L�1 for MIB and geo-

smin, respectively). Both T&O compounds have also been

found to affect fisheries, as they are easily absorbed into the

tissues of fish and other aquatic organisms, rendering them

unfit for retail sale due to the presence of off-flavours

(Persson, 1980; Hofer, 1998).

MIB and geosmin have been shown to be biodegradable

with studies implicating a variety of organisms responsible for

their removal from water (see Tables 1 and 2). In addition,

Hoefel et al. (2006) reported a novel finding where geosmin

was co-operatively biodegraded via a consortium of three

Gram-negative bacteria, namely Sphingopyxis sp. Geo24,

Novosphingobium sp. Geo25 and Pseudomonas sp. Geo33. All

three bacteria were isolated by the authors from the biofilm of

a mature sand filter.

The susceptibility of both compounds to biodegradation is

thought to be attributed to their structures which are similar

to alicyclic alcohols and ketones (Trudgill, 1984; Rittmann

et al., 1995). To date, no definitive pathways for the biodeg-

radation of MIB have been proposed, although Tanaka et al.

(1996) identified two possible dehydration products, 2-

methylcamphene and 2-methylenebornane. Trudgill (1984)

suggested that the biodegradation pathway of MIB may be

similar to camphor, a bicyclic ketone documented to be bio-

degraded through the biological BaeyereVilliger reaction. In

this process, the ring structures of camphor are cleaved

through a sequence of intermediate reactions which are

Bacillus spp. Ishida and Miyaji (1992);

Lauderdale et al. (2004)

Bacillus subtilis Yagi et al. (1988)

Candida spp. Sumitomo (1988)

Enterobacter spp. Tanaka et al. (1996)

Flavobacterium spp. Egashira et al. (1992)

Flavobacterium multivorum Egashira et al. (1992)

Pseudomonas spp. Izaguirre et al. (1988);

Egashira et al. (1992);

Tanaka et al. (1996)

Pseudomonas aeruginosa Egashira et al. (1992)

Pseudomonas putida G1 Oikawa et al. (1995);

Eaton and Sandusky (2009)

Rhodococcus ruber T1 Eaton and Sandusky (2009)

Rhodococcus wratislaviensis

DLC-cam

Eaton and Sandusky (2010)

Table 2 e Organisms implicated in the biodegradation ofgeosmin.

Organisms Reference(s)

Arthrobacter atrocyaneus Saadoun and El-Migdadi (1998)

Arthrobacter globiformis Saadoun and El-Migdadi (1998)

Bacillus cereus Silvey et al. (1970); Narayan and

Nunez (1974)

Bacillus subtilis Narayan and Nunez (1974);

Yagi et al. (1988)

Chlorophenolicus strain

N-1053

Saadoun and El-Migdadi (1998)

Chryseobacterium sp. Zhou et al. (2011)

Pseudomonas sp. SBR3-tpnb Eaton and Sandusky (2010)

Rhodococcus moris Saadoun and El-Migdadi (1998)

Rhodococcus wratislaviensis

DLC-cam

Eaton and Sandusky (2010)

Sinorhizobium sp. Zhou et al. (2011)

Sphingopyxis sp. Geo48 Hoefel et al. (2009b)

Stenotrophomonas sp. Zhou et al. (2011)

Fig. 2 e Molecular structure of microcystin-LR, highlighting

Adda and the variable amino acids leucine and arginine.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 81538

catalysed by mono-oxygenase enzymes. Oikawa et al. (1995)

confirmed this by excising the entire cam operon from

a camphor-degrading Pseudomonas putida G1, where its

subsequent transformation into Escherichia coli demonstrated

the acquired ability of that E. coli to degrade MIB. Similarly,

Eaton and Sandusky (2009) demonstrated that after enrich-

ment with camphor, three known camphor-degrading

isolates, including P. putida G1, produced metabolites result-

ing from hydroxylation at all of the three available secondary

carbons on the six-member ring of MIB.

No definitive pathways have been elucidated for the

biodegradation of geosmin. Saito et al. (1999) identified four

possible biodegradation products of geosmin, two of which

were 1,4a-dimethyl-2,3,4,4a,5,6,7,8-octahydronaphthalene

and enone. Trudgill (1984) suggested that geosmin may be

biodegraded by a pathway similar to that of cyclohexanol. The

author documented that strains of Acinetobacter and Nocardia

were capable of degrading cyclohexanol via mono-oxygenase

enzymes, similar to the biological BaeyereVilliger reaction.

Recently, Eaton and Sandusky, (2010) demonstrated geosmin

to be biodegraded by two terpene-degrading bacteria, Rhodo-

coccus wratislaviensisDLC-camand Pseudomonas sp. SBR3-tpnb.

However, both isolates were unable to degrade geosmin

without being induced by the addition of either camphor or

terpinene. Nonetheless, the authors showed that once

induced, both isolates readily degraded geosmin, affording

various ketogeosmins as by-products.

There are conflicting reports on the biodegradation rates of

MIB and geosmin. Westerhoff et al. (2005) conducted batch

studies in lake water and modelled MIB and geosmin biodeg-

radation as a pseudo-zero-order reaction. In contrast,

Rittmann et al. (1995) determined thatMIB and geosminwould

be utilised as secondary substrates in natural water, due to the

presence of natural organic material (NOM) which is present

at much higher concentrations than the T&O compounds.

Consequently, they determined the biodegradation ofMIB and

geosmin in natural water to be a second-order reaction. More

recently, a few studies have demonstrated that the rate of

biodegradation followed pseudo-first-order reaction (Ho et al.,

2007b; Zhou et al., 2011).

2.2. Microcystins

The microcystins are the most commonly reported of the

cyanotoxins worldwide and are predominantly produced by

Microcystis spp. although other cyanobacteria including Ana-

baena, Nostoc, Hapalosiphon, Anabaenopsis and Planktothrix

(Oscillatoria) have been known to produce these cyanotoxins

(Sivonen and Jones, 1999; Falconer, 2005). They are cyclic

heptapeptides consisting of seven amino acid groups, two of

which are variable. Minor variations to the other amino acids

are also seen in some variants. Over 70 variants of this cya-

notoxin have been identified to date. Themost common of the

variants, microcystin-LR (MCLR), incorporates leucine (L) and

arginine (R) in the variable positions (Fig. 2). Whilst this

variant is the most common, the majority of the cyanobacte-

rial blooms producing microcystins will produce a range of

variants. Some blooms have been found to contain no MCLR,

while others have some of the other variants as the major

components. Therefore, any investigation into the effect of

water treatment processes on microcystins should include

a range of the most commonly found variants.

The toxicity of the microcystins is generally associated

with the conjugated diene on the Adda (3-amino-9-methoxy-

2,6,8-trimethyl-10-phenyldeca-4,6-dienoic acid) amino acid

(see Fig. 2). Once absorbed by organisms, they are accumu-

lated in the liver which can lead to haemorrhage and even

death within a few hours. The potency of the microcystins in

humans was demonstrated in 1996 when 52 patients at two

dialysis centres in Caruaru, Brazil died as a result of acute

hepatic failure. It was discovered that the water used for

dialysis had been contaminated with microcystins with

concentrations of up to 600 ng mg�1 detected in the liver of

victims (Yuan et al., 2006). As a result of the concerns about

the effect of microcystins, a guideline value of 1 mg L�1 for

MCLR in drinking water has been issued by the WHO.

Of all the cyanotoxin biodegradation studies, most have

focused on the microcystins, a consequence of their biode-

gradability in drinking water sources (Jones and Orr, 1994;

Rapala et al., 1994; Cousins et al., 1996; Christoffersen et al.,

2002; Holst et al., 2003) and more recently wastewater

lagoons and stabilization ponds (Ho et al., 2010). Furthermore,

microcystins have also been documented to be biodegradable

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 8 1539

in soils and sediments (Miller and Fallowfield, 2001; Chen

et al., 2010b; Grutzmacher et al., 2010).

There is a growing number of isolated organisms reported

as having the ability to degrade microcystin in water and so

far the majority appears to belong to the family Sphingomo-

nadaceae (see Table 3). However, other studies have reported

microcystin degradation by bacteria other than the Sphingo-

monadaceae. Manage et al. (2009) identified three isolates,

Arthrobacter sp., Brevibacterium sp. and Rhodococcus sp., as

having the capability to degrade microcystin. Similarly, Hu

et al. (2009) isolated a Methylobacillus sp. from a cyanobacte-

rial sludge which could effectively degrade two microcystin

variants. Furthermore, studies have also demonstrated pro-

biotic bacteria as having the ability to biologically remove

microcystins (Meriluoto et al., 2005; Nybom et al., 2007, 2008;

Surono et al., 2008). In particular, Lactobacillus rhamnosus

strains GG and LC-705, Bifidobacterium longum 46, Bifidobacte-

rium lactis strains 420 and Bb12, were all able to remove

various microcystin variants (Meriluoto et al., 2005; Nybom

et al., 2007, 2008) as were Lactobacillus plantarum strains IS-

10506 and IS-20506 (Surono et al., 2008). However, the pro-

biotic bacteria were unable to completely remove the micro-

cystins in contrast to many of the Sphingomonadaceae which

were able to degrade microcystins to concentrations below

that required for analytical detection.

Almost all of the genotypic studies on microcystin degra-

dation have focused on the Sphingomonadaceae as many

within this group have been shown to contain specific genes

required for microcystin degradation (see Table 3). Within the

genome of the first isolatedmicrocystin-degrading bacterium,

Sphingomonas sp. ACM-3962, Bourne et al. (1996, 2001) identi-

fied a gene cluster, mlrA, mlrB, mlrC and mlrD, involved in

the degradation of MCLR. The authors determined that the

Table 3 e Organisms implicated in the degradation of microcy

Organisms

Arthrobacter sp. Manage et al. (

Brevibacterium sp. Manage et al. (

Burkholderia sp. Lemes et al. (2

Methylobacillus sp. Hu et al. (2009

Morganella morganii Eleuterio and B

Paucibacter toxinivorans Rapala et al. (2

Poterioochromonas sp. Ou et al. (2005

Pseudomonas aeruginosa Takenaka and

Ralstonia solanacearum Yan et al. (200

Rhodococcus sp. Manage et al. (

Sphingomonas sp. 7CY Ishii et al. (200

Sphingomonas sp. ACM-3962 Jones et al. (19

Sphingomonas sp. B9 (Sphingosinicella sp.) Harada et al. (2

Tsuji et al. (200

Sphingomonas sp. CBA4 Valeria et al. (2

Sphingomonas sp. MD-1 Saitou et al. (2

Sphingomonas sp. MDB2 (Sphingosinicella sp.) Maruyama et a

Sphingomonas sp. MDB3 (Sphingosinicella sp.) Maruyama et a

Sphingomonas sp. Y2 (Sphingosinicella

microcystinivorans)

Park et al. (200

Sphingopyxis sp. LH21 Ho et al. (2007

Sphingopyxis sp. USTB-05 Wang et al. (20

Stenotrophomonas sp. EMS Chen et al. (20

mlrA gene encoded an enzyme responsible for the hydrolytic

cleaving of the cyclic structure of MCLR. The resultant linear

MCLR molecule was then sequentially hydrolysed by pepti-

dases encoded by the mlrB and mlrC genes. The final gene,

mlrD, encoded for a putative transporter protein that may

have allowed for active transport of microcystin and/or its

degradation products into or out of the cell.

Since then, various studies have designed qualitative

polymerase chain reaction (PCR) assays for the detection of

these genes, in particular, mlrA, the first gene involved in

cleaving the cyclic structure of microcystin (Saito et al., 2003;

Ho et al., 2006, 2007d). More recently, Hoefel et al. (2009a)

designed and optimized a quantitative real-time PCR (qPCR)

assay for the detection of the mlrA gene within the biofilm of

sand filters. Ho et al. (2010) utilised this qPCR to investigate the

abundance of microcystin-degraders during the biodegrada-

tion of MCLR in a tertiary treated wastewater. Both studies

showed that the biodegradation of MCLR was directly related

to the abundance of microcystin-degrading bacteria, based

on mlrA gene copy numbers, suggesting that MCLR may be

a primary substrate for the proliferation of microcystin-

degrading organisms.

While it is clear that microcystin can be degraded by

organisms via the pathway originally proposed by Bourne

et al. (1996, 2001), Manage et al. (2009) have indicated that

there may be other pathways (and possibly genes) involved in

microcystin degradation as the authors were unable to detect

the mlr genes in their isolates.

Studies have demonstrated that the biodegradation of

microcystins does not yield toxic by-products. Bourne et al.

(1996) and Harada et al. (2004) identified two intermediate

products from the bacterial degradation of MCLR by Sphingo-

monas sp. ACM-3962 and Sphingomonas sp. B9, respectively.

stins.

Reference(s) mlrA gene detected

2009); Lawton et al. (2011) No

2009); Lawton et al. (2011) No

008) Unknown

) Unknown

atista (2010) Unknown

005) Unknown

); Zhang et al. (2008) Unknown

Watanabe (1997) Unknown

4) Unknown

2009); Lawton et al. (2011) No

4) Unknown

94); Bourne et al. (1996, 2001) Yes

004); Imanishi et al. (2005);

6); Kato et al. (2007)

Unknown

006) Unknown

003); Saito et al. (2003) Yes

l. (2006) Unknown

l. (2006) Unknown

1); Maruyama et al. (2003, 2006) Yes

a, 2007d); Hoefel et al. (2009a) Yes

10); Zhang et al. (2010) Unknown

10a) Yes

Table 4 e Organisms implicated in the degradation ofnodularin.

Organisms Reference(s)

Arthrobacter sp. Manage et al. (2009); Lawton et al. (2011)

Brevibacterium sp. Manage et al. (2009); Lawton et al. (2011)

Paucibacter toxinivorans Rapala et al. (2005)

Rhodococcus sp. Manage et al. (2009); Lawton et al. (2011)

Sphingomonas sp. B9

(Sphingosinicella sp.)

Harada et al. (2004); Imanishi et al. (2005);

Tsuji et al. (2006); Kato et al. (2007)

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 81540

Both studies identified linearisedMCLR (NH2-Adda-Glu-Mdha-

Ala-Leu-MeAsp-Arg-OH) and a tetrapeptide (NH2-Adda-Glu-

Mdha-Ala-OH) as the intermediate products,withHarada et al.

(2004) also isolating Adda as one of the final degradation

products. The authors determined that these intermediate

products were less active than the parent MCLR using protein

phosphatase inhibition and mouse bioassays. In addition, Ho

et al. (2007a) observed no hepatotoxic or cytotoxic by-

products from the biodegradation of MCLR and microcystin-

LA by Sphingopyxis sp. LH21 using a protein phosphatase 2A

inhibition assay and a cell-based cytotoxicity assay.

2.3. Nodularin

Nodularin (NOD) is a cyclic pentapeptide hepatotoxin that is

predominantly produced in brackish waters by Nodularia

spumigena. Its structure (see Fig. 3) and biological activity

closely resembles the microcystins. NOD has been docu-

mented to be biodegraded by bacteria with the Adda amino

acid shown to be a by-product (Imanishi et al., 2005; Edwards

et al., 2008; Toru�nska et al., 2008; Mazur-Marzec et al., 2009).

Many of the reported NOD-degrading organisms also have the

ability to degrade microcystins (see Table 4). It is believed that

this may be due to the enzymes (eg. mlrA) acting similarly for

both cyanotoxins by cleaving their cyclic structures at the

Adda-Arg peptide bond (Kato et al., 2007; Edwards et al., 2008).

Kato et al. (2007) and Edwards et al. (2008) provided evidence to

support this contention through detection of NOD biodegra-

dation by-products, including linear NOD (NH2-Adda-Glu-

Mdhb-MeAsp-Arg-OH).

2.4. Cylindrospermopsin

Cylindrospermopsin (CYN) is an alkaloid cytotoxin produced

mainly by the freshwater cyanobacteria Cylindrospermopsis

raciborskii, Umezakia natans, Anabaena bergii, Aphanizomenon

ovalisporum, Aphanizomenon flosaquae and Raphidiopsis curvata

(Falconer, 2005). The presence of high levels of CYN in

drinking water can cause liver, kidney and gastrointestinal

damage (Falconer, 2005). In addition, studies have also

shown that this cyanotoxin inhibits protein synthesis, is

genotoxic and potentially carcinogenic (Froscio et al., 2001,

2003; Falconer, 2005). While no official guideline value

exists for CYN, the WHO is considering a proposed 1 mg L�1

level due to concerns regarding the potential effect of CYN

Fig. 3 e Molecular structure of nodularin.

on human health (Rodriguez et al., 2007a). Fig. 4 shows the

molecular structure of CYN. Two other variants of this cya-

notoxin exists and are known as deoxycylindrospermopsin

and 7-epicylindrospermopsin.

Originally thought to be mainly an issue in tropical areas,

this cyanotoxin is now reported regularly in more temperate

climates (Padisak, 1997; Stirling and Quilliam, 2001; Falconer,

2005; Fastner et al., 2007; Rucker et al., 2007) suggesting that

CYN-producing cyanobacteria are highly adaptive. Further-

more, many studies have demonstrated that a large propor-

tion of CYN concentrations in the environment occur in

extracellular form (Griffiths and Saker, 2003; Chiswell et al.,

1999) which has important implications for water authorities

worldwide.

The most well documented incident where CYN had been

implicated in affecting human health occurred in 1979 in Palm

Island, Queensland, Australia (Hawkins et al., 1985). Over 120

people were reportedly poisoned by CYN after treatment of

C. raciborskii blooms in a drinking water source with copper

sulphate which caused the cyanobacterial cells to lyse,

releasing large amounts of the cyanotoxin into the water

body.

In the peer-reviewed literature only studies by Senogles

et al. (2002) and Smith et al. (2008) have demonstrated that

CYN could be biodegraded in natural water bodies. Further-

more, Smith et al. (2008) found that the concentration of CYN

influenced biodegradation with a near linear relationship

existing between the biodegradation rate and the initial CYN

concentration. Temperature and the presence of copper-

based algicides were also shown to affect the biodegradation

of CYN. Klitzke et al. (2010) only observed biodegradation of

CYN in sediments where pre-conditioning of the sediments

resulted in enhanced biodegradation of CYN. No studies to

date have isolated any definitive CYN-degrading organisms.

Fig. 4 e Molecular structure of cylindrospermopsin.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 8 1541

2.5. Saxitoxins

The saxitoxins are a group of potent alkaloid neurotoxins

produced by cyanobacteria and dinoflagellates in freshwater

and marine environments, respectively. In freshwater, the

saxitoxins were originally shown to be produced by strains of

Aphanizomenon flos-aquae (Jackim and Gentile, 1968; Alam

et al., 1978). Since then, saxitoxins have been shown to be

produced by several freshwater cyanobacterial species

including Anabaena, Raphidiopsis, Lyngbya, Planktothrix (Oscil-

latoria) and Cylindrospermopsis (Sivonen and Jones, 1999;

Murray et al., 2011). Originally isolated and characterised in

shellfish, hence their more common name, paralytic shellfish

poisons, these cyanotoxins act to block nerve cell sodium

channels and can cause death if consumed in sufficient

quantity (Kao, 1993). There are approximately 27 variants of

the saxitoxins, with the most common in drinking water

sources being the C-toxins, gonyautoxins (GTX) (including

decarbamoyl variants) and saxitoxin (STX). Of these three

classes, the doubly-sulphated C-toxins are the least toxic,

followed by the more potent singly-sulphated GTX variants,

and finally STX which is nonsulphated and the most toxic.

Although no official guideline value exists for the saxitoxins,

a provisional health alert value of 3 mg L�1 (as saxitoxin

toxicity equivalents, or STX-eq) has been suggested by

Fitzgerald et al. (1999) for the ADWG. The structure and rela-

tive toxicities of the variants are shown in Fig. 5.

Biotransformation of saxitoxin variants has been docu-

mented to occur in the marine environment, predominantly

within the tissues of marine shellfish and finfish (Shimizu and

Yoshioka, 1981; Sullivan et al., 1983; Bricelj et al., 1991;

Cembella et al., 1993; Jones and Negri, 1997). Furthermore,

studies by Kotaki and co-workers (Kotaki et al., 1985a, 1985b;

Kotaki, 1989) have suggested that marine bacteria were

capable of transforming saxitoxins; in particular, they showed

that a Pseudomonas sp. and aVibrio sp., both isolated from coral

crabs and marine snails, transformed GTX variants to more

toxic STX. The authors also showed that unidentified bacteria,

isolated from a freshwater source, were also capable of

transforming GTX2 and GTX3 to STX. Kayal et al. (2008)

assessed the fate of five saxitoxin variants (C1, C2, GTX2,

GTX3 and STX) through biologically active filters containing

N

N

R4O

H2N+

N

N

+NH 2

R2 R3

OHOH

H

N

N

N

R

R1

R1 R2 R3 Net Charge Relative Toxicity R4 = CONH2 (carbamate toxins) STX H H H +2 1.000 GTX2 H H OSO3

- +1 0.359 GTX3 H OSO3

- H +1 0.638 R4 = CONHSO3

- (n-sulfocarbamoyl (sulfamate) toxins) C1 H H OSO3

- 0 0.006 C2 H OSO3

- H 0 0.096

Fig. 5 e Structural variations and characteristics of the

saxitoxin class of cyanotoxins.

sand and anthracite media. Decreases in the concentration of

the less toxic variants (C1 and C2) coincided with increases in

the concentrations of the more toxic variants (GTX2, GTX3

and STX) through the filters containing anthracite while no

changes in variant concentrations were evident through

parallel filters containing sand alone. The authors suggested

that organisms within the biofilm of the anthracite filters

possessed the ability to biotransform the saxitoxin variants.

Studies on the potential for the saxitoxins to biodegrade in

water are sparse with only Donovan et al. (2008) providing

evidence that seven unidentified bacteria, isolated from the

digestive tract of bluemussels, were able to reduce the overall

toxicity of a saxitoxin mixture by 90% in 3 d. The authors

demonstrated that most of the bacterial isolates completely

degraded STX and neo-STX within 1e3 d, with the rate of

degradation found to follow first-order kinetics.

2.6. Anatoxin-a

Anatoxin-a is a low molecular weight neurotoxic alkaloid

which can be produced by many different genera of cyano-

bacteria including, Anabaena, Aphanizomenon, Microcystis,

Planktothrix, Raphidiopsis, Arthrospira, Cylindrospermum, Phor-

midium, Nostoc and Oscillatoria (Osswald et al., 2007). The

structure of anatoxin-a is shown in Fig. 6. Another variant of

this cyanotoxin exists and is known as homoanatoxin-a, and

this differs from anatoxin-a by an additional methylene unit

on the side chain. Anatoxin-a acts to mimic the effect of

acetylcholine and since it is not degraded by acetylcholines-

terase, can result in permanent stimulation of muscles

leading to paralysis and possibly death due to respiratory

arrest (Osswald et al., 2007).

There are many reports of animal fatalities attributed to

anatoxin-a, the first possible occurrence documented in

Canada by Gorham et al. (1964) where two cows died as

a result of drinking water contaminated with this cyanotoxin.

Anatoxin-a has also been referred to as Very Fast Death Factor

Fig. 6 e Molecular structure of anatoxin-a.

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 81542

(VFDF) because of its high lethality in mice (death in 1e4 min

after intraperitonial injection). To date, no official guideline

level exists for anatoxin-a; however, Fawell et al. (1999) sug-

gested that a 1 mg L�1 level would provide a significant safety

margin of approximately 3 orders of magnitude in relation to

drinking water.

There is a scarcity of literature regarding the biodegrada-

tion of anatoxin-a. Kiviranta et al. (1991) isolated a Pseudo-

monas sp. which could biodegrade anatoxin-a at a rate of

2e10 mg mL�1 d�1 Rapala et al. (1994) reported biodegradation

of anatoxin-a by organisms in sediments with a reduction of

25e48% after 22 d. Another study did implicate biodegradation

of anatoxin-a through granular activated carbon (GAC) filters,

although the authors were unable to conclusively demon-

strate that removal was actually due to biodegradation rather

than adsorption (UKWIR, 1996).

3. Factors affecting biodegradation ofcyanobacterial metabolites

From the literature it is evident that many of the cyano-

bacterial metabolites are susceptible to biodegradation in

water supplies (see previous section). However, there are

conflicting reports regarding the efficacy of the biodegradation

of these metabolites in water bodies. For example, while

Smith et al. (2008) documented biodegradation of CYN to only

occur in water supplies which had a history of toxic C. raci-

borskii blooms, Wormer et al. (2008) did not observe any CYN

degradation during a 40 d study, despite the water body

having been exposed to CYN-producing cyanobacteria. This

indicates that there are factors which influence biodegrada-

tion of cyanobacterial metabolites, including the dependency

of the types of degrading organisms present, in addition to

other environmental conditions.

Water temperature can be a critical factor influencing

biodegradation of cyanobacterial metabolites. Information in

the literature suggests that the temperature range for the

effective biodegradation of MIB, geosmin and microcystin is

approximately between 11 and 30 �C (Christoffersen et al.,

2002; Elhadi et al., 2006; Ho et al., 2007b, 2007c, 2007d; Hoefel

et al., 2009b). From the limited data available on the effect of

temperature on the biodegradation of CYN, the maximum

degradation rate appears to occur at approximately 25 �C, with

a much slower rate at 20 �C (Smith et al., 2008). Little to no

information is available at present regarding the effect of

temperature on the biodegradation of the other cyanobacte-

rial metabolites. The consensus from a majority of these

studies is that the lower temperature generally reduces the

rate of degradation of the metabolites, most likely due to the

organisms (or enzymes produced by the organisms) respon-

sible for the degradation being inhibited by the lower

temperatures.

An additional factor affecting biodegradation efficiency is

the abundance of the bacteria directly capable of degrading

the cyanobacterial metabolites. Research conducted to date

has demonstrated a direct relationship between the abun-

dance of degrading organisms and the rate of degradation of

MIB and geosmin (Ho et al., 2007b; Hoefel et al., 2009b), and

microcystin (Hoefel et al., 2009a; Ho et al., 2010). Another

major factor affecting the rate of cyanobacterial metabolite

biodegradation in source waters is the concentration of the

actual cyanobacterial metabolite present (Ho et al., 2007d;

Smith et al., 2008; Hoefel et al., 2009b). This finding is crucial

for the water industry as it suggests that the presence of very

high cyanobacterial metabolite concentrations may well be

met with a more rapid rate of in situ biodegradation.

Furthermore, there is currently a knowledge gap regarding

the characteristics and concentrations of other organicmatter

within source waters and the impact of those upon the

biodegradation of cyanobacterial metabolites. This may be of

significance to the water industry as the biodegradation of

cyanobacterial metabolites has been suggested to be as

a secondary substrate, where alternative organic matter,

which is usually present at mg L�1 concentrations, may be the

primary substrate for indigenous organisms (Rittmann et al.,

1995; Ho et al., 2007b). In contrast, Ho et al. (2010) suggested

that microcystin-degrading bacteria were capable of utilising

MCLR as a primary substrate at mg L�1 levels in a tertiary

treated wastewater. It is possible that cyanobacterial metab-

olites form complexes with organic matter resulting in their

enhanced mineralisation as has been observed for bacterial

degradation of nonylphenol in the presence of humic acids

(Li et al., 2007).

Klitzke et al. (2010) showed that in sediments, the presence

of aquatic dissolved organic matter yielded higher CYN

degradation rates than dissolved organicmatter released from

lysed cyanobacterial cells, suggesting that the character of

organic matter may influence biodegradation through some

sort of substrate specificity. This is highlighted in some of the

MIB and geosmin biodegradation studies, where specific

organisms were unable to degrade the T&O compounds

without being initially induced by the presence of other

organic compounds including camphor and terpinene (Eaton

and Sandusky, 2009, 2010).

The presence of other water constituents may also influ-

ence biodegradation of organic compounds as metals such as

cadmium, nickel and zinc have been shown to inhibit

biodegradation of aromatic hydrocarbons (Amor et al., 2001).

These heavy metals can inhibit the degrading organisms by

blocking critical functional groups or interfering with the

incorporation of essential metal ions with biological mole-

cules (Gadd and Griffiths, 1978; Wood and Wang, 1983;

Doelman et al., 1994).

It is apparent that all of these factors can be significantly

impacted by global warming and climate change. The chal-

lenge will be to understand whether these factors interact

synergistically or antagonistically with regards to the

biodegradation of cyanobacterial metabolites. Perhaps the

best way forward is to understand how these environmental

factors affect the degrading organisms at the genetic level, ie.

whether they induce or inhibit the transcription of the genes

which synthesize the enzymes involved in the degradation

process. This will also assist in further understanding the

origins of the lag period prior to the onset of cyanobacterial

metabolite degradation, which has been documented to occur

in may studies (Rapala et al., 1994; Christoffersen et al., 2002;

Senogles et al., 2002; Holst et al., 2003; Ho et al., 2006, 2007b,

2007d; Smith et al., 2008). Many of these studies have shown

that the lag periods could be substantially reduced, in some

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 8 1543

cases eliminated, when the organisms had been pre-exposed

to the cyanobacterial metabolites. For example, Senogles et al.

(2002) showed that CYN could be degraded by natural aquatic

bacteria in surface water. The authors found that after 30 d,

CYNwas degraded to below detection limitwith a lag period of

approximately 15 d. Upon re-addition of CYN to the sample

they observed no lag period. Likewise, Smith et al. (2008)

documented biodegradation of CYN to only occur in water

supplies which had a history of toxic C. raciborskii blooms and

found that repeated exposure of the endemic organisms to

CYN resulted in substantial decreases in the lag periods

similar to Senogles et al. (2002).

In order to comprehend why a decrease in the lag period

occurs upon cyanobacterial metabolite re-exposure, it is

important to understand why the lag period exists. There is

conjecture as to the origins of the lag period. Some studies

have suggested that the lag period may be due to the time

required for small populations of the organism(s) responsible

for the degradation to become sufficiently large to commence

degradation (Spain et al., 1980; Ventullo and Larson, 1986;

Wiggins et al., 1987; Klitzke et al., 2010). Studies have also

suggested that the reduction of the lag period may be due to

the length of time required for the enzymes responsible

for degradation to be induced (Torstensson et al., 1975;

Stephenson et al., 1984). Smith et al. (2008) inferred CYN to

be an inducer for its subsequent degradation and provided

evidence that a minimum concentration of the inducer would

be required to activate the genes involved in the biodegrada-

tion of CYN. This supports the contention that lag periodsmay

be due to the time required for the degrading organisms to

reach a specific concentration, and hence produce sufficient

quantities of the enzymes required for degradation.

Other possible explanations for the lag period include an

insufficient supply of inorganic compounds, the preferential

assimilation of other organic compounds before the target

compound of interest, or the time required for acclimatisation

to - or removal of - inhibitors present in the environment

(Vashon et al., 1982; Kuiper and Hanstveit, 1984; Lewis et al.,

1986).

It is likely that the lag period exists due to a combination of

the aforementioned factors. Nonetheless, the lag period is

a major hindrance for the application of biological treatment

processes, particularly for the removal of transient cyano-

bacterial metabolites. Further work is required to provide

definitive explanations as to the origins of lag period and it is

envisaged that this informationwill enable for the elimination

of lag periods, perhaps through some sort of environmental

and/or operational manipulation.

4. Implementation of biological filtrationprocesses

Perhaps the best way to utilise organisms for the biodegra-

dation of cyanobacterial metabolites is through biological

filtration processes. Most WTPs employ a filtration process of

some kind and it may be feasible to tailor these filters in such

a way that they not only become biologically active, but also

have the ability to harbour organisms capable of degrading

cyanobacterial metabolites. Of the biological filtration studies

conducted on cyanobacterial metabolites, a majority has been

on the microcystins, MIB and geosmin, and many of these

studies have been via sand media (Lundgren et al., 1988;

Sherman et al., 1995; Grutzmacher et al., 2002; Bourne et al.,

2006; Ho et al., 2006, 2007b, 2007c, 2007d, 2010; McDowall

et al., 2007a, 2007b, 2009; Hsieh et al., 2010). Bank filtration

has also shown promising results for the removal of cyano-

bacterial metabolites (Juttner, 1995; Miller and Fallowfield,

2001).

Biological removal of cyanobacterial metabolites has also

been reported in GAC filters (Yagi et al., 1988; UKWIR, 1996;

Newcombe et al., 2003; Elhadi et al., 2006; Ho and

Newcombe, 2007; Wang et al., 2007; Zhou et al., 2011). GAC

filters offer the advantage of two removal mechanisms,

adsorption and biodegradation, and thus are an attractive

treatment option for effective removal for organic contami-

nants. Only Wang et al. (2007) were able to discriminate

between the adsorption and biodegradation mechanisms for

microcystin removal during GAC filtration. Other media have

also been employed for biological filtration of cyanobacterial

metabolites including glass beads, porous ceramic and plastic

media (Namkung and Rittmann, 1987; Egashira et al., 1992;

Hrudey et al., 1995; Terauchi et al., 1995; Sugiura et al., 2003).

A caveat to the implementation of biological filtration for

the removal of cyanobacterial metabolites is the saxitoxins.

While studies have shown that the by-products ofmicrocystin

biodegradation are non-toxic (Bourne et al., 1996; Harada

et al., 2004; Ho et al., 2007a), experiments have indicated

that in some systems biological transformation of saxitoxins

may occur during biological filtration, rendering the filtered

water more toxic than the influent water (Kayal et al., 2008).

The operational conditions for biological filtration are

important for the successful removal of cyanobacterial

metabolites. The particle size, the chemical composition and

the roughness, or topography of the surface of media have

been identified as important factors influencing biofilm

growth and biological removal of cyanobacterial metabolites

(Hattori, 1988; Elhadi et al., 2006; Wang et al., 2007; McDowall

et al., 2007a, 2009). The effective size of the filter media has

been documented to influence removal of MIB and geosmin,

with smaller particles resulting in greater removals; a conse-

quence believed to be due to the greater surface area per unit

filter media for biofilm attachment (McDowall et al., 2007a).

Studies have suggested GAC to be a better substrate for

bacterial attachment than sand (Hattori, 1988; Wang et al.,

2007). Wang et al. (2007) attributed this to the rougher struc-

ture of GAC which contains crevasses and ridges that could

help protect newly attached bacteria from shear forces that

may be a hindrance for efficient biofilmdevelopment. Another

factor which should be taken into consideration is the surface

characteristics of the media, in particular, the presence of

extracellular polymeric substances (EPS). While McDowall

et al. (2009) implied that the physical and chemical composi-

tion of the media surface (in their case sand) could affect

biofilm attachment, the major influence was the presence of

EPS on the media surface which enhanced bacterial attach-

ment and consequently geosmin degradation. Other studies

have also documented the importance of EPS for bacterial

adhesion on biofilms on surfaces including filter media (Leon-

Morales et al., 2004; Alpkvist et al., 2006; Liu and Li, 2008).

wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 5 3 6e1 5 4 81544

The filter contact time and hydraulic loadingmay also have

a major impact on biological filtration processes. Increasing

the contact time through a filter should theoretically increase

biodegradation. However, Ho et al. (2006) showed no differ-

ence in the biodegradation ofmicrocystin through a sand filter

at contact times of 30min and 4min. It should be noted that at

both contact times, no microcystin was detected in the

effluent which may have masked any differences. Nonethe-

less, the study did demonstrate that biodegradation of

microcystin could occur not only through slow sand filters but

also through rapid sand filters. Likewise, McDowall et al.

(2007b) documented removal of MIB and geosmin to concen-

trations below analytical detection through a full-scale rapid

sand filter, confirming that efficient biodegradation of cya-

nobacterial metabolites could indeed occur at low filter

contact times.

Pre-chlorination prior to any biological filtration process

should not be employed as the presence of the disinfectant

can severely compromise the biofilm on the filter media,

which in turn, will reduce or eliminate any biodegradation of

the cyanobacterial metabolites (Metz et al., 2006; McDowall

et al., 2007a, 2007b). Evidence of the impact of adding a disin-

fectant into biological filters has been demonstrated at a full-

scale WTP in South Australia, where the filters were previ-

ously removing MIB and geosmin concentrations to below the

level of analytical detection; however, upon the addition of

monochloramine into the filter backwash water, removals of

MIB and geosmin dramatically decreased through the filters.

This was subsequently shown to be due to the inactivation of

the biofilm and biodegrading bacteria (McDowall et al., 2007b).

It is also intuitive to expect the type and concentration of

organisms to influence biological filtration of cyanobacterial

metabolites as discussed in Sections 2 and 3. Simply because

a biological filter may contain an active flourishing biofilm

does not necessarily suggest that biodegradation of the cya-

nobacterial metabolites will occur, as biodegradation is

dependent upon organisms with specific genes. A method for

enhancing the removal of cyanobacterial metabolites through

filters is to artificially inoculate or “seed” filters with organ-

isms capable of degrading the metabolites as this approach

may minimise the lag period and also increase removals.

McDowall et al. (2009) demonstrated enhanced geosmin

removals after seeding sand filters with geosmin-degrading

bacteria. Such an approach may also be a viable option for

the removal of other cyanobacterial metabolites. However,

this approach has not yet been fully optimised at the labora-

tory scale and significant advancements in the seeding

process suitable for pilot/full-scale filters are required.

5. Summary and conclusions

The cyanobacterial metabolites discussed in this review are

susceptible to biodegradation, withmany instances where the

organisms responsible for the degradation have been isolated

and characterised. However, efficient biodegradation, through

processes such as biological filtration, appears to be site

specific and dependent upon a range of factors which this

review has discussed accordingly. Much of the work in the

past has focused on the engineering aspects of biological

filtration with such assumptions as expecting that the pres-

ence of a biofilm should result in efficient biodegradation of

the cyanobacterial metabolites. This assumption is not always

valid which suggests that there may need to be a paradigm

shift in the way biological filtration is conceptualised and

what future work should be conducted. Moreover, genetic

technology has allowed for significant improvements in the

understanding of many biodegradation processes. Conse-

quently, further work is required to ensure that biological

treatment, such as biological filtration, can be confidently

applied for the removal of thesemetabolites. To date, only the

genes responsible for microcystin degradation have been

identified; hence, future work at the genetic level should focus

on this cyanobacterial metabolite since the molecular plat-

form and tools for this have been developed and optimised.

This includes understanding why the lag period, prior to

biodegradation commencing, exists and how this lag period

can be eliminated through manipulation of either environ-

mental or operational factors. The advancements in molec-

ularmicrobiology and genetic technology should also promote

future work in other cyanobacterial metabolites, in terms of

gene discovery; however, the lack of key findings in this area

to date (apart from the microcystin-degrading genes) indi-

cates that this is not a trivial exercise.

While it is evident that climate change is predicted to

increase the intensity of cyanobacterial blooms and subse-

quently, the production of these unwanted metabolites, it is

clear that climate change could also significantly influence the

biological processes which can be implemented to remove

them. This conundrum should foster research into further

understanding such biological processes, particularly with the

recent shift away from chemical- and energy-intensive

processes towards carbon and climate neutrality.

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