Oxidative elimination of cyanotoxins: Comparison of ozone, chlorine, chlorine dioxide and permanganate

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41 (2007) 3381 – 3393

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Oxidative elimination of cyanotoxins: Comparison of ozone, chlorine, chlorine dioxide and permanganate Eva Rodrı´gueza, Gretchen D. Onstadb,1, Tomas P.J. Kullc, James S. Metcalfd, Juan L. Aceroa,, Urs von Guntenb,e, a

Departamento de Ingenieria Quimica y Quı´mica Fı´sica, Faculdad de Ciencias, Universidad de Extremadura, Avda. Elvas s/n, 06071 Badajoz, Spain b Eawag, Swiss Federal Institute of Aquatic Science and Technology, Ueberlandstrasse 133, P.O. Box 611, CH-8600 Duebendorf, Switzerland c ˚ bo Akademi University, BioCity, Tykisto¨katu 6A, 20520 Turku, Finland Departments of Biology, Biochemistry and Pharmacy, A d Division of Environmental and Applied Biology, College of Life Sciences, University of Dundee, Dundee DD1 4HN, Scotland, UK e Institute of Biogeochemistry and Pollutant Dynamics, ETH Zurich, 8092 Zurich, Switzerland

art i cle info

ab st rac t

Article history:

As the World Health Organization (WHO) progresses with provisional Drinking Water

Received 6 December 2006

Guidelines of 1 mg/L for microcystin-LR and a proposed Guideline of 1 mg/L for cylindros-

Received in revised form

permopsin, efficient treatment strategies are needed to prevent cyanotoxins such as these

19 March 2007

from reaching consumers. A kinetic database has been compiled for the oxidative

Accepted 20 March 2007

treatment of three cyanotoxins: microcystin-LR (MC-LR), cylindrospermopsin (CYN), and

Available online 20 June 2007

anatoxin-a (ANTX) with ozone, chlorine, chlorine dioxide and permanganate. This kinetic

Keywords:

database contains rate constants not previously reported and determined in the present

Cyanotoxins

work (e.g. for permanganate oxidation of ANTX and chlorine dioxide oxidation of CYN and

Oxidation kinetics

ANTX), together with previously published rate constants for the remaining oxidation

Drinking water

processes. Second-order rate constants measured in pure aqueous solutions of these toxins

Disinfection by-products

could be used in a kinetic model to predict the toxin oxidation efficiency of ozone, chlorine,

Ozone

chlorine dioxide and permanganate when applied to natural waters. Oxidants were applied

OH radicals

to water from a eutrophic Swiss lake (Lake Greifensee) in static-dose testing and dynamic

Chlorine dioxide

time-resolved experiments to confirm predictions from the kinetic database, and to

Chlorine

investigate the effects of a natural matrix on toxin oxidation and by-product formation.

Permanganate

Overall, permanganate can effectively oxidize ANTX and MC-LR, while chlorine will oxidize CYN and MC-LR and ozone is capable of oxidizing all three toxins with the highest rate. The formation of trihalomethanes (THMs) in the treated water may be a restriction to the application of sufficiently high-chlorine doses. & 2007 Elsevier Ltd. All rights reserved.

Corresponding author. Tel./fax: +34924289385. Also to be corresponded to. Eawag, Swiss Federal Institute of Aquatic Science and Technology, Ueberlandstrasse 133, P.O. Box 611,

CH-8600 Duebendorf, Switzerland. Tel.: +41448235270; fax: +41448235210. E-mail addresses: [email protected] (E. Rodrı´guez), [email protected] (G.D. Onstad), [email protected] (T.P.J. Kull), [email protected] (J.S. Metcalf), [email protected] (J.L. Acero), [email protected] (U. von Gunten). 1 Current address: Department of Environmental and Occupational Health Sciences, School of Public Health and Community Medicine, University of Washington, Box 357234, Seattle, WA 98195-7234, USA. 0043-1354/$ - see front matter & 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2007.03.033

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1.

Introduction

1.1.

Cyanotoxins in drinking water

41 (2007) 3381– 3393

cause strand breakage and genotoxicity (Banker et al., 2001; Humpage et al., 2000). ANTX mimics acetylcholine, containing both amine and carbonyl moieties, to overstimulate muscle cells and cause paralysis (Carmichael et al., 1992). When choosing oxidants for the removal of the biological effects of these compounds from drinking water, the active chemical moieties must be kept in mind.

The widespread occurrence of cyanotoxins in water resources and finished drinking waters throughout the world has led to not only livestock deaths but also several cases of human hepatoenteritis and even deaths (Byth, 1980; Francis, 1878; Gugger et al., 2005; Jochimsen et al., 1998). For this reason, the World Health Organization (WHO) has set a provisional Drinking Water Guideline of 1 mg/L for microcystin-LR (MCLR) (WHO, 2004). This hepatotoxin is one of the most commonly occurring and toxic microcystins (MCs), and it originates from the cyanobacteria Microcystis, Anabaena, Planktothrix, Nostoc and Anabaenopsis. Cylindrospermopsin (CYN, see Fig. 1) is often associated with blooms of Cylindrospermopsis, Anabaena and Aphanizomenon, and is also common in occurrence (Falconer, 2005). The WHO has responded by proposing a 1 mg/L guideline for CYN, due to its hepatotoxicity, cytotoxicity and genotoxicity. Less common in occurrence, the neurotoxin anatoxin-a (ANTX) still requires further toxicity studies to establish a guideline value. It can be generated by Anabaena, Oscillatoria, and Aphanizomenon (Edwards et al., 1992). Some specific functional groups of these three cyanotoxins (Fig. 1) are responsible for their toxicities. When bonded to the peptide ring, the ADDA side chain of MC-LR (containing conjugated double bonds) is responsible for MC-LR hepatotoxicity (Fig. 1). The uracil side chain of CYN is the active moiety which inhibits protein translation to promote hepatotoxicity, or it binds to DNA to

1.2.

The most common chemical oxidants applied during drinkingwater treatment, namely ozone, hydroxyl radicals, chlorine, chlorine dioxide, chloramine and permanganate favor reaction with certain functional groups of micropollutants. Ozone attacks double bonds, activated aromatic systems and neutral amines with great specificity while hydroxyl radicals (dOH), which are formed from ozone decomposition in aqueous solutions, randomly attack carbon–hydrogen bonds in organic molecules (von Gunten, 2003). Chlorine reacts with the same moieties as ozone except for double bonds and also with much lower rates. Chlorine attack leads to less oxidation and more chlorine substitution, leading to halogenated organic compounds (e.g., chloroform and other trihalomethanes (THMs), haloacetic acids (HAAs), etc.) (Singer and Reckhow, 1999). Both ozone and chlorine unfortunately can react with bromide to form HOBr and generate brominated by-products (especially chlorine) and bromate (only ozone), which can have potentially worse health effects than their chlorinated counterparts. Chlorine dioxide reacts with tertiary amines and activated aromatic systems and does not promote brominated

HO

Microcystin-LR

Chemical oxidation

O CH3 N

ADDA

HN

NH

H3C O

O

H3C

CH3

CH3

H 2N

O

O

H 3C

NH

N

N H

N

+

Cylindrospermopsin

O O HO

Uracil

H

H

O

HN H N

O

HO

S

CH3

H N

O

O NH

-

CH2

O

CH3

NH

O

O

CH3

Mimic of Acetylcholine O

HN

H3 C

O

H N

O

NH

CH3

O

Anatoxin-a

Fig. 1 – Structures of cyanotoxins: microcystin-LR, cylindrospermopsin and anatoxin-a.

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4 1 (200 7) 338 1 – 339 3

Table 1 – Literature survey of cyanotoxin reactivity with oxidants in drinking-water treatment Cyanotoxin reactivity with oxidants Microcystins (mainly MC-LR) ClO2 k measured Cl2 k measured, dep. on HOCl speciation O3 k measured, MCLR4CYN4ANTX MnO 4 k measured (MC variants) O34MnO 4 4ClO2 Cl24NH2Cl Cl2 best at pHo8 X0.5 mg/L res. Cl2, after 30 min contact Cl2 reactivity of MC variants 50% oxidized by MnO 4 O3 releases and oxidizes intracellular toxins MnO 4 does not oxidize released toxins O3 dose dependence on DOC, treatment step O3 oxidizes MC-LR 4 ANTX O3-MC-LR review Cylindrospermopsin (CYN) Cl2 k measured MnO 4 k measured O3 k measured Chlorination product studies Chlorine4chloramine4chlorine dioxide Anatoxin-a (ANTX) Cl2 k measured O3 k measured Permanganate4ozonation O3 oxidizes MC-LR 4 ANTX

Water resource, matrix studied

References

Pure water; lake water: Finland Pure water

Kull et al. (2004, 2006) Acero et al. (2005)

Pure water; lake water: Switzerland

Onstad et al. (2007)

Pure water Reservoir and river: UK (England) Lake water: Australia Pure water, lake water: Australia Lake water: Australia

Chen et al. (2005); Rodrı´guez et al. (2007a) (Fawell et al. (1993); Hall et al. (2000) Nicholson et al. (1994) Acero et al. (2005); Nicholson et al. (1994) Nicholson et al. (1994)

Reservoirs: Poland, Australia Lake waters: USA (Wisconsin) Reservoir: Germany

Acero et al. (2005); Ho et al. (2006) Karner et al. (2001) Schmidt et al. (2002)

Reservoir: Germany

Schmidt et al. (2002)

Reservoir, river, lake: Australia; artificial lake water: Switzerland; reservoir: UK Reservoir, river, lake: Australia; lake water: Finland

Fawell et al. (1993); Hoeger et al. (2002); Rositano et al. (2001) Keijola et al. (1988); Rositano et al. (2001) Falconer (2005); Hitzfeld et al. (2000)

Pure water Pure water Pure water; lake water: Switzerland Australia, Israel Israel

Rodrı´guez et al. (2007b) Rodrı´guez et al. (2007b) Onstad et al. (2007) Banker et al. (2001); Falconer (2005); Senogles-Derham et al. (2003) Banker et al. (2001)

Pure water Pure water; lake water: Switzerland River: England Reservoir, river, lake: Australia; lake: Finland

Rodrı´guez et al. (2007b) Onstad et al. (2007) Hall et al. (2000) Keijola et al. (1988); Rositano et al. (2001)

by-products (Hoigne´ and Bader, 1994). Its reactivity is frequently found on a scale between ozone and chlorine. The application of chlorine dioxide is regulated at minimal doses to prevent accumulation of chlorite and chlorate as oxidation by-products. Chloramines are undoubtedly poor oxidants, but are used frequently as residual disinfectants to minimize total THM (TTHM, four chloro/bromo analogues) formation. Permanganate mainly reacts with double bonds by donating oxygens, but it can also abstract hydride ions, electrons or hydrogen atoms (Stewart, 1964). In a review of permanganate rate constants for reaction with unsaturated contaminants, Waldemer and Tratnyek (2006) reported values in the range of 1–100 M1 s1. They are highest for chlorine-substituted phenols and ethenes, and data for amine moieties were notably lacking. Permanganate does not promote the formation of chlorinated or brominated by-products.

1.3.

Oxidative treatment of cyanotoxins

Ozonation and activated carbon are reported to be the besttreatment options for microcystins (MCs), as opposed to

conventional flocculation–filtration–chlorination treatment (Himberg et al., 1989). Oxidative treatment studies in the literature are dominated by MC-LR (Hitzfeld et al., 2000), with few studies on CYN and ANTX (see Table 1 for literature overview). Studies of the effects of water quality on cyanotoxin removal have focused on pH and DOC concentration, which also depend on the application point of the oxidant in the treatment train (Acero et al., 2005; Hoeger et al., 2002; Rositano et al., 2001). Previous studies comparing oxidants mostly focused on the percent oxidation of a toxin by a given oxidant dose or the residual oxidant level necessary for complete toxin oxidation. While these studies are valuable for specific sites with challenging water quality and treatment issues, the results often depend on oxidant consumption by the natural water matrix and cannot be applied to other waters. Although source waters varied in water quality, different research groups came to common conclusions concerning the oxidative treatment of cyanotoxins (Table 1). Rositano et al. (2001) determined that MC-LR was more susceptible to oxidation by ozone than ANTX (also found by Keijola et al., 1988). Fawell et al. (1993) and

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41 (2007) 3381– 3393

Hall et al. (2000) found that the necessary oxidant dose for MC-LR oxidation decreases in the order chlorine dioxide 4 permanganate 4 ozone. Nicholson et al. (1994) found that chlorine was much better than chloramine for oxidation of MC-LR. Hall et al. (2000) observed that permanganate was more efficient than ozone for oxidation of dissolved ANTX. Reaction mechanism and product studies have been limited mainly to MC-LR. Lawton and Robertson (1999) present an overview of potential reaction schemes for MC-LR oxidation by ozone, permanganate and chlorine. Tsuji et al. (1997) found dihydroxy-MC-LR as a chlorination product of MC-LR, which is of lower toxicity. Banker et al. (2001) found chlorination products of CYN (attack on uracil moiety) that were no longer toxic in a mouse bioassay. Senogles-Derham et al. (2003) found that THMs and HAAs formed from chlorination of MCLR and CYN were below the recommended drinking-water guideline levels. Non-toxic MC-LR oxidation products have been identified after treatment with ozone by Hoeger (2003), and with chlorine dioxide by Kull et al. (2004, 2006). The reactivities of different microcystin variants with chlorine have been investigated by Acero et al. (2005) and Ho et al. (2006). Very few studies have been published (outside of reports and Newcombe and Nicholson, 2004) on CYN oxidation by ozone, permanganate, chlorine dioxide, or chloramine, or on ANTX oxidation by chlorine or chloramine (Rodrı´guez et al., 2007b). The objective of this study is to compare oxidants on a quantitative basis, for their ability to oxidize three cyanotoxins (MC-LR, CYN, ANTX) in their dissolved form (extracellular) during drinking-water treatment. First, a compilation of the second-order rate constants for reactions of oxidants with cyanotoxins are presented for comparison, with some recently reported data (Acero et al., 2005; Kull et al., 2004; Onstad et al., 2007; Rodrı´guez et al., 2007a, b). Second, static-dose testing is applied to a Swiss lake water (Lake Greifensee) to determine the necessary oxidant dose for toxin oxidation and the resulting oxidation by-product formation. In cases, where decrease of the toxin and the oxidant can be measured simultaneously over time, dynamic timeresolved experiments are presented together with by-product formation.

2.

Materials and methods

2.1.

Cyanotoxins

MC-LR was isolated and purified from Microcystis and Anabaena cultures (Meriluoto and Codd, 2005), CYN from Cylindrospermopsis raciborskii (Metcalf et al., 2002) and ANTX from Phormidium (Edwards et al., 1992). Standardization and handling followed a manual of standard operating procedures within the European Union project ‘‘TOXIC’’ (Meriluoto and Codd, 2005). The purity of MC-LR was estimated as 498% based on reversed-phase analytical HPLC with UV absorbance detection at 238 nm. Similarly, ANTX and CYN were purified to 499% according to HPLC-UV. Purified toxin samples were stored freeze dried until use.

2.2.

Aqueous chemical kinetics

Rate constants for the reaction of CYN and ANTX with chlorine dioxide were determined by pseudo-first-order kinetics as in Huber et al. (2005). Due to the limited availability and the need to conserve cyanotoxins, pseudo-first-order kinetics were only conducted with the oxidant in excess. In batch kinetic experiments, stock solutions of chlorine dioxide were dosed into buffered cyanotoxin-containing MilliQ water solutions with a gas-tight syringe during mixing (glass bottles were fitted with a screw top dispenser for sample collection). Rate constants for reaction of ANTX with permanganate were determined using a continuous-quenched flow system, with 25 mL syringes, with permanganate in excess (Dodd et al., 2005). Removal of colloidal MnO2 (product of the permanganate reaction) was not possible by filtration (0.2 mm), but still necessary because MnO2 can act as a catalyst in the permanganate reaction with amines (Rosenblatt et al., 1968). Therefore, ascorbic acid was used as a quenching agent for both permanganate and MnO2. Very little information exists in the literature for rate constants concerning the reaction between ascorbic acid and permanganate (Kumar et al., 1996; Rao et al., 1982), but it was successful in this application and has been used previously to quench MnO2 (Zhang and Huang, 2003). The continuousquench flow system (applied to ANTX-MnO 4 ) consisted of three gas-tight syringes (25 mL each), one containing ANTX (3 mM), one containing dilute permanganate (50–150 mM) and one containing the quenching agent ascorbic acid (3 mM). The first two reactants were mixed in an eight-way valve (KinTek Corp, Clarence, PA, USA) within reaction loops varying in volume from 15.6 to 241.6 mL. A syringe pump maintained flow (2.5–12.5 mL/min) to the mixing valve. Quenched samples were collected in amber vials for HPLC-DAD analysis of ANTX. This system was operated following methods of Dodd et al. (2005).

2.3.

Comparison of oxidants applied to natural water

Experiments were performed with water from Lake Greifensee, a eutrophic lake in Switzerland (3.6 mg/L DOC, 3.6 mM bicarbonate alkalinity, 100 mg/L ammonia and 50 mg/L bromide—after spiking 40 mg/L) with the goal of comparing the efficiency of the investigated oxidants on the elimination of toxins. The experiments were carried out at 20 1C and pH 8 (borate buffer, 10 mM). Additionally, experiments of MC-LR chlorination at pH 6 and 7 (phosphate buffer, 10 mM) were performed. The initial concentration of toxin spiked into filtered (0.45 mm) water was 1 mM (1 mg/L MC-LR, 166 mg/L ANTX and 415 mg/L CYN). Static-dose testing was performed in separate amber glass vials to which different oxidant doses were added. The initial oxidant concentrations were 0–2.0 mg/L ozone, 0–1.5 mg/L permanganate and 0–4 mg/L chlorine. The experiments lasted until the concentrations of the oxidants were below the detection limits. Residual toxin concentrations and TTHM or bromate formation were analyzed. Dynamic time-resolved experiments were performed to determine the evolution of concentrations of oxidants (chlorine and permanganate), toxins and TTHMs until total oxidant

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4 1 (200 7) 338 1 – 339 3

consumption. At each selected time, one sample was withdrawn from the reactor for direct measurement of the residual chlorine or permanganate and a second one for toxin and TTHM analysis after quenching the reaction with thiosulfate.

2.4.

3385

are given in Table 2 for comparison of the relative magnitude and importance of reactions between oxidants and cyanotoxins, especially for Lake Greifensee water. Furthermore, the half-lives are presented in Table 2 at a given oxidant concentration of 1 mg/L, predicting the order of reactions that could occur between cyanotoxins and different oxidants.

Oxidant solutions and analytical methods 3.1.1.

Stock solutions of chlorine were prepared by diluting a commercial solution of sodium hypochlorite (4% active chlorine, Aldrich) and standardized spectrophotometrically in the presence of excess iodide to form triiodide (Bichsel and von Gunten, 1999). Chlorine dioxide stock solutions of 3 mg/L were prepared by carefully combining aqueous solutions of potassium peroxodisulfate with sodium chlorite and collecting the gas in ice-cooled MilliQ water, following the method of Gates (1998). Chlorine dioxide stock concentrations were measured spectrophotometrically at 359 nm (e ¼ 1200 M1 cm1), and stock solutions stored at 4 1C for up to 3 months, as in Huber et al. (2005). Chlorine and chlorine dioxide concentrations in reaction samples were measured by the ABTS method (Pinkernell et al., 2000). Permanganate stock solutions (10 mM) were prepared by dissolving 158 mg of potassium permanganate in 100 mL MilliQ water. Dilute solutions of permanganate were prepared in buffered MilliQ water (pH 6–10). Permanganate was analyzed with DPD (Clesceri et al., 1999) after filtering the sample through 0.22 mm nylon filters to remove manganese oxide. Aqueous ozone stock solutions (1–1.6 mM) were prepared by bubbling O3-containing oxygen through ice-cooled MilliQ water. Ozone stock concentrations were measured spectrophotometrically at 358 nm (e ¼ 3000 M1 cm1) and ozone residual concentrations were measured by conversion of cinnamic acid to benzaldehyde and quantified by HPLC-DAD at 254 nm, as described in Onstad et al. (2007). Toxins were analyzed by HPLC-DAD (Meriluoto and Codd, 2005). Total THMs were measured by HS-GC-ECD (Golfinopoulos et al., 2001). Bromate was measured by Dionex IC-PCR-UV (Salhi and von Gunten, 1999).

3.

Results and discussion

3.1.

Oxidation kinetic database

Table 2 shows the database of aqueous kinetics for the oxidation of MC-LR, CYN and ANTX, developed during the European Union project ‘‘TOXIC’’. The data underlined were recently reported (Acero et al., 2005; Kull et al., 2004; Onstad et al., 2007; Rodrı´guez et al., 2007a, b), the other data were measured in this study. Species-specific second-order rate constants highlight the influence of solution pH on the overall reaction. Important parameters are the pKa of HOCl for the reaction with MC-LR and CYN, the pKa of CYN for the reactions with ozone and chlorine and the pKa of ANTX for the reactions with ozone and permanganate. Fig. 2 shows a plot of the pH dependence of these reactions according to the data in this paper and previously reported data (Acero et al., 2005; Kull et al., 2004; Onstad et al., 2007; Rodrı´guez et al., 2007a, b). Apparent second-order rate constants (kapp) at pH 8

Ozone and hydroxyl radicals

Ozone reacts with the double bonds in MC-LR and ANTX, and the deprotonated amine moieties in ANTX and CYN (Onstad et al., 2007). The reactivity of ozone with the double bonds of MC-LR is not pH dependent, while for the reaction with the amines the pH dependence is consistent with the pKa values of the amine moieties (Fig. 2a–c). The reactivity of ozone with CYN is pH dependent, consistent with the pKa (8.8 in Table 2) of its side-chain uracil. At pH 8, the cyanotoxins react with ozone in the order MC-LR4CYN4ANTX (Table 2). This is also true with respect to hydroxyl radicals (Table 2), which are generated during ozone decomposition in natural waters (Onstad et al., 2007).

3.1.2.

Chlorine and chloramine

Second-order rate constants for the oxidation of MC-LR with chlorine have been determined under pseudo-first-order conditions (Acero et al., 2005). The apparent second-order rate constant at 20 1C and pH 8 was found to be 33 M1 s1 (Table 2). A similar experimental approach was applied to determine the rate constants for the reaction of chlorine with CYN and ANTX (Rodrı´guez et al., 2007b); the values obtained at pH 8 are included in Table 2. It was found that an increase of pH had a negative effect on the MC-LR oxidation rate (Acero et al., 2005), while a maximum oxidation rate of CYN was observed at pH 7 (Rodrı´guez et al., 2007b). The kinetic approach used to determine the reactivity of chlorine with CYN was used to identify the pKa (6.5 in Table 2) defining its pH dependence. The pH profiles for the apparent secondorder rate constants are plotted in Fig. 2a–c. At circumneutral pH, the reactivity order is CYN 4 MC-LR b ANTX. The oxidation of the three investigated toxins with monochloramine are very slow processes, with second-order rate constants o1 M1 s1. From these results, it can be concluded that chlorination is a suitable option for oxidation of MC-LR and CYN during drinking-water treatment, whereas the oxidation of ANTX by chlorine is too slow.

3.1.3.

Chlorine dioxide

The second-order rate constant for reaction of ClO2 with MCLR was determined previously by pseudo-first-order batch kinetic methods (Kull et al., 2004). As shown in Table 2, the apparent rate constant at pH 8 was low (1 M1 s1) with a small pH dependence (Kull et al., 2004). The reactivity of CYN with ClO2 was similar at pH 8 (0.970.1 M1 s1) but decreased with decreasing pH (kapp ¼ 0.370.1 M1 s1 at pH 6). The reactivity of ANTX with ClO2 was low and the second-order rate constant was not measureable. This observation for ANTX, and the magnitude of kapp for MC-LR and CYN shown in Fig. 2a,b, suggests that ClO2 is not an appropriate oxidant for oxidative removal of these cyanotoxins during drinkingwater treatment.

3386

Table 2 – Kinetic database for the reaction of oxidants with cyanotoxins: second-order rate constants, apparent rate constants (pH 8) and half-livesa (pH 8) at 20 1C Oxidant

Microcystin-LR (MC-LR)

2.07  10 116 6.78

pKa 6.5

kapp (M1 s1)

t1/2a

kHT+ (M1 s1)

33

24.8 min

o1 4.1  105 1.1  1010 357 1

4 14 h 0.08 s 5 minc 5.2 min 13.1 h

(HOCl)

or 8:8 ðO3 Þ kT 1 1

(M

s

)

pH 8

pH 8

kapp (M1 s1)

t1/2a

490

1.7 min

o1 3.4  105 5.5  109 0.3 0.9

4 14 h 0.10 s 10.5 minc 4.2 d 14.4 h

pKa 9.4 kHT+ (M1 s1)

kT 1 1

(M s )

pH 8

pH 8

kapp (M1 s1)

t1/2a

o1

4 14 h

o1 6.4  105 3.0  109 2.3  104d Low

4 14 h 0.52 s 19 minc 4.8 s

7b

4.1  105 1.1  1010 357

38

1960

E40

E2.5  106

2.8  104

E8.7  105

2.1  104

4.3  104

Note: Underlined values determined in previous studies (Acero et al., 2005; Kull et al., 2004; Onstad et al., 2007; Rodrı´guez et al., 2007a b). All other values determined in this study. N Determined at 25 1C. a Assuming a constant oxidant concentration of 1 mg/L. b Third-order rate constants with units of M2 s1. c Assuming Rct ¼ 108, where Rct ¼ [dOH]/[O3]. For 1 mg/L O3: [dOH]ss ¼ 2  1013 M. Rct depends on natural water matrix with typical Rct values 108–106. d Accounts only for the reaction of ANTX with permanganate, neglecting autocatalysis by MnO2.

41 (2007) 3381– 3393

H +HOCl HOCl OCl NH2Cl O3 d OH MnO 4 ClO2

pH 8

ARTICLE IN PRESS

+

pH 8

Anatoxin-a (ANTX) WAT E R R E S E A R C H

kT (M1 s1)

Cylindrospermopsin (CYN)

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4 1 (200 7) 338 1 – 339 3

3.1.4.

O3

log kapp (M-1s-1)

6 5 4

MnO4-

3 2

chlorine

1 ClO2

0

ANTX þ 1:5 MnO 4 ! products þ xMnO2 .

-1 4

6

8

10

CYN

7

12

O3

6 log kapp (M-1s-1)

Permanganate

The reactivity of permanganate with MC-LR was not pH dependent (Table 2), kapp ¼ 357 M1 s1 at pH 8 (Rodrı´guez et al., 2007a), consistent with other MC variants (Chen et al., 2005). Permanganate was not very reactive with CYN (kapp ¼ 0.3 M1 s1, (Rodrı´guez et al., 2007b)), similar to the low reactivity of uracil with permanganate (Freeman et al., 1981). The second-order rate constant for the reaction between ANTX and permanganate was higher than could be measured by batch kinetic methods. For this reason, a continuous-quenched flow system was applied for the determination of rate constants (see experimental section). The stoichiometry with respect to permanganate was determined to be 1.5, as in the following reaction:

MC-LR

7

5 4 3

3387

chlorine

2

(1)

At pH 7, the first-order rate constant for ANTX was dependent on the permanganate concentration (Fig. 3a), the slope of this line estimates the apparent second-order rate constant of 2.1(70.3)  104 M1 s1 (n ¼ 5, 95% confidence interval) and the log–log plot of this data could be used to determine the reaction order of 1 with respect to permanganate (Fig. 3b). The following equations describe the dependence of the rate on the concentrations of ANTX and permanganate, the reaction order with respect to permanganate and the dependence of k0 on initial permanganate concentration, which was used to determine the apparent kapp at pH 6–10. 

d½ANTX = ¼ kapp ½ANTX½MnO 4  ¼ k ½ANTX, dt

(2)

1 ClO2

0

MnO4-

k= ¼ kapp ½MnO 4 .

4

From pH 6 to 8, the reactivity of ANTX with permanganate was constant at 2.1(70.1)  104 M1 s1 (n ¼ 5). A very clear pH dependence was observed for kapp between pH 8 and 10, consistent with the pKa ¼ 9.36 for the protonated secondary amine in ANTX (Koskinen and Rapoport, 1985). The plot of kapp vs. a (degree of ANTX dissociation) in Fig. 3c estimates k00T ¼ 4.3(70.3)  104 M1 s1 (n ¼ 7) for the neutral ANTX species (Table 2). This completes the model equation to predict kapp over the entire pH range (6–10) shown in Fig. 3d by the following equation:

3

kapp ¼ ð1  aÞk00HT þ ak00T .

2

3.2.

-1 4

6

8

ANTX

7

10

12

O3

6 log kapp (M-1s-1)

(3)

MnO4-

5

1 chlorine

0 -1 4

6

8

10

12

pH Fig. 2 – pH dependence of the reactions between oxidants and cyanotoxins at 20 1C: (a) MC-LR, (b) CYN and (c) ANTX.

(4)

Comparison of oxidants in natural water

Based on results from kinetic studies (see Fig. 2), ozone, permanganate and chlorine were chosen as the three oxidants for comparison of cyanotoxin oxidation in water from Lake Greifensee (Fig. 4). As shown in Figs. 2 and 4, ozone is the best oxidant for oxidation of MC-LR, followed by permanganate and chlorine. Ozone is also the best choice for oxidation of CYN, followed by chlorine, with poor oxidation by permanganate. Table 2 predicts that ozone is the best oxidant for oxidation of ANTX, followed closely by permanganate, with minimal oxidation by chlorine. However, Fig. 4 reveals greater oxidation of ANTX by permanganate rather than ozone (similar to Hall et al., 2000). This is likely

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2.5

0.4 y = 0.931 x + 4.125

y = 2.12x104 x +0.06

2.0

0.2

1.5

log k′

k′ (s-1)

0.3

R2 = 0.996

R2 = 0.993

0.1 1.0 0 0.5

-0.1

0.0

-0.2 25

50

75

100

-4.7 -4.6 -4.5 -4.4 -4.3 -4.2 -4.1 log [KMnO4]o

5

5

4

4

y = 2.2 x + 2.1 R2 = 0.995

kapp (x 104 M-1s-1)

kapp (x 104 M-1s-1)

μM) [KMnO4]o (μ

3 2 1 0 0.0

-4

ANTX+

ANTX

1 0.8

Model 3

Data

0.6

2

0.4

1

0.2

0 0.2

0.4

0.6

0.8

1.0

ANTX speciation

0

0 4

α

6

8 pH

10

12

Fig. 3 – Rate constant determination for reaction of permanganate with ANTX: (a) first-order rate constants (k0 ) of ANTX at pH 7 with varying permanganate concentrations, slope ¼ kapp , (b) log–log plot of (a) where the slope ¼ permanganate reaction order, (c) dependence of the apparent second-order rate constant on the degree of ANTX dissociation (pH 6–10), where intercept ¼ k00HT and slope+intercept ¼ k00T and (d) pH dependence of the reactivity of ANTX with permanganate, overlayed with ANTX amine speciation. [ANTX]o ¼ 1.5  106 M.

due to the higher stability of MnO 4 in Greifensee water, which yields a higher CT (oxidant exposure) for a given dose. Despite the influence of the natural matrix of Lake Greifensee on the oxidation of ANTX by permanganate, the order of oxidants for toxin oxidation corresponds roughly to the relative half-life times in the kinetic database (Table 2).

present at a reasonable concentration of 100 mg/L in Lake Greifensee water. It was shown in a previous study that such ammonia levels can reduce bromate by about 50% relative to the ammonia-free water (Pinkernell and von Gunten, 2001). Therefore, even in an ammonia-free water a safety factor of at least 5 can be expected.

3.2.1.

3.2.2.

Oxidation with ozone

Due to the magnitude of the apparent rate constants for the reactions of ozone with cyanotoxins (Table 2), static-dose testing was applied for comparison of the efficiency of oxidation of the three cyanotoxins (Fig. 5). Approximately 95% oxidation was achievable by 0.25 mg/L O3 for MC-LR, 0.38 mg/L O3 for CYN and 0.75 mg/L O3 for ANTX. This order confirms that predicted by kapp and half-life times at pH 8 for ozone and hydroxyl radicals (Table 2), and the work of Rositano et al. (2001). The formation of the ozonation byproduct bromate was only detected at doses higher than necessary for toxin oxidation (Fig. 5). Even at a 2 mg/L ozone dose, bromate levels were maintained below the WHO guideline value of 10 mg/L (WHO, 2004). The spiked level of 50 mg/L bromide is undoubtedly relatively low, however, bromate formation for complete oxidation of ANTX was only about 1 mg/L. Part of this low-bromate formation can be explained by the suppressing effect of ammonia, which is

Oxidation with permanganate

Fig. 4 compiles the results obtained in the oxidation of MC-LR, ANTX and CYN in experiments performed with different permanganate concentrations under static-dose testing conditions. The required permanganate doses for cyanotoxin oxidation were greatest for CYN followed by MC-LR and then ANTX, confirming the order of reactivity with permanganate determined by measurement of the second-order rate constants (Table 2). Thus, 1.5 mg/L of permanganate could only remove 10% of CYN, while 1.5 and 0.5 mg/L were enough to completely remove MC-LR and ANTX, respectively. Therefore, even low permanganate doses can partially remove MC-LR and especially ANTX. Higher permanganate doses lead to higher permanganate exposure (CT) and therefore greater toxin removal. From these results, it can be deduced that permanganate is a good oxidant for the elimination of MC-LR and ANTX in natural waters.

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2.0

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0.0 0.00 0.13 0.25 0.38 0.50 0.75 1.00 2.00

1.0

Ozone Dose (mg/L) MnO4-

Fig. 5 – Relative decrease of cyanotoxin and increase of bromate formation as a function of the ozone dose. Error bars indicate standard deviation. Experimental conditions: pH 8, 20 1C, 3.6 mg/L DOC, 100 lg/L ammonia, 3.6 mM  alkalinity as HCO 3 and 50 lg/L Br .

0.6 chlorine 0.4

0.0 0.0

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[permanganate], mg/L

O3

ANTX/ANTX0

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1

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permanganate

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MC-LR, experimental MC-LR, calculated

0.6

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0.2 0

0

0.6

0.8 [toxin], mg/L

CYN/CYN0

0.8

0

50

100

time, min

0.4

O3

0.2

Fig. 6 – MC-LR oxidation by permanganate (1 mg/L) in Lake Greifensee water. Experimental conditions: pH 8, 20 1C, 3.6 mg/L DOC, 100 lg/L ammonia, 3.6 mM alkalinity as HCO 3 and 50 lg/L Br.

MnO4-

0.0 0.0

0.5

1.0

1.5

2.0

Oxidant Dose (mg/L) Fig. 4 – Required oxidant doses for toxin oxidation (1 lM) in Lake Greifensee water: (a) MC-LR, (b) CYN, and (c) ANTX. Symbols: K ozone (+ dOH), m permanganate, E chlorine. Experimental conditions: pH 8, 20 1C, 3.6 mg/L DOC, 100 lg/L  ammonia, 3.6 mM alkalinity as HCO 3 and 50 lg/L Br .

The results obtained in a dynamic time-resolved experiment performed with MC-LR by analyzing permanganate and toxin concentrations are shown in Fig. 6. The initial concentration of permanganate was 1 mg/L. An initial fast decrease of the oxidant concentration was observed due to fast reactions of permanganate with highly reactive organic matter present in the natural water. Then, the oxidant

decrease becomes slower until complete depletion, which was reached after 2 h. A reduction of the MC-LR concentration higher than 90% was achieved. Natural water quality parameters such as pH usually play an important role for the oxidant efficiency. However, pH is not a key factor for the elimination of MC-LR by permanganate, as was observed in experiments performed with 0.75 mg/L of permanganate at pH values from 6 to 8, which led to toxin removal within the range of 85–90%. These results corroborate our predictions that permanganate is a suitable oxidant for MC-LR removal during drinking-water treatment (Rodrı´guez et al., 2007a). Similar experiments were performed with ANTX and CYN by applying 1 mg/L of permanganate to Lake Greifensee water. As expected from the second-order rate constants (Table 2), the oxidation of ANTX was extremely fast and could not be

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120

Oxidation with chlorine

A Chlorine dose of 3 mg/L was necessary for the almost complete MC-LR oxidation in Lake Greifensee water (Fig. 7a). CYN presented the highest reactivity with chlorine at pH 8, where 1.5 mg/L of chlorine was enough to eliminate this toxin completely (Fig. 7b). Only 15% of ANTX could be transformed by applying a chlorine dose of 3 mg/L (data not shown). This reactivity order also agrees with the values of the secondorder rate constants determined for the chlorine oxidation of these three cyanotoxins at pH 8 (Table 2). In the presence of bromide, hypobromous acid is formed from the reaction between chlorine and bromide. Then, hypobromous acid would react rapidly with NOM and ammonia present in this lake water or with NH2Cl to form NHClBr. Therefore, when bromide is present at low concentration, the amount of hypobromous acid available to react with the toxins must be low. However, in waters with high bromide levels, bromine reactions might become important. A negative consequence of chlorine application for oxidation purposes is the formation of THMs from the reaction of chlorine with the natural organic matter (NOM) present in the natural water, which depends on the chlorine dose and on the character and concentration of NOM. The formation of TTHM during the elimination of MC-LR and CYN is also shown in Fig. 7. As can be observed at a chlorine dose of 3 mg/L which was required for almost complete MC-LR elimination, around 100 mg/L of TTHM were formed, equal to the drinking water standard of the European Union (EU, 1998). Since the elimination of CYN required only 1.5 mg/L of chlorine, the final TTHM formation was around 45 mg/L, well below the standard value of 100 mg/L. However, if ANTX had to be removed by chlorine, TTHM formation would be significantly above this standard value. The dominant THM was chloroform, followed by bromodichloromethane. The presence of chlorodibromomethane and bromoform was not significant although the initial concentration of bromide was 50 mg/L. The pH dependence of TTHM formation and MC-LR oxidation is shown in the Supporting Information (Fig. SI1). Formation of THMs would be different in other natural waters. Therefore, similar experiments must be performed with each particular water in order to determine the formation of TTHM with the chlorine dose needed to reach the required toxin elimination. The change in chlorine, cyanotoxin and TTHM concentrations was also followed over time. Fig. 8 shows the results obtained during the oxidation of MC-LR with 3 mg/L of chlorine. An initial fast decrease of the chlorine concentration can be observed due to fast reactions of chlorine with highly

0.8 100 0.6

80 60

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40 0.2

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0 0

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chlorine dose, mg/L

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3.2.3.

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TTHM, µg/L

followed within the first reaction time points. Therefore, a preoxidation step with permanganate is well suited for the elimination of ANTX in contaminated surface water. In contrast, the oxidation of CYN was very slow which corresponds to the low second-order rate constant for its reaction with permanganate (Table 2), leading to a cyanotoxin elimination after complete permanganate depletion (5 h) of around 5%. The permanganate exposure required for CYN elimination would be very high and not achievable under realistic treatment conditions. As a consequence, permanganate is not a feasible option for the removal of CYN from surface waters.

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0 0

0.2 0.4 0.5 0.6 0.8

1

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2

chlorine dose, mg/L Fig. 7 – Oxidation of cyanotoxins with chlorine in Lake Greifensee water: (a) MC-LR and (b) CYN. The bars stand for the normalized cyanotoxin concentration and the line for TTHM formation. Experimental conditions: pH 8, 20 1C, 3.6 mg/L DOC, 100 lg/L ammonia, 3.6 mM alkalinity as HCO 3 and 50 lg/L Br. reactive organic moieties (such as resorcinol) and inorganic compounds (e.g., ammonia) present in Lake Greifensee water. Thereafter, the chlorine decrease became slower until complete depletion, which was reached after 24 h. MC-LR elimination was fast, being almost completed after 2 h in this surface water. TTHM formation is also depicted in Fig. 8, resulting in a continuous increase up to about 100 mg/L when chlorine was completely consumed. This TTHM formation is close to the EU standard value of 100 mg/L in drinking water. However, the TTHM concentration was only about 70 mg/L after 2 h when MC-LR was completely oxidized. Therefore, chlorine is a suitable oxidant for MC-LR removal during drinking-water treatment as long as the TTHM formation is below the standard value. A linear correlation of TTHM formation and MC-LR decrease during chlorination of MCLR-containing Lake Greifensee water is shown in the SI (Fig. SI2). A dynamic time-resolved experiment was also performed with ANTX, with an initial chlorine concentration of 3 mg/L

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chlorine

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time, min

Fig. 8 – MC-LR oxidation by chlorine (3 mg/L) in Lake Greifensee water. Experimental conditions: pH 8, 20 1C, 3.6 mg/L DOC, 100 lg/L ammonia, 3.6 mM alkalinity as HCO 3 and 50 lg/L Br.

Fig. 10 – CYN oxidation by chlorine (1 mg/L) in Lake Greifensee water. Experimental conditions: pH 8, 20 1C, 3.6 mg/L DOC, 100 lg/L ammonia, 3.6 mM alkalinity as HCO 3 and 50 lg/L Br.

3 2.5

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chlorine

2 ANTX, experimental

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0.6 0.4

[ANTX], µM

[chlorine], mg/L

3.3.

1

0.2

0.5 0 0

100

200 300 time, min

400

0 500

Fig. 9 – ANTX oxidation by chlorine (3 mg/L) in Lake Greifensee water. Experimental conditions: pH 8, 20 1C, 3.6 mg/L DOC, 100 lg/L ammonia, 3.6 mM alkalinity as HCO 3 and 50 lg/L Br.

(Fig. 9). The degradation of ANTX was very slow which corresponds to the low reactivity of this toxin with chlorine. Only 18% of ANTX could be oxidized after complete chlorine consumption (24 h), with TTHM formation of around 100 mg/L. Therefore, the elimination of ANTX by chlorine would require high-chlorine doses and stability in the surface water, which would lead to TTHM formation significantly above the EU drinking-water standard, together with a long reaction time (not available in waterworks). As a consequence, chlorine is not a feasible option for ANTX removal. Fig. 10 shows the results obtained in a dynamic experiment performed with CYN and an initial chlorine dose of 1 mg/L. The oxidation of CYN was very fast, with almost complete reaction after 30 min. The reactivity of CYN with chlorine is very high, especially at pH 7 (Fig. 2b), indicating that this cyanotoxin can be easily oxidized in surface water using a low-chlorine dose. Therefore, the final TTHM formation is expected to be well below the EU drinking-water standard.

Kinetic modeling

In the present study, a kinetic model was applied to predict the elimination of toxins in natural waters by permanganate and chlorine in dynamic time-resolved experiments. It is based on a previously applied model for oxidation of micropollutants during ozonation processes (Acero et al., 2000). For that, in addition to the rate constants, the decrease of permanganate or chlorine in the natural water must be known. With the experimental oxidant concentrations, the oxidant exposure (CT value), defined as the integral of oxidant concentration over the reaction time, can be determined by the following equation: Z t ½oxidant dt. (5) CT ¼ 0

Then, with the CT value and the rate constants, the concentration of toxin can be calculated from Eq. (6), where k is the second-order rate constant for the reaction of the oxidant with the toxin under consideration at the pH of the water. ½toxin ¼ ½toxin0 expðkCTÞ.

(6)

Following this procedure, the removal of MC-LR, CYN or ANTX during the oxidation experiments was calculated. The results are included in Fig. 6, 8, 9 and 10 (continuous lines). Good agreement between the experimental and predicted values was observed, which validates our kinetic model. Therefore, this kinetic model can be applied to predict the elimination of any other toxin during permanganate or chlorine application to drinking-water production if the rate constants for its reactions with the oxidants are known.

4.

Conclusions

According to the results of this study, permanganate is a feasible option for the elimination of ANTX and MC-LR, while chlorine is a possible oxidant for the oxidation of CYN and MC-LR, and ozone can effectively oxidize all three toxins. Moreover, the chlorine dose applied to Lake Greifensee water

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must be below 3 mg/L to avoid a TTHM concentration in the treated water above the standard value of 100 mg/L. In the presence of bromide (50 mg/L), bromate formation due to ozonation is far below the 10 mg/L bromate drinking-water standard. Bromate formation may be an issue at higher bromide levels in order to ensure complete ANTX oxidation by ozone. Overall, static oxidant dose experiments in water from Lake Greifensee were consistent with reactivities displayed in the kinetic database. Furthermore, in dynamic time-resolved experiments, a kinetic model was successfully applied to predict and confirm toxin oxidation with a given oxidant consumption curve.

Acknowledgments This study was performed within the framework of ‘‘TOXIC: Barriers Against Cyanotoxins in Drinking Water,’’ European Union Project EVK1-CT-2002-00107. Financial support for GDO was provided by the Swiss Federal Department for Education and Science (Bundesamt fu¨r Bildung und Wissenschaft). The cyanotoxins were kindly provided by the laboratories of Jussi ˚ bo Akademi University (AAU), and Geoffrey Codd, Meriluoto, A University of Dundee (UD). MC-LR was purified by Lisa Spoof, AAU. CYN and ANTX were purified by James Metcalf and Marianne Reilly, UD. We appreciate the valuable technical advice and laboratory assistance provided by Michael Dodd, Elisabeth Salhi, Marc Huber, Marc-Olivier Buffle and Manuel Sanchez-Polo of EAWAG; and by Ana Sordo and Maria E. Majado of UEx. ER and GDO contributed equally to this work.

Appendix A.

Supporting Information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.watres.2007.03.033.

R E F E R E N C E S

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