The use of ozone and associated oxidation processes in drinking water treatment

PII: S0043-1354(98)00130-4

Wat. Res. Vol. 32, No. 11, pp. 3208±3222, 1998 # 1998 Elsevier Science Ltd. All rights reserved Printed in Great Britain 0043-1354/98 $19.00 + 0.00

REVIEW PAPER THE USE OF OZONE AND ASSOCIATED OXIDATION PROCESSES IN DRINKING WATER TREATMENT M V. CAMEL** and A. BERMOND

Laboratoire de Chimie Analytique, Institut National Agronomique Paris±Grignon, 16 rue Claude Bernard, 75231 Paris Cedex 05 France (First received April 1997; accepted February 1998) AbstractÐThis paper summarizes the main applications of ozonation and associated oxidation processes in the treatment of natural waters (surface and ground waters) for drinking water production. In fact, oxidants may be added at several points throughout the treatment: pre-oxidation, intermediate oxidation or ®nal disinfection. So, the numerous e€ects of chemical oxidation are discussed along the water treatment: removal of inorganic species, aid to the coagulation-¯oculation process, degradation of organic matter and disinfection. Of prime importance in potable water production is the removal of organic matter (natural humic substances, as well as micropollutants, especially pesticides) to avoid degradation of the distributed water (mainly bad odors and tastes; formation of disinfection by-products such as trihalomethanes; microbial regrowth in the distribution system). Consequently, this point has been particularly detailed in this paper. As a matter of fact, complete mineralization hardly occurs during the process; as a consequence, further treatment (i.e. sand or granular activated carbon ®ltration) is required to improve the distributed water quality, and to meet the drinking water regulations. # 1998 Elsevier Science Ltd. All rights reserved Key wordsÐozone, hydroxy, radical, advanced oxidation processes, drinking water, oxidation by-products, trihalomethanes, bromate

INTRODUCTION

NOMENCLATURE A=

speci®c coecient of lethality of the microorganisms (mg lÿ1 minÿ1) AOP= advanced oxidation processes AOXFP= adsorbable organic halogen formation potential BAC= biological activated carbon BDOC= biodegradable dissolved organic carbon c= ozone dose in water treatment d= dilution factor 2,4-D= 2,4-dichlorophenoxyacetic acid DBP= disinfection by-products DOC= dissolved organic carbon GAC= granular activated carbon kinetic rate constant for the reaction between O3 and kO3= M (Mÿ1 sÿ1) kinetic rate constant for the reaction between OHÇ and kOHÇ = M (Mÿ1 sÿ1) M= micropollutant MCPA= 4-chloro-2-methylphenoxyacetic acid N= number of microorganisms at time t initial number of microorganisms N0= PAH= polynuclear aromatic hydrocarbons PCB= polychlorinated biphenyls 2,4,5-T= 2,4,5-trichlorophenoxyacetic acid Ti= titane THM= trihalomethane THMFP= trihalomethane formation potential TOC= total organic carbon t= contact time UV= ultraviolet radiations

*Author to whom all correspondence should be addressed. [Fax: 331-44081653].

The production of drinking water from natural waters necessitates the removal of numerous compounds present (mainly inorganic species, humic substances, toxic micropollutants). Due to its high oxidation potential, ozone has been widely used during the past few years (Rice and Netzer, 1982; DoreÂ, 1989; Masschelein, 1991; Langlais et al., 1991). However, due to the refractory character of some pollutants or the formation of by-products that are not further oxidized (Rice and Cotruvo, 1978), a more thorough oxidation is generally preferable, by means of the generation of hydroxyl radicals (using advanced oxidation processes (AOPs)). The main possible points of chemical oxidant introduction are pre-oxidation, intermediate oxidation and ®nal disinfection. Usually, pre-oxidation leads to the elimination of mineral compounds, color, turbidity and suspended solids, bad tastes an odors; in addition this step partly degrades natural organic matter and inactivates microorganisms; ®nally, this treatment generally enhances the coagulation-¯oculation-decantation step. Intermediate oxidation is intended to degrade toxic micropollutants, remove trihalomethanes precursors, and

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The use of ozone in drinking water treatment

3209

Table 1. Reported treatments of natural waters with ozone or associated oxidation processes, and their resulting e€ects Treatment Pre-oxidation

Water

Oxidant

Resulting e€ects

References

surface waters

O3

THMFP increase or decrease formation of aldehydes and carboxylic acids decrease of THMFP GAC ®ltration improvement, colour removal, low DOC removal TOC and DOC removal, ¯oculation improvement aldehydes formation BDOC formation aid in coagulation-¯oculation slight reduction in THMs, brominated THMs formation iron removal metals (Fe, Mn, Pb, Cu, As, Cd, Zn) removal tastes and odors removal, TOC reduction pesticides degradation control of DBPs formation of aldehydes THMs precursors removal bromate formation

Rice, 1980 Van Hoof et al., 1986

surface waters (brominated) ground waters

O3/H2O2 O3 O3 O3/H2O2

surface and ground waters

Intermediate oxidation

surface and ground waters (brominated) surface waters

O3 O3/H2O2 O3 O3

O3/H2O2 O3, O3/H2O2 surface waters (brominated)

Disinfection

O3

color reduction (57%), turbidity decrease (40%), AOXFP decrease (48%), THMFP decrease, biodegradability increase color removal, pesticides oxidation, THMFP reduction BDOC formation

Graham et al., 1994 Odegaard et al., 1986 Casellas et al., 1983 Schechter and Singer, 1995 Albidress et al., 1995 Paode et al., 1995 Amy et al., 1991 Cromley and O'Connor, 1976 Nieminski and Evans, 1995 Martin, 1993 Paillard et al., 1992 Shukairy and Summers, 1992 Glaze et al., 1989 Duguet et al., 1985 Lefebvre et al., 1995 Kainulainen et al., 1995

Hart et al., 1995

Lefebvre and CroueÂ, 1994 Lefebvre and CroueÂ, 1995 BDOC increase, formation of oxalate, Prevost et al., 1995 aldehydes and pyruvic acid BDOC formation, THMFP reduction, Wricke et al., 1995 bromate formation pesticides degradation Paillard et al., 1990 Prados et al., 1995a pesticides degradation Meijers et al., 1995a,b Lambert et al., 1996 bromoform and bromate formation, Baumgardt et al., 1995 THMFP reduction bromate formation Kruithof et al., 1995 chlorinated hydrocarbons removal SchwaÈmmlein and Leitzke, 1995 formation of aldehydes, aldo and keto Jammes et al., 1994, 1995 acids bromate formation Lefebvre et al., 1995

O3, O3/H2O2 ground water O3/UV, O3/H2O2 surface and ground O3, O3/H2O2 waters surface and ground O3 waters (brominated bromate formation O3, O3/H2O2 surface waters O3/H2O2 inactivation of Giardia muris cysts O3 inactivation of Cryptosporidium oocysts

increase biodegradability (for complete removal of organics upon subsequent sand or granular activated carbon (GAC) ®ltration). Thereafter, ®nal disinfection should eliminate all the remaining microorganisms with subsequent minimization of disinfection by-products formation. Table 1 illustrates the possible e€ects of ozone and associated processes (i.e. O3/H2O2 and O3/UV) in drinking water treatment. In the present paper the di€erent objectives of chemical oxidation will be discussed, namely: (1) removal of inorganic species, (2) aid to the coagulation-¯oculation process, (3) oxidation of natural organic matter, (4) oxidation of micropollutants and (5) disinfection.

Legube et al., 1995 Wolfe et al., 1989 Reading and Bell, 1995

REMOVAL OF INORGANIC SPECIES

Inorganic species are most of the time eliminated using pre-oxidation (Rice and Gomez-Taylor, 1986). However, pre-ozonation must be followed by a ®ltration or a coagulation-¯oculation-decantation step. In that way, metallic ions can be removed, as they form insoluble species upon oxidation (Bourbigot, 1983; Nieminski and Evans, 1995). Similarly, ammoniacal nitrogen is slowly oxidized into nitrate ion by ozone; consequently, it should be eliminated through a subsequent biological nitri®cation (on a sand or granular activated carbon ®lter). Yet, in presence of bromide, ammonia can be decomposed to nitrogen gas by ozonation (Haag et

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V. Camel and A. Bermond

al., 1984; Somiya et al., 1995; Yang et al., 1997). In this process, Brÿ is rapidly oxidized to HOBr by ozone, which further reacts with ammonia, leading to the formation of N2 and Brÿ; bromide can be oxidized again by ozone, thereby allowing rapid elimination of ammonia. The behaviour of bromide in water upon pre-ozonation is preoccupant as it leads to the formation of hypobromous acid/hypobromite (HOBr/BrOÿ), and bromate for longer oxidation times (Lefebvre et al., 1995), a potential carcinogen (Kurokawa et al., 1990; Miller, 1993). As a consequence, minimization of its production must be achieved (the temporary limit ®xed by the World Health Organization is currently 25 mg Lÿ1 in potable water). This may be realized by carefully optimizing the ozonation conditions that in¯uence the formation of bromate: ozone dosage (noted c); contact time (noted t), due to competition reactions between oxidation of organics and bromide; c
t value; bromide concentration; pH, as the formation of bromate depends on the hypobromite ions available in the medium; temperature (Kruithof and Schippers, 1992; Koudjonou et al., 1994; Von Gunten and HoigneÂ, 1994; Kruithof et al., 1995; Legube et al., 1995; Croue et al., 1995, 1996; Song et al., 1996; Ozekin and Amy, 1997). Also, the addition of ammonia reduces the oxidation of hypobromite, as it reacts with hypobromous acid to form monobromamines (Haag and HoigneÂ, 1983; Koudjonou et al., 1994; Von Gunten and HoigneÂ, 1994; Joost et al., 1995; Siddiqui and Amy, 1995; Kozasa et al., 1997). Finally, the use of AOPs may be another viable strategy to limit the formation of bromate, due to either the consumption of ozone by OHÇ radicals (Miller, 1993; Koudjonou et al., 1994; Grini and Iozzelli, 1996), or the reduction of hypobromous acid/hypobromite in presence of H2O2 (leading back to Brÿ) (Von Gunten and Oliveras, 1997). Thus, bromate formation can be restricted by application of high H2O2/O3 ratios (Kruithof et al., 1997). Very recently, the addition of an heterogeneous catalyst (TiO2) to ozone also minimized the formation of bromate from brominated surface waters (possibly due to either a faster reaction between molecular ozone and natural organic matter in presence of the catalyst (limiting the oxidation of bromide), or the reaction of molecular ozone on the surface catalyst (leading to OHÇ that are less ecient to oxidize bromide)) (Ciba et al., 1995). Finally, limiting the presence of bromate may also be achieved by adding a carbon adsorption step after the oxidation (Joost et al., 1995; Siddiqui et al., 1995). AID TO THE COAGULATION-FLOCULATION PROCESS

A low ozone dose (0.5 to 1 g mÿ3) is sucient to enhance the coagulation-¯oculation process (Bourbigot, 1983). Several explanations have been

proposed: the formation of smaller molecules with a higher hydrophobic anity, the formation of oxygenated functional groups (such as carboxylic acids) or a polymerization e€ect (DoreÂ, 1990). In fact, a few studies gave evidence of the possible induced polymerization upon oxidation. Thus, insoluble polar polymers (hexamer size) were formed during the ozonation of 2,4-dichlorophenol under the conditions of potable water treatment (Duguet et al., 1986a). Other studies suggested that polymerization occurred during ozonation of simple phenolic compounds (Chrostowski et al., 1983) or humic substances (Farvardin and Collins, 1989). On the opposite, some results found pre-ozonation to be detrimental to the coagulation. Thus, Becker and O'Melia (1995, 1996) noted a signi®cant reduction in the molecular weight distribution upon ozonation of model waters; in addition, an increased coagulant dose was required to allow optimum turbidity and total organic carbon (TOC) removal. As a matter of fact, an optimum ozone dose is required for polymerization to be induced, dosages above this value being detrimental to the coagulation by oxidizing the polymers formed into smaller molecules. So, proper ozonation conditions result in a TOC and dissolved organic carbon (DOC) decrease, as well as in ¯oculation enhancement (Casellas et al., 1983; Prados et al., 1995a); addition of hydrogen peroxide was found to decrease the optimum ozone dose required (Paode et al., 1995; Prados et al., 1995a). Very recently, it was observed that ozone alone could not induce destabilization of particles coated with natural organic matter; this e€ect could only be obtained in presence of a complexing agent (e.g. calcium) in the water (Chandrakanth and Amy, 1996). The sole mechanism responsible for the bene®cial e€ect of pre-ozonation was found to be an enhanced association of calcium with ozonated natural organic matter (due to the production of more ligand sites upon ozonation) as well as with the by-products formed (such as oxalic acid); this precluded the adsorption of organic matter onto particles, thereby leading to a reduction in particle stability through surface charge reduction.

OXIDATION OF NATURAL ORGANIC MATTER

Surface and ground waters contain substantial organic matter that may be detrimental to the water quality (i.e. color and odour of water). Also, humic substances are known to form trihalomethanes (THM) upon ®nal chlorination. In addition, presence of natural organic matter in the distributed water favours the bacterial regrowth in the network, which may cause sanitary problems. Consequently, as far as possible, natural organic matter should be removed during the treatment, mainly through chemical oxidation.

The use of ozone in drinking water treatment

Removal of color and UV-absorbance Most of the time, ozonation of humic substances leads to a quick decolorization (Killops, 1986; Lawrence et al., 1980; Gilbert, 1988) and a decrease in UV-absorbance due to a loss of aromaticity (Anderson et al., 1986; Legube et al., 1987a; Gilbert, 1988). This could be attributed to a depolymerization of the humic materials, as observed in case of natural waters containing high levels of humic substances; simultaneously, phenolic and acidic compounds were detected in the medium, as a result of the depolymerization (Mallevialle, 1975; Rice, 1980). Later, Anderson et al. (1986) gave evidence of the formation of low molecular weight compounds, suggesting that ozone initially oxidized the most reactive sites of humic acids. These observations were supported by the work of Xiong et al. (1992) who ozonated several aqueous fulvic acid solutions. They showed that, for a constant ozone dose, some speci®c sites on the fulvic acid were available for ozone (aromatic moieties or conjugated double bonds); at higher ozone doses, additional sites became available. The initial stage was mainly due to a molecular attack of ozone, while radical chain reactions took place as the reaction lasted (even at pH 2.6); as the pH was increased, the radical reactions occurred earlier in the oxidation. Glyoxalic acid and hydrogen peroxide were identi®ed as fulvic acid by-products, and were suspected of being the precursors of the radical chain reactions that took place as the oxidation lasted. Reduction in TOC and DOC Upon ozonation of humic substances, the TOC was either reduced (Mallevialle, 1975; Anderson et al., 1986; Yamada et al., 1986; Kusakabe et al., 1990), or unchanged (Killops, 1986). In fact, humic materials of di€erent origins may react di€erently with ozone (Gilbert, 1988). Yet, whatever the result on TOC and DOC, ozonation of humic substances leads to the formation of small molecules, mainly aldehydes (formaldehyde, acetaldehyde, glyoxal, methylglyoxal) and carboxylic acids (formic, acetic, oxalic, glyoxylic, pyruvic and ketomalonic acids), which accumulate in the solution due to their resistance towards ozone (Arai et al., 1986; Glaze et al., 1989; Kusakabe et al., 1990; Jammes et al., 1994, 1995; Garcia-Araya et al., 1995; Schechter and Singer, 1995; Takahashi et al., 1995; Westerho€ et al., 1995). These results are of health concern, since formaldehyde has shown evidence of mutagenicity and carcinogenicity, and other by-products may present similar properties. Additional health hazards arise from the presence of bromide, as ozonation of natural organic matter in water containing high levels of bromide leads to the formation of brominated compounds (i.e. bromohydrins, such as 3-bromo-2-methyl-2-butanol) (Cavanagh et al., 1992).

3211

In order to achieve a complete mineralization of natural organic matter, advanced oxidation processes have been investigated. Thus, the Sonozone process (i.e. O3/ultrasounds) allowed good TOC removal (90%) of synthetic fulvic acid solutions (Olson and Barbier, 1994). However, in particular cases (e.g. raw waters containing carbonates) the water should be pretreated before oxidation, otherwise the TOC removal would be inhibited (due to the scavenging e€ect of carbonate ions). Gracia et al. (1995, 1996) noted a better TOC removal during ozonation of humic substances in presence of a metal catalyst (best results being obtained with Mn(II) and Ag(I)) which should initiate the ozone decomposition. However, mineralization was uncomplete, as they identi®ed more than a hundred by-products (phthalates and dicarboxylic acids). Presence of a solid catalyst composed of TiO2 (the Catazone process) was also ecient in reducing the DOC of a synthetic solution of fulvic acids (Volk et al., 1997). Increase in biodegradability Ozonation of natural waters leads to low molecular weight compounds which should be better adsorbed onto activated carbon (Odegaard et al., 1986); however, the polarity of organic compounds is increased by ozonation, which results in a decrease of the adsorbability on GAC (De Laat et al., 1991; Joost et al., 1995). Yet, the O3/GAC combination has prooved to be very ecient in reducing DOC before the ®nal disinfection step because of a biodegradability enhancement upon ozonation (Lawrence et al., 1980; Rice, 1980; Yamada et al., 1986; Van Leeuwen, 1987; Gilbert, 1988; Lefebvre and CroueÂ, 1994, 1995; Takahashi et al., 1995). Indeed, it was observed that low molecular weight aldehydes and carboxylic acids, formed upon ozonation, could be easily removed using a medium that allowed the development of biological activity (i.e. GAC or slow sand ®lters). For example, Jammes et al. (1994) investigated water in various treatment plants that include an ozonation step, which is followed in some of them by a GAC ®ltration. They noted an increase in aldehydes content after the ozonation step, most of them (60± 75%) being removed upon subsequent GAC ®ltration. Other studies gave evidence of a higher biological activity across a GAC ®lter receiving ozonated humic or fulvic solutions than in another one receiving non-ozonated solutions (De Laat et al., 1991; Kainulainen et al., 1995). Finally, the O3/ GAC system outperformed the GAC system as the biodegradability increase largely compensated for the loss of DOC adsorbability on GAC. In addition, the O3/GAC process (also called the biological activated carbon (BAC) process) leads to a biologically stable water. By studying the long-term performances (up to two years) of a GAC ®lter, Dussert and Kovacic (1995) reported the GAC ®lter

3212

V. Camel and A. Bermond

to be only slightly a€ected by residual ozone, thereby con®rming the great potential of this process in potable water treatment. Recent results suggested that AOPs (e.g. O3/H2O2 and O3/g-irradiation) lead to a more important biodegradability increase than ozone alone (Karpel Vel Leitner et al., 1997). Another process, that combines ozonation, adsorption on powdered activated carbon and ultra®ltration, has recently been reported: the Cristal process (Baudin et al., 1995). It was ecient in removing tastes and odors, limiting the formation of disinfection by-products as well as the biodegradability. Reduction in trihalomethane formation potential (THMFP) To date, trihalomethanes constitute the major disinfection by-products known to be formed upon ®nal chlorination (mainly chloroform, bromodichloromethane, dibromochloromethane and bromoform). Fulvic and humic acids are known to be the predominant haloform precursors in raw waters as their oxidation by-products lead to the formation of THMs upon chlorination (Rook, 1976; Oliver and Lawrence, 1979). In addition, ozonation of aqueous fulvic acid solutions containing bromide results in the formation of brominated compounds upon ®nal chlorination (Xie and Reckhow, 1993). As chemical oxidation can not remove THMs once they are formed, their precursors must be degraded before ®nal disinfection. Several studies report possible reduction in THMFP upon ozonation due to degradation of humic substances into low molecular weight compounds that are less reactive towards chlorine (Amy et al., 1986; Amy et al., 1991; Graham et al., 1994). However, at the same time, bromide present in water was oxidized into hyprobromite which further led to brominated compounds (Graham et al., 1994). In fact, the use of ozone alone may not be a universal process to remove all these precursors. AOPs, such as O3/UV (Prengle, 1983) and O3/H2O2 (Duguet et al., 1985, 1986a) are usually more ecient. Nevertheless, chemical oxidation may have a two-fold action (Dore et al., 1978): on one hand destroying the THMs precursors and, on the other hand, forming new THMs precursors. As a consequence, the production of THMs is strongly dependent on the degree of oxidation; in addition, it is also correlated with the pH of chlorination (Reckhow et al., 1986). Release of entrapped compounds The structure of fulvic and humic acids (phenolic and benzenecarboxylic acids that are hydrogen bonded into a molecular sieve matrix) contains numerous holes that can retain or ®x organic molecules (Schnitzer and Kahn, 1972; Lawrence et al., 1980; Rice, 1980; Ebenga et al., 1986). As a matter

of fact, it is now well established that humic substances interact with metals and pesticides; besides, they may contain volatile aromatic compounds entrapped in the polymeric network. Thus, these compounds might be released during chemical oxidation, thereby resulting in TOC or/and enhanced toxicity. Thus, liberation of Fe, Mn (Mallevialle, 1975), volatiles (Killops, 1986), phthalates and fatty acid esters (Lawrence et al., 1980) were observed. Conclusions It appears that oxidation of natural organic matter inevitably leads to oxidation by-products, as well as possible release of entrapped compounds. For that reason, it seems preferable to combine chemical oxidation with GAC ®ltration, to minimize the presence of these undesirable compounds in the distributed water. Besides, this process should avoid the formation of THMs upon ®nal disinfection (by removing the THM precursors). OXIDATION OF MICROPOLLUTANTS

In raw waters, the total organic carbon is mainly constituted of humic substances. Yet, micropollutants may also be present, especially in surface waters. Unfortunately, their degradation by chemical oxidation is always a complicated process, as its eciency depends on the nature of the micropollutants as well as on the quality of the water (i.e. presence of natural organic matter or carbonates). In systems using both molecular ozone and OHÇ oxidation (such as the O3/H2O2 combination), the overall degradation of a micropollutant M may be described by the following kinetics: ÿd[M]/ dt = kO3
[O3]
[M] + kOHÇ
[OHÇ ]
[M], where kO3 and kOHÇ are the kinetic constants between M and molecular ozone or radicals, respectively. These constants have been evaluated for numerous organic micropollutants. While kO3 values are generally low (around 5±100 Mÿ1 sÿ1), kOHÇ are several orders higher (in the range 107±1010 Mÿ1 sÿ1), due to the higher oxidative power of OHÇ . Consequently, in the above equation, the ®rst term may be often neglected. Organic micropollutants that are frequently encountered in natural waters have been largely studied. We report in this article results for the most common compounds. Chlorobenzenes Oxidation of chlorobenzenes by molecular ozone is rather slow at pH 2 (kO3=0.06±3 Mÿ1 sÿ1) (Hoigne and Bader, 1983a; Yao and Haag, 1991). As illustrated for 1,3,5-trichlorobenzene, reaction rates increase at higher pH due to OHÇ formation (Masten et al., 1997). However, it is highly recommended, in practice, to use AOPs systems, such as O3/H2O2, for their ecient degradation (Cortes et al., 1996). In that way, kinetic rate constants

The use of ozone in drinking water treatment

average 4±5  109 Mÿ1 sÿ1 (Haag and Yao, 1992; Kochany and Bolton, 1992). Polychlorinated biphenyls (PCBs) Polychlorinated biphenyls exhibit very low reaction rate constants with molecular ozone (kO3<0.9 Mÿ1 sÿ1), due to the inactivation of the aromatic rings by the chlorine substituents (Yao and Haag, 1991). Besides, they lead to ring-cleavage by-products that may be of health signi®cance since they still contain chlorine atoms (Glaze, 1986). Fortunately, these compounds are more reactive with hydroxyl radicals (kOHÇ = 4.3±8  109 Mÿ1 sÿ1) (Sedlak and Andren, 1991; Haag and Yao, 1992). Oxidation reaction proceeds through hydroxylation at one of the nonhalogenated sites, the more chlorinated compounds having the lower reactivity. Polynuclear aromatic hydrocarbons (PAHs) Trapido et al. (1995) recently investigated the ozonation of seven PAHs. All compounds were e€ectively degraded by ozone (pH 3±6.5) in a few minutes, with the following reactivity order: benzo[a]pyrene (5.3  104 Mÿ1 sÿ1)>pyrene (3.6  4 ÿ1 ÿ1 10 M s )>anthracene (2.7  104 Mÿ1 sÿ1)> phenanthrene (1.0 104 Mÿ1 sÿ1) r ¯uoranthene (9.5  103 Mÿ1 sÿ1)>benzo[ghi]perylene (8.4  103 Mÿ1 sÿ1)>¯uorene (4.2  103 Mÿ1 sÿ1). Due to a higher reaction rate in neutral and acidic media than in basic ones, and a slower degradation using advanced oxidation processes, they concluded that the reaction proceeds mostly by molecular ozone. These results are consistent with previous constants evaluated for naphthalene: 550 Mÿ1 sÿ1 at pH 5.6 (Legube et al., 1986a,b), 3000 Mÿ1 sÿ1 at pH 2 (Hoigne and Bader, 1983a). Similarly, Corless et al. (1990) observed complete destruction of pyrene by ozone (after only 1 min) in experimental conditions similar to those encountered in drinking water treatment. Consequently, ozonation appears as the best process to eciently degrade PAHs. Pesticides Ozonation of pesticides was reviewed few years ago (Reynolds et al., 1989). So, the main points are exposed below, along with recent reported results. Most of the time, the removal of pesticides by ozonation may be related to their water solubility (Shimazaki et al., 1995). However, faster degradations are usually obtained with AOPs, such as O3/H2O2 (Prados et al., 1995a) or O3/UV (Gahr and Niessner, 1995). Organochlorinated pesticides. Molecular ozone is rather unreactive towards pesticides that contain several chlorine atoms or sterically hindered unsaturations (kO3<0.04 Mÿ1 sÿ1 for lindane, endrin and chlordane) (Yao and Haag, 1991). Presence of accessible unsaturated cycles on the molecule leads to more reactivity (kO3=270 Mÿ1 sÿ1 for methoxychlor) (Yao and Haag, 1991). However kinetic con-

3213

stants remain unsucient for an e€ective degradation using ozone alone. Radicals have to be involved (kOHÇ =2.7±170  108 Mÿ1 sÿ1 for lindane, endrin and chlordane; 2  1010 Mÿ1 sÿ1 for methoxychlor) (Haag and Yao, 1992). Hence, lindane could be partially degraded using O3/H2O2; the identi®ed by-products suggested dehydrochlorination as the primary reaction (Prados et al., 1995b). Yet, complete degradation could not be achieved, as well as for endosulfan; thus a complementary treatment should be used to remove these compounds (mainly GAC ®ltration). It must be noted that pentachlorophenol is much more reactive towards ozone (kO3>>3  105 Mÿ1 sÿ1) (Hoigne and Bader, 1983b). Organophosphorus pesticides. Organophosphorus pesticides are easier degraded upon ozonation than organochlorine compounds. As an example, Roche and Prados (1995) reported ecient removal of malathion (>99%) and methyl-parathion (91.2%) upon ozonation at pH 8.3. Similar results were observed for dimethoate and diazinon (Meijers et al., 1995b). Very recently, the O3/H2O2 combination has been found to almost completely degrade (>95%) monocrotophos within 20 min (Ku and Wang, 1997). Carbamates. Mason et al. (1990) investigated the removal of four carbamates (aldicarb, methomyl, carbaryl and propoxur) by ozone during the disinfection step. A faster degradation was observed for the oxime carbamates (aldicarb and methomyl) as compared to the aromatic carbamates. Aldicarbsulfoxide was identi®ed as one of the by-products of thio-carbamates. The mechanism involved could be an electrophilic attack of ozone on the sulfur atom. These results are consistent with further work of Yao and Haag (1991) who determined the rate constant for the reaction of ozone and aldicarb (kO3=4.4  104 Mÿ1 sÿ1) and proposed the sulfur atom to be the main reaction site. On the other hand, for oxamyl, the rate constant is three orders of magnitude lower (kO3=620 Mÿ1 sÿ1) (the sulfur lone pair is less available for reaction than in aldicarb, as it is conjugated with the carbon-nitrogen double bond). Naturally, these compounds react rapidly with OHÇ (kOHÇ =2  109 and 8.1  109 Mÿ1 sÿ1 for oxamyl and aldicarb, respectively) (Haag and Yao, 1992). s-Triazines. The oxidation of s-triazines by molecular ozone is rather slow, with the following reactivity order: simazine (11.9 Mÿ1 sÿ1)>terbutylazine (8.9 Mÿ1 sÿ1)>atrazine (7.9 Mÿ1 sÿ1); at pH 1, no reaction took place, possibly due to the high level of protonation of triazines (Legube et al., 1987b; Brambilla et al., 1995). Ozonation by-products are N-dealkylated and acetamido-s-triazines; they are further degraded into complete N-dealkylated, deamined, dehalogenated and hydroxylated s-triazines, as illustrated in Fig. 1 (Adams and Randtke, 1992; De Laat et al., 1994a, 1995; Brambilla et al., 1995).

Fig. 1. Proposed pathways for the degradation of atrazine: (a) dealkylation, (b) side chain oxidation and (c) dechlorination.

3214 V. Camel and A. Bermond

The use of ozone in drinking water treatment

The degradation is greatly enhanced at elevated pH, in presence of H2O2, under UV-radiations, or in presence of catalytic amounts of Mn(II), due to the OHÇ attack (Benitez et al., 1995; Gahr and Niessner, 1995; Lai et al., 1995a,b; Meijers et al., 1995a,b; Nurizzo et al., 1995; Prados et al., 1995a,b; Roche and Prados, 1995; Lambert et al., 1996; Ma and Graham, 1997; Prados and Ciba, 1997). Other AOPs were also ecient: the Fenton's reagent (Arnold et al., 1995), H2O2/UV (Beltran et al., 1993, 1994a), Fe3+/UV (Larson et al., 1991). Yet, Paillard et al. (1990, 1992) reported O3/H2O2 to be the most ecient process (as compared with direct UV photolysis, H2O2/UV and O3/UV). The following reactivity order towards OHÇ was reported: ametryne (>2.6  1010 Mÿ1 sÿ1) 1 simetryne (>2.6  1010 Mÿ1 sÿ1)>simeton (4.7  109 Mÿ1 sÿ1)>atraton (3.3  109 Mÿ1 sÿ1)>simazine (3.1  109 Mÿ1 sÿ1)> terbutylazine (2.8  109 Mÿ1 sÿ1)>atrazine (2.4  109 Mÿ1 sÿ1) > cyanazine (1.9  109 Mÿ1 sÿ1) >propazine (1.8  109 Mÿ1 sÿ1) (De Laat et al., 1994b, 1995). Reactivity of by ÿ products formed towards OHÇ has also been evaluated: hydroxyatrazine (2.6±2.9  109 Mÿ1 sÿ1)>deisopropylatrazine (1.9±2.2  109 Mÿ1 sÿ1)>deethylatrazine (1.2.±1.6  109 Mÿ1 sÿ1)>deethyldeisopropylatrazine (<5± 6  107 Mÿ1 sÿ1)>cyanuric acid (<<2  107 Mÿ1 sÿ1) (Beltran et al., 1994a; Laplanche et al., 1994; De Laat et al., 1995). As a consequence of their low reactivity, the two latter products are refractory to radical oxidation; therefore they accumulate in solution. So, the s-triazine ring remains unaltered through all the oxidative processes. Substituted phenylureas. Phenylurea herbicides are, on the whole, well degraded by ozone alone (Meijers et al., 1995a,b; Roche and Prados, 1995). The following reactivity order has been reported: isoproturon (141 Mÿ1 sÿ1)>>chlortoluron (50.5 Mÿ1 sÿ1)>diuron (15.5 Mÿ1 sÿ1)>linuron ÿ1 ÿ1 (3.1 M s ) (Prados et al., 1995b; De Laat et al., 1996). This di€erent behavior should be attributed to the substituents nature on the aromatic ring (± Cl, ±CH3 or ±CH(CH3)2, the latter two increasing the reactivity, in agreement with an electrophilic attack of molecular ozone). On the contrary, these compounds present similar high reactivity towards OHÇ (4.3±5.2  109 Mÿ1 sÿ1) (De Laat et al., 1996); so, ecient removal of linuron and isoproturon could be achieved using AOPs (Allemane et al., 1994, 1995; Prados et al., 1995b; Roche and Prados, 1995). By-products identi®ed for isoproturon (phenylated and/or nitrated compounds) indicated that the initial attack of OHÇ occurred either on a C±N bond or a C±H bond. Further oxidation led to oxygenated compounds (alcohols and carbonyls). Acetamides. Acetamides are even less reactive towards molecular ozone than phenylureas, with the following reactivity order: alachlor (3.4±3.8 Mÿ1 sÿ1)1 metolachlor (3.0 Mÿ1 sÿ1)> ÿ1 ÿ1 propachlor (0.94 M s ) (Yao and Haag, 1991; De

3215

Laat et al., 1996). This reactivity order could be related to the presence of methyl substituents on the aromatic ring. On the opposite, similar behaviour for these pesticides towards OHÇ was noted (4.3± 7  109 Mÿ1 sÿ1) (Haag and Yao, 1992; De Laat et al., 1996). Phenoxycarboxylic acids. The most frequently encountered phenoxycarboxylic acids have shown to be degraded by ozone alone. Hence, Beltran et al. (1994b) performed mecoprop ozonation at several pH; for pH < 7, the direct pathway predominates (kO3=37.9 Mÿ1 sÿ1 at pH 2), while at pH 12 mecoprop is oxidized by radicals (kOHÇ =9.1  109 Mÿ1 sÿ1). Ozonation of MCPA (4chloro-2-methylphenoxyacetic acid) proceeds through the aromatic ring cleavage, leading to aliphatic acids (kO3=30±44 Mÿ1 sÿ1 depending on the pH) (Benitez et al., 1991). On the opposite, during ozonation under UV-radiations, rapid side-chain oxidation was observed (Benoit-Guyod et al., 1986). For 2,4-D and 2,4,5-T (2,4-dichloro- and 2,4,5-trichlorophenoxyacetic acids) the ozone reactive site is still the aromatic cycle, but lower reactivity than MCPA was noted (kO3=1±2.3 and 8.9 Mÿ1 sÿ1 for 2,4-D and 2,4,5-TP, respectively) (Yao and Haag, 1991). Ecient removal of these two pesticides was achieved in presence of OHÇ (kOHÇ =4± 5  109 Mÿ1 sÿ1) (Haag and Yao, 1992; Meijers et al., 1995a,b; Scheuer et al., 1995). Conclusions Due to the uncomplete mineralization of most micropollutants (especially pesticides) and the formation of by-products (most of the time of unknown toxicities), it is highly recommended to use chemical oxidation before a sand or GAC ®ltration, in order to remove the remaining organic compounds before ®nal disinfection, either by adsorption or biodegradation (probably both mechanisms are involved) (Paillard et al., 1990, 1992; Lambert et al., 1996). Thus, the combination of O3/ H2O2 and GAC ®ltration (the Ozocarb process) is used on site in several french plants (Paillard et al., 1992; Martin, 1993; Prados et al., 1995a). Another process has shown promising results: the addition of an heterogeneous catalyst (mainly TiO2) to O3 or O3/H2O2; it allowed a much better removal of oxalic acid (Paillard et al., 1991) and endosulfan (Prados et al., 1995a). This process involves oxidation reactions at the catalyst surface. Yet, at the present time, it is still at the stage of pilot studies. DISINFECTION

The ®rst experiments about the disinfection of water with ozone were carried out by De Meritens in 1886. Since then, there has been a great interest in the use of ozone as a viable alternative to chlorine in the disinfection of water, to eliminate the formation of trihalomethanes and organochlorine

3216

V. Camel and A. Bermond

compounds. Yet, oxidation products are known to appear, sometimes increasing the mutagenic activity of the water. Thus, Cooper et al. (1986) reported ozonation to lead to the possible formation of bromoform in water containing bromide, even without a postchlorination. Indeed, e€ects of ozonation depend on the experimental conditions (i.e. ozone dose and contact time), as mutagenicity variations are correlated to the degradation of the initial solutes present in water and the formation of byproducts (Duguet et al., 1983). Very often, a high ozone dose or/and a long contact time should lead to an advanced oxidation and a lower toxicity (Bourbigot et al., 1983; Kool and Hrubec, 1986). However, eciency of the disinfection step is strongly dependent on the pre-treatments (i.e. a good removal of suspended solids, which are good supports for microorganisms, and organic materials, as they consume oxidants and lead to biodegradable DOC which results in bacterial regrowth in the distribution network) (Glaze, 1987; Levi, 1995). Inactivation of microorganisms Ozone has reported to be e€ective in killing bacteria, viruses, as well as certain forms of algae (Glaze, 1987). So, its use as a disinfectant seems promising and, a few years ago, Drapeau and Paquin (1980) reviewed its possible e€ects. In fact, it is generally accepted that molecular ozone is a more e€ective biocide than hydroxyl radicals, since the latter are very short-lived and non-selective species. The resistance of microorganisms follows the increasing order: bacteria, viruses, and cysts (Duguet, 1995). For bacteria, Escherichia coli have undoubtly been the strains the most investigated (Bringmann, 1954; Holluta and Unger, 1954; Katzenelson et al., 1974; Burleson et al., 1975; Hunt and Marinas, 1997). Ishizaki et al. (1987) found that ozone penetrates the cell membrane and reacts with cytoplasmic substances. In addition, chromosomal desoxyribonucleic acids may be degraded, being one of the factors responsible of the cell killing. According to Mc Guire and Davis (1988), O3/H2O2 is as e€ective as ozone for removing Escherichia coli bacteria. Ozone was found e€ective to inactivate poliomylitis viruses (Coin et al., 1964, 1967; Suckhov, 1964; Gevaudan et al., 1971; Katzenelson et al., 1974; Majumdar et al., 1973). In fact, the mechanism of virus inactivation can be due to either the damage of the protein coat, or the direct damage of the nucleic acids. In case of tobacco mosaic virus, ozone attacked both protein coat and ribonucleic acids (Shinriki et al., 1988). As reported by Wickramanayake et al. (1984), ozone seems to be more ecient than free chlorine for cysts inactivation; however, most of the protozoan cysts are more resistant than bacteria and viruses to ozone. Labatiuk et al. (1992) studied fac-

tors a€ecting the inactivation of Giardia muris cysts using ozone. Cysts were rapidly inactivated and contact time beyond 2 min resulted in little additional inactivation. For surface waters, 5 min were required for sucient inactivation. Wolfe et al. (1989) reported O3/H2O2 to be also an ecient disinfectant (at a 0.2 ratio of H2O2/O3, it was e€ective against Giardia muris cysts). Kinetics of disinfection The kinetics of disinfection has been expressed by Chick (1908) and Watson (1908): dN/ dt = ÿ A
cd
N, where N represents the number of microorganisms at time t; N0 the initial number of microorganisms; t the contact time; A the speci®c coecient of lethality of the microorganisms (mg lÿ1 minÿ1), c the oxidant concentration and d the dilution factor. Speci®c coecient of lethality for several microorganisms have been reported with ozone as a disinfectant: bacteria 500 mg lÿ1 minÿ1; amibe cysts 0.5 mg lÿ1 minÿ1; viruses 5 mg lÿ1 minÿ1; spores 2 mg lÿ1 minÿ1 (Duguet et al., 1986b, 1987). In a well-stirred batch reactor, the above equation can be expressed as follows: ln(N/ N0) = ÿ A
cd t. According to this relation, ln(N/N0) vs time should give a straight line. Indeed experimental laws di€er somewhat (Boisdon, 1995). As an example, for bacteria and viruses, inactivation is fast initially, and then slows down; on the opposite, for protozoan cysts, a lag is initially observed before a linear phase (Wickramanayake et al., 1984). Anyway, this relation gives evidence of the in¯uent factors for disinfection: contact time, concentration as well as nature of the oxidant (O3, OHÇ ), kinetics of competitive reactions that may consume oxidant (e.g. with humic substances), nature of the microorganisms and their physical form, temperature. Hence, Fransolet (1980) pointed out that an ecient disinfection requires optimized ozone doses and contact times, while Duguet et al. (1986b, 1987) underlined the in¯uence of disinfectant concentration, contact time, pH, temperature, presence of organic matter and microorganisms physiology (they may form aggregates, thus being more resistant to oxidation). Naturally, the transfer of ozone from the gas to the liquid phase must also be ecient to ensure disinfection (Legeron and Perrot, 1981; Bourbigot, 1983). Temperature has opposing in¯uences on ozone stability and disinfection eciency: increased temperature reduces the solubility and stability of ozone in water, but increases the reaction rate between microorganisms and ozone. So, in case of protozoan cysts, elevated temperatures enhanced ozone eciency (Wickramanayake et al., 1984). Conclusions In practice, the (c
t) term is recommended as the valuable factor to ensure an ecient disinfection for ®ltered waters; nevertheless, as discussed by

The use of ozone in drinking water treatment

Labatiuk et al. (1992) this concept should be reexamined for raw waters. Very often, an ozone dose of 0.4 mg lÿ1 during 4 min is sucient for an e€ective disinfection (for a pretreated water, i.e. the organic matter has been removed) (Legeron and Perrot, 1981). However, as ozone rapidly decomposes in water, its life time in aqueous solutions is too short (less than one hour) to ensure that a residual ozone capacity will remain throughout all the distribution system. Consequently, ozone is ecient only in particular cases (mainly a recent and short distribution system); usually it is replaced by chlorine or chlorine dioxide in the ®nal disinfection step. This requires some precaution (Bourbigot, 1983; Miller, 1993): the injection point of the disinfectant must be far enough from the ozone oxidative treatment in order to allow residual ozone to completely disappear, thereby avoiding consumption of the disinfectant by ozone. CONCLUSION

In the past few years, ozonation and associated oxidation processes have been widely used in water treatment, due to their numerous advantages over chlorine. Indeed, most applications have been developed for drinking water treatment, as concerns about the potential re-use of wastewaters is quite recent. These processes o€er the opportunity of enhancing further treatments, such as coagulation¯oculation or granular activated carbon ®ltration. Moreover, they allow reduction of the trihalomethane formation potential, and degradation of micropollutants that were refractory to chlorine, due to a higher oxidative potential. Naturally, knowledge of both the reaction rate constants and the nature of the by-products are of prime importance to ensure an ecient process (i.e. degradation of the toxic pollutants initially present in the water, with subsequent minimization of the by-products formed). Yet, as pointed out, degradation is seldom complete and numerous compounds generally remain after oxidation; this may cause several problems such as a bacterial regrowth in the distribution network and/or an increase in mutagenic activity. Consequently, chemical oxidation should be followed by a biological step (either sand or GAC ®ltration) before the ®nal disinfection and the distribution. REFERENCES

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