Ozonation of Drinking Water Part II

8/11/2019 Ozonation of Drinking Water Part II

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Water Research 37 (2003) 14691487

Review

Ozonation of drinking water: Part II. Disinfection andby-product formation in presence of bromide,

iodide or chlorineUrs von Gunten*

Swiss Federal Institute for Environmental Science and Technology, EAWAG, Ueberlandstr 133, CH-8600 D .ubendorf, Switzerland

Abstract

Ozone is an excellent disinfectant and can even be used to inactivate microorganisms such as protozoa which are veryresistant to conventional disinfectants. Proper rate constants for the inactivation of microorganisms are only availablefor six species ( E. coli , Bacillus subtilis spores, Rotavirus , Giardia lamblia cysts, Giardia muris cysts, Cryptosporidium parvum oocysts). The apparent activation energy for the inactivation of bacteria is in the same order as most chemicalreactions (3550 kJ mol 1), whereas it is much higher for the inactivation of protozoa (80 kJ mol 1). This requiressignicantly higher ozone exposures at low temperatures to get a similar inactivation for protozoa. Even for theinactivation of resistant microorganisms, OH radicals only play a minor role. Numerous organic and inorganicozonation disinfection/oxidation by-products have been identied. The by-product of main concern is bromate, whichis formed in bromide-containing waters. A low drinking water standard of 10 mg l 1 has been set for bromate.Therefore, disinfection and oxidation processes have to be evaluated to full these criteria. In certain cases, whenbromide concentrations are above about 50 mg l 1, it may be necessary to use control measures to lower bromateformation (lowering of pH, ammonia addition). Iodate is the main by-product formed during ozonation of iodide-containing waters. The reactions involved are direct ozone oxidations. Iodate is considered non-problematic because itis transformed back to iodide endogenically. Chloride cannot be oxidized during ozonation processes under drinkingwater conditions. Chlorate is only formed if a preoxidation by chlorine and/or chlorine dioxide has occured.r 2002 Elsevier Science Ltd. All rights reserved.

Keywords: Ozone; Disinfection; Disinfection by-products; Bromate; Chlorate; Iodate

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1470

2. Disinfection . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14702.1. Action of ozone . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14702.2. Role of OH radicals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1472

3. Oxidation/disinfection by-products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14723.1. By-product formation in absence of bromide . . . . . . . . . . . . . . . . . . . . . 14723.2. By-product formation in the presence of bromide . . . . . . . . . . . . . . . . . . . 1473

3.2.1. Bromate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14733.2.2. Bromo-organic compounds . . . . . . . . . . . . . . . . . . . . . . . . . . 1480

*Corresponding author. Tel.: +41-1-823-52-70; fax: +41-1-823-52-10.E-mail address:

[email protected] (U. von Gunten).

0043-1354/03/$ - see front matter r 2002 Elsevier Science Ltd. All rights reserved.PII: S 0 0 4 3 - 1 3 5 4 ( 0 2 ) 0 0 4 5 8 - X


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1. Introduction

This is the second part of an article on drinking waterozonation processes dealing with disinfection andformation of disinfection by-products. The rst partfocussed on the oxidation kinetics and product forma-tion for the oxidation of inorganic and organiccompounds [1]. For a general introduction into ozona-

tion processes including all aspects of oxidation,disinfection and by-product formation see [1].Ozone is an excellent disinfectant and is able to

inactivate even more resistant pathogenic microorgan-isms such as protozoa (e.g. Cryptosporidium parvumoocysts) where conventional disinfectants (chlorine,chlorine dioxide) fail. However, the ozone exposurerequired to inactivate these microorganisms is quitehigh. This may lead to the formation of excessconcentrations of undesired disinfection by-products,in particular bromate, which is considered to be apotential human carcinogen. Bromate is particularlyproblematic because unlike many other organic by-products it is not biodegraded in biological lters whichusually follow an ozonation step.

The present paper discusses the available kinetic datafor the inactivation of microorganisms and how it can beapplied to assess disinfection processes that occur duringozonation. The role of OH radicals for disinfection willalso be assessed. Furthermore, kinetic and mechanisticinformation on the formation of disinfection/oxidationby-products with an emphasis on halogenates (bromate,iodate, chlorate) will be provided. In the case of bromate, an overview on bromate formation in water-treatment plants and possible options for minimization

and elimination after its formation are presented.

2. Disinfection

2.1. Action of ozone

Ozone has been applied for water disinfectionpurposes for almost a century. A recent survey inSwitzerland among water works that apply ozone hasshown that for 90% of these plants the main reason forits application is disinfection [2]. Because monitoring for

every pathogenic microorganism is not feasible, many

countries have adopted the concept of indicator micro-organisms [3]. For this purpose, fecal coliforms and E.coli were chosen because their presence indicates thatwater may be contaminated with human and/or animalwastes. Using indicator microorganisms to assess disin-fection works well as long as the inactivation of theundesired pathogens is at least as efcient as theinactivation of the chosen indicator microorganisms.

However, if the inactivation of indicator microorgan-isms is more efcient than the inactivation of pathogens,the absence of indicator microorganisms does notnecessarily mean that the water is safe to drink. Toovercome this problem, a process-oriented approach canbe based on an estimation of the disinfectant exposure ina disinfection reactor ( ct -concept). The disinfectantexposure ( ct) is calculated as the time-dependentconcentration of the disinfectant ( c f t) integratedfor the time t of its action. It has been found that thelogarithmic relative decrease of vital microorganisms isproportional to cnt; according to the ChickWatsonequation (see [4] and references therein).

log N =N 0 kcnt; 1

where N 0 is the number of microorganisms at time t 0;N the number of vital microorganisms at time t; k therate constant for the inactivation of a particularmicroorganism; c the concentration of a disinfectant; tthe contact time; n the tting parameter for non-rst-order behavior.

In many cases, n is equal to 1 and therefore theinactivation of the corresponding microorganisms is arst-order process [5,6]. Hence, the logarithmic decreaseof the relative inactivation of microorganisms is a

function of a pseudo-second-order inactivation rateconstant k and the ct-value. Typically, the assessmentof disinfection is based on published ct-values whichwere measured for an inactivation of a specic micro-organism by several orders of magnitude [7,8]. Becausethe integral ct-value is not easily accessible in non-ideallybehaving reactors during drinking water treatment, ctcalculations are usually performed with a conservativeapproach. The disinfectant concentration, c, is measuredat the outlet of the reactor and multiplied with thecontact time t 10 , corresponding to the time required for10% of an applied conservative tracer to travel through

the reactor. However, in ozonation this approach is very

3.3. By-product formation in iodide-containing waters . . . . . . . . . . . . . . . . . . . 14813.4. Chlorine-derived by-products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14813.5. Effect of ozonation on precursors for chlorination by-products . . . . . . . . . . . . 1482

4. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1483

Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1483

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1483

U. von Gunten / Water Research 37 (2003) 14691487 1470


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conservative and may lead to a signicant underestima-tion of the true integral ct in the reactor. This may leadto excessive ozonation and correspondingly high by-products formation (see below). An overdosage of ozone

is also not cost effective and should be avoided indrinking water treatment.One way to overcome this problem is a more dynamic

approach where reactor hydraulics are coupled withoxidation and disinfection kinetics [911]. With thisapproach it is possible to account for the real ozoneexposure which allows a more precise assessment of thedisinfection process.

While a large kinetic database for ozone reactions isavailable, proper kinetics of the inactivation of micro-organisms have been measured only in recent years.Table 1 gives the rate constants for ozone-baseddisinfection processes available to date. For microor-ganisms which are more difcult to inactivate (e.g.Bacillus subtilis spores ( B. subtilis ), Cryptosporidium parvum oocysts ( C. parvum )), a temperature-dependentlag-phase for disinfection is observed ( ct lag ) which has tobe taken into account [12,13,6] . In these cases, theinactivation can be modeled by a delayed ChickWatsonapproach described by Rennecker et al. [4]. The rateconstants for the inactivation of Giardia lamblia cystsgiven in the literature vary by a factor of two ( Table 1 ).This degree of variability has to be accounted for instudies of the inactivation of microorganisms and maybe due to differences in viability assays, different

ozonation protocols and differences in the strains [15].The data for C. parvum inactivation shown in Table 1

were determined by an in vitro excystation method [14].Inactivation data for C. parvum from animal infectivitydeviate somewhat from the data shown in Table 1 [16].A recent pilot-scale study in which the viability wastested with animal infectivity revealed a rate constant forthe inactivation of C. parvum similar to the value shownin Table 1 (0.86 l mg 1 min 1, [17]).

In Table 1 , the apparent activation energies for theinactivation of microorganisms are also shown. Theactivation energies for bacteria are in a similar range as

chemical reactions with ozone (3550kJ mol1

, [18]),

whereas for protozoa ( G. muris , C. parvum ) they aremuch higher ( E 80kJmol 1). No activation energies forG. lamblia cysts are available as yet. To assessdisinfection for variable temperatures, these activationenergies have to be compared to the activation energyfor ozone decay. The activation energy for the decom-position of ozone in various water matrices has beendetermined as 6570 kJ mol 1 [19]. This means that for

the same ozone dose, the degree of disinfection of bacteria and bacterial spores will be higher at lowertemperatures, while it will be smaller for protozoa. Thisis illustrated in Fig. 1 for an ozonation experiment of River Seine water with an ozone dosage of 2 mgl 1 atpH 8 and varying temperatures [20]. The results areshown for a complete depletion of ozone. The inactiva-tion of B. subtilis spores was calculated from the kineticsgiven by Driedger et al. [6], by including the temperaturedependence of the lag-phase and the rate constants. Themaximum inactivation is a result from the combinationof the temperature-dependent lag-phase and the inacti-

vation rate constant. The inactivation of C. parvum

Fig. 1. Calculated temperature dependence of inactivation of B. subtilis spores and Cryptosporidium parvum oocysts in River

Seine water after an ozone dosage of 2 mg l1

at pH 8. Waterquality: DOC 2.4 mgl 1, alkalinity 3.8mM, pH 8.

Table 1Kinetics of the inactivation of microorganisms with ozone at pH 7

Microorganism k O 3 (1mg1 min 1) ctlag (mgminl

1) k O3 (M1 s 1) E act (kJmol

1) T (1C) Refs.

E. coli 130 1.04 105 37 20 [5]B. subtilis spores 2.9 2.9 2.3 103 42 20 [6]

Rotavirus 76 6 104

20 [8]Giardia lamblia cysts 29 2.3 104 25 [135]

12a 9.5 103 a 22 [15]Giardia muris cysts 15.4 a 1.2 104 a 80a 25 [136]Cryptosporidium parvum oocysts 0.84 0.83 6.7 102 81 20 [14]

a Estimated value.



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oocysts was calculated in a similar way with thetemperature-dependent kinetics given by Driedger et al.[6]. The temperature-dependent inactivation differs forthe two microorganisms. Whereas for B. subtilis thegeneral trend shows a less efcient inactivation forincreasing temperatures, for C. parvum an increase intemperature leads to an increasing degree of inactiva-tion. This inverse behavior is a result of the differentactivation energies for the inactivation of these micro-organisms. The results shown in Fig. 1 also demonstratethat the use of B. subtilis spores as surrogate for thebehavior of C. parvum oocysts is linked with severalproblems and that this approach can only be applied inspecial cases. The inactivation of E. coli was not plottedin Fig. 1 because it would have been out of range bymany orders of magnitude due to its easier inactivation.This exemplies the problems associated with the use of E. coli to assess the efciency of disinfection processes.

2.2. Role of OH radicals

In the drinking water community, there is a frequentdebate on the role of OH radicals for disinfectionprocesses. While some authors state that ozone is themain disinfectant [2124] others suggest that OHradicals may play an important role for disinfection[25,26] . The kinetic data shown in Table 1 together withthe expected range of R c values (ratio of the concentra-tions of OH radicals and ozone Rc

d

OH =O3 10 62 10 9; [1]) allow an estimation of the role of OHradicals during disinfection. Taking both ozone and OHradicals into account for disinfection, the following rateequation can be formulated:

log N =N 0 k O3 k OH R cZ O3 d t; 2where k O3 is the rate constant for inactivation of microorganisms with ozone; k OH : rate constant forinactivation of microorganisms with OH radicals.

An R c of 108 is typical for the secondary phase of an

ozonation during which most of the disinfection occurs[19]. According to Eq. (2) the rate constant k OH forinactivation of C. parvum with OH radicals would have

to be 7 1010 M 1 s 1 for a similar disinfection ef-ciency as with ozone ( k O3 k OH R c). This rate constantis about a factor of 10 above rate constants typicallyobserved for OH radical reactions. Because the rateconstants for the inactivation of the other microorgan-isms in Table 1 by ozone are even higher than for C. parvum , k OH would have to be substantially higher. Inadvanced oxidation processes (AOP) where an Rc of 10 6 can be expected, OH radicals could theoreticallyplay a certain role for C. parvum oocysts and B. subtilisspores inactivation. For the other microorganismsshown in Table 1 , even in ozone-based AOPs, OH

radicals will have no effect on disinfection. Therefore,

the effect of OH radicals in disinfection processes cangenerally be neglected. In addition to these kineticcalculations, one also has to consider that the maintarget for inactivation of microorganisms is the DNAand not the cell wall. OH radicals would be scavenged inthe cell wall and their journey to the DNA would behampered by other cell constituents. The travelingdistance of OH radicals in a cell can be estimated tobe 69 nm [27].

3. Oxidation/disinfection by-products

Chemical disinfection processes in drinking watertreatment lead to the formation of disinfection by-products that are undesired due to their potentialchronic toxicity [28]. Typically, they result from theoxidation of matrix components in the water. There is

an ongoing debate on the relevance of disinfection by-products compared to the risks of waterborne diseases[29,30] . Recently, the risk of infection by C. parvum andthe development of renal cell cancer due to bromateformation in bromide-containing waters were comparedfor ozonation processes. It was concluded that thebenets of preventing gastroenteritis by proper disinfec-tion outweighs health losses by premature death fromrenal cell cancer by a factor of 10 [31]. In developingcountries, the emphasis on disinfection by-products canbe detrimental, as demonstrated by the cholera epidemicin Peru in the early 1990s [32].

A recent study proposed using the by-productformation during the disinfection process to assess thedisinfection efciency [33]. This concept is based on thefact that the concentration of by-products increases by aknown function with increasing exposure of a disin-fectant ( ct ). Because the logarithmic disinfection ef-ciency is also proportional to the disinfectant exposure(ct ), the measurement of the concentration of, e.g.,bromate allows us to calculate the extent of inactivationof a certain microorganism. However, because disinfec-tion efciency is measured by a logarithmic scalewhereas the build-up of by-products is measured on alinear scale, reactor hydraulics must be taken into

account for this approach.

3.1. By-product formation in absence of bromide

During ozonation, a variety of organic by-products isformed from the oxidation of natural organic matter(NOM). Aldehydes, ketones, keto aldehydes, carboxylicacids, keto acids, hydroxy acids, alcohols and estershave been reported ( [3438] and references therein). Formany of the polar non-volatile by-products there is stilla lack of analytical tools [39]. Up to now, there has beenno systematic study to assess the role of ozone and OH

radicals in the formation of organic by-products. Also,



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information on the kinetics of their formation duringozonation is scarce. Some data suggests that most of theorganic by-products are formed during the initial phaseof an ozonation process [40]. Most of the organic by-products are readily biodegradable and may lead to anincrease of the concentration of the assimilable organiccarbon (AOC) [34,26,4143] . Only a small fraction of the AOC (ca. 30%) could so far be explained by specicorganic compounds [39]. To avoid fouling in thedistribution system, a biological ltration step isadvisable after ozonation for waters containing a DOCconcentration of approximately >1 mg l 1 [43]. Biolo-gical lters after ozonation are routinely applied inmany drinking water treatment plants throughout theworld [4451]. Higher molecular weight oxidation by-products might not be readily reduced during biologicalltration. However, there is only limited informationabout these products. Bromate is not degraded under

typical oxic ltration conditions. Therefore, onceformed it will remain in the water all the way to thetap (see below).

3.2. By-product formation in the presence of bromide

In the presence of bromide (for typical bromide levelsin natural waters see below), many bromo-organic by-products such as bromoform, bromopicrin, dibromoa-cetonitrile, bromoacetone, bromoacetic acid, bromoalk-anes, bromohydrins, etc. have been identied ( [5254]and references therein). They are all formed by thereaction of hypobromous acid (product of the reactionbetween bromide and ozone) with NOM ( [55] seebelow). However, experience in many water worksshows that the concentrations of these bromoorganiccompounds are usually far below the current drinkingwater standards. This can be explained by a competitionkinetics approach [56]. Therefore, the main by-productof concern in the presence of bromide is bromate(BrO 3 ). It has been found that bromate is a genotoxiccarcinogen inducing, for example, renal cell tumors inrats [57]. The specic mechanism by which bromateproduces tumors is not known [58]. Bromate is thoughtto produce its toxic response through damage that

results from increased levels of lipid peroxides (LPO) or

from oxygen radicals that are generated from LPO andinduce DNA damage. Based on these ndings, bromatewas declared a potential human carcinogen whichrecently led to stringent drinking-water standards. TheWorld Health Organisation (WHO) issued a guidelinevalue of 25 mg l 1, both the European Union and theUSEPA established a maximum contaminant level of 10 mg l 1 [7,59,3] . Further toxicological studies have yetto show whether there is a threshold concentrationabove the implemented 10 mg l 1 BrO 3 which could beconsidered safe for drinking waters.

3.2.1. Bromate3.2.1.1. Occurrence of bromide and bromate in ozonation plants . Bromide levels in natural waters are highlyvariable in a range of 101000 mg l 1 (Table 2 , [60]).Both natural processes (salt water intrusion, water fromspecial geological formations) and anthropogenic activ-

ities (potassium mining, coal mining, chemical produc-tion, etc.) may contribute to increased bromide levels innatural waters. In many waters, low levels of bromide(o 20 mg l 1) are found which are unproblematic forbromine-derived by-products. For bromide levels in therange 50100 mg l 1, excessive bromate formation mayalready become a problem. However, in these waters,optimization and control options can be applied.Depending on the treatment goals, bromate formationcan become a serious problem for bromide levels above100 mg l 1.

Surveys on bromate formation in ozonation plants innumerous European countries [6164] and the USA [65]have been carried out for water works under standardtreatment conditions (grab samples). The disinfectiontarget may vary from one treatment plant to the other.Table 2 gives a summary of the ve studies performed sofar. In European and US American water works,bromate formation was generally below 10 mg l 1 underthese operation conditions. Only in about 6% of themore than 150 investigated water works, bromateconcentrations were above 10 mg l 1 . In some of thesecases, optimization of the ozonation can lead to asubstantial lowering of bromate formation. At the timeof these studies, the target of the disinfection was

inactivation of virus and/or G. lamblia . If a treatment

Table 2Bromate formation in full-scale ozonation plants in Europe and the USA

Country Number of plants Bromide range ( mgl 1) Bromate range ( mgl 1) Number of water works >10 mg l 1 a Refs.

France 10 12658 o 219 2 [61]France 32 o 20200 o 219.6 2 [63]Germany 4 30150 o 112 1 [62]Switzerland 86 o 550 o 0.520 2 [64]USA 24 2180 0.140 3 [65,137,138]

a

Once or more during survey.



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goal for inactivation of C. parvum oocysts wasimplemented, the number of plants with bromate levelsabove 10 mg l 1 could increase signicantly. Further full-scale tests will be necessary to draw rm conclusions.

3.2.1.2. Bromate formation during ozonation of bromide-containing waters . Bromate is formed in ozonationprocesses from the oxidation of bromide through acombination of ozone and OH radical reactions. Itsformation includes up to six oxidation states of bromine(Table 3 ). Because both oxidants can act simultaneouslyor in sequence on various oxidation levels ( Table 3 ), thewhole reaction system is extremely complicated andhighly non-linear. In the following, the various oxida-tion reactions are discussed in detail.

Ozone reactions . The oxidation of bromide duringozonation processes was rst investigated in detail byHaag and Hoign

!

e [55]. This study only considered direct

ozone reactions and came up with the mechanism shownin Fig. 2a . The rst step is an oxidation of bromide (Br )

by ozone via an oxygen atom transfer to producehypobromite (OBr ). For this reaction, a second-orderrate constant k 160 M 1 s 1 was determined by theseauthors. Recently, a second-order rate constantk 258M 1 s 1 was measured [66]. The two valuesdeviate signicantly. However, because bromate forma-tion is not only controlled by this reaction, the absolutevalue of this rate constant is not decisive for bromateformation (see below).

The reaction of bromide with ozone is a reversibleprocess and occurs via a BrOOO intermediate whichcan react further by two mechanisms [66]:

BrOOO - OBr O2 oxygen atom transfer 3

BrOOO - Brd

Od

3 electron transfer 4

The oxygen atom transfer reaction (3) is preferredover the electron transfer (4) by 402 kJmol 1 and

therefore is the dominant pathway [66]. In reaction (3)only about 56% of the formed oxygen is singlet oxygen

Table 3Bromine species formed during bromate formation, oxidation states and important oxidants

Species Chemical formula Bromine oxidation state Controlling oxidizing species

Bromide Br I O 3, d

OHBromine radical Br

d

0 O3Hypobromous acid HOBr +I d OHHypobromite OBr +I O 3,

d OH ; COd

3Bromine oxide radical BrO

d

+II

Bromite BrO 2 +III O 3Bromate BrO 3 +V

Fig. 2. Reaction scheme for bromate formation during ozonation of bromide-containing waters. (a) reactions with ozone and (b)reactions with ozone and OH radicals. The bold lines show the main pathway during the secondary phase of an ozonation process.

Adapted from von Gunten and Hoign

!

e [68].



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[O2 (1Dg)] [67]. This can be explained by the hypothesisthat the BrOOO adduct is sufciently long-lived toreduce the singlet oxygen yield by conversion due to aheavy atom effect [67]. In the case of iodide, this is evenmore pronounced (see below). The hypobromite formedis in equilibrium with hypobromous acid (HOBr) ( pKa(HOBr)=8.89 [55,68]. Therefore, the equilibriumHOBr/OBr lies on the side of HOBr for typicaldrinking water treatment conditions (pH=6.58). Thekinetics of protonation reactions are fast with second-order rate constants of the order of 2 1010 M 1 s 1.Considering the equilibrium constant of 10 9 for HOBr,this results in a deprotonation rate constant of B 20 s 1.For the further reaction of OBr with ozone(400 M 1 s 1, eqs. (5)(7)) a maximum rate constant of 8 10 3 s 1 can be calculated for an ozone concentra-tion of 1mgl 1 (2 10 5 M). Therefore, the HOBr OBr equilibrium reaction is faster than its further

oxidation by several orders of magnitude.The oxidation of bromide by ozone to HOBr isdirectly applied for swimming pool water disinfection.Because ozone is unstable in water, bromide is addedduring ozonation to produce HOBr. In excess of bromide, the further oxidation of HOBr/OBr can besuppressed and the resulting HOBr acts as a long-termdisinfectant. OBr undergoes two reactions: (i) attack of ozone on the oxygen atom (5) and (6), (ii) attack of ozone on the bromine atom (7):

OBr O3- OOBr O 2

k 330 M 1 s 1 5

OOBr - Br O 2 6

OBr O3- OBrO O 2

k 100 M 1 s 1 7

The OOBr

intermediate from the attack of ozoneon the oxygen atom leads back to bromide whereasreaction (7) leads to OBrO (bromite). The rateconstants for reaction (5) is 330 M 1 s 1 and for reaction(7) it is 100M 1 s 1 [55]. This means that onlyapproximately 1/4 of the oxidized OBr leads to

BrO 2 and eventually to BrO 3 .The successive oxidation of bromite to bromate by

ozone is very fast ( k > 105 M 1 s 1, [55],8.9 104 M 1 s 1, [69]). Whereas Haag and Hoign

!

e[55] describe it as a oxygen atom transfer reaction,recent studies by Nicoson et al. [69] suggest an electrontransfer reaction. At a pH of 7, only about 1% of theHOBr species is present as OBr . This leads to anapparent second-order rate constant of k 1 M 1 s 1

for its oxidation by ozone. The formation of bromitewill subsequently proceed with a rate constant of only0.25M 1 s 1. This is considerably smaller than the

second-order rate constants for the oxidation of both

bromide and bromite. Therefore, the rate limiting stepfor the oxidation of bromide to bromate by ozone indrinking waters is the oxidation of OBr , and it can beexpected and is observed that HOBr will build-up duringozonation of bromide-containing waters [55,56].

OH radical reactions . In addition to the mechanismshown in Fig. 2a , there are numerous oxidation stepswhich occur via OH radical oxidation ( Table 3 ). Ozoneand OH radicals react in parallel or in succession leadingto a strongly non-linear behavior of the system. There-fore, it is impossible to predict bromate formationintuitively. Fig 2b shows an extended reaction schemeincluding direct ozone reactions and secondary oxidants(OH radicals and carbonate radicals). This mechanismhas been established in many studies including experi-mental and modeling work [52,56,58,7079] . For acomplete compilation of all reactions and their corre-sponding rate constants see, e.g., von Gunten and

Oliveras [78], Pinkernell and von Gunten [56].According to Fig. 2b and Table 3 , OH radicals playan important role in the oxidation of bromide andHOBr/OBr . The oxidation of bromide occurs accord-ing to the following reaction sequence [80]:

Br d

OH " BrOH

k 1:06 108 M 1 s 1

k 3:3 107 s 1 8

BrOH H " Brd

H 2O

k 1:06 108 M 1 s 1

k 1:36 s 1 9For typical drinking water conditions, the reaction

leading from bromide to the bromine radical (Brd

) canbe described by an overall rate constant of 1.1 109 M 1 s 1 [76].

In ozonation processes, the bromine radical can befurther oxidized by ozone to the BrO

d

radical or reactwith bromide to Br 2

d

and eventually to HOBr [81,78]:

Brd

O 3- BrOd

O2 k 1:5 108 M 1 s 1 10

Brd

Br - Brd

2 k 1010 M 1 s 1 11

Brd

2 Brd

2 - Br Brd

3 k 2 109 M 1 s 1 12

Brd

3 H 2O - HOBr 2Br H 13

For a bromide concentration of 40 mg l 1 (0.5 mM) andan ozone concentration of 1 mg l 1 (20 mM), about 40%of the Br radical react through reaction (10) and 60%through reaction (11). If bromide levels are higher and/or ozone concentrations lower, the fraction of bromineradicals reacting over reaction (11) increases. Becausehigh ozone concentrations are mostly encounteredduring the initial phase of an ozonation process,

reaction (10) is most important during this phase (see



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below control strategies for bromate minimization).BrO

d

radicals are unstable and disproportionate intoHOBr/OBr and BrO 2 . The latter species is quicklyoxidized to bromate by ozone.

The BrOd

radical is only formed from Brd

in thepresence of ozone, i.e., it is not an intermediate inoxidation processes where OH radicals are the onlyoxidants (e.g., UV/H 2O 2 or g-irradiation, [78]). In OHradical-based oxidation processes, Br

d

reacts with Br(reaction (11)), eventually leading to HOBr/OBr as adecisive intermediate. This has consequences for bro-mate formation during hydrogen-peroxide-based ad-vanced oxidation processes (see below).

In contrast to the oxidation by ozone, which canonly oxidize OBr , OH radicals oxidize both HOBrand OBr to BrO

d

with similar rate constants(k HOBr 2 109 M

1 s 1, k OBr 4:5 109 M 1 s 1,[82]). Therefore, the pH-dependent speciation of

HOBr/OBr can only change the rate of this reactionby a factor of two at the most. The oxidation of HOBr/OBr with carbonate radicals to BrO

d

is not wellestablished and has only been assessed qualitatively sofar [68].

Because we know the kinetics of all involvedreactions, we can estimate the fraction of Br andHOBr/OBr oxidized by ozone and OH radicals,respectively, if we know the ratios of the concentrationsof OH radicals and ozone during an ozonation process.For the oxidation of bromide the following rateequation can be formulated:

dBr =d t k 1Br O3 k 2Br d

OH 14The ratio Rc=[

d

OH]/[O 3] can be used to express the[

d

OH] as a function of the [O 3] [1]. This yields:

dBr =d t k 1Br O3 k 2R cBr O3 Br O 3 k 1 k 2 Rc 15

Thus, the fraction f OH ;Br of bromide reacting with OHradicals can be calculated as

f OH ;Br k 2Rc=k 1 k 2R c 16

An analogous calculation can be made for HOBr/OBr .However, in this case also the pH-dependent speciation

of HOBr has to be considered [56]. Table 4 shows the

results of such calculations. For the oxidation of bromide, the ozone reaction is the preferred pathwayfor a wide range of R c values. Only for high R cX 10 7 aconsiderable fraction of bromide will be oxidized by OHradicals. Independent of the pH, such Rc values cantypically be measured during the initial phase of anozonation process [6] or in ozone-based advancedoxidation processes. For the oxidation of HOBr/OBr ,a pH-dependent behavior for the importance of ozonerelative to OH radicals can be calculated. Whereas atlow pH oxidation occurs almost entirely by OH radicals,at higher pH values and lower R c values, the fraction of HOBr/OBr oxidized by ozone can be as high as 90%.

However, as the pH increases, so does the Rc. Foraverage waters, this leads to a X 50% fraction for theoxidation of HOBr/OBr by OH radicals at pH 8. Thesecalculations show clearly that the oxidation pathway forbromide and hypobromous acid/hypobromite is given

largely by the ratio of the concentrations of OH radicalsand ozone. For the secondary reaction phase of anozonation process, this ratio can be strongly inuencedby pH changes, i.e. the Rc can be lowered by up to afactor of 10 if the pH is lowered from 8 to 6. It has beenfound for River Seine water that the fraction f O3 reactingwith ozone for HOBr decreased from p 40% at pH 8 top 15% at pH 7 and p 5% at pH 6 [56]. The effect of pHchanges on bromate formation is discussed in moredetail in the context of control options for bromateformation (see below). The resulting dominant bromateformation pathway during the secondary reaction phaseof an ozonation process is highlighted in Fig. 2b bythicker arrows. The rst oxidation step (Br - OBr )occurs via ozone, the second step (HOBr/OBr - BrO

d

)via OH radicals, the third step is a disproportionation (2BrO

d

- OBr +BrO 2 ), and the fourth step is anoxidation by ozone (BrO 2 - BrO 3 ).

In addition to the reactions shown in Fig. 2b , thereare other reactions which interfere with bromateformation. First of all, the Br atom can react withNOM to form bromo-organic compounds or bromide.The rate constant for this reaction was estimated to be109 M 1 s 1 for a NOM concentration of 2 10 6 M(based on DOC: 12 mg l 1=1 mM, [56]). For this

scenario, the reaction of Brd

with NOM would be of

Table 4Fraction of ozone ( f O 3 ) and OH radicals ( f OH ) pathway for the oxidation of bromide and HOBr/OBr (towards bromate)

Bromide Hypobromous acid/ Hypobromite

pH 6 pH 7 pH 8

[OHd

]/[O3] f O3 (%) f OH (%) f O 3 (%) f OH (%) f O3 (%) f OH (%) f O 3 (%) f OH (%)

10 7 60 40 0.08 99.92 0.8 99.2 6 9410 8 94 6 0.8 99.2 7 93 40 6010 9 99.3 0.7 8 92 44 56 87 13



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the same order of magnitude as its reaction withbromide or ozone.

In addition, there are three types of reactionsinterferring on the level of HOBr/OBr :

HOBr =OBr NOM - Br NOM 17

HOBr =OBr NOM - NOM ox Br 18

HOBr HO 2 - Br H 2O O2

k 7:6 108 M 1 s 1 19

HOBr NH 3- NH 2Br H 2O

k 8 107 M 1 s 1 20

The reaction of HOBr/OBr with NOM is of interest inconnection with the potential formation of bromo-

organic compounds. It may also lead back to bromide,resulting in a catalytic destruction of ozone. The fastreaction (19) of HOBr with HO 2 (H 2O 2) is of particularimportance in hydrogen-peroxide-based advanced oxi-dation processes such as O 3/H 2O 2 and UV/H 2O2 [83].Reaction (20) is important in ammonia-containingwaters. The reaction product of the fast reactionbetween HOBr and NH 3 is monobromamine which isoxidized by ozone to bromide and nitrate [84]. There-fore, addition of ammonia can be used to reducebromate formation (see control options).

Modeling of bromate formation . Because bromateformation during ozonation of bromide-containingwaters is a highly non-linear process, kinetic modelinghas been applied to improve mechanistic understandingand to predict bromate formation [70,79,85,68,56] . Thefull model consists of more than 50 coupled kineticequations which can be solved simultaneously with acomputer code such as ACUCHEM [86]. However, themain problem of these models is the unknown behaviorof NOM both in ozone decomposition and bromateformation. To overcome the problem on the oxidantside, the experimental ozone and OH radical concentra-tions were introduced into the model by the Rc value.This allowed modeling bromate formation during

ozonation of natural waters in laboratory systemswithout introducing any additional tting parameters[56]. In an other study, kinetic modeling was applied toinvestigate parts of the bromate formation mechanism[78]. However, the predictive capabilities of such modelsfor the ozonation of any water should not be over-estimated. Due to the fact that the rate constants for thereactions in the bromate formation mechanism stemfrom various studies, predictions should preferably beused to investigate trends for changing water treatment/quality parameters. A simulation within a factor of tworelative to the measured bromate concentration has to

be considered satisfactory.

3.2.1.3. Bromate formation during oxidation of micropollutants in advanced oxidation processes . In advancedoxidation processes (AOPs), bromate formation isgoverned by oxidation processes with OH radicals. Ithas been shown in g-radiolysis experiments that forsystems where OH radicals are the only oxidants, HOBris a decisive intermediate [78]. Since H 2O 2 is animportant chemical in many AOPs (UV/H 2O 2, O 3/H 2O 2, Fenton reaction), the reaction of HOBr withHO 2 back to bromide (reaction (19)) is particularlyimportant because it is fast [83].

In the UV/H 2O 2 process, OH radicals are the onlyoxidants and they are produced in solutions in presenceof a high excess of H 2O 2. The OH radical oxidation of bromide leads to Br

d

which eventually forms HOBr (seeabove, reactions (8), (9), (11)(13)). HOBr is then againreduced by hydrogen peroxide to Br . Since hydrogenperoxide is present in a large excess, the further

oxidation of HOBr by OH radicals is outcompeted byits reduction. Even if a steady-state OH radicalconcentration in the UV/H 2O 2 process of 10

10 M isassumed, only a fraction (approx. 3.6%) of the totalHOBr would be further oxidized and nally lead tobromate for a H 2O 2 concentration of 10 mg l

1. There-fore, no bromate can be detected in the AOP UV/H 2O 2[77,78] . However, relatively high hydrogen peroxidelevels need to be applied (520 mg l 1) because of the lowefciency of H 2O 2 photolysis. High concentrations of H 2O 2 are undesired from a toxicological point of view.In addition, hydrogen peroxide reacts quickly withchlorine applied for the distribution system. Therefore,biological ltration is necessary to remove excess H 2O 2.

In contrast to the UV/H 2O 2 process, in the O 3/H 2O 2-based AOP, bromate is still formed even if very highH 2O 2 levels are applied [78]. In this process, part of thebromate formation occurs via the Br atom which isoxidized to the BrO

d

radical by ozone (reaction (10)).This species is not reduced by H 2O 2 [78]. Hence, in theAOP O 3/H 2O 2 there is a trade-off between a moreefcient oxidation of the target micropollutant andbromate formation.

Table 5 shows an example of the oxidation of atrazinein a full-scale water treatment plant [75] and methyl tert-

butyl ether (MTBE) in a laboratory study [87] and theconnected bromate formation. Results are shown forconventional ozonation or the AOP O 3/H 2O 2 for waterscontaining about 50 mg l 1 bromide.

The two experiments shown for the full-scale watertreatment plant demonstrate the benecial effect of anaccelerated O 3/H 2O 2-based AOP. For the same ozonedose, the oxidation of atrazine increases from 42% to75% if the process is changed from conventionalozonation to the AOP. This is due to a fastertransformation of ozone into OH radicals. In theconventional ozonation process, the ozone residual

concentration at the outlet of the reactor was



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0.65mgl 1 , whereas it was 0.1 mg l 1 for the AOP. Dueto this higher degree of ozone transformation, the OHradical oxidation capacity for the AOP is signicantlyhigher. Bromate formation decreases drastically from 19to 4 mg l 1 from conventional ozonation to the AOP.

This decrease is due to the reduction of HOBr byH 2O 2 /HO 2 (Eq. (19)). However, even in presence of H 2O 2 , bromate formation is not zero. This is due to theoxidation of the Br atom by ozone (reaction (10)).

The laboratory experiments for MTBE oxidation inthe two waters in Table 5 basically show similar features.For a constant ozone dose and a complete transforma-tion of ozone, the addition of H 2O 2 slightly increases theelimination percentage of MTBE and decreases bromateformation. A higher ozone dose increases the elimina-tion percentage of MTBE but also increases bromateformation. Both waters shown in Table 5 (MTBE) haveapproximately the same OH radical scavenging rateconstant. It is composed of the OH radical scavenging

by NOM and carbonate/bicarbonate. This is reected insimilar oxidation percentage of MTBE for similar ozonedosages. It is an indication that the OH radicaloxidation in both waters is fairly similar. For an ozonedose of 2 mgl 1, approximately 40% of the MTBE canbe oxidized, whereas for 4 mg l 1 it is approximately70%. As a negative trade-off, bromate formationincreases signicantly if the ozone dose is increased. Inaddition, the absolute level of bromate is markedlyhigher in the high-alkalinity water (well water) than inthe water with the lower alkalinity (lake water). This isdue to the reaction of carbonate radicals with OBr [68].

Therefore, the application of the AOP O 3/H 2O 2 has to

be carefully evaluated for high-alkalinity waters to avoidexcessive bromate formation.

Both the oxidation of atrazine and MTBE haveshown that it is possible to oxidize micropollutantsefciently with the O 3 /H 2O 2-based AOP. However, inturn, the process must be optimized with regard tobromate formation. This is of particular importance if the rate constants for the micropollutants that need tobe oxidized are relatively low. This is certainly the casefor MTBE ( k OH 1:9 109 M 1 s 1, [87]). For micro-pollutants with higher second-order rate constants forreactions with OH radicals, the oxidation can be carriedout to a higher degree without violating the drinkingwater standard for bromate. This is shown with theoxidation of atrazine ( k OH 3 109 M

1 s 1, [88])where a lower ozone dose already leads to a similardegree of oxidation (75%). In this case the bromate levelstill remains low (4 mgl 1).

3.2.1.4. Balance between bromate formation and disinfection in bromide-containing waters . The occurrence of more resistant pathogens such as C. parvum oocystsleads to a demand for increased cts for their inactiva-tion. In waters with bromide levels above 50 mg l 1,bromate formation may exceed the drinking waterstandard of 10 mg l 1 under certain treatment conditions[89,90] . One of the decisive factors is the temperature of the treated water. Both the efciency of inactivation of microorganisms ( Table 1 ) and bromate formationincrease with increasing temperature. Therefore, it hasto be tested whether lower or higher temperatures lead

to a better performance with regard to the overall

Table 5Oxidation of micropollutants and bromate formation

Oxidation of atrazine in a full-scale plant [75]( pH 7.6, DOC 2mgl 1, alkalinity 4mM)

ozone 1.6 mg l 1 ozone/H 2O 2 1.6/0.5 mg l1

% elim. atrazine 42 75Bromate a (mg l 1) 19 4

Oxidation of MTBE in laboratory system [87]ozone 2mgl 1 ozone/H 2O 2 2/0.7mgl

1 ozone/H 2O 2 4/1.4mgl1

Lake Z.

urich water( pH 7, DOC 1.4mgl 1; alkalinity 2.5mMOH scavenging rate constant : 5.6 104 s 1)

% elim. MTBE 39 46 71Bromate (b) (mgl 1) 15.1 8.8 16.8

Well water Porrentruy( pH 7, DOC : 0.8mgl 1; alkalinity 5mM)

OH scavenging rate constant : 6.2 104 s 1)% elim. MTBE 28 37 65Bromate b (mg l 1) 20.7 12.5 33.5

Initial bromide concentrations:a 45 mg l 1.b 50 mg l 1.



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process (sufcient disinfection with low bromate). Fig. 3shows a comparison of bromate formation and inactiva-tion of B. subtilis spores for 5 1C, 20 1C and 30 1C [6].

It can be seen that the initial phase is decisive for theoverall process. During this phase, bromate increasesquickly because of the importance of OH radicalreactions. However, inactivation of B. subtilis does notoccur due to the lag-phase. Therefore, the critical factorwill be the bromate formed during the lag-phase of theinactivation. In the example shown in Fig. 3 , highertemperatures are more critical because initial bromateformation under these conditions is very fast. Thisoutrules the shorter lag-phase for inactivation at thesetemperatures. To overcome these problems, it might benecessary to use control options to minimize bromateformation.

3.2.1.5. Strategies for bromate minimization . To solvethe problem of excessive bromate formation, severaltreatment options have been tested, namely ammoniaaddition, pH depression, OH radical scavenging andscavenging or reduction of HOBr [91,92,56] . It turns outthat only the rst two options, namely ammoniaaddition and pH depression, are feasible in drinkingwater treatment [93,56]:

* Ammonia addition does not alter the ozone stabilityand therefore, oxidation and disinfection processesremain unchanged. Ammonia only interferes byreacting with HOBr (reaction (19)), the decisiveintermediate of the bromate formation mechanism[94]. Fig. 4a shows the effect of ammonia on bromateformation in Lake Z .urich water which was spiked toa bromide level of 50 mg l 1. It is shown that this

measure has no inuence on bromate formation

during the initial phase of ozonation (ozone exposurep 2 m g l 1 min). As discussed above, during thisozonation phase a large fraction of bromide isoxidized by OH radicals leading to the bromine

atom which is further oxidized to BrOd

by ozone.BrO

d

does not react with ammonia and therefore, thereaction proceeds towards bromate. Only part of thispathway leads to the formation of HOBr/OBrwhich is necessary for the effeciency of the ammoniaaddition. Fig. 4a also shows that ammonia additionis only effective up to a certain concentration (in thiscase 200 mg l 1). A further increase of ammoniabeyond this concentration does not show an im-provement in bromate minimization [56]. It has beensuggested that the reaction between HOBr and NH 3is a base-catalyzed equilibrium reaction which still

leaves a fraction of HOBr/OBr to be further

Fig. 4. Bromate minimization during ozonation processes: a)Lake Z

.

urich (LZ) water (Switzerland), addition of ammonia, b)River Seine (RS) water (France) pH depression (86). Waterquality parameters and experimental conditions: LZ: DOC1.4mgl 1, alkalinity 2.5 mM, pH=8, T =20 1C, [O 3]o =1.5mgl 1, [Br ]o =50 mgl

1; RS: DOC 2.4 mgl 1, alkalinity3.8mM, T =10 1C, [O 3]o =2 mgl

1, [Br ]o =50 mg l1,

[NH 4+ ]=126 mg l 1. Adapted from Pinkernell and von Gunten

[56].

Fig. 3. Inactivation of B. subtilis spores and bromate formationduring ozonation of Lake Z .urich water at temperatures 5 1C,201C and 30 1C ([Br ]o =30 mgl

1, pH=7, DOC=1.4 mgl 1,alkalinity 2.5 mM).



12/19

oxidized to bromate [56]. Therefore, this method isnot efcient in waters that already contain medium tohigh levels of ammonia. Because even small ammoniaconcentrations show a positive effect, this could bean economical way of minimizing bromate forma-tion.

* Lowering the pH is another effective method forbromate minimization ( Fig. 4b ). This inuencesbromate formation by shifting the HOBr/OBrequilibrium towards HOBr and increasing the ozoneexposure relative to OH radical formation. As aconsequence, for a certain ozone exposure the OHradical exposure decreases with decreasing pH. Thismeans that for similar disinfection targets the overalloxidant exposure (ozone and OH radicals) will besmaller at a lower pH, leading to a lower bromateformation. Because the oxidation of HOBr/OBr isdominated by OH radicals under almost all treat-

ment conditions, a decrease of pH will lead to lessformation of bromate if normalized to a certainozone exposure. The effect of pH on the contributionof direct ozone reaction is expected to be small.Similar to the addition of ammonia, this measurealso does not lead to a reduction of the initialbromate formation. The initial fast transformation of ozone into OH radicals is almost independent of thepH. This leads to a similar initial bromate formationfor various pH values ( Fig. 4b ). Due to the longerlifetime of HOBr under depressed pH conditions,there may be a trade-off between lower bromateformation and higher total organic bromine (TOBr)formation [93]. In high-alkalinity waters, pH depres-sion might be expensive because large amounts of carbon dioxide have to be added to the water.

Overall, ammonia addition and reasonable pHdepression may lead to a 50% decrease of bromateformation. This might be a feasible control option forwaters containing bromide levels between 50 and150 mg l 1.

3.2.1.6. Bromate elimination during drinking water treat-ment . Once bromate is formed, its elimination isdifcult. Most of the research has concentrated onbromate removal by activated carbon ltration andmicrobiological procsses such as BAC or slow sandltration. Ozonation is frequently followed by theseprocesses to remove biodegradable organic compounds.However, other process such as UV irradiation andaddition of iron were also investigated:

* It has been demonstrated in several studies that newgranular activated carbon (GAC) is able to reducebromate to bromide [9598]. However, there is aconsensus in the literature that the change from GAC

to biologically active carbon (BAC) results in a

decrease of the rate of bromate reduction. Also thepresence of NOM or high concentrations of otheranions decrease the ability of GAC to reducebromate. Bromate elimination in BAC was observedin the laboratory under conditions of low oxygenconcentrations ( o 2 m g l 1, [99]). However, this effectwas not observed in full-scale plants, mainly due tothe high oxygen concentrations after ozonation. Theoxygen concentrations are especially high if ozone isproduced from oxygen. Therefore, bromate reduc-tion in activated carbon lters has only limitedapplicability after ozonation.

* The reduction of bromate by iron(II) was tested inlaboratory systems [100,101] . Bromate was reducedby Fe(II) to bromide in the pH range of drinkingwaters. As expected, the rate of bromate reductionincreases with decreasing pH. Relatively high Fe(II)doses (>10 mg l 1) are required to obtain a sufcient

reduction rate. Iron (II) adsorbed to iron(III)(hydr)-oxides enhances the rate of reduction of bromatesignicantly under anoxic conditions [102]. However,for both the homogeneous and the heterogeneousprocess, the presence of dissolved oxygen is proble-matic because it competes with bromate as anoxidant for Fe(II). Therefore, the addition of Fe(II)does not appear to be a feasible option for bromatereduction after ozonation.

* It was demonstrated that UV irradiation of bromate-containing solutions with low-pressure mercurylamps ( l 255 nm) leads to its reduction to hypo-bromous acid and nally bromide [100]. However,the required UV exposures were much higher thanthose typically required for disinfection processes. Itwas recently demonstrated that the inactivation of C. parvum oocysts by two and three orders of magnituderequired a UV exposure of 10 and 25 mJ cm 2,respectively [103]. To achieve a reduction of 50 mg l 1

bromate to approximately 25 mg l 1, the UV exposureranged from 250 to 550 mJ cm 2 depending on thelamp type (low vs. medium pressure, [104]). Toreduce bromate from 50 to o 10 mg l 1, exposuresabove 1000mJ cm 2 were required [100]. This is abouta factor of 100 higher than doses required for

disinfection. Therefore, this process will be restrictedfor economical reasons.

3.2.2. Bromo-organic compoundsOnly few studies have been carried out to establish the

kinetics of reactions (17) and (18) [105,106,56] .The formation of bromo-organic compounds is

controlled by the competition of reactions (17) and(18) with the oxidation of HOBr/OBr with ozone andOH radicals. In addition, HOBr/OBr and O 3 competefor similar structures in the NOM. A comparison of thekinetics of the reaction of HOBr and ozone with organic

model compounds in combination with the fact that the



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concentration of ozone is always much higher than theconcentration of HOBr allows the conclusion that only aminor fraction of HOBr, reacts via reactions (17) and(18) [56]. In addition, we know from the analogousreaction of chlorine with NOM that only a minorfraction (a few percent) of the halogen adds to NOM[107]. Therefore, it is expected that the formation of bromo-organic compounds during ozonation is of minorimportance. This is in agreement with ndings fromwater works, where bromoform was o 5 mg l 1 in 37 outof 38 ozone treatment plants [63]. This is far below thedrinking water standards of the European Union(S THM 100 mg l 1) and the USEPA ( S THM 80 mg l 1) [7,3]. In the same study, in one water withextremely high concentrations of DOC>6 mg l 1, bro-moform was found to be between 30 and 73 mg l 1. Partof the organic bromine will also bind to highermolecular weight organic material from NOM. Westerh-

off et al. [85] have shown in laboratory experiments thatthis fraction can make up between 1% and 8% of theinitial bromide concentration. However, these highermolecular weight bromine compounds are currently notregulated.

3.3. By-product formation in iodide-containing waters

Iodide concentrations in natural waters are usuallyfairly low ( o 10 mg l 1). However, due to specialgeological formations or seawater intrusion, iodideconcentrations can reach levels of X 50 mg l 1 [114].The oxidation of iodide by ozone is very fast with asecond-order rate constant k 1:2 109 M 1 s 1 [66]and occurs via an oxygen atom transfer. It leads tohypoiodite (OI ) which is in equilibrium with hypoio-dous acid (HOI) ( Fig. 5 ). In the case of iodide, onlyabout 12% of the formed oxygen is singlet oxygen [O 2(1Dg)] which conrms the explanation of a long-livedintermediate adduct (I-OOO ). This reduces the singletoxygen yield by conversion due to a heavy atom effect[67]. For drinking waters, OI accounts only for a verysmall fraction of the overall HOI concentration( pK a 10:4; [108]). Similar to the case of bromine, theOI /HOI intermediate can undergo several reactions. It

is shown in Fig. 5 that it can be further oxidized toiodate, disproportionate to iodate and iodide, and it canreact with NOM to form iodo-organic compounds[1 0 9]. Disproportionation of HOI is a very slow processunder drinking-water treatment conditions. Therefore,this process is insignicant for the lifetime of HOI [108].Among the iodinated organic compounds, iodoform ismost undesired because of its low organoleptic thresholdconcentration which is between 0.3 and 1 mg l 1

[110,111] . Iodoform leads to an unpleasant medicinaltaste and odor in drinking waters. Under typicaldrinking water treatment conditions, HOI does not

react with ammonia [109].

For drinking waters, iodate is the desired sink. Unlikebromate it is not considered a carcinogenic risk[112] because it is quickly reduced to iodide in thebody [113]. In contrast to aqueous bromine andchlorine, both HOI ( k 3:6 104 M 1 s 1) and OI(k 1:6 106 M 1 s 1) are quickly further oxidized to

iodate by ozone [114]. Therefore, OH radical oxidationscan be neglected in the case of iodine. The oxidation of the I(+I) species by ozone occurs predominantly viaHOI for pH p 8. Only at pH>8 the oxidation of OIcontributes signicantly to the overall oxidation process.The lifetime of HOI/OI in ozonation processes is in theorder of ms. Therefore, the reaction of HOI with NOMcan be ruled out.

In the case of chlorination and especially chloramina-tion processes, HOI is sufciently long lived to reactwith NOM. The further oxidation of HOI is eitherslow (chlorine) or does not occur (chloramin)[115,116,109,117] . Therefore, the risk for iodoformformation is substantially higher under these treatmentconditions.

3.4. Chlorine-derived by-products

Chloride is not oxidized by ozone. Also, the oxidationof chloride to Cl radicals by OH radicals can beneglected under typical drinking water conditions(circumneutral pH, [82]). The oxidation of chloride byOH radicals is analogous to the oxidation of bromide.The formation of Cl . and OH is given by a pH-dependent equilibrium [80]:

Cl d

OH " ClOH

k 4:3 109 M 1 s 1

k 6 109 s 1 21

ClOH H " Cld

H 2O

k 2 1010 M 1 s 1

k 1:6 105 s 1 22

Reaction (21) has a very fast back reaction which leadsto small steady-state concentrations of ClOH . Inaddition, its further reaction (22) to Cl

d

and H 2O

depends on the pH. Taking both reactions (21) and (22)

Fig. 5. Iodate formation during ozonation of iodide-containingwaters.



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into account, the rate constant for the oxidation of chloride by

d

OH to Cl atoms is about 10 3 M 1 s 1 at pH7 and increases by a factor of 10 per pH unit decrease.For this reason, it only becomes important at pH o 3.

Thus, inorganic chlorine-derived by-products are onlyformed during ozonation if a water is pretreated withchlorine or chlorine dioxide prior to ozonation. Fig. 6shows the oxidation processes that can occur duringozonation of chlorine-containing waters. The processesare analogous to the bromine and iodine chemistrieswith the difference that they start from the Cl(+I)oxidation state. In the rst step of the oxidation, ozonetransfers an oxygen atom to hypochlorite. Hypochlor-ous acid which is in equilibrium with hypochlorite( pK a =7.5) is not oxidized by ozone [118]. The attack onhypochlorite can occur either at the chlorine atom or atthe oxygen atom, leading to chlorite (ClO 2 ) and chloride(Cl ), respectively. The oxidation of chlorite by ozone is

fast, with the formation of chlorine dioxide [69,119] . It isa rare case of an electron transfer reaction between aninorganic compound and ozone. In excess of ozone,chlorine dioxide is then quickly further oxidized tochlorine trioxide (ClO 3) which reacts with chlorinedioxide to chlorate (ClO 3 ) the end product of theoxidation by ozone (see bottom of Fig. 6 ). The overallsecond-order rate constant for the oxidation of hypo-chlorite by ozone is 140 M 1 s 1 with 110 M 1 s 1

leading to chloride and 30 M 1 s 1 leading to chlorite[118,119] . The rate constant for the oxidation of OClwith ozone is relatively small. Therefore, it is possible forboth ozone and chlorine to be present simultaneously.The oxidation of chlorite with ozone to chlorine dioxidehas a second-order rate constant of 8.2 106 M 1 s 1

[69]. Chlorine dioxide reacts with ozone with a second-order rate constant of 1.05 103 M 1 s 1 [120]. If ozoneis present in large excess over chlorite, ClO 3 may be themain product. It is unstable in water and dispropor-tionates to ClO 3 and ClO 4 [121].

OH radicals can also play a signicant role for theoxidation of chlorine to chlorate ( Fig. 6 ). The second-order rate constant for the oxidation of OCl with OHradicals to the ClO radical is 9 109 M 1 s 1 [122]. Novalue is available in literature for HOCl. Considering

only the oxidation of OCl by ozone and OH radicals,the fraction f O3 of the total HOCl reacting with O 3 canbe calculated to be 60% ( R c=[

d

OH]/[O 3]=108). This

fraction is independent of the pH and correspondsapproximately to the f O 3 for HOBr at pH 8 ( Table 3 ).If there is a fast reaction of HOCl with OH radicals, f O3 will decrease and will become pH dependent. Asshown in Fig. 6 , the further oxidation of chloritewith OH radicals ( k OH ;ClO 2 4:2 10

9 M 1 s 1)leads to chlorine dioxide and then to chlorate(k OH ;ClO 2 4:2 10

9 M 1 s 1) [120]. For conventionalozonation processes these reactions are not important.

The direct reactions with ozone are much faster. Based

on this nding the reaction between ozone and chlorinewas used to calibrate a full-scale ozonation reactorduring conventional ozonation [9]. A shock prechlorina-tion was applied and the chlorate formation wasmeasured in the ozonation reactor. A combination of hydraulic modeling with reaction kinetics allowed theassessment of the ozone concentration and, in turn, thedisinfection and oxidation processes in this particularreactor [9].

Similar to HOBr, HOCl reacts quickly with NH 3 toform monochloramine, which is oxidized by ozone tonitrate and chloride. Because chloride is not oxidizedduring ozonation, the presence of ammonia completelyprevents the formation of chlorate ( Fig. 6 ).

Possible adverse health effects from chlorate are notwell established to date. The WHO gives no guidelinevalue for drinking water, however, it does suggest thatchlorate should be kept as low as possible until moretoxicological information is available [59]. In Switzer-land, the drinking water standard for chlorate is

200 mg l 1 [123].

3.5. Effect of ozonation on precursors for chlorination by- products

The reactions of ozone with NOM may lead to apartial removal of precursors of trihalomethanes(THM), total organically bound halogens (TOX),trichloroacetic acid and dichloroacetonitrile. After pre-ozonation, these post-chlorination by-products may bereduced by 575% ( [124,125,8] and references therein).However, precursors of dichloroacetic acid were not

altered and precursors for 1,1,1-trichloroacetone even

Fig. 6. Oxidation of chlorine species to chlorate duringozonation. (a) processes involving hypochlorite and (b) oxida-tion of chlorine dioxide.



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Ozonation of Drinking Water Part II