1 Ozone oxidation for the alleviation of membrane fouling by natural organic matter:
2 A review
3
4
5 Steven Van Geluwe a
; Leen Braeken a,b
; Bart Van der Bruggen a
.
6
7 a
Laboratory of Applied Physical Chemistry and Environmental Technology
8 Department of Chemical Engineering
9 K.U. Leuven
10 W. de Croylaan 46
11 B-3001 Leuven (Heverlee)
12
13 b
Department of Industrial Sciences and Technology
14 KHLim
15 Universitaire Campus Gebouw B, bus 1
16 B-3590 Diepenbeek
17
18 E-mail addresses of the authors: steven.vangeluwe@cit.kuleuven.be;
19 leen.braeken@khlim.be; bart.vanderbruggen@cit.kuleuven.be
20
1
21 Corresponding author:
22 Steven Van Geluwe
23 Tel: +32 16 322 341
24 Fax: +32 16 322 991

25 steven.vangeluwe@cit.kuleuven.be
26
27 Word count (not including Abstract or References): 7999
28
29 Abbreviations:
30 AOP: advanced oxidation process
31 COD: chemical oxygen demand
32 DBP: disinfection by-product
33 DOC: dissolved organic carbon
34 FTIR: Fourier transform infrared
35 GAC: granular activated carbon
36 HAA: haloacetic acid
37 IEP: isoelectric point
38 MF : microfiltration
39 MWCO: molecular weight cut-off
40 NF: nanofiltration
41 NMR: nuclear magnetic resonance
42 NOM: natural organic matter
43 RO: reverse osmosis
44 SBH: Staehelin, Bühler and Hoigné
45 THM: trihalomethane
46 UF: ultrafiltration
47 USEPA: United States Environmental Protection Agency
48 UVA: absorbance (optical density) of UV irradiation
49
2

3
50 ABSTRACT
51
52 Membrane fouling by natural organic matter is one of the main problems that slow
53 down the application of membrane technology in water treatment. O
3
is able to efficiently
54 change the physico-chemical characteristics of natural organic matter in order to reduce
55 membrane fouling. This paper presents the state-of-the-art knowledge of the reaction
56 mechanisms between natural organic matter and molecular O
3 or .
OH radicals, together
57 with an in-depth discussion of the interactions between natural organic matter and
58 membranes that govern membrane fouling, inclusive the effect of O
3
oxidation on it.
59
60 Key words : humic acids; hydrophobicity; electrostatic interactions; molecular mass;
61 aggregation; hydrogen peroxide.

4
62 INTRODUCTION
63
64 Membrane technology has become well established in water treatment, and the
65 demand for membranes increases yearly by 8% (Leiknes, 2009). The most important type
66 of membranes are pressure-driven processes, including microfiltration (MF), ultrafiltration
67 (UF), nanofiltration (NF) and reverse osmosis (RO). Typical values of the main membrane
68 characteristics, i.e. water permeability, operating pressure, pore size and retention
69 characteristics for these four membrane types are provided in Table 1. Because of the
70 large pores of the MF and UF membranes, the water flux is high while the transmembrane
71 pressure is low. MF is used for the removal of suspended particles, turbidity and various
72 micro-organisms (Yuan and Zydney, 1999), while UF removes viruses (van Voorthuizen
73 et al., 2001), colloids and the high-molecular mass fraction of natural organic matter
74 (NOM) as well (Siddiqui et al., 2000; Lee et al., 2005(a), Kennedy et al., 2005). NF
75 membranes have smaller pores, but still maintain a fairly high flux at a reasonable
76 pressure. NF is very effective in the removal of the medium- and lower-molecular mass
77 fraction of NOM (Siddiqui et al., 2000; Shon et al., 2004; Meylan et al., 2007; de la Rubia
78 et al., 2008), and emerging micropollutants such as pesticides, pharmaceuticals and
79 endocrine disrupting chemicals (Kimura et al., 2003; Nghiem et al., 2004; Yoon et al.,
80 2006, Verliefde et al., 2007). The retention of inorganic ions by NF membranes is strongly
81 dependent on the charge of the ions. The retention of divalent ions ranges between 50 and
82 100%. It is much higher than the retention of monovalent ions, which is usually lower than
83 40%, because of Donnan exclusion (de la Rubia et al., 2008; Ouyang et al., 2008). RO is
84 commonly used for desalting brackish water and seawater, but operates under very high
85 transmembrane pressures and a low permeate flux compared to the other pressure-driven
86 membranes. However, RO shares about 45% of the global production capacity of

5
87 desalinated water, because of its lower energy consumption compared to multistage flash
88 evaporation (Darwish and Al-Najem, 2000; Eltawil et al, 2009).
89
90 In spite of the excellent retention characteristics of membrane filtration in water
91 treatment, there are still problems that slow down its growth. The most well-known
92 problem is fouling of the membrane, which results in a reduction in water flux, and thus
93 leads to higher operating costs. Over time, fouling and subsequent cleaning of the
94 membranes causes deterioration of membrane materials, resulting in a compromised
95 permeate water quality and ultimately, a shorter membrane lifetime (Košutić and Kunst,
96 2002; Seidel and Elimelech, 2002; Al Amoudi and Lovitt, 2007). Membrane fouling is
97 usually minimized by an excessive pretreatment or a very conservative membrane flux
98 needs to be used. Consequently, the capital cost is high, which makes membrane filtration
99 less competitive against conventional water treatment technologies (such as coagulation or
100 activated carbon) in certain cases (Pianta et al., 2000).
101
102 The emerging use of O
3
oxidation in water treatment offers new opportunities,
103 because O
3
is able to decompose certain membrane foulants very efficiently. The present
104 paper is a critical review of literature concerning the fouling potential of NOM in water
105 purification and the use of O
3
oxidation for the alleviation of membrane fouling by NOM.
106 The effect of O
3
oxidation on membrane fouling is difficult to predict due to the complex
107 nature of NOM, the strong variability of the NOM characteristics and the water matrix
108 with location, season and weather (Lowe and Hossain, 2008), and the major effect of the
109 water matrix on the conformation of NOM and the decomposition of O
3
. This review
110 paper presents the reaction mechanisms and its products when NOM solutions are treated
111 by O
3
or O
3
+ H
2
O
2
, and the various interactions that exists between NOM components
112 and the membrane surface. The relation between the complex fouling behaviour of NOM,

6
113 before and after its reaction with O
3
, is discussed in a systematic and detailed way, in
114 order to better understand the mechanisms behind the fouling reduction in water treatment
115 by O
3
treatment.
116
117

7
118 THE CHEMICAL COMPOSITION OF DIFFERENT NOM FRACTIONS
119
120 NOM is a complex heterogeneous mixture of organic material, such as humic
121 substances, polysaccharides, aminosugars, proteins, peptides, lipids, small hydrophilic
122 acids, and others (Frimmel et al., 2002). As a first approach to separate the different
123 components in NOM, it is divided into two major classes. The first class, i.e.
124 autochthonous NOM, is derived from extracellular macromolecules of micro-organisms in
125 the water body and carbon fixation by algae and aquatic plants. The second class,
126 allochthonous NOM, is derived from the decay of plant and animal residues in the
127 watershed (Frimmel et al., 2002). It is usually referred to as humic substances, and this
128 term will be used in the remaining text.
129
130 Although the correct chemical structure of humic substances still remains
131 unknown, they consist of a skeleton of alkyl and aromatic units, cross-linked by a variety
132 of functional groups. Humic substances are high in aromatic carbon and have a negative
133 charge. This charge is primarily contributed by their three main functional groups, namely
134 carboxylic acids, methoxyl carbonyls and phenolic functional groups (Thurman, 1986).
135 The chemical properties of humic substances are succinctly explained by McDonald et al.
136 (2004) and Sutzkover-Gutman et al. (2010). The humic substances present in natural
137 waters are traditionally divided into two categories, namely humic and fulvic acids. Humic
138 acids have a higher molecular mass (2000-5000 g mol
-1
) than fulvic acids (500-2000 g
139 mol
-1
) (Her et al., 2003). They have a lower oxygen content and are more hydrophobic
140 than fulvic acids (Thurman, 1986).
141

8
142 The knowledge of the structural chemistry of humic substances is like that of
143 proteins in the middle of the last century. Although the main building blocks are known,
144 there are no conclusive studies about the long-range conformational structure of humic
145 substances. This is probably because of the difficulty of obtaining reliable structural data,
146 due to the size and number of stereochemical isomers (Jansen et al., 1996). In addition, the
147 conformation of humic substances may vary significantly due to changes in pH (humic
148 substances are highly deprotonated is most aquatic environments), cation concentration
149 (humic substances form strong complexes with metals and other cations) and the great
150 stabilization effect on the electrostatic energy by the presence of water molecules (Kubicki
151 and Apitz, 1999). Therefore, the complexity of humic substances in its natural
152 environment is too high for the application of modelling techniques on an atomic scale.
153
154 Autochthonous NOM includes a large number of relatively simple compounds of
155 known structures: carbohydrates, aminosugars, proteins, peptides, small organic acids,
156 etc.. The main functional groups in autochthonous NOM are carboxylic acids, alcohols
157 and amines, which make these compounds hydrophilic, in contrast to the more
158 hydrophobic humic substances. The molecular mass of these hydrophilic compounds
159 shows a great variety. The simple organic molecules have a molecular mass of a few 100 g
160 mol -1 , and have thus an apparently smaller molecular size than humic substances, while
161 the biopolymers, such as polysaccharides and proteins, have a molecular mass between
162 10,000 and 30,000 g mol
-1
(Lee et al., 2004), which is about one order of magnitude
163 higher than the molecular mass of humic substances. Polysaccharides present in surface
164 water have a diameter in the range between 2 and 20 nm. They have a linear molecular
165 structure and can be hundreds of times longer than wide and can be branched (Leppard,
166 1997).
167

9
168 THE DECOMPOSITION OF NOM BY OZONE AND HYDROXYL RADICALS
169
170 3.1 Ozone reacts selectively with certain functional groups in NOM
171 O
3
is a powerful oxidant and its high reactivity can be attributed to the electronic
172 configuration of this molecule. The O
3
molecule represents a hybrid, formed by the four
173 possible resonance structures shown in Figure 2. The structures I and II contribute the
174 most to the actual electron distribution in O
3
, because all the oxygen atoms attain a noble
175 gas configuration, and the oxygen atoms with a positive and negative formal charge are
176 closest to each other. The absence of electrons in the central oxygen atom in resonance
177 structures I and II, explains the electrophilic character of O
3
. Conversely, the excess
178 negative charge present in one of the terminal atoms imparts a nucleophilic character.
179 These properties make O
3
an extremely reactive compound (Beltrán, 2004).
180
181 The electrophilic character of O
3
accounts for the very fast reaction of O
3
with
182 unsaturated bonds (von Gunten, 2003(a)). The fast reaction of O
3
with double bonds and
183 aromatic rings present in NOM molecules, is manifested by a sharp decrease of the optical
184 density at 254 nm (UVA
254
) during ozonation. For instance, Song et al. (2010) reported a
185 UVA
254 reduction of 71% in surface water treatment, at a O
3
dosage of 3.0 mg L
-1
186 (oxidation time: 10 minutes). Wang and Pai (2001) reported a reduction of 40% for
187 biologically treated wastewater effluents, and Lee et al. (2009) observed a reduction of
188 55% for RO concentrates in wastewater reclamation, at the same O
3
dosage and oxidation
189 time. Although O
3
oxidation is able to efficiently remove unsaturated bonds, it shows only
190 a minor dissolved organic carbon (DOC) removal under acceptable economic conditions.
191 Typical reductions of DOC achieved by ozonation in drinking water plants, with O
3
doses
192 between 2 and 5 mg L
-1
, is only about 10 to 20% (Can and Gurol, 2003).
The DOC

10
193 removal in the experiments of Song et al. (2010) was approximately 10%. Wang and Pai
194 (2004) reported 15% DOC removal, while Lee et al. (2009) observed 5% DOC removal,
195 for the same O
3
dosage and oxidation time as mentioned above.
196
197 Von Gunten (2003) reports that O
3
preferentially oxidizes unsaturated bonds to
198 oxygenated saturated functional groups, such as aldehydic, ketonic and especially
199 carboxylic groups. This can be founded on the results of different spectroscopic
200 techniques. Nuclear magnetic resonance (NMR) spectroscopy by Westerhoff et al. (1999),
201 used for investigating the oxidation of surface waters by O
3
, found a depletion of aromatic
202 against aliphatic moieties. The O
3
consumption was positively correlated with aromatic
203 carbon content, especially electron enriched aromatics, and inversely correlated with
204
205 aliphatic carbon content. Fluorescence spectra of NOM solutions, before and after O
3 oxidation (Świetlik and Sikorska, 2004; Zhang et al., 2008), revealed a reduction of the
206 number of aromatic rings and conjugated bonds, and the decomposition of condensed
207 aromatic moieties to smaller molecules. The number of electron withdrawing groups, such
208 as carboxyl, carbonyl, hydroxyl, alkoxyl and amino groups, increased during ozonation.
209 Mass spectrometry analysis of Suwannee River fulvic acids, by These and Reemtsma
210 (2005), showed that O
3
removes preferentially molecules with a low oxidation state (low
211 O/C ratio) and a high degree of unsaturation (low H/C ratio). They also observed that
212 molecules with a more extended carbon skeleton and less carboxylate substituents showed
213 higher reactivity, whereas some highly unsaturated molecules did not show measurable
214 removal up to a specific O
3
dose of 2.5 mg per mg DOC. The reaction products were
215 characterized by a very high number of carboxylate groups, i.e. the O/C ratio increased
216 from 0.2 to 0.7.
217

11
218 Ozonation products can contain generally alcoholic, carbonyl and carboxyl groups
219 (von Gunten, 2003(a)). The main reaction products after ozonation consist mainly of
220 short-chain (< C5) carboxylic acids, such as formic, acetic and particularly oxalic acid,
221 and aldehydes, such as formaldehyde, acetaldehyde, glyoxal and methylglyoxal (Xiong et
222 al., 1992; Nawrocki et al., 2003; Hammes et al., 2006; Wert et al., 2007). Oxalic acid is is
223 formed mainly by the destruction of aromatic rings by O
3
(Kusakabe et al., 1990). The
224 amounts of carboxylic acids generated upon ozonation are usually much higher, i.e.
225 approximately one order of magnitude, than those of aldehydes and ketones (Nawrocki et
226 al., 2003; Xie, 2004). Can and Gurol (2003) observed that a high O
3
dose results in a
227 decline in the concentration of aldehydes, due to their oxidation to carboxylic acids. The
228 saturated reaction products accumulate in the solution and are not mineralized, even after
229 long oxidation times (Oh et al., 2003; von Gunten, 2003(a); Van Geluwe et al., 2010(a),
230 Van Geluwe et al., 2010(b)). The inefficient reaction between O
3
and the saturated
231 reaction products can be demonstrated by the low rate constants between O
3
and these
232 molecules, which are given in Table 2. With the exception of formate, which reacts
233 relatively well with O
3
, the rate constants range between 10 -5 and 10 1 M -1 s -1 , while rate
234 constants with olefins and aromatic rings can achieve values of 10
6
and 10
9
M
-1
s
-1
,
235 respectively (Williamson and Cvetanovic, 1970; Hoigné and Bader, 1983(b)). This
236 explains why O
3
oxidation can only achieve a small DOC removal, as mentioned above.
237
238 The vast abundance of unsaturated bonds in humic substances facilitates the
239 efficient decomposition of these compounds by O
3
(Van Geluwe et al., 2009; Van Geluwe
240 et al., 2010(a)). The unsaturated bonds in these molecules are transformed to oxygenated
241 saturated bonds. This is schematically represented in Figure 1, where a model structure of
242 a humic acid molecule is drawn, before and after O
3
oxidation. The decomposition of

12
243 proteins and polysaccharides by O
3
was investigated by Cataldo (2003) and Wang et al.
244 (1999), respectively. Cataldo (2003) found that only the aromatic amino acids tryptophan,
245 tyrosine and phenylalanine are oxidized, as well as cysteine. The reaction schemes
246 between these amino acids and O
3
are presented in Figure 3. The most reactive amino acid
247 is tryptophan, followed by tyrosine while phenylalanine appears much less reactive
248 towards O
3
. Histidine and methionine should probably react quite well with O
3
, but there
249 is no direct evidence of this claim. Concerning cysteine, O
3
oxidates the thiol group, with
250 the consequent formation of disulfide bonds and crosslinks between proteins containing
251 cysteine residues. The polyamide bond of the main chain of the protein is not degraded by
252 the action of O
3
, even after prolonged exposure. However, O
3
causes denaturation of
253 proteins, i.e. introduces changes in their secondary and tertiary structure (Cataldo, 2003).
254
255
Wang et al. (1999) showed that O
3
depolymerizes polysaccharides by reacting with the glycosidic linkages in those molecules. It selectively oxidates

-D-glycosidic linkages to
256
257 aldonic esters, as shown in Figure 4. Ozonolysis proceeds under strong stereo-electric control. However, the oxidation of

-D-glycosidic linkages during O
3
oxidation is also
258 possible, but slow, and is caused by several side reactions with radicals and acid
259 hydrolysis.
260
261 3.2 Which functional groups in the NOM can act as a promoter or inhibitor of O
3
262 decomposition?
263 O
3
may react directly with dissolved substances in water, or it may decompose to
264 form radical species, which themselves react with these substances. This corresponds to
265 direct oxidation by molecular O
3
and indirect oxidation, respectively. The most important
266 radical species is the .
OH radical, because of its high standard reduction potential (2.80
267 volts), which is even higher than the standard reduction potential of O
3
(2.07 volts)

13
268 (Beltrán, 2004). In contrast to O
3
, .
OH is believed to be a non-selective oxidant, which
269 reacts very fast with the vast majority of organic and inorganic compounds in water (von
270 Gunten, 2003(a)).
271
272
273
274
The decomposition of O
3
in pure water is well-described in literature. The SBH
(Staehelin, Bühler, Hoigné) model is widely used for predicting the lifetime of O
3
in natural waters (Beltrán, 2004). This model is represented in Figure 5 and a brief discussion
275 is given below.
276
277
278
The decomposition of O
3
in water is a radical-type chain reaction, where various solutes can act as initiators, promoters or inhibitors (Staehelin and Hoigné, 1985).
279
280
Initiation step: The decomposition of O
3
is initiated by OH
-
ions (reaction 1), and this leads to the formation of one superoxide anion ( .
O
2
-
) and one hydroperoxyl radical
281 ( .
HO
2
), which are in an acid-base equilibrium (pK a
= 4.8):
282 O
3
+ OH
.
O
2
-
+ .
HO
2
(k = 70 M
-1
s
-1
) (1)
283 In addition, the reaction of unsaturated bonds in NOM with O
3
can lead to the
284 consumption of O
3
(reaction 2), or the production of an ozonide ion radical ( .
O
3
-
) by an
285 electron transfer reaction (reaction 3):
286
287
O
3
+ NOM
O
3
+ NOM
NOM ox
.
NOM
+
+ .
O
3
-
(2)
(3)
288 The direct reactions (2) and (3) between O
3
and NOM are generally attributed to double
289 bonds, electron-rich aromatic systems, amines and sulphides (von Gunten, 2003(a)). These
290 direct reactions control the decomposition of O
3
during the initial phase of ozonation (t <
291 20 s), in which very high yields of .
OH radicals are generated, i.e. the concentration .
OH is
292 about 10
-6
-10
-8
times the concentration of O
3
(von Gunten, 2003(a)), and the presence of

14
293 radical scavengers does not exert any significant effect on the O
3
consumption (Xiong and
294 Legube, 1991). During the second phase (t > 20 s), when the most reactive moieties of
295 NOM have already reacted with O
3
, the O
3
decomposition is mostly controlled by radical
296 chain reactions instead of direct reaction with NOM, and the .
OH concentration is about
297 ten times lower than during the initial phase of ozonation (Buffle and von Gunten, 2006).
298
299
300
Propagation step: .
O
2
-
is a highly selective catalyst for the decomposition of O
3
in water. The rate constant with which .
O
2
-
reacts with O
3
molecules is very high and results
301 in the formation of .
O
3
-
:
302 .
O
2
-
+ O
3
.
O
3
-
+ O
2
(k = 1.6 .
10
9
M
-1
s
-1
) (4)
303 .
O
3
decomposes upon protonation into .
OH radicals:
304
305 .
.
O
3
-
+ H
+
HO
3
.
HO
3
(k = 5 .
10
10
M
-1
s
-1
)
.
OH + O
2
(k= 1.4 .
10
5
M
-1
s
-1
)
306
307
Even in solutions where the DOC concentration is low, such as drinking water, the .
OH radicals react with a solute before they encounter another radical (Staehelin and Hoigné,
(5)
(6)
308 1985). Typical rate constants for reactions of .
OH radicals with organic solutes are in the
309 range 10
6
– 10
10
M
-1
s
-1
(Buxton et al., 1988; Lal et al., 1988; Mao et al., 1991).
310
311 Some functional groups present in NOM are known to react with .
OH, and this can
312 lead to the formation of carbon centred radicals (see Figure 5, reaction 7). The reaction of
313 these carbon centred radicals with O
2
(reaction 8) subsequently leads to the elimination of
314 .
HO
2
/ .
O
2
-
in a base catalysed reaction (reaction 9). Due to the high selectivity of .
O
2
-
for
315 O
3
, the conversions of the less ozone-selective .
OH radical into .
O
2
-
promotes the chain
316
317 reaction. This explains how certain functional groups of NOM act as promoters for the O
3 decomposition (Staehelin and Hoigné, 1985).

15
318
319
320
Termination step:
Many other solutes do not liberate .
O
2
-
radicals after reaction with .
OH (reaction
321 10). These substances are called inhibitors or radical scavengers.
322
323
324
The question that remains is “which functional groups of NOM promote the decomposition of O
3
, and which functional groups act as inhibitors?” Primary alcohols
325 (such as methanol, glycol, glycerine, glucose and other carbohydrates) promote the
326
327 decomposition of O
3
, although their direct reaction with O
3
is rather slow, due to the absence of unsaturated bonds (Staehelin and Hoigné, 1985). Aromatic rings can also be
328 regarded as promoters for O
3 decomposition, because the attack of O
3
and .
OH cleave
329 aromatic rings into olefins, which immediately react with O
3
, with the subsequent
330
331
332 formation of H
2
O
2
(Pi et al., 2005). Glyoxalate can act both as a promoter and initiator
(Staehelin and Hoigné, 1985). Alkyl groups and carboxylic acids (with the exception of formic acid, which is a promoter) are inhibitors (Staehelin and Hoigné, 1985). .
OH can
333
334 also subtract a H atom from higher alkyl alcohols than methanol, such as tert-butyl alcohol, that is not in a position

to the -OH group. These reactions are not followed by
335 the formation of .
O
2
-
radicals, so these compounds act as scavengers. The relative
336
337 importance of scavenging to promotion for aliphatic alcohols increases with the length of the alkyl chain in these compounds, relative to the amount of H atoms in a position

to
338 the -OH group (Staehelin and Hoigné, 1985).
339
340 3.3 Reaction mechanisms of NOM with .
OH radicals
341 Although the .
OH radical is thought to be an unselective oxidant, organic
342 molecules with unsaturated bonds react faster with .
OH radicals than saturated C-C bonds

16
343 or C-H bonds (Westerhoff et al., 1999). Reactions of .
OH radicals with olefins and
344 aromatic molecules occur at nearly diffusion-controlled limits (10
10
M
-1 s
-1
), whereas the
345 slowest reaction rates with .
OH radicals were observed for aliphatic polyhalogenated
346 compounds such as haloacetic acids (HAA) and trihalomethanes (THM) (10
7
- 10
8
M
-1 s
-1
)
347 (Buxton et al., 1988; Lal et al., 1988; Mao et al., 1991). .
OH can be regarded as an
348 electrophilic oxidant like O
3
, because electron donating groups, such as hydroxyls and
349 amines, enhance the reactivity of adjacent carbon bonds, while electron withdrawing
350 groups, such as carboxyl groups, lessen its reactivity (Song et al., 2008). A few
351 researchers determined the second rate constant of .
OH reaction with NOM (3.6 .
10
8
M
-1
352 s
-1
(Westerhoff et al., 1999)) or humic acids (1.9 .
10
7
M
-1
s
-1
(Liao and Gurol, 1995), 8.1 .
353 10 8 M -1 s -1 (Westerhoff et al., 1999), 2.3 - 3.2 .
10 9 M -1 s -1 (Goldstone et al., 2002)), at
354 which the concentration of organic matter is expressed as mol C per liter.
355
356 Reactions of .
OH radicals with organic compounds fall into two basic mechanisms,
357 i.e. addition, generally to an aromatic ring (hydroxylation), or abstraction of a H atom.
358 Both mechanisms can lead to the formation of low-molecular mass acids. Hydroxylation
359 of aromatic moieties in NOM, followed by ring opening, can produce both mono- and
360 dicarboxylic acids (Gopalan and Savage, 1994). Abstraction of .
H from an unsaturated C-
361 C bond leads to the formation of a carbon-centred radical, followed by reaction with O
2
to
362 form a peroxyl radical, and subsequent decomposition to a carboxylic acid (Goldstone et
363 al., 2002).
364
365 3.4 The addition of H
2
O
2
slightly improves the mineralisation of organic matter
366 during ozonation.

17
367 The contribution of the radical pathway can strongly be enhanced by adding H
2
O
2
368 to the solution. In this case, the main O
3
decay path is the reaction between O
3
and the
369
370
371 conjugated base of H
2
O
2
, i.e. HO
2
-
(pK a
= 2.24 .
10
-12
(Paillard et al., 1988)), leading to a higher production rate of .
OH radicals (Staehelin and Hoigné, 1982):
O
3
+ HO
2
.
OH + .
O
2
-
+ O
2
(k = 2.8
.
10
6
M
-1 s
-1
) (10)
372 This leads to an advanced oxidation process (AOP). Indeed, in pure water systems, two
373 .
OH radicals are formed per three O
3
molecules consumed, while in the presence of H
2
O
2
,
374 one .
OH radical is formed per O
3
molecule consumed (Acero and von Gunten, 2001). The
375 main advantages of AOPs are the oxidation of ozone-resistant compounds. However, the
376 higher formation rate of .
OH is at the cost of a higher O
3
consumption, i.e. the
377
378 decomposition of O
3
is accelerated in the presence of H
2
O
2
at pH values higher than 5
(Staehelin and Hoigné, 1982). In addition, .
OH reacts rather unselective, and thus only a
379 small fraction of these radicals reacts with the target pollutant, which severely reduces the
380 efficiency of AOPs (von Gunten, 2003(a)).
381
382 Several researchers reported that O
3
+ H
2
O
2
systems better remove DOC, because
383 of the ability of .
OH radicals to react with the unsaturated reaction products formed during
384 O
3
oxidation. The effectiveness of .
OH radicals for the mineralisation of these compounds
385 can be assessed by comparing the rate constants with O
3
and .
OH radicals (see Table 2).
386 The rate constants for the reaction with .
OH radicals is evidently higher than for the
387 reaction with molecular O
3
. However, the concentration of .
OH radicals is typically 10
7
to
388 10
9
times lower than the dissolved O
3
concentration (von Gunten, 2003(a)). If the
389 proportion of the two rate constants is lower than 10
8
, an increase of the .
OH concentration
390 will not be able to improve the decomposition of these saturated compounds. This
391 proportion is calculated for each compound in the last column of Table 2, and it can be

18
392 seen that it is higher than 10 8 for all compounds, with the exception of formate, which
393 already reacts relatively well with O
3
. Goldstone et al. (2002) reported that reactions
394 between humic substances and .
OH radicals produced CO
2
with a high efficiency of ~ 0.3
395 mole CO
2
per mole .
OH. This efficiency stayed approximately constant from the early
396 phases of oxidation until complete mineralization of the DOC, and was not significantly
397 altered within the range of pH values found in natural waters (pH 4 – 10). However, they
398 reported that the CO
2
production from .
OH reaction with low-molecular mass acids is an
399 insignificant fraction of the overall CO
2
formation rate, indicating that most of the
400 mineralization of humic substances does not occur via low-molecular mass acid
401 intermediates.
402
403
404
Acero and von Gunten (2001) reported that, when the organic matter content in the water is low (< 1 mg L
-1
DOC), the addition of H
2
O
2
considerably enhances the oxidation
405 capacity by
.
OH radicals. However, if the NOM concentration in natural waters is higher
406 than 3 mg L
-1
DOC, O
3
decomposition is controlled by the promoting or inhibiting effect
407 of certain functional groups in the DOC on the O
3
decomposition. Therefore, the effect of
408 H
2
O
2
addition in order to raise the
.
OH concentration, will be negligible in that case. This
409 was also observed by the following researchers, who observed that O
3
+ H
2
O
2
performed
410 well, but only slightly better than pure O
3
oxidation. Irabelli et al. (2008) reported for
411 surface water (DOC > 3 mg L
-1
) a DOC removal of 61% by O
3
+ H
2
O
2
(0.7 mg H
2
O
2
was
412 added prior to oxidation) compared to 53% by O
3
(O
3
dosage: 2.0 mg L
-1
, reaction time:
413 13.2 minutes). Van Geluwe et al. (2010(a)) found that the continuous addition of H
2
O
2
414 between 10 and 30 minutes oxidation time (0.25 mole H
2
O
2 per mole O
3
), could enhance
415 the removal of UV absorbing compounds at 280 nm, from 42% (if no H
2
O
2
was added) to
416 54%, for the treatment of humic acid solutions. If the same experimental conditions were

19
417 applied to NF concentrates, obtained after filtration of surface water, a reduction of 74%
418 could be achieved by O
3
+ H
2
O
2
, compared to 68% by O
3
(Van Geluwe et al., 2010(b)).
419
420 3.5 Guidelines for finding the optimal dose of H
2
O
2
in water treatment
421 The dosage of H
2
O
2
have a significant effect on the mineralization of organic
422 compounds. There exist an optimal H
2
O
2
dosage because this compound acts not only as a
423 generator of .
OH radicals, but also as a scavenger of .
OH radicals at high H
2
O
2
424 concentrations. This is attributed to the formation of much less powerful HO
2
.
radicals by
425 the reaction between H
2
O
2
and .
OH (reaction 11), and the subsequent reaction of HO
2
.
426 with .
OH (reactions 12 and 13)(Weinstein and Bielski, 1979; Buxton et al., 1988):
427
428
429
H
2
O
HO
2
.
2
2 HO
+
+
2
.
.
.
OH
OH
H
2
O + HO
2
.
(k = 2.7 .
10 7 M -1 s -1 )
H
2
O + O
2
(k = 6.0
.
10
9
M
-1
s
-1
)
H
2
O
2
+ O
2
(k = 7.6
.
10
5
M
-1
s
-1
)
(11)
(12)
(13)
430
431 Multiple addition of H
2
O
2
at different stages is better for DOC removal than single
432 addition of H
2
O
2
only at the inlet of the reactor (Kosaka et al., 2001). Higher H
2
O
2
doses
433 are required at later stages, in which low reactivity compounds with .
OH radicals are
434 predominant (Kosaka et al., 2001). It is also important to maintain a low concentration of
435 dissolved O
3
for efficient DOC removal, because O
3
consumption is required for the rapid
436 formation of .
OH radicals (Kosaka et al., 2001). Once H
2
O
2
is completely consumed, the
437 AOP comes to a halt, achieving a similar COD conversion if compared to pure O
3
438 oxidation (Rivas et al., 2009)
439
440 The optimal H
2
O
2
/O
3
ratio is dependent on the composition of the water solution,
441 such as the type and concentration of the solutes (Kosaka et al., 2001), and the alkalinity

20
442 of the water (Rosal et al., 2009). In most full-scale plants, the optimal ratio of the H
2
O
2
443 dosage to the O
3
dosage, appeared to lie between 0.25 and 1.0. There are no formulas
444 available yet for predicting the optimal value, so it should be determined experimentally
445 for each raw water and O
3
installation (Acero and von Gunten, 2001). Paillard et al. (1988)
446 determined the optimal conditions for removal of oxalic acid by O
3
+ H
2
O
2
. Oxalic acid is
447 a major reaction product that accumulates during O
3
oxidation of natural waters. The
448 continuous introduction of H
2
O
2
improved the degradation of oxalic acid, compared to
449 single addition of H
2
O
2
at the start of the experiment. The optimal pH was equal to 7.5 and
450 the optimal dosage of H
2
O
2
ranged between 0.6 and 0.7 .
10
-4
M, for an initial oxalic acid
451 concentration of 2 . 10
-4
M. This corresponded to a ratio of 0.5 mole H
2
O
2
consumed per
452 mole O
3
. The presence of bicarbonates did not influence the optimal conditions, but if the
453 solution contained bicarbonates, a significant reduction of the degradation rate for oxalic
454 acid was observed.
455
456 3.6
A short note on the health hazard of O
3
in water treatment
457 3.6.1.
Bromate formation
458
459
The formation of the suspected carcinogenic bromate (BrO
3
-
) in waters containing
Br
-
is a serious concern in O
3
oxidation. Br
-
is a constituent of all natural waters and its
460 concentration depends mainly on the geochemistry of the watershed (Magazinovic et al.,
461
2004). It ranges between 10 and 1000 µg L -1
(von Gunten, 2003(b)). Currently, the U.S.
462 Environmental Protection Agency (USEPA) and the European Union have set a maximum
463
464 contaminant level of 10 µg L -1
for BrO
3
-
in drinking water (Agus et al., 2009). At Br
concentrations as low as 50 µg L -1
, excessive BrO
3
-
formation can already become a
465 problem. Details about the formation mechanisms of bromate and strategies to control it
466 can be found in von Gunten (2003(b)).

21
467
A reduction of the O
3
dose by two could reduce the formation of BrO
3
-
by a factor 468
469 of more than ten (Meunier et al., 2006). However, the reduction of O
3
doses decreases the
470 inactivation of resistant protozoa, such as Cryptosporidium parvum and Giardia lamblia .
471 The lower residual O
3
concentration in AOPs and the reduction of HOBr by H
2
O
2
reduce
472 BrO
3
formation in O
3
-based AOPs in comparison to conventional O
3
oxidation (von
473 Gunten, 2003(b)). BrO
3
-
formation is noticeably higher in high-alkalinity waters than in
474 waters with lower alkalinity. This is due to the reaction of carbonate radicals with OBr
-
.
475 Therefore, the application of O
3
oxidation has to be carefully evaluated for high-alkalinity
476 waters to avoid excessive BrO
3
-
formation (von Gunten, 2003(b)). Another decisive factor
477 is the temperature of the treated water: BrO
3
formation increases with increasing
478 temperature (von Gunten, 2003(b)).
479
480 3.6.2. Trihalomethanes and haloacetic acids
481 The reaction of NOM with chlorine causes the formation of a variety of potentially
482 carcinogenic DBPs such as THMs and HAAs. The USEPA has currently set maximum
483
484 contamination levels of total THMs and HAAs at 80 µg L -1
and 60 µg L
-1
, respectively.
The European Union has only a limit for THMs, namely 100 µg L -1
(see Agus et al., 2009
485 for more details).
486
487 The hydrophobic NOM fraction is the main precursor of DBP formation, because it
488 has the highest tendency to react with chlorine (Chang et al., 2002). In section 4.1, it will
489 be shown that O
3
is able to transform these hydrophobic components into hydrophilic
490
491 products, so that O
3
oxidation will decrease the formation potential of THMs and HAAs quite efficiently (Chang et al., 2002; Gallard and von Gunten, 2002; Chin and Bérubé,

22
492 2005; Li et al., 2008). Kleiser and Frimmel (2000) and Meunier et al. (2006) found that
493 even a small O
3
dose of 1.5 and 2.5 mg per mg DOC respectively, reduced the THM
494 formation potential by about 70% in surface water, whereas almost no DOC was removed.
495

23
496 THE STRUCTURAL CHANGES OF NOM BY O
3
OXIDATION
497
498 NOM fouling is attributed to the accumulation of molecules which are retained on
499 the membrane surface, forming a cake or gel layer, and the adsorption of non-retained
500 molecules in the inner pores of the membrane, leading to constriction and blocking of the
501 inner pores of the membrane. The fouling potential of NOM is defined by different types
502 of chemical and physical interactions between the NOM and the membrane surface, such
503 as hydrophobic interactions, hydrogen bridges and electrostatic interactions, besides the
504 size and conformation of the NOM molecules. NOM fouling in membrane filtration is
505 already reviewed by several authors: Zularisam et al., 2006; Al-Amoudi and Lovitt, 2007;
506 Amy, 2008; Al-Amoudi, 2010; Sutzkover-Gutman et al., 2010. Therefore, this section will
507 only discuss the effect of O
3
oxidation on the different fouling mechanisms. A summary of
508 recent studies on the effect of O
3
oxidation on membrane fouling, both in surface water
509 applications and municipal wastewater treatment, is given in Table 3.
510
511 4.1 Hydrophobic interactions between NOM and membrane surfaces
512 The electrophilic nature of O
3
results in the transformation of unsaturated bonds in
513 aromatic moieties, into hydrophilic reaction products, such as carboxylic acids, as stated in
514 the previous section. This could be confirmed by several researchers. Świetlik et al. (2004)
515 treated sand-filtered groundwater with 1.0 mg O
3
per mg DOC, for 10 minutes, and
516 observed that the hydrophobic DOC (i.e. the DOC fraction retained by the Amberlite
517 XAD 8 resin or a similar resin) decreased from 54 to 5% of the total DOC. Song et al.
518 (2010) reported that a O
3
dose of 0.5 mg O
3
per mg DOC, applied for 10 minutes, to
519 surface water, reduced the hydrophobic DOC by ~ 45%, while the total DOC dropped by
520 only 10%. Van Geluwe et al. (2010(b)) showed that the hydrophobic fraction of the

24
521 chemical oxygen demand (COD) decreased by 86%, while the overall COD decreased by
522 only 22%, while treating the concentrate stream after NF of surface water, with 2.0 mg O
3
523 per mg COD for 20 minutes. The transformation of the hydrophobic part of NOM into
524 hydrophilic reaction products is schematically shown in Figure 1, where a model structure
525 of a humic acid molecule is drawn, before and after O
3
oxidation. Most of the aromatic
526 rings in the humic acid molecule are cleaved by O
3
, and the molecule is enriched in
527 oxygenated functional groups. The hydrophilic reaction products have a lower propensity
528 for adsorption on membrane surfaces. This is an important reason for the increase of the
529 membrane flux after O
3
oxidation.
530
531
532 4.2 Molecular size and (dis)aggregation of NOM
533 4.2.1 Results obtained by high-performance size exclusion chromatography (HPSEC)
534 The size of NOM is a critical parameter for membrane fouling. The occurrence of
535 different modes of fouling is related to foulants size relative to membrane pore size
536 (Katsoufidou et al., 2005; Huang et al., 2008; Lin et al., 2009). Substances close to the size
537 of the membrane pores can cause pore blocking, which provokes a severe increase of the
538 filtration resistance (Lohwacharin and Takizawa, 2008). This is the case for the high-
539 molecular mass molecules, such as polysaccharides and proteins, constricting and
540 blocking the pores in loose membranes (MF and UF). Substances much larger than the
541 membrane pores lead to cake formation, that is more readily removed. This is the case for
542 the vast majority of the NOM components with tight membranes (NF and RO). In this
543 case, substances with a high molecular mass have a small back diffusion rate during cross-
544 flow filtration, so that a thick cake or gel layer is formed on the membrane surface (Lee et
545 al., 2004).

25
546
547 Various researchers observed that ozonation is able to decompose NOM molecules
548 into smaller fragments. HPSEC analysis before and after ozonation, shows a clear shift
549 toward longer retention times, corresponding to a decrease in the average molecular mass.
550 This is observed for synthetic solution of humic substances (Kerc et al., 2004; These and
551 Reemtsma, 2005; Van Geluwe et al., 2009), as well as for natural waters and wastewater
552 effluents (Nissinen et al., 2001; Nawrocki et al., 2003; Lehman and Liu, 2009, Song et al.,
553 2009). Similar observations were made in studies where the NOM reacted with
.
OH
554 radicals instead of molecular O
3
(Le-Clech et al., 2006; Liu et al., 2008). The cross-flow
555 operation can flush away loose fragments of the cake layer. This decreases the thickness
556 of the foulant layer and thus flux decline (Bhavana et al., 2005; Lee et al., 2005(b)).
557
558 Jansen et al. (2006) investigated the molecular mass distribution of concentrated
559 solutions of humic substances with a low polydispersity (1.23) during O
3
oxidation. The
560 average molecular mass shifted from 2600 to 1500 g mol
-1
and the polydispersity of the
561 peak decreased remarkably from 1.23 to 1.09, after a O
3
consumption of 0.65 mol O
3
per
562 mol C. Another peak emerged at longer retention times, because small compounds were
563 formed during the ozonation process. The size of the reaction products was restricted to
564 only two distinct peaks throughout the reaction. The shape of the original peak did not
565 broaden toward longer elution times in almost all experiments. This indicates that no
566 random splitting of the humic molecules occurs, since that would result in a very broad
567 range of molecular sizes. Only small molecular fragments are split off from the periphery
568 of the larger humic molecules. The main structure reduces size during ozonation, but
569 remains intact. Cleavage occurs mainly on the periphery of the molecules. Steric
570 obstruction related to the coiled structure of the humic substances, prevents O
3
and .
OH

26
571 radicals from cleaving bonds in the core of the molecules (Jansen et al., 2006; Nawrocki,
572 2007).
573
574 4.2.2. Aggregation of humic substances by calcium and magnesium
575 The size of the humic substances is also influenced by aggregation/disaggregation
576 forces, which in turn are dependent on the pH of the solution. Humic substances have been
577 regarded for a long time as polymeric molecules, having a relatively high molecular mass.
578 Molecular masses of several 10,000 g mol
-1
have been reported (Sutton and Sposito,
579 2005). Nowadays, humic substances are regarded as supramolecular associations of many,
580 relatively small (<1000 g mol
-1
) and chemically diverse bio-organic molecules, that are
581 hold together by different kinds of weak interactions, such as Van der Waals forces and
582 hydrogen bridges (Piccolo, 2002; Sutton and Sposito, 2005; Ahn et al., 2008). Therefore,
583 it is appropriate to study the aggregation kinetics of humic substances, and the effect of O
3
584 oxidation on it.
585
586 Aggregation is promoted by the presence of divalent inorganic cations, such as
587 calcium and magnesium. These ions can bind to carboxylate groups present in certain
588 fractions of the NOM (Hong and Elimelech, 1997; Ahn et al., 2008). Charge screening by
589 the adsorbed cations decreases the net negative charge of the NOM molecules, and thus
590 the electrostatic repulsions with the membrane surface (Engebretson and von Wandruszka,
591 1998). This phenomena can be observed even at low concentrations of divalent ions, i.e.
592 lower than 0.5 .
10
-3
M (Shao et al., 2010).
593
594 If the concentration of divalent ions is higher than 0.5 .
10 -3 M, these cations act as
595 bridges between carboxylic groups of two adjacent molecules, increasing the attractive

27
596 forces between them (Shao et al., 2010). This bridging function of divalent cations has
597 been proven with atomic force microscopy (AFM) (Li and Elimelech, 2004; Nguyen et al.,
598 2007). This phenomena was observed for all NOM fractions, but was especially
599 significant in the presence of hydrophobic NOM, i.e. humic substances (Makdissy et al.,
600 2002). Due to the formation of complexes between NOM and divalent ions, the size of the
601 humic aggregates becomes larger (Shao et al., 2010). It is important to note that these
602 conclusions are all obtained without the effect of O
3
treatment on NOM. The impact of
603 divalent ions on the complexation of ozonated NOM is not investigated yet, but it is an
604 important issue for further study.
605
606 Presently, there is little agreement on the effect of complexation on membrane
607 fouling. Shao et al. (2010) stated that the porous character of the flocculated aggregates
608
609 enhances the permeability of the cake layer formed on the membrane surface, and a decreased flux decline was observed (see Figure 9). The aggregate porosity (
 agg
) is related
610 to the primary particle radius (r prim
), aggregate radius (r agg
), and the fractal dimension D
611 (Jiang and Logan, 1991):
612

1 agg
 r agg r prim


D

3

613 Fractial dimensions are related to the aggregate structure, e.g. D = 3 for a sphere, D = 2 for
614 a flat sheet, D ≈ 1.8 – 2.4 for typical aggregates (Waite et al., 1999). The equation shows
615 that, for any other shape than perfect spheres, the aggregate porosity increases as r agg
/ r prim
616 becomes larger.
617
618 In addition, the aggregates can act as a secondary membrane, which removes a
619 significant portion of the small particles, before they reach the polymeric membrane

28
620 surface. The removal of the small particles prevents the obstruction of membrane pores or
621 the formation of a tightly packed cake layer on the membrane surface (Kuberkar and
622 Davis, 2000). Membrane fouling will be strongly alleviated if NOM aggregation causes a
623 change of the fouling mechanisms, i.e. from internal pore constriction to the formation of
624 a permeable cake on the membrane surface (Lohwacharin and Takizawa, 2008).
625
626 Other researchers said that the complexation causes severe flux decline, because the cross-
627 linked aggregates form a dense, compact cake layer on the membrane surface, with a high
628
629 fouling resistance (Hong et al. and Elimelech, 1997, Yuan and Zydney, 1999; Cho et al.,
1999; Cho et al., 2000; Schäfer et al., 2002; Costa et al., 2006; Tang et al., 2007;
630 Katsoufidou et al., 2008, Mattaraj et al., 2008 ; Zhu et al., 2010), or internal deposition of
631 aggregates reduces the pore size of the membrane (Aoustin et al., 2001).
632
633 4.3 Electrostatic interactions and hydrogen bridges
634 The surface of many membranes carries a negative charge at neutral pH, because
635 of the presence of carboxylic and amine groups (Childress and Elimelech, 1996).
636 Electrostatic repulsions between the negatively charged groups in NOM and the
637 membrane surface decreases the fouling tendency. For instance, Shao et al. (2010)
638 observed a lower flux decline when humic acids were filtered by a charged regenarated
639 cellulose membrane, in comparison to a neutral membrane of the same material.
640
641 Due to the absence of ionisable groups in polysaccharides, electrostatic interactions
642 do not play a role for these molecules. Proteins are nearly all negatively charged at neutral
643 pH, because of the preponderance of weakly acid residues in almost all proteins. The IEP
644 of natural proteins lies in many cases between 5 and 6 (Scopes, 1994). The presence of

29
645 carboxylic (pK a
~ 3 - 5) and phenolic (pK a
~ 8 - 11) acid groups in humic substances,
646 make them anionic at neutral pH (Buffle, 1991; Milne et al., 2001). The mole fractions of
647 either carboxylic or phenolic sites in humic substances are usually markedly unequal, with
648 carboxylic acids typically dominating (Buffle, 1991). For instance, Suwannee River humic
649 acids contain 7.85 mole carboxylic groups per kg (dry weight), while only 1.86 mole
650 phenolic groups per kg (IHSS, 2001).
651
652 The number of carbonyl and carboxyl groups increases after O
3
oxidation, as stated
653 previously. Carbonyl compounds carry a negative charge at neutral pH, which is repulsed
654 by the negatively charged membrane. However, carboxyl groups can form strong
655 hydrogen bridges with hydrophilic membranes, such as polyamide and regenerated
656 cellulose membranes.
657
658 Van Geluwe et al. (2010(b)) took Fourier transform infrared (FTIR) spectra of the
659 NF 270 membrane, after 40 hours filtration of concentrated surface water. He observed
660 that carbonyl compounds were able to adsorb onto the membrane, if the solution was
661 treated with O
3
(O
3
dosage: 1.0 mg O
3
per mg COD; oxidation time: 10 minutes)).
662 However, the membrane had a significant lower fouling rate after O
3
oxidation, although
663 the COD of the solution did not decrease. The mean water flux, during the first 40 hours
664 filtration, increased from 73 to 88% of the pure water flux.
665
666 An estimation of the strength of these hydrogen bridges is provided by Nie et al.
667 (2005). They calculated the strength of hydrogen bridges between the carboxyl group (of
668 butyric acid) and methanol, via ab initio methods. The methanol molecule can be
669 considered as an alcoholic group on the membrane surface. Table 4 shows the dissociation

30
670 energies of several hydrogen bridges, with water as solvent. The notation, H O -C=O,
671 represents that the hydroxyl O atom of the carboxyl group forms a hydrogen bridge with
672 an alcoholic group on the membrane surface.
673
674 The strength of the hydrogen bridges is compared to the electrostatic repulsion
675 energy between two negative charges, e.g. a carboxylate group in the NOM and a negative
676 charge on the membrane surface. The calculated hydrogen bridge strengths strongly
677 depend on the specific type of hydrogen bridgeing interactions. The formation of
678 hydrogen bridges between carboxyl groups and alcoholic groups are all exothermic, with
679 the exception of H O -C=O. The hydrogen bridges are relatively weak, and weaker than
680 those formed between water molecules (11 kJ mol -1 ). The strongest hydrogen bridges are
681 formed with the acid hydrogen of the carboxyl group ( H O-C=O). This type of bond does
682 not occur frequently at neutral pH, because most carboxyl groups are dissociated at this
683 pH. The other hydrogen bridges are weaker than the electrostatic repulsion forces between
684 two elementary negative charges. These results show that the hydrogen bridges between
685 the carboxylate groups in the NOM, formed during O
3
oxidation, with the membrane
686 surface are in most cases relatively weak, and are of comparable strength as the
687 electrostatic repulsion forces between charges with the same sign. If the increase of the
688 number of carboxylic groups by O
3
oxidation promotes the adsorption of NOM on the
689 membrane surface or not, is an open question at this moment, and it would be interesting
690 to study this in the future.
691
692

31
693 CONCLUSION
694
695 The electrophilic character of O
3
accounts for the fast reaction of O
3
molecules
696
697 with unsaturated bonds. The possible reaction pathways between NOM and O
3
were described within the framework of the SBH model (Staehelin, Bühler, Hoigné). The direct
698 reaction of unsaturated bonds in NOM with O
3
can lead to the consumption of O
3
, or the
699 production of an ozonide ion radical ( .
O
3
-
), which decomposes upon protonation into a
700 hydroxyl ( .
OH) radical. The .
OH radical is a strong oxidizing agents that can react with
701 NOM molecules. This is referred to as the indirect oxidation pathway. Although the .
OH
702 radical is thought to be an unselective oxidant, it can be regarded as electrophilic oxidant.
703
704 These direct reactions between NOM and O
3
control the decomposition of O
3
705 during the initial phase of ozonation (t < 20 s), in which very high yields of .
OH are
706 generated, i.e. the concentration .
OH is about 10
-6
-10
-8
times the concentration of O
3
, and
707 the presence of radical scavengers does not exert any significant effect on the O
3
708 consumption. During the second phase (t > 20 s), the most reactive moieties of NOM have
709 reacted with O
3
, so that O
3
decomposition is mostly controlled by a radical chain reaction,
710 and the .
OH concentration is about ten times lower than during the initial phase. During
711 the second phase, some functional groups in the NOM, such as aromatic rings and primary
712 alcohols (abundant in polysaccharides), act as promoters for the decomposition of O
3
.
713 Alkyl groups and carboxylic acids act as scavengers of .
OH radicals.
714
715
716
Although O
3
oxidation without addition of H
2
O
2
is not considered as an AOP, the concentration of .
OH radicals during ozonation of NOM solutions, can be as high as in
717 AOPs at neutral pH (>> 10
-12
M). The contribution of the radical pathway can strongly be

32
731
732
733
734
735
736
737
738
718 enhanced by adding H
2
O
2
to the solution. The addition of H
2
O
2
improves the
719 mineralisation of the saturated reaction products, mostly carboxylic acids, that are formed
720 during O
3
oxidation.
721
722 Several researchers proved that the application of O
3
oxidation of the feed water prior
723 to membrane filtration, resulted in a significant decrease in membrane fouling, although
724 only a minor DOC removal (10 – 20%) could be achieved. This is explained by the fact
725 that O
3
causes substantial structural changes to the NOM present in the feed water, of
726 which the most important are:
727 an efficient removal of aromatic rings, which decreases the adsorption of NOM by
728 hydrophobic interactions
a significant increase of the number of carboxylic functions, which are repelled by 729
730 the negative membrane surface. These repulsion forces have a comparable strength as the hydrogen bridges that carboxylic groups can form with the membrane surface.
decomposition of molecules into smaller fragments, whereby small molecular fragments are split off from the periphery of the larger molecules that remain intact.
a higher propensity for complexation of humic substances with divalent ions, if the concentration of divalent ions is larger than 0.5 .
10
-3
M. However, it is important to note that this figure is obtained with non-ozonated NOM.

33
739
740
741
ACKNOWLEDGMENTS
Steven Van Geluwe is grateful to IWT-Vlaanderen (Institute for the Promotion of
742 Innovation by Science and Technology in Flanders) for providing a fellowship.
743

34
744 REFERENCES
745
746 Acero, J.L., von Gunten, U., 2001. Characterization of oxidation processes: ozonation and
747 the AOP O
3
/H
2
O
2
. J. Am. Water Works Ass. 93(10), 90-100.
748
749 Agus, E., Voutchkov, N., Sedlak, D.L., 2009. Disinfection by-products and their potential
750 impact on the quality of water produced by desalination systems: A literature review.
751 Desalination 237, 214-237.
752
753 Ahn, W.Y., Kalinichev, A.G., Clark, M.M., 2008. Effects of background cations on the
754 fouling of polyethersulfone membranes by natural organic matter: Experimental and
755 molecular modeling study. J. Membr. Sci. 309, 128-140.
756
757 Al-Amoudi, A.S., 2010. Factors affecting natural organic matter (NOM) and scaling
758 fouling in NF membranes: A review. Desalination 259, 1-10.
759
760 Al-Amoudi, A., Lovitt, R.W., 2007. Fouling strategies and the cleaning system of NF
761 membranes and factors affecting cleaning efficiency. J. Membrane Sci. 303, 4-28.
762
763 Amy, G., 2008. Fundamental understanding of organic matter fouling of membranes.
764 Desalination 231, 44-51.
765
766
Aoustin, E., Schäfer, A.I., Fane, A.G., Waite, T.D., 2001. Ultrafiltration of natural organic
767 matter. Sep. Purif. Technol. 22-23, 63-78.
768

35
769 Beltrán, F.J., 2004. Ozone reaction kinetics for water and wastewater systems, CRC Press,
770 Boca Raton (FL, USA).
771
772 Buffle, J., 1991. Complexation reactions in aquatic systems: an analytical approach, Ellis
773 Horwood, Chichester (United Kingdom).
774
775 Buffle, M.O., von Gunten, U., 2006. Phenols and amine induced HO .
generation during
776 the initial phase of natural water ozonation. Environ. Sci. Technol. 40, 3057-3063.
777
778 Buxton, G.V., Greenstock, C.L., Helman, W.P., Ross, A.B., 1988. Critical review of data
779 constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals in
780 aqueous solutions. J. Phys. Chem. Ref. Data 17, 513-886.
781
782 Can, Z.S., Gurol, M., 2003. Formaldehyde formation during ozonation of drinking water.
783 Ozone Sci. Engin. 25(1), 41-51.
784
785 Cataldo, F., 2003. On the action of ozone on proteins. Polym. Degrad. Stabil. 82, 105-114.
786
787 Chang, C.N., Ma, Y.S., Zing, F.F., 2002. Reducing the formation of disinfection by-
788 products by pre-ozonation. Chemosphere 46, 21-30.
789
790 Childress, A.E., Elimelech, M., 1996. Effect of solution chemistry on the surface charge of
791 polymeric reverse osmosis and nanofiltration membranes. J. Membrane Sci. 119, 253-268.
792

36
793 Chin, A., Bérubé, P.R., 2005. Removal of disinfection by-product precursors with ozone-
794 UV advanced oxidation processes. Water Res. 39, 2136-2144.
795
796 Cho, J., Amy, G., Pellegrino, J., 1999. Membrane filtration of natural organic matter:
797 initial comparison of rejection and flux decline characteristics with ultrafiltration and
798 nanofiltration membranes. Water Res. 33, 2517-2526.
799
800 Cho, J., Amy, G., Pellegrino, J., 2000. Membrane filtration of natural organic matter:
801 factors and mechanisms affecting rejection and flux decline with charged ultrafiltration
802 (UF) membrane. J. Membrane Sci. 164, 89-110.
803
804 Costa, A.R., de Pinho, M.N., Elimelech, M, 2006. Mechanisms of colloidal natural
805 organic matter fouling in ultrafiltration. J. Membr Sci. 281, 716-725.
806
807 Darwish, M.A., Al-Najem, N.M., 2000. Energy consumption by multi-stage flash
808 evaporation and reverse osmosis desalters. Appl. Therm. Eng. 20, 399-416.
809
810 de la Rubia, A., Rodríguez, M., León, V.M., Prats, D., 2008. Removal of natural organic
811 matter and THM formation potential by ultra- and nanofiltration of surface water. Water
812 Res. 42, 714-722.
813
814 Engebretson, R.R., von Wandruszka, R., 1998. Kinetic aspects of cation-enhanced
815 aggregation in aqueous humic acids. Environ. Sci. Technol. 32, 488-493.
816

37
817 Eltawil, M.A., Zhengming, Z., Yuan, L., 2009. A review of renewable energy
818 technologies integrated with desalination systems. Renew. Sust. Energ. Rev. 13, 2245-
819 2262.
820
821 Frimmel, F.H., Abbt-Braun, G., Heumann, K.G., Hock, B., Luedemann, H.D., 2002.
822 Refractory organic substances in the environment, Wiley VCH, Weinheim (Germany).
823
824 Gallard, H., von Gunten, U., 2002. Chlorination of natural organic matter: kinetics of
825 chlorination and of THM formation. Water Res. 36, 65-74.
826
827 Goldstone, J.V., Pullin, M.J., Bertilsson, S., Voelker, B.M., 2002. Reactions of hydroxyl
828 radical with humic substances: bleaching, mineralization and production of bioavailable
829 carbon substrates. Environ. Sci. Technol. 36, 364-372.
830
831 Gopalan, S., Savage, P.E., 1994. Reaction mechanisms for phenol oxidation in
832 supercritical water. J. Phys. Chem. Phys. 98, 12646-12652.
833
834
Hammes, F., Salhi, E., Köster, O., Kaiser, H.P., Egli, T., von Gunten, U., 2006.
835 Mechanistic and kinetic evaluation of organic disinfection by-product and assimilable
836 organic carbon (AOC) formation during the ozonation of drinking water. Water Res. 40,
837 2275-2286.

38
838
839
Hoigné, J., Bader, H., 1983(b).
Rate constants of reactions of ozone with organic and
840 inorganic compounds in water - II: Dissociating organic compounds. Water Res. 17, 185-
841 194.
842
843 Hong, S., Elimelech, M., 1997. Chemical and physical aspects of natural organic matter
844 (NOM) fouling of nanofiltration membranes. J. Membrane Sci. 132, 159-181.
845
846
Huang, H., Spinette, R., O’Melia, C.R., 2008. Direct-flow microfiltration of aquasols: I.
847 Impacts of particle stabilities and size. J. Membr. Sci. 314, 90-100.
848
849 I.H.S.S. Standard and Reference Collection, International Humic Substance Society, 2001.
850
851 Irabelli, A., Jasim, S., Biswas, N., 2008. Pilot-scale evaluation of ozone vs. peroxone for
852 trihalomethane formation. Ozone Sci. Engin. 30, 356-366.
853
854 Jansen, S.A., Malaty, M., Nwabara, S., Johnson, E., Ghabbour, E., Davies, G., Varnum,
855 J.M., 1996. Structural modeling in humic acids. Mater. Sci. Eng. C 4, 175-179.
856
857 Jansen, R.H.S., Zwijnenburg, A., van der Meer, W.G.J., Wessling, M., 2006. Outside-in
858 trimming of humic substances during ozonation in a membrane contactor. Environ. Sci.
859 Technol. 40, 6460-6465.
860

39
861 Jansen, R.H.S., Zwijnenburg, A., van der Meer, W.G.J., Wessling, M., 2006. Response to
862
863 comment on “Outside-in trimming of humic substances during ozonation in a membrane contactor.” Environ. Sci. Technol. 41, 5162-5164.
864
865 Jiang, Q., Logan, B.E., 1991. Fractal dimensions of aggregates determined from steady-
866 state size distributions, Environ. Sci. Technol. 25, 2031-2038.
867
868 Karnik, B.S., Davies, S.H.R., Chen, K.C., Jaglowski, D.R., Baumann, M.J., Masten, S.J.,
869 2005. Effects of ozonation on the permeate flux of nanocrystalline ceramic membranes.
870 Water Res. 39, 728-734.
871
872 Katsoufidou, K., Yiantsios, S.G., Karabelas, A.J., 2005. A study of ultrafiltration
873 membrane fouling by humic acids and flux recovery by backwashing: experiments and
874 modeling. J. Membr. Sci. 266, 40-50.
875
876 Katsoufidou, K., Yiantsios, S.G., Karabelas, A.J., 2008. An experimental study of UF
877 membrane fouling by humic acid and sodium alginate solutions: the effect of backwashing
878 on flux recovery. Desalination 220, 214-227.
879
880 Kennedy, M.D., Chun, H.K., Quintanilla Yangali, V.A., Heijman, B.G.J., Schippers, J.C.,
881 2005. Natural organic matter (NOM) fouling of ultrafiltration membranes: fractionation of
882 NOM in surface water and characterisation by LC-OCD. Desalination 178, 73-83.
883
884 Kerc, A., Bekbolet, M., Saatci, A.M., 2004. Effects of oxidative treatment techniques on
885 molecular size distribution of humic acids. Water Sci. Technol. 49(4), 7-12.

886
887 Kim, J., Davies, S.H.R., Baumann, M.J., Tarabara, V.V., Masten, S.J., 2008. Effect of
888 ozone dosage and hydrodynamic conditions on the permeate flux in a hybrid ozonation-
889 ceramic ultrafiltration system treating natural waters. J. Membrane Sci. 311, 165-172.
890
891 Kim, J.O., Jung, J.T., Yeom, I.T., Aoh, G.H., 2007. Effect of fouling reduction by ozone
892 backwashing in a microfiltration system with advanced new membrane material.
893 Desalination 202, 361-368.
894
895 Kimura, K., Amy, G., Drewes, J.E., Heberer, T., Kim, T.U., Watanabe, Y., 2003.
896 Rejection of organic micropollutants (disinfection by-products, endocrine disrupting
897 compounds, and pharmaceutically active compounds) by NF/RO membranes. J.
898 Membrane Sci. 227, 113-121.
899
900 Kleiser, G., Frimmel, F.H., 2000. Removal of precursors for disinfection by-products
901 (DBPs) - differences between ozone- and OH-radical-induced oxidation. Sci. Total
902 Environ. 256, 1-9.
903
904 Kosaka, K., Yamada, H., Shishida, K., Echigo, S., Minear, R.A., Tsuno, H., Matsui, S.,
905 2001. Evaluation of the treatment performance of a multistage ozone/hydrogen peroxide
906 process by decomposition by-products. Water Res. 35, 3587-3594.
907
908 Košutić, K., Kunst, B., 2002. RO and NF membrane fouling and cleaning and pore size
909 distribution variations. Desalination 150, 113-120.
910
40

41
911 Kuberkar, V.T., Davies, R.H., 2000. Modeling of fouling reduction by secondary
912 membranes. J. Membr. Sci. 168, 243-258.
913
914 Kubicki, J.D., Apitz, S.E, 1999. Models of natural organic matter and interactions with
915 organic contaminants. Org. Geochem. 30, 911-927.
916
917 Kusakabe, K., Aso, S., Hayashi, J.I., Isomura, K., Morooka, S., 1990. Decomposition of
918 humic acid and reduction of trihalomethane formation potential in water by ozone with
919 UV irradiation. Water Res. 24, 781-785.
920
921 Lal, M., Schöneich, C., Mönig, J., Asmus, K.D., 1988. Rate constants for the reactions of
922 halogenated organic radicals. Int. J. Radiation Biol. 54, 773-785.
923
924 Le-Clech, P., Lee, E.K., Chen, V., 2006. Hybrid photocatalysis/membrane treatment for
925 surface waters containing low concentrations of natural organic matters. Water Res. 40,
926 323-330.
927
928
Lee, N., Amy, G., Croué, J.P., Buisson, H., 2004. Identification and understanding of
929 fouling in low-pressure membrane (MF/UF) filtration by natural organic matter (NOM).
930 Water Res. 38, 4511-4523.
931
932
Lee, N., Amy, G., Croué, J.P., Buisson, H., 2005(a). Morphological analysis of natural
933 organic matter (NOM) fouling of low-pressure membranes (MF/UF). J. Membrane Sci.
934 261, 7-16.
935

42
936 Lee, S., Lee, K., Mohtar Wan, W., Choi, Y., 2005(b). Comparison of membrane
937 permeability and a fouling mechanism by pre-ozonation followed by membrane filtration
938 and residual ozone in membrane cells. Desalination 178, 287-294.
939
940 Lee, L.Y., Ng, H.Y., Ong, S.L., Hu, J.Y., Tao, G., Kekre, K., Viswanath, B., Lay, W.,
941 Seah, H., 2009. Ozone-biological activated carbon as a pretreatment process for reverse
942 osmosis brine treatment and recovery. Water Res. 43, 3948-3955.
943
944 Lehman, S.G., Liu, L., 2009. Application of ceramic membranes with pre-ozonation for
945 treatment of secondary wastewater effluent. Water Res. 43, 2020-2028.
946
947 Leiknes, T., 2009. The effect of coupling coagulation and flocculation with membrane
948 filtration in water treatment: A review. J. Environ. Sci. 21, 8-12.
949
950 Leppard, G. G., 1997. Colloidal organic fibrils of acid polysaccharides in surface waters:
951 electron-optical characteristics, activities and chemical estimates of abundance. Colloids
952 and Surfaces A: Physicochem. Eng. Aspects, 120, 1-15.
953
954 Li, Q., Elimelech, M., 2004. Organic fouling and chemical cleaning of nanofiltration
955 membranes: measurements and mechanisms. Environ. Sci. Technol. 38, 4683-4693.
956
957 Li, J., Liu, H., Zhao, X., Qu, J., Liu, R., Ru, J., 2008. Effect of preozonation on the
958 characteristic transformation of fulvic acid and its subsequent trichloromethane formation
959 potential: Presence or absence of bicarbonate. Chemosphere 71, 1639-1645.
960

43
961 Lin, C.F., Lin, A.Y.C., Sri Chandana, P., Tsai, C.Y., 2009. Effects of mass retention of
962 dissolved organic matter and membrane pore size on membrane fouling and flux decline.
963 Water Res. 43, 389-394.
964
965 Liao, C.H., Gurol, M.D., 1995. Chemical oxidation by photolytic decomposition of
966 hydrogen peroxide. Environ. Sci. Technol. 29, 3007-3014.
967
968 Liu, S., Lim, M., Fabris, R., Chow, C., Chiang, K., Drikas, M., Amal, R., 2008. Removal
969 of humic acid using TiO
2
photocatalytic process – Fractionation and molecular weight
970 characterisation studies. Chemosphere 72, 263-271.
971
972 Lohwacharin, J., Takizawa, S., 2009. Effects of nanoparticles on the ultrafiltration of
973 surface water. J. Membrane Sci. 326, 354-362.
974
975 Lowe J., Hossain Md. M., 2008. Application of ultrafiltration membranes for removal of
976 humic acid from drinking water. Desalination 218, 343-354.
977
978 Magazinovic, B., Nicholson, B.C., Mulcahy, D.E., Davey, D.E., 2004. Bromide levels in
979 natural waters: its relationship to levels of both chloride and total dissolved solids and the
980 implications for water treatment. Chemosphere 57, 329-335.
981
982
Makdissy, G., Croué, J.P., Amy, G., Lee, N., Habarou, H., Buisson, H, 2002. Advance in
983 the knowledge of the fouling of an ultrafiltration membrane, in: Proceedings of the
984 AWWA Water Quality and Technology Conference, Seattle.
985

44
986 Mao, Y., Schöneich, C., Asmus, K.D., 1991. Identification of organic acids and other
987 intermediates in oxidative degradation of chlorinated ethanes on TiO
2
surfaces en route to
988 mineralization. A combined photocatalytic and radiation chemical study. J. Phys. Chem.
989 95, 10080-10089.
990
991 Mattaraj, S., Jarusutthirak, C., Jiraratananon, R., 2008. A combined osmotic pressure and
992 cake filtration model for crossflow nanofiltration of natural organic matter. J. Membrane
993 Sci. 322, 475-483.
994
995 McDonald, S., Bishop, A.G., Prenzler, P.D., Robards, K., 2004. Analytical chemistry of
996 freshwater humic substances. Anal. Chim. Acta 527, 105-124.
997
998 Meunier, L., Canonica, S., von Gunten, U., 2006. Implications of sequential use of UV
999 and ozone for drinking water quality. Water Res. 40, 1864-1876.
1000
1001 Meylan, S., Hammes, F., Traber, J., Salhi, E., von Gunten, U., Pronk, W., 2007.
1002 Permeability of low molecular weight organics through nanofiltration membranes. Water
1003 Res. 41, 3968-3976.
1004
1005 Milne, C.J., Kinniburgh, D.G., Tipping E., 2001. Generic NICA-Donnan model
1006 parameters for proton binding by humic substances. Environ. Sci. Technol. 35, 2049-2059.
1007
1008
Nawrocki, J., Świetlik, J., Raczyk-Stanisławiak, U., Dąbrowska, A., Biłozor, S., Ilecki,
1009 W., 2003. Influence of ozonation conditions on aldehyde and carboxylic acid formation.
1010 Ozone Sci. Eng. 25, 53-62.

45
1011
1012
1013
Nawrocki, J, 2007. Comment on “Outside-in trimming of humic substances during ozonation in a membrane contactor.” Environ. Sci. Technol. 41, 5161.
1014
1015
Nghiem, L.D., Schäfer, A.I., Elimelech, M., 2004. Removal of natural hormones by
1016 nanofiltration membranes: Measurement, modeling, and mechanisms. Environ. Sci.
1017 Technol. 38, 1888-1896.
1018
1019 Nguyen, T.H., Chen, K.L., 2007. Role of divalent cations in plasmid DNA adsorption to
1020 natural organic matter-coated silica surface. Environ. Sci. Technol. 41, 5370-5375.
1021
1022 Nie, B., Stutzman, J., Xie, A., 2005. A vibrational spectral maker for probing the
1023 hydrogen-bonding status of protonated Asp and Glu residues. Biophys. J. 88, 2833-2847.
1024
1025 Nissinen, T.K., Miettinen, I.T., Martikainen, P.J., Vartiainen, T., 2001. Molecular size
1026 distribution of natural organic matter in raw and drinking waters. Chemosphere 45, 865-
1027 873.
1028
1029 Oh, B.S., Park, S.J., Lee, H.G., Kim, K.S., Lee, K.H., Kang, J.W., 2003. Application of
1030 ozone/UV process for the reclamation of sewage treatment plant effluent. J. Water
1031 Environ. Technol. 1, 141-153.
1032
1033 Ouyang, L., Malaisamy, R., Bruening, M.L, 2008. Multilayer polyelectrolyte films as
1034 nanofiltration membranes for separating monovalent and divalent cations. J. Membrane
1035 Sci. 310, 76-84.

46
1036
1037 Paillard, H., Brunet, R., Dore, M., 1988. Optimal conditions for applying an ozone-
1038 hydrogen peroxide oxidizing system. Water Res. 22, 91-103.
1039
1040 Pi, Y., Schumacher, J., Jekel, M., 2005. Decomposition of aqueous ozone in the presence
1041 of aromatic organic solutes. Water Res. 39, 83-88.
1042
1043
Pianta, R., Boller, M., Urfer, D., Chappaz, A., Gmünder, A., 2000. Costs of conventional
1044 vs. membrane treatment for karstic spring water. Desalination 131, 245-255.
1045
1046 Piccolo, A., 2002. The supramolecular structure of humic substances. A novel
1047 understanding of humus chemistry and implications in soil science. Advan. Agron. 75, 57-
1048 134.
1049
1050 Rivas, J., Gimeno, O., Beltrán, F., 2009. Wastewater recycling: Application of ozone
1051 based treatments to secondary effluents. Chemosphere 74, 854-859.
1052
1053 Rivera-Utrilla, J., Sánchez-Polo, M., Méndez-Díaz, J.D., Ferro-García, M.A., Bautista-
1054 Toledo, M.I., 2008. Behavior of two different constituents of natural organic matter in the
1055 removal of sodium dodecylbenzenesulfonate by O
3
and O
3
-based advanced oxidation
1056 processes. J. Colloid Interf. Sci. 325, 432-439.
1057
1058
Rosal, R., Rodríguez, A., Perdigón-Melón, J.A., Petre A. and García-Calvo E., 2009.
1059 Oxidation of dissolved organic matter in the effluent of a sewage treatment plant using
1060 ozone combined with hydrogen peroxide (O
3
/H
2
O
2
). Chem. Eng. J. 149, 311-318.

47
1061
1062 Sartor, M., Schlichter, B., Gatjal, H., Mavrov, V., 2008. Demonstration of a new hybrid
1063 process for the decentralised drinking and service production from surface water in
1064 Thailand. Desalination 222, 528-540.
1065
1066 Schäfer, A.I., Mastrup, M., Lund Jensen, R., 2002. Particle interactions and removal of
1067 trace contaminants from water and wastewaters. Desalination 147, 243-250.
1068
1069 Schlichter, B., Mavrov, V., Chmiel, H., 2004. Study of a hybrid process combining
1070 ozonation and microfiltration/ultrafiltration for drinking water production from surface
1071 water. Desalination 168, 307-317.
1072
1073 Scopes, R.K, 1994. Protein purification: principles and practice (3 rd
edition), Springer
1074 Verlag, New York (NY, USA), pp. 73.
1075
1076 Seidel, A., Elimelech, M., 2002. Coupling between chemical and physical interactions in
1077 natural organic matter (NOM) fouling of nanofiltration membranes: Implications for
1078 fouling control. J. Membrane Sci. 203, 245-255.
1079
1080 Shao, J., Hou, J., Song, H., 2010. Comparison of humic acid rejection and flux decline
1081 during filtration with negatively charged and uncharged ultrafiltration membranes. Wat.
1082 Res., doi:10.1016/j.watres.2010.09.006
1083

48
1084 Shon, H.K., Vigneswaran, S., Kim, I.S., Cho, J., Ngo, H.H., 2004. Effect of pretreatment
1085 on the fouling of membranes: Application in biologically treated sewage effluent. J.
1086 Membrane Sci. 234, 111-120.
1087
1088 Siddiqui, M., Amy, G., Ryan, J., Odem, W., 2000. Membranes for the control of natural
1089 organic matter from surface waters. Water Res. 34, 3355-3370.
1090
1091 Song, W., Ravindran, V., Pirbazari, M., 2008. Process optimization using a kinetic model
1092 for the ultraviolet radiation- hydrogen peroxide decomposition of natural and organic
1093 compounds in groundwater. Chem. Eng. Sci. 63, 3249-3270.
1094
1095 Song, Y., Dong, B., Gao, N., Xia, S., 2010. Huangpu River water treatment by
1096 microfiltration with ozone pretreatment. Desalination 250, 71-75.
1097
1098
Staehelin, J., Hoigné, J., 1982. Decomposition of ozone in water: rate of initiation by
1099 hydroxide ions and hydrogen peroxide. Environ. Sci. Technol. 16, 676-681.
1100
1101
Staehelin, J., Hoigné, J., 1985. Decomposition of ozone in water in the presence of organic
1102 solutes acting as promoters and inhibitors of radical chain reactions. Environ. Sci.
1103 Technol. 19, 1206-1213.
1104
1105 Sutton, R., Sposito, G., 2005. Molecular structure in soil humic substances: the new view.
1106 Environ. Sci. Technol. 39, 9009-9015.
1107

49
1108 Sutzkover-Gutman, I., Hasson, D., Semiat, R., 2010. Humic substances fouling in
1109 ultrafiltration processes. Desalination 261, 218-231
1110
1111
Świetlik, J., Sikorska, E., 2004. Application of fluorescence spectroscopy in the studies of
1112 natural organic matter fractions reactivity with chlorine dioxide and ozone. Water Res. 38,
1113 3791-3799
1114
1115 Tang, C.Y., Kwon, Y.N., Leckie, J.O., 2007. Fouling of reverse osmosis and nanofiltration
1116 membranes by humic acid – effects of solution composition and hydrodynamic conditions.
1117 J. Membrane Sci. 290, 86-94.
1118
1119 These, A., Reemtsma, T., 2005. Structure-dependent reactivity of low molecular weight
1120 fulvic acid molecules during ozonation. Environ. Sci. Technol. 39, 8382-8387.
1121
1122 Thurman, E.M., 1986. Organic Geochemistry of Natural Waters, Martinus Nijhoff
1123 Publishers, Dordrecht (The Netherlands).
1124
1125 Van Geluwe, S., Braeken, L., Vinckier, C., Van der Bruggen, B., 2009. Ozonation and
1126 perozonation of humic acids in nanofiltration concentrates. Desalin. Water Treat.
6, 217-
1127 221.
1128
1129 Van Geluwe, S., Vinckier, C., Bobu, E., Trandafir, C., Vanelslander, J., Braeken, L., Van
1130 der Bruggen, B., 2010(a). Eightfold increased membrane flux of NF 270 by O
3
oxidation
1131 of natural humic acids without deteriorated permeate quality. J. Chem. Technol. Biot. 85,
1132 1480-1488.

50
1133
1134 Van Geluwe, S., Vinckier, C., Braeken, L., Van der Bruggen, B., 2010(b). Ozone
1135 oxidation of nanofiltration concentrates alleviates membrane fouling in drinking water
1136 industry. Submitted for publication.
1137
1138 van Voorthuizen, E.M., Ashbolt, N.J., Schäfer, A.I., 2001. Role of hydrophobic and
1139 electrostatic interactions for initial enteric virus retention by MF membranes. J. Membrane
1140 Sci. 194, 69-79.
1141
1142 Verliefde, A., Cornelissen, E., Amy, G., Van der Bruggen, B., van Dijk, H. (2007) Priority
1143 organic micropollutants in water sources in Flanders and the Netherlands and assessment
1144 of removal possibilities with nanofiltration. Environ. Pollut. 146, 281-289.
1145
1146 von Gunten, U., 2003(a). Ozonation of drinking water: Part I. Oxidation kinetics and
1147 product formation. Water Res. 37, 1443-1467.
1148
1149 von Gunten U., 2003(b). Ozonation of drinking water: Part II. Desinfection and by-
1150 product formation in presence of bromide, iodide and chlorine. Water Res. 37, 1469-1487.
1151
1152
Waite, T.D., Schäfer, A.I., Fane, A.G., Heuer, A., 1999. Colloidal fouling of ultrafiltration
1153 membranes: impact of aggregate structure and size, J. Colloid Interface Sci. 212, 264-274.
1154
1155 Wang, Y., Hollingsworth, R.I., Kasper, D.L., Kasper, D.L., 1999. Ozonolytic
1156 depolymerization of polysaccharides in aqueous solution. Carbohyd. Res. 319, 141-147.
1157

51
1158 Wang, G.S, Pai, S.Y., 2001. Ozonation of dissolved organic matter in biologically treated
1159 wastewater effluents. Ozone Sci. Eng. 23, 351-358.
1160
1161 Wang, X., Wang, L., Liu, Y., Duan, W., 2007. Ozonation pretreatment for ultrafiltration of
1162 the secondary effluent. J. Membrane Sci. 287, 187-191.
1163
1164 Weinstein, J., Bielski B.H.J., 1979. Kinetics of the interaction of HO
2
and O
2
- radicals with
1165 hydrogen peroxide. The Haber-Weiss reaction. J. Am. Chem. Soc.
101, 58-62.
1166
1167 Wert, E.C., Rosario-Ortiz, F.L., Drury, D.D., Snyder, S.A., 2007. Formation of oxidation
1168 byproducts from ozonation of wastewater. Water Res. 41, 1481-1490.
1169
1170 Westerhoff, P., Aiken, G., Amy, G., Debroux, J, 1999. Relationships between the structure
1171 of natural organic matter and its reactivity towards molecular ozone and hydroxyl radicals.
1172 Water Res. 33, 2265-2276
1173
1174 Williamson, D.G., Cvetanovic, R.J., 1970. Rates of ozone-paraffin reactions in carbon
1175 tetrachloride solution. J. Am. Chem. Soc. 92, 2949-2952.
1176
1177 Xie, Y.F., 2004. Desinfection by-products in drinking water, Lewis, Boca Raton (FL,
1178 USA).
1179
1180
Xiong, F., Croué, J.P., Legube, B., 1991. Enhancement of radical chain reactions of ozone
1181 in water in the presence of an aquatic fulvic acid. Ozone Sci. Eng. 13, 349-363
1182

52
1183 Xiong, F., Croué, J.P., Legube, B., 1992. Long-term ozone consumption by aquatic fulvic
1184 acids acting as precursors of radical chain reactions. Environ. Sci. Technol. 26, 1059-
1185 1064.
1186
1187 Yuan, W., Zydney, A.L, 1999. Humic acid fouling during microfiltration. J. Membrance
1188 Sci. 157, 1-12.
1189
1190 Yoon, Y., Westerhoff, P., Snyder, S.A., Wert, E.C., 2006. Nanofiltration and ultrafiltration
1191 of endocrine disrupting compounds, pharmaceuticals and personal care products. J.
1192 Membrane Sci. 270, 88-100.
1193
1194 You, S.H., Tseng, D.H., Hsu W.C., 2007. Effect and mechanism of ultrafiltration
1195 membrane fouling removal by ozonation. Desalination 202, 224-230.
1196
1197 Zhang, T., Lu, J., Ma, J., Qiang, Z., 2008. Fluorescence spectroscopic characterization of
1198 DOM fractions isolated from a filtered river water after ozonation and catalytic ozonation.
1199 Chemosphere 71, 911-921.
1200
1201 Zhu, H., Wen, X., Huang, X., 2010. Membrane organic fouling and the effect of pre-
1202 ozonation in microfiltration of secondary effluent organic matter. J. Membrane Sci. 352,
1203 213-221.
1204
1205 Zularisam, A.W., Ismail, A.F., Salim, R., 2006. Behaviours of natural organic matter in
1206 membrane filtration for surface water treatment – a review. Desalination 194, 211-213.