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The widespread applications of ozone technologies are established on the basis of large-scale manufacture of ozone generator and chemical reactivity of ozone. It is hence necessary to summarize the principles of ozone generation and to analyze the physicochemical properties of ozone, which are of fundamental significance to indicate its technical developments and practical applications. This review presents a summary concerning ozone generation mechanisms, the physicochemical properties of ozone, as well as the applications of ozone in water treatment. Ozone can be produced by phosphorus contact, silent discharge, photochemical reactions, and electrochemical reactions, principally proceeding by the reaction of oxygen atom with oxygen molecule. There are side reactions to the generation of ozone, however, which are responsible for ozone depletion including thermal decomposition and quenching reactions by reactive species. The solubility of ozone in water is much higher than that of oxygen, suggesting that it may be reliably applied in water and wastewater treatment. Based on the resonance structures of ozone, one oxygen atom in ozone molecule is electron-deficient displaying electrophilic property, whereas one oxygen atom is electron-rich holding nucleophilic property. The superior chemical reactivity of ozone can also be indirectly revealed by radical-mediated reactions initiated from homogenous and heterogeneous catalytic decomposition of ozone. Owing to the reliable generation of ozone and its robust reactive properties, it is worthy to thoroughly elaborate the applications of ozone reaction in drinking water disinfection and pre- or post-treatment of industrial wastewater including cyanide wastewater, coking wastewater, dyeing wastewater, and municipal wastewater. The structural characteristics of ozone reactors and energy requirement of applied technologies are evaluated. In addition, future directions concerning the development of ozone generation, ozone reactivity, and industrial wastewater ozonation have been proposed.
Content may be subject to copyright.
Rev Chem Eng 2016; aop
*Corresponding author: Chaohai Wei, School of Environment and
Energy, South China University of Technology, Guangzhou, 510006,
P.R. China, e-mail:
Fengzhen Zhang, Yun Hu and Chunhua Feng: School of Environment
and Energy, South China University of Technology, Guangzhou,
510006, P.R. China
Haizhen Wu: School of Bioscience and Bioengineering, South China
University of Technology, Guangzhou, 510006, P.R. China
Chaohai Wei*, Fengzhen Zhang, Yun Hu, Chunhua Feng and Haizhen Wu
Ozonation in water treatment: the generation,
basic properties of ozone and its practical
DOI 10.1515/revce-2016-0008
Received February 10, 2016; accepted May 18, 2016
Abstract: The widespread applications of ozone technolo-
gies are established on the basis of large-scale manufac-
ture of ozone generator and chemical reactivity of ozone.
It is hence necessary to summarize the principles of ozone
generation and to analyze the physicochemical proper-
ties of ozone, which are of fundamental significance to
indicate its technical developments and practical applica-
tions. This review presents a summary concerning ozone
generation mechanisms, the physicochemical properties
of ozone, as well as the applications of ozone in water
treatment. Ozone can be produced by phosphorus contact,
silent discharge, photochemical reactions, and electro-
chemical reactions, principally proceeding by the reac-
tion of oxygen atom with oxygen molecule. There are side
reactions to the generation of ozone, however, which are
responsible for ozone depletion including thermal decom-
position and quenching reactions by reactive species. The
solubility of ozone in water is much higher than that of
oxygen, suggesting that it may be reliably applied in water
and wastewater treatment. Based on the resonance struc-
tures of ozone, one oxygen atom in ozone molecule is elec-
tron-deficient displaying electrophilic property, whereas
one oxygen atom is electron-rich holding nucleophilic
property. The superior chemical reactivity of ozone can
also be indirectly revealed by radical-mediated reactions
initiated from homogenous and heterogeneous catalytic
decomposition of ozone. Owing to the reliable generation
of ozone and its robust reactive properties, it is worthy to
thoroughly elaborate the applications of ozone reaction
in drinking water disinfection and pre- or post-treatment
of industrial wastewater including cyanide wastewater,
coking wastewater, dyeing wastewater, and municipal
wastewater. The structural characteristics of ozone reac-
tors and energy requirement of applied technologies are
evaluated. In addition, future directions concerning the
development of ozone generation, ozone reactivity, and
industrial wastewater ozonation have been proposed.
Keywords: energy consumption; ozone generation; ozone
reactor; physicochemical properties; water treatment.
1 Introduction
The electronic and ionic excitations of molecular oxygen
by physical field including electric discharge, ultraviolet
irradiation, and electrolysis are the preliminary reactions
for ozone generation. The electric discharge (e.g. silent
discharge) is the most reliable approach for ozone genera-
tion in both laboratory and industrial scale owing to the
convenience and operational efficiency in different ozone
concentration requirement (Nomoto et al. 1995, Salam
etal. 2013). A configuration of ozone generator adopting
silent discharge was first proposed by Siemens in 1857,
characterized by the presence of a dielectric layer on the
inner surface of electrodes. The electrochemical reactions
proceeded at anode in a variety of aqueous media could
also be capable of generating ozone, whose advantages
comprise low voltage operation, no gas fed of any descrip-
tion, and the possibility to produce high concentration
of ozone in gaseous and aqueous phase (Christensen
etal. 2009). It was found that (˙OH)ads and Oads generated
by means of electrolysis were intermediates involved in
ozone production (Da Silva etal. 2006, Kim and Korshin
Due to the instability of ozone, only ozone generator
can be operated practically and stably; it is possible to
elucidate the physical and chemical properties of ozone,
focusing on its solubility and chemical reactivity, which
will widen its application in water and wastewater treat-
ment. Ozone is generally colorless and less soluble in
water, with special pungent odor at ordinary tempera-
ture, from which its name was derived (Guzel-Seydim
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etal. 2004). On the other hand, ozone is a chemical reac-
tive reagent exhibiting electrophilic and nucleophilic
characteristics, which are closely related to its resonance
structures. Kalemos and Mavridis (2008) reported that the
ground state of ozone is a “closed-shell singlet resulted
from O2(1Δg) and O(1D),” demonstrating that its electronic
structure corresponds to two resonant structures with one
O-O single bond and one O=O double bond, which could
be merged into one closed-shell structure by attributing
one and a half bond to each O-O interaction. Compara-
tively, Miliordos and Xantheas (2014) indicated that an
open-shell structure of ozone could be established with
a single bond and two lone electrons on each terminal
atom, based on ground-state multireference configura-
tion interaction wave functions. Although ozone is a pow-
erful oxidant, it reacts less efficiently with some organic
(e.g. saturated aliphatic acid) and inorganic (e.g. NH4+)
compounds. In order to accelerate the mineralization
efficiency, advanced oxidation processes such as O3/UV,
O3/H2O2, catalytic ozonation and photocatalytic ozonation
stimulating the generation of more robust and stronger
radical species, have been suggested as potential alter-
natives. However, owing to the inhibited effect of radical
scavengers (i.e. carbonate and bicarbonate) on ˙OH-medi-
ated reactions, it is encouraged to establish novel reaction
systems based on liquid-liquid or liquid-solid extraction
of ozone and organic substances from aqueous solution
to organic solvent, followed by the reactions proceeded in
non-aqueous phase (Kasprzyk-Hordern etal. 2003).
In view of its strong oxidation ability with a selection
of pollutants, ozone can be potentially used as a broad-
spectrum disinfectant and a powerful oxidant for waste-
water treatment. A large variety of micro-organisms could
be effectively inactivated by ozone through cell lysis and
nucleic acids destruction (von Gunten 2003a, Kingsbury
and Singer 2013, Tachikawa and Yamanaka 2014). Even
though the production of halogenated compounds can be
avoided by ozone disinfection, the vast majority of oxygen
atom containing byproducts including bromate, aldehyde,
ketone, esters, and keto acids may be generated (Richard-
son etal. 1999). Ozone has also been extensively applied
for further treatment of industrial wastewater such as
cyanide containing wastewater, coking wastewater, oil
refining wastewater, and pharmaceutical wastewater,
through the selective oxidative reactions with unsaturated
and conjugated matrix components (Laera et al. 2012,
Chen et al. 2014a, Lin etal. 2014a). It had been estab-
lished that ozone-based technologies (e.g. non-catalytic
ozonation, ozonation catalyzed by ultraviolet irradiation,
hydrogen peroxide and various heterogeneous catalysts)
can be the alternatives or the pre-treatment methods for
biological treatment of actual agro-industrial wastewater,
which were indicated by the improvement of biodegrada-
bility and toxicity removal (Martins and Quinta-Ferreira
While previous publications have provided a wealth
information on ozone generation (Wei et al. 2014), the
basic properties of ozone (Sonntag and von Gunten
2012), the mechanisms of ozonation and catalytic ozona-
tion (Kasprzyk-Hordern et al. 2003, Hübner etal. 2015),
the disinfection of drinking water, and mineralization of
industrial wastewater (Langlais etal. 1991, Barndõk etal.
2014, Zhang etal. 2014a), an integrated overview concern-
ing ozone production, physico-chemical properties of
ozone, and its practical application is still missing. A criti-
cal review that combines the technical developments of
ozone generation, the analysis of its basic properties, and
the general mechanisms of ozone reaction will be helpful
to the implementation of its application in large scale. The
objectives of this review are to summarize ozone genera-
tion mechanisms and its generator as a whole; to analyze
the fundamental characteristics of ozone, so as to discuss
the pollution control for which ozone-based technologies
applied; and to evaluate the energy consumption of ozo-
nation processes in terms of ozone generation as well as
contaminants removal.
2 The generation of ozone and its
2.1 The principle of ozone generation
2.1.1 Phosphorus contact ozone generation
The existence of oxygen atom produced from excitation of
oxygen molecule is the prerequisite reaction for various
ozone generation techniques. The white phosphorus
contact with wet air, electrolysis of acidic aqueous solu-
tion, and photosensitive reaction of mercury were respon-
sible for ozone generation during early days of its discovery
(Fallon and Vanderslice 1960). The white phosphorus
contact was the most prevalent approach at that moment
due to its simplicity and low investment requirement. In
general, the reaction of white phosphorus with wet air can
be expressed in Eqs. (1–5) (Andrews and Withnall 1988),
with the generation of oxygen atom being the critical
step. Then the resulting mixtures containing ozone and
phosphorus oxides were purified by water. It was found,
however, that ozone productivity by phosphorus reaction
was unsatisfactory, suffering from variable yield results
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and low efficiency, although maximum ozone concentra-
tion of 2.5 mg/l could be attained by an optimized proce-
dure (Leeds 1879, Rubin 2002). Studies dedicated in other
alternative technologies for improved ozone generation
as well as basic mechanisms behind its generation were
therefore conducted.
+→ +
42 4
+→ +=
4n 24n1
PO OPOO (n 19)
P20O PO 10O
OO MOM, H144.8 kJ/mol,
MN, or O
++→+ =
= (6)
Fundamentally, the first preliminary reaction for ozone
generation is the excitation of molecular oxygen by high
energy field, resulting in the generation of oxygen atom.
Ozone can be subsequently produced by the essential
Three-body Collision Reaction involving an oxygen atom,
an oxygen molecule, and a third body as shown in Eq. (6)
(Rubin 2002). In the electronic point of view, ozone forma-
tion can be achieved through the spin-pairing of antibo-
nding πg electrons of the ground state molecular oxygen
with two unpaired electrons of oxygen atom (Harcourt
etal. 1986), which is widely accepted by scientific com-
munities. These achievements underpin the popularity
of silent discharge, photochemical synthesis, and elec-
trolysis techniques, which are in essence proceeded with
energy transferred from physical fields.
2.1.2 Gaseous discharge for ozone generation
The basic principle of ozone generation by means of
gaseous discharge comprises the creation of electrons
in discharge gap as well as the excitation of oxygen mol-
ecule, eventually leading to ozone appearance. Silent
discharge, also known as dielectric barrier discharge
(DBD), is the non-equilibrium discharge process carried
on between discharge gap in which the dielectric barrier
is intercalated (Yehia 2015). Typically, as illustrated in
Figure1, the configurations of dielectric layers are either
covered on the surface of electrodes inside discharge gap
or suspended between them. In general, ozone generation
by silent discharge involves three following steps: (1) a
breakdown of the gap occurs in a few nanoseconds upon
the load of applied voltage, (2) the production of current
pulse occurs as a result of increasing density of charged
particles and then the transfer of electric charge within
10 ns, (3) the ozone generation reactions take place in
H.V. H.V.
Figure 1:The schematic configurations of silent discharge electrodes assembled by plated electrode [(A)–(C)] and coaxial tube electrode
[(D)–(F)]. (Common dielectric materials consisted of glass, ceramics, enamel, and mica; the average thickness of these dielectrics is
0.5–3mm. The gray color parts are referred to electrodes, while the blue color parts are assigned to insulating dielectric materials).
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micro-discharge current channel (Eliasson et al. 1987).
The chemical reactions consisting of electron-impact
reactions, neutral-neutral reactions, and charged parti-
cle reactions mainly contribute to ozone generation, with
some exemplary reactions discussed below (Kitayama and
Kuzumoto 1999, Lee etal. 2004). When air or pure oxygen
gas is fed through the discharge gap, O2 can be excited
from its ground state
(( ))OX g to metastable state
OB , (( )( )
by inelastic collision with high-
energy electrons in <3 ns, producing O (3P) and O (1D), as
shown in Eqs. (7–9) (Kitayama and Kuzumoto 1997). Ozone
can be then produced through the well-established Three-
body Collision Reaction [Eq. (6)]. Besides, ozone can also
be produced via metastable reaction of oxygen molecule
with +
)OA u [Eq. (10)] in <3 μs (Benson 1959). In con-
trast to the significant contribution of high-energy elec-
tron, low-energy electron is capable of exciting oxygen
molecule into its negative ion state [Eq. (11)], which imme-
diately reacts with oxygen molecule to produce ozone as
illustrated in Eq. (12) (Benson 1959). On the other hand,
the formed ozone may be inevitably destroyed by oxygen
negative ion, oxygen atom, and electron [Eqs. (13–16)],
where M is assigned to any particles present in the gas.
Also, heat accumulated from discharge gap could promote
thermal decomposition of ozone with a reaction rate con-
stant of 4.61±0.25×1012 exp(-24,000/RT) l/(mol·s) (Jones
and Davidson 1962).
eOXeOB eO(P)O(D) Reaction Threshold: 8( .)(
gu (7)
+→ +
)O O( P)
eOXeOA eO(P)O(P)Reaction Threshold: 6()
gu (9)
u (10)
eaction Threshold: 3.6 eV+→+ (11)
322 m
OO OO H-82 kcal/mol (13)
+→ +
+→ +
+→++ =
32 m
eO OO eH-10 kcal/mol
Based on the concept that discharge area is essential
for ozone generation, modified discharge assembly of
DBD including surface discharge and hybrid discharge
is widely employed for ozone generation in a variety of
applications. The invention of surface dielectric barrier
discharge (SDBD) enables propagation of micro-discharge
in a non-uniform field along dielectric surface and a
larger discharge area than channel diameter (Nassour
etal. 2016). Typically, there is a dielectric surface in SDBD
with thin parallel electrode strips on one side of dielec-
tric barrier, while a metallic electrode is covered on the
reverse side of barrier, leaving the gas gap unfixed as
depicted in Figure 2A–B (Pavon etal. 2007). An extended
discharge area on dielectric barrier surface can be devel-
oped as a result of the unfixed gas gap in SDBD, which is
confirmed by stronger emission discovered from the edges
in the form of long filaments (streamers) with increasing
airflow velocity (Figure 2C). Similarly, Xu (2001) reported
the extended discharge area of SDBD and indicated that it
depended on the accumulated surface charge during dis-
charge development, starting on the dielectric surface and
flowing towards the surface electrode in case of positive
The ozone generation can be improved by the hybridi-
zations of different discharge types. Nomoto etal. (1995),
for instance, specially designed a discharge unit inte-
grating silent discharge with surface discharge for ozone
generation, which was assembled of inner coil electrode
and external copper cylinder electrode (Figure 2D). The
length and diameter of inner coil electrode were 80mm
and 0.25 mm, respectively, keeping in contact with inner
wall of coaxial glass tube (inner diameter was 12.2 mm,
thickness was 1.4 mm). The inner diameter of grounded
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Airflow stripes
Dielectric barrier
External copper cylinder electrode
Silent discharge gap
O2, O3
O2, O3
Coil electrode
Surface discharge area
Coaxial glass tube
Airflow // stripes
Figure 2:The configurations of surface dielectric barrier discharge electrode (A–B) and silent-surface hybrid discharge (D) for ozone
generation, and CCD images of the whole discharge area for surface dielectric barrier discharge with the isentropic flow (corresponding to
velocity of airflow) increasing from left to right (C). [(A) to (C) were adapted from Pavon etal. 2007, (D) was adapted from Nomoto etal. 1995].
outer brass cylinder electrode was 17.0 mm, thereby
leaving 1-mm gap between outer side of coaxial glass
tube and inner side of copper cylinder. The combination
of different electrode types made it possible that silent
discharge occurred in the gap, while surface discharge
proceeded along the coil electrode. The authors revealed
that in the course of increasing applied voltage, surface
discharge was developed first along inner coil electrode
when voltage was <6.0 kV. And then silent discharge
became active in space gap as applied voltage increased
above 8.0 kV. Under optimized operation conditions, the
yield efficiency of ozone using silent-surface hybrid dis-
charge reached 274 g/kWh, which was almost two times
higher than that obtained using silent discharge (roughly
130 g/kWh). It was probably owed to the synergistic effect
of enhanced surface discharge driven by charge created
from silent discharge (Nomoto etal. 1995).
The evaluation of other discharge techniques such as
pulse streamer discharge, plasma jet discharge, pulse DBD,
and pulse corona discharge (PCD) is of significant impor-
tance for rising discharge efficiency of ozone generation
(Masuda etal. 1988, Robinson etal. 1998, Samaranayake
etal. 2011). The pulse dielectric barrier discharge (PDBD), for
example, is equipped with pulsed power with voltage pulse
duration ranging from 50 to 300 ns, improving the genera-
tion rate of ozone, and avoiding massive loss of energy in
heat (Kornev etal. 2006). Nevertheless, the preliminary dis-
charge pathways responsible for ozone generation by PDBD
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6C. Wei etal.: Ozonation in water treatment
remain unchanged in comparison to conventional DBD,
i.e. ozone generation is continued through three follow-
ing steps described earlier. The PCD (the configurations of
corona electrodes and a PCD ozone generator are displayed
in Figure 3A–D) is an attractive approach in atmospheric
pressure and non-uniform electric filed (Lukes etal. 2005).
Principally, electrons can be created by both negative and
positive corona in a typical PCD process. The created elec-
trons leave ionizing region and move to plate electrode,
completing the circuit while producing pulse current fila-
ments and micro-current column (Šimek and Člupek 2002).
It has to be emphasized that three advantages of PCD rela-
tive to conventional DBD should be taken into account:
(1) the breakdown of gas gap is not a requisite as corona
onset voltage is usually below breakdown voltage, (2) the
dielectric barrier between electrodes is not needed due to
the limited development of spark, and (3) the humidity of
feed gas is not strictly restricted as a result of limited spark
development and perhaps the absence of surface discharge
(Hegeler and Akiyama 1997).
The presence of inert gas in feed air also exhibits a
significant influence on the mechanism of ozone genera-
tion. It was reported that, for instance, low concentration
of N2 (110%, volume ratio) in feed air may promote the
generation of ozone (Kogelschatz 2003, von Sonntag and
von Gunten 2012). This was explained by the fact that part
of oxygen atom and electron is capable of reacting with
metastable N2 through the excitation and dissociation of
ground state N2, as shown in Eqs. (17–19). The storage and
liberation of reactive species (i.e. oxygen atom and high-
energy electron) can be therefore accomplished via series
of electron transfer and radical reactions [Eqs. (20–27)],
reducing the probability of ozone depletion with these
species (Tokuaga and Suzuki 1984). However, the con-
centration of NOx produced from above reactions reaches
a level at which the recombination reactions of oxygen
atoms [Eqs. (28–30)] might be dominated or ozone could
be depleted by NOx [Eqs. (31–34)], adversely affecting
the generation of ozone (Toby 1984, Robert et al. 1988,
Eliasson and Kogelschatz 1991). Moreover, the catalytic
Pulsed power
Outer metal
Cooling water
Gas flow
Gas flow High-voltage
wire electrode
Figure 3:The schematic configurations of corona electrodes (A), (B), and (C) are assigned for wire-cylinder, needle-plate, and wire-plate
modus, respectively, and the coaxial pulsed corona ozone generator (D).
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cycles [Eqs. (29) and (33)] also remarkably contribute
to ozone destruction (Yagi and Tanaka 1979). In order
to obtain the influence mechanisms of other inert gas
(e.g. He, Ar, Kr, Xe) on ozone generation, the readers are
referred to Wei etal. (2014).
O( P) NA NO()N
u (17)
O( D) NA NO()N
u (18)
+→ +→ +
B, be2N e
u (19)
+→ +
22 2
u (22)
(, 2
u (23)
+→ +
eN N2e
+→ +
+→ +
+→ +
(ONOB)NO O (32)
+→ +
2.1.3 Photochemical ozone generation
The photochemical reactions are potentially capable of
exciting molecular oxygen, producing oxygen atom, and
thus accelerating ozone generation (Hashem etal. 1997).
Mechanistic studies indicated that ultraviolet (UV) irra-
diation favors the excitation of singlet or triplet oxygen
atom from ground state oxygen molecule [Eqs. (35–36)]
(Salvermoser etal. 2008). Also, it has been reported that
the triplet oxygen atom participating in ozone generation
can be produced from photo-excitation of NO2 in feed gas
[Eq. (37)] (Calvert 1976, Schnell etal. 2009). Other nitrogen
containing species are also found to be beneficial to ozone
generation. The excitation of oxygen atom, for example,
was promoted when TiO2 surfaces were exposed to NOx
(e.g. NO3
-) under illumination via charge transfer reactions
(Monge etal. 2010). Similar with the quenching reactions
of ozone by oxygen atoms [Eqs. (13–16)] excited from elec-
trons, however, the photons with different wave length as
well as oxygen atoms existed in the photoreaction system
could adversely exhaust ozone via Eqs. (38–41) (Levy 1971,
Matsumi and Kawasaki 2003).
OhvO(P)O(D)175 nm
+→ <
Ohv2O( P) 20 0 nm 242 nm
NO hv O( P) NO 420 nm
< (37)
OhvO O( D) 200 nm 310 n
λ< (38)
() 11 nm
g (39)
()OhvO O( P) 612 nm+→
< (40)
+→ <
OhvOXO(P)118 m() 0 n
g (41)
2.1.4 Electrochemical ozone generation
The electrochemical reactions in aqueous solution are
generally believed to be able to produce ozone with several
advantages involving no need of gas feed; low voltage oper-
ation; simple system design; and, in particular, reduced
loss of ozone by thermal decomposition during handling
(Foller and Kelsall 1993). Principally, ozone can be gener-
ated from electrolysis of water by employing anode mate-
rials with high oxygen evolution potential (e.g. graphite,
glassy carbon, and lead alloys) through the Six Electron
Transfer mechanism, as illustrated in Eq. (42), with stand-
ard redox potential of 1.51 V (Da Silva etal. 2006). In fact,
as standard redox potential for the oxidation of water into
oxygen (1.23 V) is lower than that of Eq. (42), molecular
oxygen can be simultaneously produced with ozone gen-
eration. Detailed mechanism courses for ozone generation
have been proposed consisting of water spilt and oxygen
atom transfer reactions [Eqs. (43–46)] on the basis of the
assumption that the generation of ˙OH is an essential
step (Feng etal. 1994). As the significant role of oxygen
atom responsible for ozone generation in gaseous phase
has already been discussed previously, the intermediate
oxygen atoms produced by electrolysis also participate in
ozone production in aqueous phase. Ozone generation can
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be thus additionally achieved by the reaction of oxygen
atoms affiliated on anode surface, which is derived from
the decomposition of ˙OH [Eq. (47)] with dissolved oxygen
(Kötz and Stucki 1987). The delivery of ozone from ozone
generator to treated solution can be avoided by employ-
ing the in situ generation of ozone using electrolysis in
aqueous solution, in which ozone generation and decon-
taminant reaction are supposed to proceed.
→+ +
3H OO6H 6e
HO (OH) He
+ (43)
2ads ad
3ads 3
HO HO e()
→++ (47)
2.1.5 Summary
The generation of ozone can be reliably achieved through
the reaction of oxygen molecule with oxygen atom, which
can be technically produced by means of gas discharge,
photochemical excitation, and electrochemical reaction.
The silent discharge is widely adopted in ozone generator,
owing to its effectiveness and operational stability. The
ozone produced from silent discharge is, however, inevi-
tably consumed by thermal decomposition and side reac-
tions, thus limiting its wide practical application. In order
to reach a higher ozone yield efficiency, hybrid discharge
combinations such as integrating silent discharge with
surface discharge are superior to the individual technique,
as a result of the synergistic effect of different discharge
patterns. Moreover, PCD, which does not require electrode
insulation, could be an alternative option due to its ultra-
short pulsed power; simple electrode configuration; and,
most importantly, the possibility to operate with humid
gas. Particularly, it is the ability to work with humid gas
that makes PCD promising for in situ application of ozone
without transferring ozonated gas. Nevertheless, if PCD is
employed solely for ozone generation without watering
the electrodes, its ozone productivity per unit of energy
does not differ too much from silent discharge.
2.2 The ozone generator
Almost all ozone generators are practically operated with
silent discharge principle. The ozone generator funda-
mentally consisted of discharge unit, cooling system,
feed gas supplement, applied power, and corresponding
controlling system (Kitayama and Kuzumoto 1999, Alonso
et al. 2005), among which the design and operation of
discharge unit are of significant importance for ozone
generation. The ozone generator assembled using silent
discharge technique is commonly composed of coaxial
glass tube, whose outside wall of outer tube and inner
side wall of inner tube are uniformly coated by tin foil,
connecting to high-voltage power (Nassour et al. 2016).
Generally, the typical industrial ozone generators in early
times were categorized by plate-based, tube-based, and
other complex structural-based equipment, as compared
in Table 1. The plate-based Otto discharge cell (Figure4),
for example, was structured by water-cooled hollow
boxes working as high-voltage electrodes and parallel
glass plates acting as insulator. Technically, in case of
parallel connection of 48 discharge units, the maximum
ozone productivity was expected to be 18kg/h with power
efficiency of 20–22 kWh/ kg O3 (Information Institute of
Shanghai Science and Technology, 1976).
Currently, the most commonly used silent discharge
unit for ozone generation is usually assembled by several
water-cooled discharge tubes installed in parallel, as shown
in Figure 5 (Kogelschatz 2003). The cylindrical discharge
tubes with diameter of 20–50mm and length of 1–3m are
made of borosilicate glass, closed at one side, and mounted
inside slightly wider stainless steel tubes to remain annular
gap of 1–3mm (Figure 5). Metal coatings on the inner wall of
inside glass tubes serve as high-voltage electrodes, whereas
outer steel tubes function as ground electrodes. Some large-
scale ozone generators comprise several hundred discharge
tubes inside big steel tanks in order to provide required dis-
charge area, with maximum ozone productivity >800kg/h
applied in Japan for drinking water treatment and 230kg/h
applied in United States for municipal water treatment
(Loeb etal. 2012). The main obstacles hindering the devel-
opment of large-scale ozone generator are probably the
side ozone depletions and its thermal decomposition in
discharge gap, thus requiring further research efforts.
3 The physicochemical properties
of ozone
3.1 The physical properties of ozone
The physical properties of ozone, including odor, color,
and solubility, are the bases for its wide applications in
water and wastewater treatment. Ozone is an unstable
and colorless gas with pungent fishy smell under ambient
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Table 1:The compositional structure and operation conditions of ozone generators in early period.a
Applied powerSize of discharge unit Discharge
gap (mm)
Energy consumption
per area electrode
Ozone yield
Otto plate  Glass plate area,  mm; central
pore size, . mm; thickness of
plate, –. mm
. kg/hcWater cooled
Block plate  Glass plate diameter,  mm;
central pore size, . mm;
thickness of plate, – mm
.. kg/hdOil cooled
Van der Made
Thickness of outer tube, .mm;
diameter, . mm; length,
mm; diameter of inner tube,
. mm; length,  mm
<. g/heWater cooled
Siemens & Halske
,Thickness of dielectric glass,
.mm; diameter,  mm; length,
 mm
. .. kg/hfWater cooled
Welsbach tube  –Thickness of outer tube, . mm;
diameter,  mm; length,  mm
.–.. kg/hgWater cooled
Lowther grid plate– . . kg O/dAir cooled
aData from Translations of Ozone and Its Applications, Series 2. Information Institute of Shanghai Science and Technology, 1976.
bData were obtained by air as feed gas.
cParallel connection of 48 discharge units.
dParallel connection of 64 discharge units.
eParallel connection of 30 discharge tubes.
fParallel connection of 6 discharge tubes.
gParallel connection of 204 discharge tubes.
Figure 4:The schematic configuration of Otto plate based ozone
generator. (A: grounding electrode of water cooled hollow box;
B:dielectric glass; C: discharge gap; D: high-voltage electrode of
water cooled hollow box; E: transformer. The size of discharge cell
could be referred to Table 1).
pressure and temperature (Khadre etal. 2001). The special
odor may be possibly used to indicate the presence of
ozone and to provide a warning of toxic exposure in
advance. The other physical properties of ozone such as
melting point, boiling point, and dipole moment are listed
in Table2. It can be seen that, for example, ozone is a weak
polar molecule taken into account of its dipole moment,
demonstrating that ozone may exhibit various properties
(i.e. solubility and chemical reactivity) in water solution
(Beltrán et al. 2007). The most distinguishable physical
property of ozone concerning its potential application in
water treatment is its solubility. By conducting iodometric
titration, solubility of ozone in water was first accurately
determined to be 0.834 l/l (1°C, 741.5mm Hg), which was
13 times as much as that of oxygen in temperature range of
0–30°C (Rubin 2002), although some other different solu-
bility data had been reported (Rice etal. 1981, Roth and
Sullivan 1981).
The dissolution equilibrium of ozone in aqueous solu-
tion is basically influenced by the partial pressure of ozone
in gaseous phase, the temperature of water, and its bubble
size distributed in solution. The dissolved concentration
of ozone at fixed temperature is linearly proportional to
its partial pressure, following the famous Henry’s law
(Biń 2004). Therefore, the solubility of ozone in aqueous
solution can be increased as partial pressure of ozone in
gaseous phase is elevated, based on which the increase
of gas pressure in ozone contactor will be advantageous
to ozone dissolution (see the discussion in Section 4.6).
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Discharge gap
Discharge gap
HV fuse
Cooling water Outer steel tube
Glass tube
Gas flow
Metal plating
Figure 5:The schematic configuration (A) and the real figure (B) of the modern industrialized tube ozone generator. Adapted with permis-
sion from Kogelschatz 2003, Copyright 2003, Plenum Publishing Corporation.
Table 2:The physical properties of ozone.
Physical properties Data Reference Physical properties Data Reference
Melting point
(mm Hg)/°C
-.±.Jenkins and Dipaolo Thermal conductivity
(-°C)/cal/(s) (cm) (°C/cm)
.cWaterman etal. 
Boiling point
(mm Hg)/°C
-.±.Birdsall etal.  Dielectric constant .dThorp 
Critical temperature/°C-.aBirdsall etal.  Solubility in water/(l/l)e Roth and Sullivan 
Evaporation heat
(at b.p.)/(Kcal/mol)
. Hersh etal.  °C.
Dipole moment/Debye . Gill and Laidler  °C.
Vapor pressure/mm Hg .bBirdsall etal.  °C.
Gaseous density at NTP/(g/l). Jenkins and Dipaolo °C.
Solid density
(. K)/(g/cm)
. Jenkins and Dipaolo °C.
aData were derived from 54.6 atm.
bData are valid at 80.2K (-193°C), the triple point of ozone.
cThe pressure at which thermal conductivity was measured is the vapor pressure of liquid at that particular temperature.
dData were obtained at 1 kilocycle per second at temperature of -183°C.
eData were obtained at 1 atm.
It is well recognized that dissolution of gas into water is
an exothermic process. Bablon etal. (1991) reported that
ozone solubility increased as the temperature of water
decreased on the basis of a negative logarithmic relation-
ship between solubility ratio [Sr, referred to Eq. (48)] and
water temperature in the range of 0.5–43°C. Moreover, the
dissolution rate of ozone can be accelerated by decreasing
ozone bubble size due to improved mass transfer, higher
inner pressure in the micro-bubble, as well as longer
hydraulic retention time (HRT) relative to macro-bubble
ozone (Chu etal. 2008). Zheng etal. (2015) investigated the
influence of bubble size (micro-bubble and macro-bub-
ble) on ozone dissolution behavior. The ozone dissolution
can be defined as Eq. (49), where kL (m/s), a (m-1), kd (s-1), ρL
(mg/l), and ρL* (mg/l) were assigned to the mass transfer
coefficient from gaseous into aqueous phase, the specific
surface area for gas-liquid contact per volume, the decom-
position rate constant of ozone, the dissolved ozone con-
centration in water, and the equilibrium concentration of
ozone in aqueous solution, respectively. It was revealed
that kLa of micro-bubble ozone contactor was 2.2 times
higher than that of macro-bubble ozone contactor, which
is consistent with results reported by Chu etal. (2007)
and Liu etal. (2010). Gong etal. (2007) confirmed that the
solubility ratio was higher in the case of smaller diameter
bubble by developing the solubility ratio [Eq. (50)] from
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C. Wei etal.: Ozonation in water treatment11
Eq. (48), where jb (kg/s), tb (s), c0 (kg/m3), and Rb (m) were
the instantaneous mass transfer rate of a single bubble,
the contact time, the initial concentration of dissolved gas
inside bubble, and the bubble diameter, respectively. In
spite of the different dissolution behavior of ozone, the
dissolved ozone concentration would be equal to its solu-
bility at fixed temperature according to Henry’s law, until
equilibrium dissolution is established. In addition, chemi-
cal reactions responsible for ozone decomposition initi-
ated by dissolved organic matter (DOM), OH- in solution
(alkaline condition), and solid catalysts may considerably
influence the dissolution of ozone.
r3 3
(mg/l Oin water)/(mg /l Oin gas ph ase)S
(-) -
r3 3
(mg/l Oin water)/(mg /l Oin gas
( )
d/(4 /3)
Rcj (50)
3.2 The chemical reactivity of ozone
3.2.1 The chemical properties of ozone
The extensive applications of ozone in wastewater treat-
ment, potable water disinfection, air cleaning, and food
preservation are fundamentally determined by its active
chemical properties including strong oxidative potential
and reaction with a selection of compounds. In general,
ozone is a relatively strong oxidation reagent (Table 3),
due to +4 valence in its special molecular formula
from the perspective of valence state theory (Zhang 2008).
The selective chemical reactivity of ozone is closely
associated with its molecular and electronic structure.
Ozone molecule (O3) is composed of three oxygen atoms,
in which one atom is the center atom, while other two
atoms are connected to the central one by covalent bond,
forming an isosceles triangle-like (bond angle of center
atom is 116°49′±30, the length of two O-O bonds are both
1.278±0.003 Å, see Figure 6A) molecular structure via sp2-
hybridized rearrangement (Trambarulo et al. 1953). The
ground state of ozone, for example, can be identified
as a biradical pattern (Figure 6B) with two σ-bonds and
a long or formal π-bond associated with terminal atoms
(Hay etal. 1975, Harcourt etal. 1986). The four closed-shell
resonance structures (Figure 6C) of ozone molecule were
proposed according to its electronic configuration and
spatial model (Langlais etal. 1991, Kalemos and Mavridis
Table 3:The standard redox potential of common oxidizing
OxidantEθ(V) Redox potential
relative to ozone
Oxidant Eθ (V) Redox potential
relative to ozone
F. . MnO
-. .
SO˙-b .–. .–. ˙OHads
d>. >.
c.–. .–. ClO. .
O. . ClO-. .
O. . Cl. .
HO. . Br. .
HO˙. . O. .
aMost data were derived from Lin and Yeh (1993).
bData were derived from Cong etal. (2015).
cThe redox potential of ˙OHfree depends on solution pH. In acid
solution, Eθ(˙OH, H+/H2O)=2.4–2.7 V derived from Wang and Liang
(2014), while Eθ(˙OH/OH-)=1.89–2.0 V in alkaline condition referred
from Neta etal. (1988).
dData were derived from Lawless etal. (1991).
2008). Taking into account these resonance structures,
one oxygen atom is electron-deficient exhibiting electro-
philic property, suggesting that ozone can efficiently react
with unsaturated functional groups, amines, sulfides, and
other reductive compounds, whereas one oxygen atom
is electron-rich possessing nucleophilic property, thus
preferring to react with carbonyl, carbon-nitrogen bond
containing substances and strong Brønsted acids (Riebel
etal. 1960, Cocace and Speranza 1994). Therefore, ozone
is a robust reactive reagent that is capable of oxidizing a
large spectrum of pollutants by direct ozonation reactions
or by indirect reactions with oxygen-containing radicals
produced from ozone decomposition.
3.2.2 The reaction mechanism of ozone with inorganic
Ozonation of inorganic compounds is basically the redox
reactions featured with electron-transfer from reductive
substrates to ozone. Apart from some direct electron-
transfer reactions including ozone reaction with hydrogen
peroxide ions [Eq. (51)], superoxide anion radicals [Eq.
(52)], and metal-organic anion complexes [Eq. (53)] (Zuo
and Hoigné 1992), the ozonation mechanisms of most
inorganic compounds are, however, carried out by the
oxygen-atom transfer reactions (Løgager etal. 1992). The
resulting unstable intermediates can subsequently hydro-
lyze into the oxidized state. Løgager etal. (1992) inspected
the reaction mechanism of Fe2+ with ozone in acid solu-
tion by stopped-flow spectrophotometer. The ferryl ion
(FeO2+), an unstable intermediate compound, produced
from oxygen-atom transfer from ozone to Fe2+ [Eq. (54)]
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12C. Wei etal.: Ozonation in water treatment
was measured. It was then subjected to decomposition
dominant in Eq. (55) with the excess of Fe2+, whereas
the decay reaction was mainly governed by Eqs. (56–57)
with the excess amount of ozone. By contrast, ferrous
ion was found to be oxidized by ozone directly via elec-
tron transfer, as shown in Eq. (58) (Holgné etal. 1985), as
Fe2+ could initiate chain reactions responsible for ozone
decomposition, which will be explained later. Concerning
cyanide oxidation, it was indicated that ozonation of CN-
was achieved by forming cyanate (CNO-) via oxygen-atom
transfer mechanism (Parga et al. 2003). CNO- was then
easily further converted into CO3
2-, NH4+-N, NO3
- by ozona-
tion or hydrolysis reactions depending on the introduced
ozone concentration and solution pH value.
+→+ (51)
+→+ (52)
24 24 3
…+→… + (53)
+→ +
OFeFeO O (54)
22 3
+→++ (56)
+→ + (57)
+→+ (58)
3.2.3 The reaction mechanism of ozone with organic
The ozonation of organic contaminants is basically limited
by selective oxidation characteristics of ozone with spe-
cific functional groups. In general, ozonation mecha-
nisms of organic compounds only comprise 1,3-dipolar
cycloaddition, electrophilic addition, and nucleophilic
Figure 6:The spatial structure (A), biradical pattern (B), and closed-shell resonance structures (C) of ozone molecule. Original drawing of
(A) was done by us using ChemBioOffice software and (B) was adapted with permission from Hay etal. (1975). Copyright 2008, AIP Publish-
ing LLC., and Harcourt etal. (1986), Copyright 1969, Royal Society of Chemistry. (C) was adapted from Langlais etal. (1991).
Figure 7:The mechanistic model for the ozonation of olefins
(adapted from Beltrán etal. 2007, von Gunten 2003b).
addition (non-aqueous solution), responsible for degrad-
ing olefins, aromatic compounds, and carbon-nitrogen
bond containing compounds, respectively (Riebel et al.
1960, von Sonntag and von Gunten 2012). The 1,3-dipolar
cycloaddition of olefins subjected to ozone attack, for
example, is shown in Figure 7 (von Gunten 2003b). It
involves the generation of highly unstable trioxolanes
[i.e. ozonide (II) or (III)] through Criegee mechanism, after
which carbonyl compound (VI) and hydroxyhydroperox-
ide (V) can be produced from (III) decomposition. The
intermediate (V) will be then decompose into a carbonyl
compound (VI) and hydrogen peroxide (VII) (Dowideit
and von Sonntag 1998). On the other hand, the cycload-
dition for some substituted olefins does not necessarily
result in the production of ozonide, possibly owing to
sterical hindrance effect as revealed by partial oxidation
of carbon-carbon double bond (Bailey and Lane 1967).
The electrophilic addition is the primary principle of
initial reaction of ozone with electron-rich organic com-
pounds, including phenols/phenolates, aromatic rings
substituted with electron-donating groups, and polyaro-
matic compounds (Larcher et al. 2012). The ozonation
of phenol, for example, is carried out preliminarily by
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C. Wei etal.: Ozonation in water treatment13
formation of carbon cation with substitution on ortho- and
para- positions of electron donating -OH group (Beltrán
etal. 2007). The ozone adduct could be stabilized by the
delocalization of electrons in aromatic ring, allowing four
resonance structures as shown in Figure 8A. Then carbon
cation can be transformed into catechol, hydroquinone,
and benzoquinone via SN1 reaction and deprotonation
reaction. The resulting electrophilic substitution interme-
diates are further decomposed through cycloaddition and
hydrolysis, as illustrated in Figure 8B (Eisenhauer 1968).
In addition, the nitrogen-containing compounds (e.g. ali-
phatic amines) can also be ozonated through electrophilic
reaction at nitrogen atom. It has been shown that N-oxides
and hydroxylamines are formed by ozone adduct of amines
at lone electron pair on nitrogen atom followed by oxygen
loss reaction. Amine radical cations may be alternatively
produced from electrophilic ozonation by abstraction
Figure 8:Ozonation mechanism of phenol. (A) The resonance
structures of carbon cations (the case of ortho position adduct, E is
assigned to ozone); (B) the mineralization pathway of phenol (AO is
assigned to abnormal process). [(A) was adapted from Beltrán etal.
2007, (B) was adapted from Eisenhauer 1968].
Figure 9:The schematic mechanisms of nucleophilic reaction of
benzaldehyde (A); alkyl isocyanides (B); and Schiff bases including
N-benzylideneaniline, N-benzylidene-m-nitroaniline, and N-ben-
zylidene-t-butylamine (C) with ozone. [(A) was adapted with permis-
sion from Bailey 1958, Copyright 1958, American Chemical Society,
(B) was adapted with permission from Feuer etal. 1958 in which R is
assigned to alkyl chain, Copyright 1958, American Chemical Society,
(C) was adapted with permission from Riebel etal. 1960, Copyright
1960, American Chemical Society].
of an ozonide radical anion, leading to the generation of
dealkylated amine, ketone, or aldehyde through rearrange-
ment of these cations (von Sonntag and von Gunten 2012).
On the basis of resonance structures, ozone may also
behave as a nucleophilic reagent, being capable of react-
ing with electron-deficient positions of carbonyl groups,
carbon-nitrogen triple or double bonds. As illustrated
in Figure 9A, benzaldehyde can be nucleophilically
ozonated to form benzoic acid by the addition of ozone
on carbon atom of carbonyl groups, followed by loss of
oxygen (Bailey 1958). The carbon atom of carbon-nitrogen
triple bond (e.g. isonitriles) and the carbon atom of car-
bon-nitrogen double bond (e.g. Schiff bases) are primary
reactive sites upon nucleophilic ozonation, as described
in Figure 9B and C, respectively. The nucleophilic reac-
tion of Schiff bases (I) with ozone, for example, may
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14C. Wei etal.: Ozonation in water treatment
yield unstable intermediate (II), resulting from reaction
at carbon atom of double bond (Riebel etal. 1960). The
produced (II) will be then decompose into oxazirane (III)
and amide (IV) by ring closure and proton shift reaction,
respectively, with the carbon-nitrogen bond undestroyed.
The cleavage of carbon-nitrogen bond accompanied by
loss of oxygen, in contrast, could lead to the formation
of benzaldehyde (V). Regarding the aromatic ring sub-
stituted with electron withdrawing groups, nucleophilic
ozonation was more favorable than electrophilic ozona-
tion due to the decreasing electron density of aromatic
ring (Hong and Zeng 2002).
3.2.4 The reaction mechanisms of catalytic ozonation
Apart from direct reactions of ozone with a selection of com-
pounds, oxidation can also be achieved by indirect reaction
with oxygen-containing radicals (mainly ˙OH) generated
upon homogenous and heterogeneous catalytic decompo-
sition of ozone. The homogenous catalytic ozone decompo-
sition mechanisms in aqueous solution involving SHB [Eqs.
(52, 59–68)] and TFG [Eqs. (51, 69–75)] models were widely
accepted half a century after Weiss (1935) first described the
decomposition mechanism of ozone catalyzed by OH- ion
(Bühler etal. 1984, Staehelin et al. 1984, Tomiyasu etal.
1985). The general features of SHB and TFG models are
chain reactions, which were initiated by OH- and propa-
gated with the involvement of O3˙-, HO2˙, HO3˙. Thus, ozone
decomposition may be significantly promoted in basic solu-
tion (7<pH<10) whereas be less effective in acidic solution
(pH<3). It should be emphasized that some acidic anions
in solution could adversely affect ozone decomposition.
For example, O3 can be depleted by Cl- in acidic solution,
producing Cl2 and ClO3
- without the involvement of chain
reactions responsible for radical generation (Levanov etal.
2012, Levanov etal. 2015). Indeed, it has been reported as
a promoting effect of Cl2 and HOCl working as additional
oxidants on oxalic acid and p-hydroxylbenzoic acid degra-
dation during ozonation (Nowell and Hoigné 1992, Wang
etal. 2016). Although there is no direct evidence supporting
ozone reaction with CO3
2-, the presence of CO3
2- can inhibit
ozone decomposition by the removal of dominant chain
carrier radical, i.e. O3˙- (Nemes etal. 2000).
Chain initiationOOH HO
0 l/(m )
̇̇ ol sk+→
Chain propagationHOH
.9 10 l/(mol s)
→+ (60)
HO HO 5 10 l/(mol )
̇̇ sk
++→ (61)
HO HO 5.2 10 l/(mol )
++→ (62)
H OHO3.3 10 s
→+ (63)
HO OOH1.s 1
→+ (64)
̇ 10k
+→ (65)
HO HO O2.s
̇̇ 8 10k
→+ (66)
4422 3
Chain terminationHOHOHO2O510 l/(mol s
̇̇ ̇)k
4322 32
HO HO HO OO 510 l/(mol
̇̇ )k+→ ++
Chain initiation OOHHOO 40 l/(mols)k+→ +=
32 2
Chain propagation O HO OH OOH2030 l/(mol·s)
̇̇ k+→++ =−
OOHHOO 610 /(mol s)
Chain termination CO OH HO CO 4.21 0l/
(mol s)
33 33
CO OOCO 610l/(mo )
HCOOHHOHCO 8.510l/(mols)
CO HCO500 sk
→+ = (75)
HO HO Hp11.6K
→+ = (76)
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C. Wei etal.: Ozonation in water treatment15
Besides the applications of physical fields (e.g. UV
irradiation and ultrasonic vibration) and other soluble
catalysts (e.g. H2O2, Fe2+, Co2+, Mn2+, Co2+, Ni2+, Cu2+) are
the alternative approaches to catalytically decomposed
ozone into radical species in homogenous solution.
Principally, the energy transfer from applied physi-
cal fields to ozone could accelerate its decomposition.
The formation of ˙OH, for example, can be remarkably
strengthened by coupling ultrasonic vibration with
ozone, primarily resulting from additional thermal
decomposition of ozone occurred in cavitation bubbles
(Hou et al. 2013). The decomposition of ozone can be
facilitated by deprotonated H2O2 (HO2
-), producing O3˙-
and HO2˙ [Eqs. (51, 76)] via electron transfer reactions,
which eventually leads to ˙OH formation (Torres-Socías
etal. 2013). The ozone reactions catalyzed by dissolved
transitional metal ions essentially comprise two mecha-
nisms: (1) the reaction catalyzed by free-state ions and
(2) the reaction strengthened by complex state of metal
ions. Basically, direct electron transfer from free-state
transitional metal ions with variable valence to ozone
will favor its decomposition. Ferrous ion, for example,
can be oxidized to its triple valence state by ozone with
the formation of O3˙- [Eq. (58)] (Holgné etal. 1985), which
is readily transformed into ˙OH [Eqs. (62, 64)]. The ˙OH
may also be produced from the hydrolysis of FeO2+ [Eq.
(56)], resulting from oxygen-atom transfer reaction [Eq.
(54)]. The oxidized Fe3+ can be then reduced to its diva-
lent state, as shown in Eq. (77), in which HO2˙ is produced
from Eqs. (59, 65, 66, 71) (Wilde etal. 2014). The coordina-
tion of metal ions with α-carbonyl or α-hydroxy contain-
ing carboxylate anions (e.g. pyruvic acid, oxalic acid,
citric acid) could, on the other hand, facilitate the cata-
lytic decomposition of ozone. This is owing to the high
electron density (donated by ligand) on the coordinated
metal site, being more favorable for electron abstraction
by ozone and leading to O3˙- formation (Zhang and Croué
2014c). Beltrán etal. (2003) studied the catalytic influ-
ence of homogenous Co2+ on ozonation of oxalic acid.
The results showed that CoC2O4 was formed via bidentate
complexation, increasing the reactivity of CoC2O4 with
ozone attributed to partial donation of electron density
from C2O4
2- to Co2+. The unstable intermediate (CoC2O4)+
could be produced [Eq. (78)] upon ozone reaction. After
that, Co2+ was regenerated by electron extraction from
carboxylic acid ligand to Co3+ [Eq. (79)], giving rise to the
appearance of superoxide ion radical and carbon dioxide
through decarboxylation reaction, as shown in Eq. (80)
(Zuo and Hoigné 1992, Pines and Reckhow 2002). Simul-
taneously, the generation of ˙OH was enhanced from O3˙-
and O2˙- via Eqs. (62, 64) and Eq. (52), respectively.
+→++ (77)
+→ + (78)
→+ (79)
+→ + (80)
22-32- -
≡+→≡ + (81)
Transitional metal oxides, reactive metal supported on
metal oxides, minerals, clays, carbon-based materials,
and ceramic honeycomb have been applied to assist
in the degradation of various organic compounds by
ozonation (Zhao et al. 2009a, Nawrocki 2013) and to
investigate the catalytic ozonation mechanisms (Ikhlaq
etal. 2013, Bing et al. 2015). The latest investigations
of various ozonation catalysts are listed in Table 4.
The mechanisms of heterogeneous catalytic ozonation
include (1) ˙OH generation from reaction of ozone with
surface hydroxyl groups, (2) ˙OH generation from reac-
tion of ozone with metal ion site with variable valence,
and (3) elimination of ozone recalcitrant compounds
via intermediate metal-organic complexes. The reac-
tion mechanism of ozone decomposition at surface
hydroxyl group sites depends on state of surface charge
(Zhang et al. 2015). The surface hydroxyl groups are
positively charged at solution pH below the point of
zero charge (pHPZC) of catalysts, promoting decomposi-
tion of ozone driven by electrostatic force and hydro-
gen bond between ozone and positively charged surface
(Zhao etal. 2009b). With the increase of solution pH,
surface hydroxyl groups are neutrally or negatively
charged at solution pH equal or above pHPZC, favor-
ing interaction of ozone with catalyst surface due to
the electrophilic nature of ozone (Jung etal. 2007). A
positive linear correlation between density of surface
hydroxyl groups and catalytic performance of catalyst,
for example, was established using surface properties
measurements (Zhao et al. 2015). In addition, metal
ion sites with variable valence present on surface of
solid catalysts are also reactive sites for ˙OH generation.
CoOx/ZrO, for instance, was confirmed to be an efficient
catalyst for ozone decomposition due to the oxidation
of surfaceCo2+ by O3, resulting in O3˙- and therefore ˙OH,
as shown in Eqs. (62, 64) (Hu etal. 2008). The electron
transfer from surface metal complexes to ozone can
accelerate its decomposition. It was observed that in
the presence of CuO and oxalate, the coordination of
Cu2+ with organic ligand could reduce the redox poten-
tial of Cu3+/Cu2+ couple (Eθ=2.3 V), making Eq. (81) pos-
sible, which would be favorable for electron extraction
by ozone (Zhang and Croué 2014c).
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16C. Wei etal.: Ozonation in water treatment
Table 4:The latest investigations for the catalytic ozonation of various contaminants.
Pollutant Catalyst Preparation methods of
the catalysts
Ozone dosage pHbResearch remarks Reference
Pharmaceuticals γ-AlOCommercially available . g/l . ml/min , ., ˙OH mediated reaction Ikhlaq etal. 
Ibuprofen FeO/AlO@
Incipient wetness
. g/l  mg/l
Ozone decomposed into oxygen atom at Al+
sites, oxygen atom transformed into ˙OH and
O˙- at Fe+ sites
Bing etal. 
Phenacetin CuFeOCo-precipitation and
 g/l .–. mg/l
.Ozone efficiently broke down the initial
model compound structure, ˙OH improved the
mineralization of intermediates
Qi etal. 
Phenol α-, β-, γ-MnOHydrothermal  g/l . mg/min Particular Mn-O bonds reacted with ozone to
generate oxygen species
Dong etal. 
Phenol ZnAlOHydrothermala g/l . mg/min .±.Surface hydroxyl groups and chemisorbed
water on catalyst were catalytic active sites
Zhao etal. 
para-chlorobenzoic acidPeroxymonosulfate
Commercially available . m . m The concomitant production of ˙OH and SO˙-
arose from synergistic effect of O/PMS
Cong etal. 
Several emerging
TiOCommercially available . g/l  mg/l
The combination of photocatalytic
oxidation and ozonation enhanced ozone
decomposition and ˙OH generation
Quiñones etal.
para-chlorobenzoic acidMultiwalled carbon
Commercially available  mg/l  μeStrong correlation between RCT
f values and
surface oxygen concentrations was found
Oulton etal. 
Oxalic acid ElectrongCommercially available  mg/l
Electro-reduction of O to ˙OH at the cathode,
and O decomposition to ˙OH at high local pH
near cathode
Wang etal. 
Indigo Carboxylated
carbon nanotubes
Chemical vapor deposition
and acidification
 mg/l . mg/l
.Functional -COOH groups directly reacted with
amino groups of indigo.
Qu etal. 
aZnAl2O4 was synthesized by a one-step hydrothermal method, followed by calcination at different temperature.
bThe pH is referred to as initial pH value without being specified.
cThe chemical formula of Peroxymonosulfate is 2KHSO5·KHSO4·K2SO4.
dSix commonly detected emerging contaminants including acetaminophen, antipyrine, bisphenol A, caffeine, metoprolol and testosterone were degraded by photocatalytic ozonation.
eThe pH value was buffered by 5 m phosphate.
fThe RCT value was defined as total exposure to ˙OH relative to total O3 exposure during ozonation.
gThe electron for catalytic ozone decomposition was provided by electrolysis of cathode (Pt or boron-doped diamond electrode).
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C. Wei etal.: Ozonation in water treatment17
3.3 Summary
Ozone exhibits the characteristics of higher solubility
relative to oxygen and active chemical reactivity asso-
ciated with its resonance structures. It mainly involves
two mechanisms concerning the removal of inorganic
and organic compounds. The first one is direct reaction
with molecular ozone proceeding via electron transfer or
oxygen-atom transfer pathways for ozonation of inorgan-
ics. The fundamental pathways for direct ozonation with
organics are the cycloaddition, electrophilic addition, and
nucleophilic addition reactions. The second one is indi-
rect oxidation carried out with oxygen-containing radicals
produced from catalytic decomposition of ozone. Owing
to its special physicochemical properties, ozone has been
potentially applied in water and wastewater treatment.
Indeed, it suffered from some limitations, such as insuf-
ficient oxidation capability, scavenging effect of inorganic
carbon, and DOM on radical-mediated reactions. All of
these definitely need to be carefully investigated in future
4 The application of ozone in water
and wastewater treatment
On account of its convenient generation and strong oxi-
dation ability, ozone is extensively applied in chemical
products manufacturing, food processing, aquaculture,
health-medical care, and atmospheric and aqueous pollu-
tion control industries. As a broad-spectrum disinfectant,
ozone is able to inactivate micro-organisms and decompose
precursors of chlorine disinfection by-products (DBPs)
with concomitant removal of odor and color in drinking
water treatment. The bio-refractory organic compounds
and reductive inorganics residual in biologically treated
industrial wastewater can be further oxidized by ozone. In
this review, the applications of ozone technology in potable
water disinfection and advanced treatment of industrial
wastewater are analyzed. The performance of ozonation
reactor in which ozone reaction occurs is assessed. The
energy consumptions in terms of ozone generation and
contaminants degradation are finally evaluated.
4.1 Ozone disinfection
Micro-organisms such as bacteria, fungi, viruses, proto-
zoa, bacterial, and fungal spores in water can be effec-
tively inactivated by ozone for disinfection purposes. The
overall disinfection mechanism is realized as cell lysis
results from reactions with ozone molecule or oxygen-
containing radicals (e.g. ˙OH, O2˙-, HO2˙) (Hunt and Marinas
1997). Generally, microbial inactivation is induced by
ozone reaction with several cellular constituents includ-
ing proteins, unsaturated lipids and respiratory enzymes
in cell membranes, peptidoglycans in cell envelopes,
and nucleic acids in cytoplasm (Khadre etal. 2001). The
destruction of lipoprotein on outer shell and lipopolysac-
charide on inner cell wall, for example, were realized by
the oxidation of double bonds in unsaturated lipids and
sulfhydryl groups in enzymes with ozone (Chang 1971).
This will consequently lead to the reduction of cell perme-
ability, eventually causing rapid cell death. The complete
inactivation of microbial cell by ozone should be even
further achieved through the damage of genetic genes,
parasitic bacteria, phages, and mycoplasma residual in
dead cell. It was indicated that the reaction of nucleic
acids inside cell with ozone was carried out by opening
the circular plasmid DNA, producing single- or double-
strand breaks in plasmid DNA and ultimately resulting
in the decrease of transcription activity (Roy etal. 1981).
The clarity of the above basic principles will underpin
the development and application of ozone disinfection in
water treatment. The recent research activities concerning
ozone disinfection are summarized in Table 5.
4.1.1 Non-catalytic ozonation for disinfection
It is well established that direct ozone disinfection mainly
focused on inactivation of micro-organisms residual in
water and elimination of DBPs. The inactivation of two
antibiotic resistance genes (ARGs) in municipal waste-
water by ozone or UV disinfection was compared (Zhuang
etal. 2015). The results showed that although ozone dis-
infection (ozonation dose of 177.6 mg/l for 1.68–2.55 log
reductions of ARGs) achieved less inactivation of ARGs
than did UV irradiation (UV dose of 12,477 mJ/cm2 for
2.48–2.74 log reductions of ARGs), the 16S rDNA was more
efficiently destroyed by ozonation. Hiragaki etal. (2015)
applied ozone foam to inactivate Pseudomonas fluores-
cens with its initial number of more than 108, suggesting
that survival rate log (N/N0) decreased obviously along the
treatment time. Zimmermann etal. (2011) assessed ozone
disinfection for full-scale municipal wastewater treat-
ment, indicating that disinfection capacity was 1–4.5log
units in terms of total cell counts (TCC) and 0.5–2.5 log
units for Escherichia coli.
Agbaba et al. (2015) elucidated the influence of
applied ozonation on natural organic matter (NOM)
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18C. Wei etal.: Ozonation in water treatment
Table 5:The recent publications in ozone disinfection.
Ozone oxidation Wastewater quality Reaction conditions Main results Reference
Ozonation integrated
with ceramic
DOC: .–. mg/l, UV:
.–. m-, Br-: . mg/l, pH:
Ozone dose:  mg/l,
Single membrane sheet:
××mm with total
filtration  m
The removal of organic matter was significantly
improved with ozonation in advance, and a close
linear relationship of DOC with the formation
potential of DBPs was established
Fan etal. 
Ozonation DOC: .–. mg/l, pH: .–
., UV: .–. cm-,
alkalinity: –mg CaCO/l
Ozone doses: .–.mg O/mg
DOC, ozone delivery rate:  l/h
Increasing ozone dose (.–.mg O/mg DOC)
generally led to reductions in DOC (–%) and
trihalomethane formation potential (–%)
Agbaba etal. 
Ozonation catalyzed
by TiO
DOC: .±. mg/l, pH: .±.,
total alkalinity: ±mg CaCO/l
Ozone doses: .–.mg O/mg
DOC, TiO concentration: . mg/l
The application of TiO enabled an improved
ozonation in terms of trihalomethane (THM)
precursor content removal (up to %)
Molnar etal. 
flotation followed by
COD:  mg/l, pH: ., Escherichia
coli:  CFU  ml-, turbidity:
 NTU
Ozone doses:  g/h, average
diameter of flocs: –μm,
rise rate of flocs: – m/h
The integrated processes resulted in disinfected
(Escherichia coli<. CFU  ml-) and clarified
water ( NTU)
Etchepare etal. 
Equivalent concentration for amino
acids: – μg N l-, pH: 
Ozone doses:  mg/l, chlorine
doses:  mol/mol precursor
The increases in the formation of chloropicrin
from ozonation-chlorination relative to
chlorination alone varied from % for
-aminophenol to % for ala-ala for the four
amine surrogates
Bond etal. 
Ozonation with
magnetic ion
exchange resin
TOC: .–. mg/l, UV: .–
. cm-, bromide: – μg/l
Resin doses: .–. ml/l, alum
flocs doses: – mg/l, ozone
doses: mg O/mg TOC
The amount of bromate produced by ozonation of
resin-treated waters was similar to or less than
that of raw water and alum-treated water
Kingsbury and Singer
Ozonation DOC:  mg/l, bromide concentration:
 μg/l, pH: .±.
Dissolved ozone:  mg/l, water
bath: ±°C
The use of ozone as a primary disinfectant may
cause a shift to more brominated DBPs during
subsequent chlorination
Mao etal. 
Ozonation DOC:  mg/l, bromide concentration:
– μg/l, pH: .–.
Ozone doses: –mg O/mg
DOC, water bath: ±.°C
The concentration of BrO
- steadily increased
with increasing O dosage at high Br-
concentration (> μg/l)
Lin etal. c
Ozonation Initial concentrations of 
β-estradiol and estrone:  ng/l
and  μg/l, turbidity: .±. NTU
Ozone doses: .– mg/l, water
bath: ±°C, pH: ±.
Both  β-estradiol and estrone have been
found to form two main by-products after ozone
disinfection of surface water, following the same
reaction pathways
Pereira Rde etal.
Ozonation TOC: .–. mg/l, UV:
.–. cm-, SUVA: .–
. l/mg m, pH: –
Ozone doses: .–.mg/l,
NaOCl:TOC= (wt),
Ozonation reduced the formation of HANs from
EOM, increased the yields of HANs from IOM
Zhu etal. 
Ozonation, or UV
COD: – mg/l, alkalinity:
– mg/l, original gene copies:
.–. copies/ml
UV irradiation: –,
 mJ/cm, UV wave length:
nm, ozone dose: . mg/l
Inactivation of antibiotic resistance genes
underwent by increased doses of disinfectors,
and S rDNA was more efficiently removed by
Zhuang etal. 
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C. Wei etal.: Ozonation in water treatment19
Ozone oxidation Wastewater quality Reaction conditions Main results Reference
Ozonation DOC: .–. mg/l, pH: .±. Ozone doses: .–. g O/g
Ozonation decreased TCC by –. log units, not
consistently correlating with ozone doses. E. coli
concentrations reduction only partially correlated
with ozone exposure
Zimmermann etal.
Ozonation/ BAC
DOC: . mg/l, dissolved organic
nitrogen (DON): . mg/l
Dissolved ozone doses:
.–.mg/l, density of BAC:
 g/l
Abrupt termination of ozonation resulted in some
N-DBP precursors and then substantial formation
of N-DBP during chlorination
Chu etal. 
Ozonation catalyzed
by Fe-Mn oxide or
DOC: .–. mg/l, UV:
.–. cm-, pH: .–.
Ozone dose: . mg/l, ozone flow
rate:  ml/min, catalyst doses:
.–. g/l
The use of both catalysts in catalytic ozonation
effectively removed the precursors of DBPs,
resulting in the formation of fewer THMs and nine
haloacetic acids (HAA).
Chen etal. b
Table 5(continued)
fractionation and DBPs formation. The results indicated
that the reactivity of all individual NOM fractions towards
trihalomethane formation decreased by 47–69% relative
to raw waters when subjected to ozone oxidation (0.8mg
O3/mg dissolved organic carbon [DOC]), which is con-
sistent with findings from Chu etal. (2012). In contrast,
the ozonation treatment of nitrogen containing organic
compounds could increase the formation potential of
DBPs. It was reported that the disinfection of alanylala-
nine with ozone prior to chlorination could accelerate the
formation of chloropicrin (Bond etal. 2014). The authors
suggested that ozone is more effective than chlorine in
mediating an oxidation step in chloropicrin formation,
most plausibly involving conversion of an amine group
to a nitro group. Similarly, Zhu et al. (2015) indicated
that haloacetonitriles (HANs) formation from extracel-
lular organic matter (EOM) on the course of ozonation
decreased from 0.73 to 0.35μg/mg C, whereas HAN yields
from intracellular organic matter (IOM) raised by 21.2%.
This was owing to the lower nitrogen-organic concentra-
tion in EOM than that in IOM, as ozone disinfection might
transform amine compounds to nitrous compounds, being
the precursors of HAN. Mao etal. (2014) also reported a
similar result concerning the increase of formation poten-
tial of chloral hydrate and trichloro nitromethane with
rising ozone doses up to 6mg/l. Moreover, the ozone dis-
infection of emerging estrogens bearing surface water will
lead to the formation of DBPs including 17 β-estradiol and
estrone (Pereira Rde etal. 2011). The authors conducted
experiments at two different concentrations of estrogens
in water (100 ng/l and 100 μg/l) and at varying ozone
dosages (0–30 mg/l). Results showed that these DBPs have
been found to be formed at both high and low concentra-
tions of estrogens and to be persistent even after exposure
to high ozone dosages.
The ozone disinfection of Br- containing water aroused
wide concern due to a potential carcinogen DBP, BrO3
recognized by the World Health Organization (WHO) and
United States Environmental Protection Agency (USEPA)
(USEPA 1998, WHO 2011). It was established that BrO3
- can
be produced from the reaction of ozone with OBr-, result-
ing from ozonation of Br- (Pinkernell and von Gunten
2001). Mao etal. (2014) found that bromate incorporation
factor values of trihalometanes (THMs), trihaloacetic acid
(THAA), dihaloacetic acid (DHAA), and dihaloacetonitrile
(DHAN) increased from 0.62, 0.37, 0.45, and 0.39 without
ozone exposure to 0.89, 0.65, 0.62, and 0.89 at O3 dose
of 6 mg/l, respectively. The production of bromate could
be significantly influenced by its precursor and trans-
ferred ozone doses. Lin etal. (2014c) demonstrated that,
for example, decreasing initial Br- concentration was an
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20C. Wei etal.: Ozonation in water treatment
effective approach in controlling the formation of BrO3
When Br- concentration was lower than 100 μg/l, while
keeping the ratio of O3 dosage equal to DOC, the BrO3
- pro-
duction was effectively suppressed.
4.1.2 Catalytic ozonation for disinfection
The application of catalysts on the course of ozonation
will generate more reactive and less selective radical
species, which are expected to accelerate the elimination
of DBP precursors as well as residual DOC. The effect of
Fe-Mn oxide and TiO2/α-Al2O3 employed as ozonation cat-
alysts on the formation of DBPs was studied (Chen etal.
2014b). Results pointed out that ozone reactions catalyzed
by Fe-Mn oxide and TiO2/α-Al2O3 effectively removed the
precursors of DBPs, improving 77.5% and 30% of THMs
removal capability compared with non-catalytic ozona-
tion, respectively. Molnar etal. (2012) evaluated the perfor-
mance of catalytic ozonation (O3/TiO2) and non-catalytic
ozonation in the content and structural change of THM
and HAN precursors in groundwater. The results indicated
that TiO2-catalyzed ozonation (18% DOC removal) favored
more effective reduction of NOM in comparison with non-
catalytic ozonation (6% DOC removal). On the other hand,
the structural changes of NOM by both treatments led to
a 70% increased proportion of hydrophilic fraction (HI),
which were the most reactive THM and HAN precursors.
4.1.3 Physical separation techniques assisted ozone
In addition to combining catalysts with ozonation for the
purpose of removing DBPs and their precursors, coupling
various separation technologies with ozone oxidation has
been suggested as promising alternatives. Fan etal. (2015)
examined the degradation of DBP precursors by integrat-
ing ozonation with ceramic ultrafiltration from micro-
polluted surface water. The integrated process enabled
73% of DOC removal, in which the reduction of 77% of tri-
halomethane precursor, 76% of haloacetic acid precursor,
83% of trichloracetic aldehyde precursor, 77% of dichloro-
acetonitrile precursor, 51% of trichloroacetonitrile precur-
sor, 96% of 1,1,1-trichloroacetone precursor, and 63% of
trichloronitromethane precursor can be achieved. The per-
formance of magnetic ion exchange (MIEX) resin followed
by ozonation in controlling bromate and chlorinated DBPs
formation was investigated (Kingsbury and Singer 2013).
Results demonstrated that MIEX resin removed 39–85%
of trihalomethane formation potential (THMFP), while
35–45% reduction of THMFP was achieved by ozonation.
This suggested that MIEX resin can be a more promis-
ing pre-treatment for ozone disinfection, as it achieves
superior THM and bromate precursors removal. Chu etal.
(2015) evaluated the N-DBP formation during ozonation
(continuously bubbling and terminated bubbling) cou-
pling with biological activated carbon (BAC) filtration. It
was indicated that more N-DBP precursors passed into
post-BAC water when ozonation was terminated, resulting
in greater formation of N-DBP when the water was sub-
sequently chlorinated, compared to a parallel BAC filter
under continuously bubbling ozonation. This resulted
from the generation of new N-DBP precursors such as
nitrogen containing soluble microbial products (SMPs),
which were likely produced from endogenous respiration
by micro-organisms when ozone bubbling was terminated
(Schiener etal. 1998, Chu etal. 2010).
In summary, inactivation of micro-organisms and
inhibition of conventional chlorinated DBPs can be effec-
tively achieved by ozone disinfection. Nevertheless, this
technology suffers from several limitations as an alterna-
tive approach to chlorine disinfection. Firstly, ozone reacts
with NOMs residual in water to produce low-molecular
weight oxygenated by-products, which are generally more
biodegradable and ozone recalcitrant. The biological
regrowth in the distribution system may be, however, stim-
ulated by these intermediates. Secondly, ozone is unstable
in water with its half-life time <1h under common condi-
tions of drinking water supplement, resulting in insuffi-
cient continuous disinfection functionality (Glaze 1987).
Thirdly, the ozonated by-products including bromate as
well as N-DBPs (e.g. N-nitrosodimethylamine, chloro-
picrin and HANs) have been recently identified during
ozone disinfection (Schwarzenbach etal. 2006, Parrino
etal. 2014, Marti etal. 2015). Moreover, ozonation of some
nitrogen-containing organic compounds will, however,
increase the formation potential of DBPs during following
chlorination. In order to overcome these deficiencies, the
pre-treatments responsible for removing DBPs precursors
such as Br- and amine group containing organics will be of
necessity for ozone disinfection.
4.2 Ozonation of cyanide containing
Cyanide in wastewater is an important issue because of its
toxic effect, which normally existed in the form of hydro-
gen cyanide, anion, and various cyanide complexes. The
free cyanide ions, including cyanide anion and hydro-
gen cyanide either in gaseous or aqueous phase, are the
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C. Wei etal.: Ozonation in water treatment21
most poisonous forms (Dash etal. 2009). The ozonation
of cyanide in various sources is promising due to the
intense reactivity of cyanide with ozone and complete
elimination of cyanide without undesirable by-products.
The preliminary product of cyanide ozonation is cyanate,
which can be further degraded via hydrolysis or ozona-
tion proceeding, however, at relatively slower reaction
rate than the oxidation of cyanide (Mert etal. 2014). The
oxidation of cyanide by ˙OH produced upon ozone decom-
position, comparatively, is carried out by producing CNO-
as the less hazardous intermediate through two pathways
[Eq. (82) and Eqs. (83–85)] (Gurol and Bremen 1985a).
Although hydrogen cyanide is less reactive with ozone, it
can be readily degraded by ˙OH-mediated reactions [Eqs.
(86–87)]. The research activities in cyanide abatement by
means of catalytic and non-catalytic ozonation are con-
cluded in Table 6.
+→ + (82)
2 CN (CN)
HO+→ +
̇ (88)
4.2.1 Non-catalytic ozonation for cyanide removal
The non-catalytic ozonation of cyanide bearing waste-
water is usually used in alkaline solutions, in which the
Table 6:The ozonation of cyanide containing wastewater.
Solution system Ozone dosage Solution pHResearch remarks Reference
O/ CAC Synthetic solution –. mg/l .CN- first adsorbed on CAC surface
reaching adsorption equilibrium
condition, then oxidation of CN- to
CNO- occurred
etal. 
Ozonation and
– mg/l .–.Ozonation pre-treatment improved
the wastewater biodegradability
Cui etal. 
Mining effluent  mg/min .Activated carbon and O
exhibited synergetic effect on the
decomposition of cyanide
etal. 
O/HO/UV Jewellery producing
–The combination of catalytic
ozonation and photolysis
promoted radical generation
Mert etal. 
O/HOAutomobile industry
– mg/l –Deprotonated HO could
effectively decompose ozone
into ˙OH at basic condition,
accelerating cyanide removal
Mudliar etal. 
O/HOSynthetic solution .–. mg/l –Indirect oxidation of CN- was more
effective than direct ozonation
Kepa etal. 
OCassava starch
.– g/haOzonation of cyanide was limited
by the liquid phase volumetric
mass transfer coefficient of ozone
etal. 
OSynthetic solution .–. mol/molbThe cyanide oxidation depended
mainly on the specific ozone dose
etal. 
OMining wastewater .–. g/min .Intermediate CNO- was less
reactive than CN- against ozone
Parga etal. 
O/TiO/UV Synthetic solution  cm/minc.Decomposition and adsorption
both accounted for CN- removal
etal. 
OSynthetic solution . g/min .The reaction stoichiometric ratio
of CN-:CNO-:O was ::.
etal. 
aThe ozone dosage was referred to ozone producing ability of an ozone generator.
bThe ozone dosage was the ratio of mol O3 feed per mol CN-.
cThe ozone dosage was a final ozone molar fraction in the gas phase of about 1550 ppmv.
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22C. Wei etal.: Ozonation in water treatment
oxidation ability of ozone can be enhanced by OH- ion. A
synthetic alkaline solution containing CN- was ozonated
in a counter-current bubble column, with special empha-
sis on specific ozone dose requirement (Barriga-Ordonez
etal. 2006). The oxidation efficiency mainly depends on
specific ozone dose, with 90% CN- removed under specific
ozone dose of 1.2 mol O3/mol CN-, which was suggested to
be the least specific ozone requirement for ozonation of
cyanide. This significant ratio of ozone with cyanide was
also reported by Gurol and Bremen (1985a). By contrast,
Zeevalkink etal. (1980) and Carrillo-Pedroza etal. (2000)
revealed that the stoichiometric ratio of cyanide ozone
reaction was equal to 1. The discrepancy may resulted
from the presence of other ozone reactive compounds and
different solution pH under which the experiments were
In case of cyanide containing industrial wastewater,
the ozonation mechanism of cyanide normally presented
in the form of metal ion-complex is more complicated than
that of free cyanide. A mechanistic study for describing
the degradation of free and copper-complexed cyanide
in a completely mixed ozone contactor was developed
(Kumar and Bose 2005). The authors proposed that copper
cyanide complexes could be ruptured due to interaction
with hydroxyl radicals [Eqs. (88)], without proper evi-
dence supporting the initial reaction of copper cyanide
complexes with ozone molecule. It was also confirmed
that the overall degradation rate of copper-complexed
cyanide was slower than that of free cyanide under identi-
cal conditions, as the breakdown of complexes consumed
extra ˙OH, which was generated from ozone decomposi-
tion. The produced copper presented in cuprous form was
oxidized to cupric form and subsequently precipitated as
Cu(OH)2 under pH value as high as 12.8. This is, however,
contradictory to the results from Gurol etal. (1985b), in
which the enhanced rate of copper cyanide degradation
as compared to that of free cyanide was due to the cata-
lytic activity of free Cu(I)/Cu(II) towards ozone decompo-
sition under pH value as high as 11.5.
4.2.2 Ozonation associated with other technologies for
cyanide removal
The combination of ozone oxidation with biological treat-
ment process has gained significant concerns in degrading
toxic and biological inert wastewater, as the biodegrada-
bility of wastewater could be improved by ozonation. The
combination of ozone reaction with biological aerated
filter (BAF) in treating cyanide containing electroplating
wastewater was employed (Cui et al. 2014). Although the
authors found the increase of m(O3):m(CN-) being benefi-
cial in reducing effluent CN- concentrations (nearly com-
plete removal), the non-catalytic ozonation remained
problematic due to high dosage of NaOH required for
achieving optimal pH condition. Aiming at this obstacle,
the ozone oxidation associated with BAF (BAF1-O3-BAF2)
was proposed for cyanide abatement, removing 99.7%
CN-, 81.7% chemical oxygen demand (COD) under rela-
tively moderate reaction conditions.
The applications of catalyst-assisted ozonation can
also be used to achieve desirable cyanide removal under
neutral conditions. The oxidation processes compris-
ing O3, O3/UV, and O3/UV/H2O2 were employed to treat
cyanide in jewellery manufacturing wastewater (Mert
etal. 2014). It was indicated that total cyanide removal
by O3 reached 86% at pH 12, while total cyanide removal
by O3/UV/H2O2 increased significantly to 99% even at pH
as lower as 10. Ozonation catalyzed by hydrogen perox-
ide was used to destruct cyanide containing automobile
industry wastewater and synthetic wastewater (Mudliar
et al. 2009). Optimized loading of 116.8×103 mg/l H2O2
and 5.0 mg/l O3 was developed, reducing CN- concen-
tration from 250 mg/l below 0.02 mg/l. The results also
indicated that CN- destruction was more efficient in
automobile industry wastewater than that in synthetic
wastewater, ascribed from abundant Fe2+ (212.84±13.8
mg/l) residual in industrial wastewater, which could
potentially accelerate ozone oxidation in homogenous
solution. The activated carbon synthesized by coconut
shell (CAC) was employed to enhance the removal of
cyanide by ozone (Sánchez-Castillo et al. 2015). Under
the optimum reaction conditions, 1200 mg/ml of cyanide
was totally eliminated after 3h by using 1 g of CAC and
2mg O3/min. Importantly, it was clearly found based on
the results that before the oxidation of CN- to CNO-, the
CN- concentration predominantly decreased to adsorp-
tion equilibrium condition on CAC surface. The reason
was that the basic typical sites on CAC surface, includ-
ing quinone and pyrone, could be oxidized by ozone,
generating positively charged sites. In this scenario, the
ion-exchange type of CN- adsorption may be extended by
electrostatic adsorption.
In summary, free cyanide ions can be effectively
abated upon ozone reaction, occurring primarily by oxy-
gen-transfer reactions. The decomplexed reaction driven
by ˙OH will be the prerequisite step for abatement of
complex state cyanide. The promoting effect of produced
metal ions on cyanide ozonation in strong alkaline solu-
tion, however, remains unclear.
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C. Wei etal.: Ozonation in water treatment23
4.3 Ozonation of coking wastewater
The coking wastewater has gained immense attention due
to its complex composition and highly toxic feature. After
biological treatment, there are still phenolic, benzene
series, polycyclic aromatic hydrocarbons (PAHs), and
nitrogen-containing heterocyclic compounds residual
in coking wastewater effluent. These non-biodegrada-
ble organic contaminants in coking wastewater can be
effectively eliminated by ozone-based technologies, as
shown in Table 7. It can be found that the most concerned
research interests lie in the reduction of potential risk of
coking wastewater effluent and the improvement of oxida-
tion capability of ozone by coupling with active catalysts.
The toxicological effect of hazardous compounds in
coking wastewater subjected to anaerobic/aerobic/ozona-
tion process on the development of maize embryos and
the activity of amylase and protease in maize seeds was
investigated (Wei etal. 2012). Obviously, after ozone oxi-
dation, the binding ability of poisonous organics in coking
wastewater with α-amylase was inhibited. This result may
be related to the degradation or structural change of toxic
compounds including quinolines, indoles, and PAHs.
Moreover, Chang etal. (2008) evaluated the contribution
of ozone pre-treatment of coking wastewater to the toxicity
reduction, demonstrating that toxic substances compris-
ing cyanide, thiocyanate, and phenol can be destroyed
by ozone and converted to less toxic or non-toxic sub-
stances. The toxicity inhibition of coking wastewater on
aerobic microorganisms after ozonation was lower than
20%. However, the authors (Chang etal. 2008) reported
that biodegradability of wastewater was not enhanced as
evidenced by biochemical oxygen demand (BOD)5/COD
ratio, decreasing from 0.52 to about 0.1, which is not con-
sistent with other studies (Bijan and Mohseni 2005, Cui
etal. 2014).
It is generally believed that the application of non-
catalytic ozonation for wastewater decontamination is
adversely affected by the insufficient oxidation ability of
ozone, making the association of ozone oxidation with
other technologies remain as fundamental interest. The
ozonation of biologically treated coking wastewater cata-
lyzed by ZnFe2O4 was conducted to elucidate the organic
compounds transformation (Zhang et al. 2015). On the
basis of ultraviolet absorption spectroscopy analysis, the
absorbance in the range of 250–300 nm was evidently
Table 7:Catalytic and non-catalytic ozonation of coking wastewater.
Wastewater Ozonation Ozone reactor Research remarks Reference
Coking wastewateraO/ZnFeOThree-necked flask The ZnFeO can accelerate the generation of
radicals upon ozone decomposition, and single
ring aromatic compounds as well as PAHs can be
effectively degraded
Zhang etal. 
Simulated coking
Rotating packed bed The BOD/COD value of simulated coking
wastewater treated by O/Fenton reached . and
was % higher than that obtained in O process
Wei etal. 
Coking wastewateraO/UV Pilot-scale FBR The ozonation of industrial wastewater highly
depended on the reductive activity sequence of
pollutants and mass transfer of ozone in FBR
Chen etal. 
Coking wastewateraO/UV Pilot-scale FBR FBR accelerated the mass transfer of ozone from
gaseous to aqueous phase, and therefore the
removal of PAHs was increased
Lin etal. a
Coking wastewaterbOCylindrical column Under the optimized conditions of ozone dosage
.mg O/mg COD, pH ., ozone combined with
BAF can effectively abate COD, NH+-N, UV
Zhang etal. b
Coking wastewatercOAirtight Pyrex columnOrganic matter and cyanide cannot be completely
degraded, due to the inadequate oxidation ability
of ozone to remove ozonated by products
Chang etal. 
Coking wastewaterdOCylindrical column Ozone can initially decompose refractory organics
such as quinolines, indoles, and PAHs, although
derivatives of pyridines and PAHs still remained.
Wei etal. 
aThe wastewater was obtained from anaerobic-aerobic-aerobic FBR, Song Shan Coking Plant, Guangdong, China.
bThe wastewater was obtained from anaerobic-anoxic-oxic process, Magang Steel Factory in Ma’anshan, China.
cThe wastewater effluent studied was obtained from a local coke plant.
dThe effluent wastewater sample was obtained from the biofilter underwent anaerobic/aerobic treatment.
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24C. Wei etal.: Ozonation in water treatment
lower for coking wastewater subjected to catalytic ozona-
tion than that treated by non-catalytic ozonation. It was
indicated that the degradation of aromatic organic com-
pounds should be more efficient by means of catalytic
ozonation than that by non-catalytic ozonation. Wei etal.
(2015) combined Fenton reagent with ozone in treating
simulated coking wastewater in a rotating packed bed,
with special emphasis on pollutants degradation and
corresponding reaction pathways. The results suggested
that removal efficiency of phenol, aniline, quinoline, and
NH4+-N by O3/Fenton reached 100%, 100%, 95.68%, and
100%, respectively, which were much superior than those
obtained by non-catalytic ozonation. The improved per-
formance of Fenton ozonation was due to the promoting
effect of Fe2+ and H2O2 on ˙OH generation on the course
of ozone decomposition. In addition, the BAF combined
with ozone oxidation in degrading bio-treated coking
wastewater was examined (Zhang etal. 2014b). The bio-
degradability was found to be substantially improved
from BOD5/COD ratio of 0.09 to 0.43 under optimized oper-
ation parameters. In comparison with BAF process, the
removal of COD and UV254 can be increased remarkably
by BAF/O3, while there appeared nearly no difference for
NH4+-N removal by these two methods. This indicated the
poor oxidizing potential of ozone with NH4+-N, which is
not structured with typical unsaturated functional groups
and conjugated structures.
To sum up, the highly toxic organic compounds in
biologically treated coking wastewater effluent, especially
the aromatic or the conjugated structural organics, will be
preferentially degraded by catalytic and non-catalytic ozo-
nation post-treatment. However, the bio-refractory com-
pounds residual in coking wastewater is normally present
in low dose, inhibiting their kinetic reactions with ozone.
On the other hand, when ozonation was employed as the
pre-treatment for biological process, the biodegradability
could be substantially increased as a result of selective
oxygen atom addition on aromatic structures, which are
more biodegradable than their precursors.
4.4 Ozonation of dyeing wastewater
The textile industry is a water-consuming and seriously
contaminated industry, in which dyeing wastewater
accounts for 80% of total pollution discharge (Lin and
Chen 1997). Dyestuffs and textile auxiliaries, being the
main components of dyeing wastewater, are potentially
toxic and less biodegradable, which result in a threat to
aquatic organisms (Sharma etal. 2007). The effluent COD
of dyeing wastewater after primary sedimentation and
biological treatment is composed of refractory compounds
and SMPs. The SMPs mainly consisting of proteins, poly-
saccharides, and organic colloids are derived from bio-
logical treatment and constituted the majority of soluble
effluent organic matters (EfOM) (Wu etal. 2016a). Table8
summarizes the latest investigations of dyeing wastewater
treatment by means of ozone oxidation.
4.4.1 Non-catalytic ozonation of dyeing wastewater
The color removal from textile dyeing wastewater efflu-
ent treated by ozonation and the evaluation of toxic inter-
mediate compounds were investigated (Mezzanotte et al.
2013). Results indicated that average color unit removal
is 0.0018±0.0004, 0.0016±0.0005, and 0.0006±0.0002 per
mg O3/l for 426, 558, and 660nm absorbance, respectively.
The aldehydes including glyoxal and methylglyoxal were
recognized as the main carbonylic by-products from ozone
reaction. Nevertheless, exposed at 60 mg O3/l, the alde-
hyde concentration ranged from 0.7 to 1.0 mg/l, complying
with the discharge standards. Interestingly, glyoxal and
methylglyoxal concentrations were directly related to color
removal, which can be developed as a simple indicator to
the toxic potential of ozone by-products in textile effluents.
The ozonation mechanisms of anionic sulphonated azo dye
were elucidated (Turhan and Ozturkcan 2013). The molecu-
lar structure of RO16 dye contains both electron-donating
groups (-OH, -NHR) and electron-withdrawing groups (-SO3
and -SO2R). In general, aromatic rings substituted with elec-
tron-donor groups are highly reactive with ozone at carbon
atoms in ortho- and para- positions due to high electron den-
sities. Thus ozone reactions may occur predominantly on
these electron-rich sites. It was found that although nearly
complete decolorization can be achieved within 8-min ozo-
nation by the breakdown of chromophore groups, there
was still 30% COD residual in wastewater. This was owed
to the presence of electron-deficient sites on RO16 molecu-
lar as well as ozonated by-products. Benincá etal. (2013)
reported an interesting phenomenon concerning the dif-
ference in color removal of azo dye solution (Ponceau 4R)
from synthetic water and industrial wastewater. The results
indicated that it took no more than 600s for complete color
removal in synthetic dye solution (pH=5.8), as ozone is able
to cleave the azo bond and electron-rich organic functional
groups in dye molecule. The color removal in real industrial
wastewater (pH=4.7), however, did not exceed 50% under
nearly identical reaction conditions. The reason for such
behavior was probably the formation of suspended solids
from matters initially dissolved, although not fully under-
stood (Lin and Lin 1993).
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C. Wei etal.: Ozonation in water treatment25
Table 8:The latest investigations of dyeing wastewater ozonation.
Ozone oxidation Wastewater quality Reaction conditions Main results Reference
Ozonation catalyzed
by sheet loaded with
Cu/Zn/Al/Zr catalyst
Concentration of basic yellow :
mg/l, pH: .
Ozone dose:  mg/l, gas
flow: . l/min, temperature:
The removal efficiency of BY  (%) and COD
(%) can be sustained with O/BY  mole
ratios of .–.
Zhu etal. 
Ozonation catalyzed
by iron shavings
COD: ± mg/l, Colour: ± times,
pH: .±., BOD: ±. mg/l
Ozone dose: . mg/l, gas
flow:  ml/min, catalyst
dose:  g/l
Catalytic ozonation effectively removed organic
pollutants with COD decreasing from  to 
mg/l, below discharge limit
Wu etal. a
Ozonation catalyzed
by copper oxide
Concentration of Reactive Black :
mg/l, pH: ., COD:  mg/l
Ozone dose: . mg/min,
catalyst dose:  g/l
COD removal reached % in catalytic ozonation
after -min reaction, while it was only % in
non-catalytic ozonation
Hu etal. 
Ozonation COD: ± mg/l, pH: .–. Ozone doses: – mg/l  mg/l color removal was complete and at
mg O/l, the final aldehyde concentration
ranged between . and . mg/l
Mezzanotte etal.
Ozonation, activated
carbon and BAF
COD: – mg/l, BOD: .–
. mg/l, color: – times
Ozone dose: . mg/mg
COD, apparent density of AC:
 kg/m
Outflows with COD of  mg/l, BOD of . mg/l,
and color of . times were obtained after the
combined treatment
Zou 
COD:  mg/l, NH+-N: . mg/l,
pH: 
Ozone doses: .–. g/l Anaerobic treatment removed COD and color.
Ozonation further reduced the organic content,
especially the aromatic compounds
Punzi etal. 
Ozonation Concentration of reactive orange
: mg/l, COD:  mg/l,
Ozone dose:  mg/l min COD of basic dyestuff wastewater was reduced
.% after ozone bubbling treatment
Turhan and
Ozturkcan 
Ozonation catalyzed
by Fe(II)
Concentration of Reactive Red :
.mmol/l, pH: 
Ozone doses: – mg/l,
flow rate: – ml/min,
Fe(II) doses: – mg/l
The maximum mineralization rate of RR dye was
about %, under optimized operation condition
of Fe(II)  mg/l, pH , ozone dose  mg/l, flow
rate  ml/min
Zhang etal. a
Ozonation COD: ± mg/l, apparent color:
mg PtCo/l, azo dye:  mg/l
Ozone flow rate: .mg/s,
pH(synthetic): .,
pH(wastewater): .–.
Azo dye aqueous solution was completely
degraded after only  s, and a substantial
reduction of TOC (around %) was noticed
Benincá etal. 
Ozonation with
COD:  mg/l, color: mg
PtCo/l, alkalinity:  mg/l, pH: .
Ozone dose: . g/h, current
density (EC): . mA/cm
% color removal, % turbidity removal, and
% COD reduction could be attained by using
the integrated process
etal. 
Ozonation catalyzed
by bio-composite
Indigo carmine concentration:
mg/l, pH: 
Ozone dose:  mg/l, catalyst
dose:  mg/l, water
bath: ±°C
Nearly complete removal of indigo carmine was
obtained after min by catalytic ozonation
Torres-Blancas etal.
Ozonation with BAF COD: – mg/l, UV: .–.,
Color: – times, BOD:
Ozone doses: – mg/l,
hydraulic retention time
(HRT) of BAF: . h
With the ozone dosage of  mg/l and HRT of
BAF of . h, COD of effluent from O-BAF process
was below  mg/l
Wu etal. b
Ozonation catalyzed
by interior micro
electrolysis (IM)
Reactive Black  solution: . g/l,
pH:., COD: . mg/l
Ozone dose:  mg/h,
 cm iron scraps filling
with  cm AC
Acute toxicity tests showed that pre-treatment
by IM generated effluents that were more toxic
compared with initial wastewater.
Guo etal. 
- 10.1515/revce-2016-0008
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26C. Wei etal.: Ozonation in water treatment
4.4.2 Catalytic ozonation of dyeing wastewater
The employment of catalytic ozonation could improve the
mineralization of simulated dyeing wastewater as a result
of radical-mediated reactions, besides nearly complete
color removal. The ozonation of Basic Yellow 87 (BY 87)
in synthetic wastewater catalyzed by Cu/Zn/Al/Zr cata-
lyst (i.e. CuO, ZnO, Al2O3, and ZrO2), which was loaded on
porous copper fiber sintered sheet, was evaluated (Zhu
etal. 2014). The removal efficiency of COD and TOC was
improved by two times and five times, respectively, by the
assistance of catalyst chip as compared to non-catalytic
ozonation. For a typical heterogeneous catalytic process,
physical adsorption for gaseous ozone took place due
to the presence of hydrophobicity surface and relatively
small pore size on catalyst chip surface. The BY 87 and its
by-products can be subsequently degraded by oxidizing