1-s2.0-S0048969719350156-main

Review
Degradation of antibiotics by advanced oxidation processes: An
overview
Jianlong Wang
a,b,
⇑
, Run Zhuan
a
a
Laboratory of Environmental Technology, INET, Tsinghua University, Beijing 100084, PR China
b
Beijing Key Laboratory of Radioactive Waste Treatment, Tsinghua University, Beijing 100084, PR China
highlights
Antibiotics are ubiquitous due to their
extensive production and
consumption.
AOPs are effective to degrade
antibiotics in water and wastewater.
The recent advance in antibiotics
degradation by AOPs was analyzed
and summarized.
Fenton, ozonation, photocatalytic,
electrochemical and ionizing
radiation were introduced.
Concluding remarks were given and
their future perspectives and
challenges were discussed.
graphical abstract
article info
Article history:
Received 6 September 2019
Received in revised form 15 October 2019
Accepted 15 October 2019
Available online 3 November 2019
Keywords:
Advanced oxidation processes
Antibiotics
Ionizing radiation
Ozonation
Photocatalytic oxidation
Fenton-like oxidation
abstract
Antibiotics are becoming emerging contaminants due to their extensive production and consumption,
which have caused hazards to the ecological environment and human health. Various techniques have
been studied to remove antibiotics from water and wastewater, including biological, physical and chem-
ical methods. Among them, advanced oxidation processes (AOPs) have received increasing attention due
to their fast reaction rate and strong oxidation capability, which are effective for the degradation of
antibiotics in aquatic environments. In this review paper, a variety of AOPs, such as Fenton or Fenton-
like reaction, ozonation or catalytic ozonation, photocatalytic oxidation, electrochemical oxidation, and
ionizing radiation were briefly introduced, including their principles, characteristics, main influencing
factors and applications. The current applications of AOPs for the degradation of antibiotics in water
and wastewater were analyzed and summarized, the concluding remarks were given and their future
perspectives and challenges were discussed.
Ó2019 Elsevier B.V. All rights reserved.
https://doi.org/10.1016/j.scitotenv.2019.135023
0048-9697/Ó2019 Elsevier B.V. All rights reserved.
⇑
Corresponding author at: Energy Science Building, INET, Tsinghua University, Beijing 100084, PR China.
E-mail address: [email protected] (J. Wang).
Science of the Total Environment 701 (2020) 135023
Contents lists available at ScienceDirect
Science of the Total Environment
journal homepage: www.elsevier.com/locate/scitotenv

Contents
1. Introduction ........................................................................................................... 2
2. Fenton and Fenton-like process. ........................................................................................... 2
2.1. Fenton-like catalysts . . . . . . . . . . . . . . . . . .............................................................................. 3
2.2. Catalyst dosage . . ................................................................................................. 3
2.3. H
2
O
2
concentration . . . . . . . . . . . . . . . . . . .............................................................................. 3
2.4. pH value. . . . . . . . ................................................................................................. 4
2.5. Antibiotics removal by Fenton and Fenton-like oxidation . . . .............................................................. 4
3. Ozonation or catalytic ozonation . . . . . . . . . . . . . . . . . . ........................................................................ 4
3.1. Ozone concentration . . . . . . . . . . . . . . . . . .............................................................................. 5
3.2. pH value. . . . . . . . ................................................................................................. 5
3.3. Mineralization of pollutants . . . . . . . . . . . .............................................................................. 5
3.4. Biodegradability improvement of pollutants . . . . . . . . . . . . . . .............................................................. 5
3.5. Antibiotics removal by ozone oxidation . .............................................................................. 5
4. Photocatalytic oxidation . . . . . . ........................................................................................... 5
4.1. Photocatalytic materials . . . . . . . . . . . . . . .............................................................................. 6
4.2. pH value. . . . . . . . ................................................................................................. 7
4.3. Catalysts dosage . ................................................................................................. 7
4.4. Catalysts stability ................................................................................................. 7
4.5. Mineralization of antibiotics. . . . . . . . . . . .............................................................................. 7
4.6. Antibiotics removal by photocatalytic oxidation. . . . . . . . . . . .............................................................. 7
5. Electrochemical oxidation . . . . . ........................................................................................... 7
5.1. Electrode materials . . . . . . . . . . . . . . . . . . .............................................................................. 8
5.2. Current density . . ................................................................................................. 8
5.3. pH value. . . . . . . . ................................................................................................. 8
5.4. Antibiotics removal by electrochemical oxidation . . . . . . . . . .............................................................. 9
6. Ionizing radiation . . . . . . . . . . . . ........................................................................................... 9
6.1. Absorbed dose . . . ................................................................................................. 9
6.2. pH value. . . . . . . . ................................................................................................ 10
6.3. Inorganic anions, organic matters and matrix. . . . . . . . . . . . . ............................................................. 10
7. Concluding remarks and perspectives. . . . . . . . . . . . . . . ....................................................................... 10
Declaration of Competing Interest . . . . . . . . . . . . . . . . . ....................................................................... 11
Acknowledgements . . . . . . . . . . .......................................................................................... 11
Appendix A. Supplementary material . . . . . . . . . . . . . . . . . ....................................................................... 11
References . .......................................................................................................... 11
1. Introduction
Antibiotics are chemical compounds which are applied to treat
microbial infectious diseases, they have widely applied for the
treatment of human and animal diseases as well as in aquaculture
and livestock feeding (Manzetti and Ghisi, 2014). The extensive use
of antibiotics, especially the overuse or abuse of antibiotics has
attracted public concern. During the production and application
of antibiotics, a large amount of antibiotics-containing wastewater
are generated and discharged into the environment, causing seri-
ous pollution (Focazio et al., 2008). The residual antibiotics are per-
sistent and difficult to degrade by conventional biological
treatment methods (Wang and Wang, 2016; Prado et al., 2009;
Kummerer et al., 2000). Therefore, antibiotic were frequently
detected in various natural environments (Wang et al., 2019c),
including river water (Huang et al., 2019), groundwater (Szekeres
et al., 2018), surface water (Danner et al., 2019), soil (Cerqueira
et al., 2019), sediment (Chen and Zhou, 2014) and drinking water
(Sanganyado and Gwenzi, 2019). The long-term occurrence of
antibiotics in the natural environments may lead to the generation
of antibiotic resistant genes (ARGs) and antibiotic resistant bacte-
ria (ARBs), accelerating the spread of antibiotic resistance, causing
threat to human health and ecological systems (Kummerer, 2009).
Various techniques have been studied for the removal of antibi-
otics from water and wastewater, including coagulation, mem-
brane separation, adsorption and biodegradation (Wang and
Wang, 2019a; 2018b; Zhuang et al., 2020, 2019a, 2019b; Wang
and Zhuang, 2019, 2017). However, they have not been widely
applied due to their low removal efficiency and high operational
cost. By contrast, advance oxidation processes (AOPs) can degrade
antibiotics or convert them to small molecule substances, which
could alleviate the inhibitive effect of antibiotics on microorgan-
isms, and enhance their biodegradability and the removal rate
(Wang and Wang, 2019b; Hernandez et al., 2002).
Advanced oxidation processes use strong oxidation agents, such
as hydroxyl radical (
OH), ozone (O
3
), superoxide radical (O
2
)to
degrade organic pollutants (Wang and Wang, 2018a; Wang and
Bai, 2017; Wang and Xu, 2012; Buxton et al., 1988). According to
the different ways used to produce oxidation agents, AOPs can be
classified into Fenton oxidation, photocatalytic oxidation, electro-
chemical oxidation and so on (Fig. 1).
In this review, the degradation of antibiotics by various
advanced oxidation processes (AOPs), including Fenton or
Fenton-like reaction, ozonation or catalytic ozonation, photocat-
alytic oxidation, electrochemical oxidation, and ionizing radiation
were briefly introduced, their principles, characteristics, main
influencing factors and applications for the degradation of antibi-
otics in water and wastewater were analyzed and summarized,
the concluding remarks and future challenges were discussed.
2. Fenton and Fenton-like process
The combination of ferrous salt and hydrogen peroxide is called
Fenton reagent (Fenton, 1894). Fenton oxidation methods are
widely used in wastewater treatment. As for Fenton oxidation
method, Fenton reagent (Fe
2+
and H
2
O
2
) are added into wastewa-
ter, which can react to form hydroxyl radicals (
OH), as Eqs. (1)–(3).
2J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023

Fe
2þ
+H
2
O
2
!Fe
3þ
+ OH + OH
—
ð1Þ
OH + H
2
O
2
! HO
2
+H
2
Oð2Þ
2
OH !H
2
O
2
ð3Þ
These radicals could oxidize or degrade antibiotics. Fenton oxi-
dation method has advantages, such as higher degradation effi-
ciency and easy operation. Various operating parameters,
including pH value, temperature, H
2
O
2
concentration and Fe
2+
con-
centration, all have influence on the treatment efficiency. However,
Fenton oxidation has several disadvantages, which is limited to the
acidic condition, and large amount of iron-containing sludge will
yield which is difficult to dispose. In order to overcome these dis-
advantages, other catalysts are used to replace Fe
2+
, which called
Fenton-like oxidation process (Wang and Wang, 2018e; Wang
et al., 2016b).
2.1. Fenton-like catalysts
Although homogeneous Fenton oxidation can effectively
degrade organic pollutants, there are some problems in practical
application. Firstly, the utilization rate of H
2
O
2
is low, causing
low decomposition rate of pollutants. Secondly, homogeneous Fen-
ton require pH at around 3, which is lower than pH of practical
wastewater. Adjusting pH value will increase the operational cost.
Finally, adding ferrous salt will cause the production of iron-
containing sludge, resulting in secondary pollution.
Heterogeneous Fenton or Fenton-like process can be performed
at a wide range of pH, the catalyst can be utilized circularly, which
can avoid the production of iron sludge (He et al., 2016; Nidheesh,
2015; Soon and Hameed, 2011). Heterogeneous Fenton catalysts
mainly include:
(1) iron minerals, such as magnetite (Xu and Wang, 2012),
goethite (Wang et al., 2015), ferrite (Liu et al., 2012), ferrihy-
drite (Barreiro et al., 2007), schorl (Xu et al., 2013);
(2) zero-valent iron (ZVI) (Xu and Wang, 2011; Zhou et al.,
2008);
(3) other single metal and metallic oxide, such as MnO
2
(Saputra
et al., 2013), TiO
2
(Zhang et al., 2016a), Pd (Yuan et al., 2011);
(4) iron- and iron oxide-loaded materials, commonly used sup-
porters include activated carbon (Sekaran et al., 2011), alu-
mina (Ghosh et al., 2012), clay (Djeffal et al., 2014), silica
(Martinez et al., 2007), zeolite (Fukuchi et al., 2014), biosor-
bents (Daud and Hameed, 2010);
(5) metal-organic frameworks (MOFs), which are crystalline
functional material composed of transition metal ions and
organic ligands (Tang and Wang, 2018b; Etaiw and El-
bendary, 2012; Lee et al., 2009).
These heterogeneous catalysts have been reported for the
degradation of antibiotics, for instance, metacycline (Qi et al.,
2019), lincomycin (Ouyang et al., 2019), enrofloxacin (Hou et al.,
2019), tetracycline (Zhang et al., 2019), oxytetracycline (Pan
et al., 2019), sulfamethazine (Tang and Wang, 2018a).
2.2. Catalyst dosage
Catalyst dosage is important in Fenton and Fenton-like oxida-
tion process, which has crucial influence on the degradation of
organic pollutants. The overdose of catalyst may scavenge hydro-
xyl radicals (
OH) and inhibit the degradation of pollutants. More-
over, excessive catalyst dosage will raise costs and limit practical
application (Wang et al., 2016b). Qi et al. (2019) observed the influ-
ence of catalyst dosage on metacycline degradation using CuCo
2
O
4
nano-catalyst for Fenton reaction. The removal efficiency increased
from 38.4% to 89.1% when the dosage increased from 5.0 to
12.0 mg. When further increasing to 15.0 mg, the removal rate only
slightly increased to 92.5%. Ouyang et al. (2019) applied iron-based
catalysts (GFe0.5) for Fenton-like oxidation of lincomycin. When
adding 0.01 g/L GFe0.5, the removal of lincomycin reached
93.85% after 90 min. When the dosages of catalyst was increased
to 0.05 and 0.1 g/L, lincomycin was completely removed within
10 min. Nasseh et al. (2019) synthesized FeNi
3
/SiO
2
magnetic
nano-catalyst and applied to degrade metronidazole by heteroge-
neous Fenton-like process. The degradation efficiency increased
from 40.96% to 84.29% when catalyst dosage increased from
0.005 to 0.1 g/L, because large amount of active sites were pro-
vided, which caused the elevation of hydroxyl radical through
the decomposition of hydrogen peroxide, and promoted the degra-
dation of organic pollutants.
2.3. H
2
O
2
concentration
H
2
O
2
plays an important role in Fenton oxidation, as the domi-
nant source of hydroxyl radicals (
OH). Insufficient H
2
O
2
dosage
will cause the lack of hydroxyl radicals (
OH) and reduce degrada-
tion efficiency. By contrast, excessive H
2
O
2
dosage is not suitable
for the degradation of pollutants (Wang et al., 2016b). The required
theoretical H
2
O
2
dosage could be calculated according to the fol-
lowing Eq. (4):
C
a
H
b
N
c
O
d
þ2aþ1
2bþ5 2cd
H
2
O
2
!aCO
2
þ2aþbþ2cdðÞH
2
OþcHNO
2
ð4Þ
Theoretically, one mole of C
a
H
b
N
c
O
d
requires
2aþ
1
2
bþ
5
2
cd moles H
2
O
2
. Usually, the actual added H
2
O
2
con-
centration should be higher than the calculated value according to
the chemical equation, which can be examined through the prelim-
inary experiments. Nasseh et al. (2019) synthesized magnetic nano
FeNi
3
/SiO
2
composite and used it as heterogeneous Fenton-like
catalyst for the oxidation of metronidazole. They found that the
degradation efficiency of metronidazole firstly increased with
increase of H
2
O
2
dosage from 50 mg/L to 150 mg/L, then it
decreased when H
2
O
2
dosage reached 200 mg/L. Qi et al. (2019)
Fig. 1. Hydroxyl radicals ( OH)-based advanced oxidation processes.
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023 3

evaluated the degradation of metacycline using CuCo
2
O
4
as cata-
lyst. The results showed that the removal of metacycline was
43.6%, 54.3% and 95.1%, respectively when H
2
O
2
dosage was 100,
300 and 500
l L.
2.4. pH value
In Fenton and Fenton-like processes, pH value is an important
parameter for effective treatment (Wang et al., 2016b). In the tra-
ditional homogeneous Fenton processes, the suitable pH value is
about 3.0, while in the Fenton-like processes, the optimal pH
depends on the reaction system, especially the reaction mecha-
nisms which rely on the catalyst performance (Wang et al., 2016b).
Elmolla and Chaudhuri (2009) evaluated the Fenton oxidation
of the antibiotics, including amoxicillin, ampicillin and cloxacillin.
After 60 min reaction time, COD of antibiotics wastewater
degraded 49.0%, 57.7%, 81.5%, 76.9% and 75.6% at pH 2.0, 2.5, 3.0,
3.5 and 4.0, respectively. While DOC degradation percent was
33.9, 43.5, 54.3, 50 and 48.4 at pH 2.0, 2.5, 3.0, 3.5 and 4.0, respec-
tively. The best decomposition of amoxicillin, ampicillin and clox-
acillin wastewater achieved at pH 3.0. The decrease of degradation
rate at pH over 3.0 may be due to the decrease in dissolved iron.
Wan and Wang (2016a, 2016b, 2016c) studied the influence of
pH on the degradation of sulfamethazine using Ce
0
/Fe
0
-RGO com-
posites as Fenton-like catalyst. The removal efficiency of sulfamet-
hazine decreased as pH increased from 6.0 to 8.3. The change of pH
value had influence on the adsorption of sulfamethazine on the
catalyst surface. When pH was over 7.42, which is the pKa
2
of sul-
famethazine, negative charged catalyst would repel anionic form
sulfamethazine, decreasing the adsorption and inhibiting the oxi-
dation reaction.
Zhang et al. (2019) investigated the degradation of tetracycline
using zero-valent iron and Fe
0
/CeO
2
as Fenton oxidation catalyst.
The results showed that the degradation efficiency of tetracycline
decreased from 93% to about 50% when pH increased from 3.0 to
5.8 and nZVI was used as catalyst. The removal efficiency of tetra-
cycline was over 93% when pH ranged from 3.0 to 5.8 and Fe
0
/CeO
2
was used, exhibiting high reactivity at a wide range of pH values.
2.5. Antibiotics removal by Fenton and Fenton-like oxidation
The degradation of antibiotics by Fenton and Fenton-like oxida-
tion were summarized in Table 1.
3. Ozonation or catalytic ozonation
Ozonation or catalytic ozonation is an environmentally-friendly
technology for wastewater treatment (Wang and Bai, 2017). Ozone
with 2.07 V oxidation potential can oxidize a variety of refractory
organic pollutants. Ozone molecule can degrade organic pollutants
directly. Moreover, ozone can react with water with the help of
catalyst to form hydroxyl radicals (
OH), which has stronger oxida-
tion capability, according to Eqs. (5)–(9) (Yargeau and Leclair,
2008).
O
3
+H
2
O!2 OH + O
2
k = 1.1 10
4
L/(mols) ð5Þ
O
3
+OH
—
!O
2—
+HO
2
k = 70 L/(mols) ð6Þ
O
3
+HO
2
!2O
2
+ OH k = 1.610
9
L/(mols) ð7Þ
O
3
+ OH !O
2
+HO
2
ð8Þ
Table 1
Antibiotics removal by Fenton and Fenton-like oxidation.
Antibiotics Catalyst (dosage); pH range Removal efficiency (%) References
Amoxicillin zero-valent iron (nZVI) (0.2–2 g/L); pH = 2–5 86.5 (Zha et al., 2014)
Fe(II) (0.32–24.3 mM); pH = 2–4 80 (Elmolla et al., 2010)
H
2
O
2
/Fe
2+
= 2.0–150; pH = 2.0–4.0 100 Elmolla and Chaudhuri (2009)
H
2
O
2
/Fe
2+
= 1–50; pH = 1–9 80.92 (Guo et al., 2015a)
Ampicillin Fe(II) (53–87
l M); pH = 2.3–5.7 90.2 (Rozas et al., 2010)
Fe(II) (0.32–24.3 mM); pH = 2–4 80 (Elmolla et al., 2010)
H
2
O
2
/Fe
+
= 2.0–150; pH = 2.0–4.0 100 Elmolla and Chaudhuri (2009)
Azithromycin H
2
O
2
/Fe
2+
= 1.75 mM; pH = 3 95 (Mackul’ak et al., 2015)
Cloxacillin Fe(II) (0.32–24.3 mM); pH = 2–4 80 (Elmolla et al., 2010)
H
2
O
2
/Fe
2+
= 2.0–150; pH = 2.0–4.0 100 Elmolla and Chaudhuri (2009)
Ciprofloxacin Fe
3
O
4
(1.0–2.5 g/L); pH = 3–11 89 (Hassani et al., 2018)
CNTs/FeS (5–35 mg); pH = 1–12 91.03 (Ma et al., 2015)
H
2
O
2
/Fe
2+
= 1.75 mM; pH = 3 95 (Mackul’ak et al., 2015)
Clarithromycin H
2
O
2
/Fe
2+
= 1.75 mM; pH = 3 95 (Mackul’ak et al., 2015)
Chlorpheniramine nZVI (11.2–28 g/L); pH = 2–5 100 (Wang et al., 2016a)
Doxycycline SBC@b-FeOOH (Zhang et al., 2016b)
Lincomycin GFe0.5 (0.01 g/L) 100 (Ouyang et al., 2019)
Metacycline CuCo
2
O
4
(0.1–0.3 g/L) 95.1 (Qi et al., 2019)
Metronidazole FeNi
3
/SiO
2
(0.005–0.1 g/L); pH = 3–11 95.32 Nasseh et al. (2019)
Nofloxacin [Fe(II)] (0.8–3 mM) (Santos et al., 2015)
Ofloxacin Alg/Fe (0.2–1.4 g/L); pH = 3 100 (Titouhi and Belgaied, 2016a)
Alg/CDTA/Fe (0.01–0.09 g); pH = 3 100 (Titouhi and Belgaied, 2016b)
CQDs/Cu-MIO (0.1–0.25 g/L); pH = 3.6–10 100 (Tian et al., 2017)
Fe-Cu@MPSi (0.5–1 g/L); pH = 3–9 100 (Zheng et al., 2017)
Oxytetracycline Cu@Fe
3
O
4
(0.1–1 g/L); pH = 3.10–9.05 100 (Pham et al., 2018)
Fe
0
(0.3 mM) 100 (Pan et al., 2019)
Sulfamethazine Fe
0
(0.3 mM) 100 (Pan et al., 2019)
CUS-MIL-100(Fe) (0.2–1.5 g/L); pH = 3–6 100 (Tang and Wang, 2018b, 2018a)
Ce
0
/Fe
0
-RGO (0.1–1 g/L); pH = 6–8 99 Wan and Wang (2016a, 2016b, 2016c)
Sulfamethoxazole Zn-Fe-CNTs (0.2–1 g/L); pH = 1.0–3.0 100 (Liu et al., 2018b)
Sulfadiazine Fe
0
(0.3 mM) 100 (Pan et al., 2019)
Sulfathiazole Fe
3
O
4
/Humic acid (0–5 g/L); pH = 3.5–9 100 (Niu et al., 2011)
Tetracycline Fe@Bacillus subtilis (0.5 g/L); pH = 4.0–6.0 100 (Zheng et al., 2016)
Fe
0
(0.3 mM) 100 (Pan et al., 2019)
CFO (0.05–0.2 g/L) 84 (Parmar et al., 2017)
Fe
0
/CeO
2
(0.01–0.2 g/L); pH = 3–7 93 (Zhang et al., 2019)
4J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023

2HO
2
!O
2
+H
2
O
2
ð9Þ
Therefore catalytic ozonation process can be used to enhance
the degradation efficiency of organic pollutants, including homo-
geneous and heterogeneous catalytic ozonation.
In homogeneous catalytic ozonation process, liquid catalysts,
especially transition metal ions are used, such as Fe
2+
,Mn
2+
,Ni
2+
,
Co
2+
,Cd
2+
,Cu
2+
,Ag
+
,Cr
3+
,Zn
2+
in reaction solution. These catalysts
can excite ozone to generate hydroxyl radicals (
OH) and improve
degradation efficiency.
In heterogeneous catalytic ozonation process, solid catalysts
such as metal oxide, activated carbon, porous materials and their
composite materials are added into reaction solution (Kasprzyk-
Hordern et al., 2003).
3.1. Ozone concentration
The ozone concentration has important influence on the degra-
dation of antibiotics. The mass transfer rate and the volumetric
mass transfer coefficient of ozone increases with increase of ozone
concentration. More ozone can be absorbed and react with antibi-
otic molecules, finally improving the decomposition of antibiotics
(Zhao et al., 2006; Kornmuller and Wiesmann, 2003).
Oh et al. (2016) studied the influence of ozone dosage on degra-
dation of antibitics, they found that tetracycline was degraded
more quickly at 7 ppm ozone exposure than at 3 ppm. Iakovides
et al. (2019) found that the elimination of antibiotics increased
when the ozone dosage increased, including ampicillin, azithromy-
cin, clarithromycin, erythromycin, ofloxacin, sulfamethoxazole,
tetracycline, trimethoprim. Paucar et al. (2019) studied the degra-
dation of ciprofloxacin, levofloxacin, clarithromycin and nalidixic
acid, they found that antibiotics degradation enhanced when the
initial ozone concentration increased. Hollender et al. (2009)
explored the effect of ozone dosage on the elimination of various
micro-pollutants. Overall, the removal efficiency of selected
micro-pollutants increased with increase of ozone dosage. De
Witte et al. (2009) investigated the ozonation of ciprofloxacin,
and found that the pseudo first-order constants increased with
the increase of ozone inlet concentration.
3.2. pH value
Generally, ozone can degrade organic pollutants through direct
oxidation by ozone molecule in acidic condition. In alkaline condi-
tion, organic pollutants are oxidized by both ozone molecule and
hydroxyl radicals (
OH) (Ikehata et al., 2006; Yargeau and Leclair,
2008). Thus, the degradation of antibiotics by ozonation depends
on the solution pH values.
Feng et al. (2016a) found that the degradation of flumequine
was faster at higher pH values. The reaction rate constant
increased from 0.3772 min
1
to 2.5219 min
1
when pH values
increased from 3.0 to 11.0. In alkaline condition, more O
3
was
transformed to
OH, and indirect oxidation of OH could be more
beneficial in decomposition of flumequine than direct oxidation
of O
3
. Moreover, the species of flumequine under different pH also
influenced the result.
Wang et al. (2012) explored the chloramphenicol (CAP) degra-
dation by ozone in aqueous solution at various initial pH values.
The removal efficiency of chloramphenicol (CAP) was 41.4 ± 1.0%
and 65.3 ± 3.0%, respectively at initial pH of 2.0 and 8.0. This result
may be attributed to more free radicals generated in alkaline con-
ditions. However, the removal rate decreased at initial pH of 10.0.
Oncu and Balcioglu (2013) investigated the influence of pH on
the ozonation of ciprofloxacin (CIP) and oxytetracycline (OTC).
Higher degradation of ciprofloxacin (CIP) and oxytetracycline
(OTC) was achieved at higher pH.
Jung et al. (2012) examined the effect of pH on the degradation
efficiency of ampicillin, the biodegradability and toxicity after
ozonation. The second-order rate constant and COD removal rate
increased with increase of pH when pH was in the range of 5–9.
A higher biodegradability and acute toxicity was observed at the
highest pH (pH 9).
On the one hand, an increase of ozone decomposition to gener-
ate
OH occurred in alkaline condition. On the other hand, non-
protonated organic amine species (–NH
2
) was more reactive
toward ozone molecules than the mono-protonated form (–NH
3
)
(Hoigne & Bader, 1983). At pH 9, non-protonated amine (–NH
2
)
was the dominant group of ampicillin, which could be attacked
by ozone more easily.
3.3. Mineralization of pollutants
Usually, the ozonation process could not totally mineralize the
antibiotics. On the one hand, carbonate (CO
3
2–
) and bicarbonate
(HCO
3
–
) formed during antibiotic decomposition process are hydro-
xyl radical scavengers, which can inhibit the removal of antibiotics.
On the other hand, solution pH decreased with the ozonation reac-
tion proceeding, which is adverse for the generation of hydroxyl
radicals (
OH). Uslu and Balcioglu (2008) observed that mineraliza-
tion rate of oxytetracycline reached 20% after 30 min ozonation at
pH 8.5. Feng et al. (2016a) found that 39.45% of TOC was removed
after the ozonation of flumequine aqueous solution. Kuang et al.
(2013) observed complete trimethoprim degradation after ozona-
tion, while no mineralization was determined. Goncalves et al.
(2012) found that TOC removal efficiency was 33.5% after
180 min ozonation of sulfamethoxazole solution.
3.4. Biodegradability improvement of pollutants
The BOD
5
/COD ratio is usually used for characterizing the
biodegradability of a pollutant or wastewater. The biodegradability
of antibiotics wastewater can be improved by ozonation due to the
generation of low molecule weight and biodegradable intermedi-
ate products.
Balcioglu and Otker (2003) observed that biodegradability of
wastewater after ozonation increased. More low-molecular weight
intermediate products that are more amenable to biodegradation
generated after ozonation (Stockinger et al., 1995).
Jung et al. (2012) found that the BOD
5
/COD ratio at 9 increased
constantly from 0 to 0.41 after 120 min of ozonation, enhancing
the biodegradability and biological treatability of ampicillin-
containing wastewater.
Dantas et al. (2008) reported that the biodegradability
increased from 0 to 0.3 during sulfamethoxazole ozonation, indi-
cating that the antibiotic was conversed to biodegradable interme-
diate product.
Uslu and Balcioglu (2008) observed that the BOD
5
/COD ratio of
synthetic oxytetracycline wastewater increased from 0.05 to 0.3
due to the formation of biodegradable intermediate products dur-
ing ozonation process.
3.5. Antibiotics removal by ozone oxidation
The degradation of various antibiotics by ozonation was sum-
marized in Table 2.
4. Photocatalytic oxidation
Photocatalytic oxidation has been extensively studied for the
degradation of organic pollutants. Semi-conductor materials, such
as TiO
2
, ZnS, WO
3
and SnO
2
are used as photo-catalyst. When
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023 5