1-s2.0-S0048969719350156-main

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
Review
Degradation of antibiotics by advanced oxidation processes: An
overview
Jianlong Wang a,b,⇑, Run Zhuan a
a
b
Laboratory of Environmental Technology, INET, Tsinghua University, Beijing 100084, PR China
Beijing Key Laboratory of Radioactive Waste Treatment, Tsinghua University, Beijing 100084, PR China
h i g h l i g h t s
g r a p h i c a l a b s t r a c t
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.
a r t i c l e
i n f o
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
a b s t r a c t
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 chemical 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 Fentonlike 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.
Ó 2019 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).
https://doi.org/10.1016/j.scitotenv.2019.135023
0048-9697/Ó 2019 Elsevier B.V. All rights reserved.
2
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
Contents
1.
2.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2
Fenton and Fenton-like process. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2
2.1.
Fenton-like catalysts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3
2.2.
Catalyst dosage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3
2.3.
H2O2 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 serious pollution (Focazio et al., 2008). The residual antibiotics are persistent 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 bacteria (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 antibiotics from water and wastewater, including coagulation, membrane 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 microorganisms, 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 (O3), 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, electrochemical 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, photocatalytic oxidation, electrochemical oxidation, and ionizing radiation
were briefly introduced, their principles, characteristics, main
influencing factors and applications for the degradation of antibiotics 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 (Fe2+ and H2O2) are added into wastewater, which can react to form hydroxyl radicals (OH), as Eqs. (1)–(3).
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
3
(4) iron- and iron oxide-loaded materials, commonly used supporters include activated carbon (Sekaran et al., 2011), alumina (Ghosh et al., 2012), clay (Djeffal et al., 2014), silica
(Martinez et al., 2007), zeolite (Fukuchi et al., 2014), biosorbents (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 Elbendary, 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
Fig. 1. Hydroxyl radicals (OH)-based advanced oxidation processes.
Fe2þ + H2 O2 ! Fe3þ + OH + OH—
ð1Þ
OH + H2 O2 ! HO2 + H2 O
ð2Þ
2 OH ! H2 O2
ð3Þ
These radicals could oxidize or degrade antibiotics. Fenton oxidation method has advantages, such as higher degradation efficiency and easy operation. Various operating parameters,
including pH value, temperature, H2O2 concentration and Fe2+ concentration, 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 disadvantages, other catalysts are used to replace Fe2+, 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 H2O2 is low, causing
low decomposition rate of pollutants. Secondly, homogeneous Fenton 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 ironcontaining 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), ferrihydrite (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 MnO2 (Saputra
et al., 2013), TiO2 (Zhang et al., 2016a), Pd (Yuan et al., 2011);
Catalyst dosage is important in Fenton and Fenton-like oxidation process, which has crucial influence on the degradation of
organic pollutants. The overdose of catalyst may scavenge hydroxyl radicals (OH) and inhibit the degradation of pollutants. Moreover, excessive catalyst dosage will raise costs and limit practical
application (Wang et al., 2016b). Qi et al. (2019) observed the influence of catalyst dosage on metacycline degradation using CuCo2O4
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 FeNi3/SiO2 magnetic
nano-catalyst and applied to degrade metronidazole by heterogeneous 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 provided, which caused the elevation of hydroxyl radical through
the decomposition of hydrogen peroxide, and promoted the degradation of organic pollutants.
2.3. H2O2 concentration
H2O2 plays an important role in Fenton oxidation, as the dominant source of hydroxyl radicals (OH). Insufficient H2O2 dosage
will cause the lack of hydroxyl radicals (OH) and reduce degradation efficiency. By contrast, excessive H2O2 dosage is not suitable
for the degradation of pollutants (Wang et al., 2016b). The required
theoretical H2O2 dosage could be calculated according to the following Eq. (4):
1
5
C a Hb Nc Od þ 2a þ b þ c d H2 O2
2
2
! aCO2 þ ð2a þ b þ 2c dÞH2 O þ cHNO2
ð4Þ
Theoretically,
one
mole
of
C a Hb N c Od
requires
2a þ 12 b þ 52 c d moles H2O2. Usually, the actual added H2O2 concentration should be higher than the calculated value according to
the chemical equation, which can be examined through the preliminary experiments. Nasseh et al. (2019) synthesized magnetic nano
FeNi3/SiO2 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 H2O2 dosage from 50 mg/L to 150 mg/L, then it
decreased when H2O2 dosage reached 200 mg/L. Qi et al. (2019)
4
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
evaluated the degradation of metacycline using CuCo2O4 as catalyst. The results showed that the removal of metacycline was
43.6%, 54.3% and 95.1%, respectively when H2O2 dosage was 100,
300 and 500 lL.
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 traditional 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 mechanisms 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, respectively. The best decomposition of amoxicillin, ampicillin and cloxacillin 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 Ce0/Fe0-RGO composites as Fenton-like catalyst. The removal efficiency of sulfamethazine 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 pKa2 of sulfamethazine, negative charged catalyst would repel anionic form
sulfamethazine, decreasing the adsorption and inhibiting the oxidation reaction.
Zhang et al. (2019) investigated the degradation of tetracycline
using zero-valent iron and Fe0/CeO2 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 tetracycline was over 93% when pH ranged from 3.0 to 5.8 and Fe0/CeO2
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 oxidation 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 oxidation capability, according to Eqs. (5)–(9) (Yargeau and Leclair,
2008).
O3 + H2 O ! 2 OH + O2 k = 1.1 104 L/(mols)
ð5Þ
O3 + OH— !O2 — + HO2 k = 70 L/(mols)
ð6Þ
O3 + HO2 ! 2O2 + OH k = 1.6109 L/(mols)
ð7Þ
O3 + OH ! O2 + HO2 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
Fe(II) (0.32–24.3 mM); pH = 2–4
H2O2/Fe2+= 2.0–150; pH = 2.0–4.0
H2O2/Fe2+ = 1–50; pH = 1–9
Fe(II) (53–87 lM); pH = 2.3–5.7
Fe(II) (0.32–24.3 mM); pH = 2–4
H2O2/Fe+ = 2.0–150; pH = 2.0–4.0
H2O2/Fe2+ = 1.75 mM; pH = 3
Fe(II) (0.32–24.3 mM); pH = 2–4
H2O2/Fe2+ = 2.0–150; pH = 2.0–4.0
Fe3O4 (1.0–2.5 g/L); pH = 3–11
CNTs/FeS (5–35 mg); pH = 1–12
H2O2/Fe2+ = 1.75 mM; pH = 3
H2O2/Fe2+ = 1.75 mM; pH = 3
nZVI (11.2–28 g/L); pH = 2–5
SBC@b-FeOOH
GFe0.5 (0.01 g/L)
CuCo2O4 (0.1–0.3 g/L)
FeNi3/SiO2 (0.005–0.1 g/L); pH = 3–11
[Fe(II)] (0.8–3 mM)
Alg/Fe (0.2–1.4 g/L); pH = 3
Alg/CDTA/Fe (0.01–0.09 g); pH = 3
CQDs/Cu-MIO (0.1–0.25 g/L); pH = 3.6–10
Fe-Cu@MPSi (0.5–1 g/L); pH = 3–9
Cu@Fe3O4 (0.1–1 g/L); pH = 3.10–9.05
Fe0 (0.3 mM)
Fe0 (0.3 mM)
CUS-MIL-100(Fe) (0.2–1.5 g/L); pH = 3–6
Ce0/Fe0-RGO (0.1–1 g/L); pH = 6–8
Zn-Fe-CNTs (0.2–1 g/L); pH = 1.0–3.0
Fe0 (0.3 mM)
Fe3O4/Humic acid (0–5 g/L); pH = 3.5–9
Fe@Bacillus subtilis (0.5 g/L); pH = 4.0–6.0
Fe0 (0.3 mM)
CFO (0.05–0.2 g/L)
Fe0/CeO2 (0.01–0.2 g/L); pH = 3–7
86.5
80
100
80.92
90.2
80
100
95
80
100
89
91.03
95
95
100
(Zha et al., 2014)
(Elmolla et al., 2010)
Elmolla and Chaudhuri (2009)
(Guo et al., 2015a)
(Rozas et al., 2010)
(Elmolla et al., 2010)
Elmolla and Chaudhuri (2009)
(Mackul’ak et al., 2015)
(Elmolla et al., 2010)
Elmolla and Chaudhuri (2009)
(Hassani et al., 2018)
(Ma et al., 2015)
(Mackul’ak et al., 2015)
(Mackul’ak et al., 2015)
(Wang et al., 2016a)
(Zhang et al., 2016b)
(Ouyang et al., 2019)
(Qi et al., 2019)
Nasseh et al. (2019)
(Santos et al., 2015)
(Titouhi and Belgaied, 2016a)
(Titouhi and Belgaied, 2016b)
(Tian et al., 2017)
(Zheng et al., 2017)
(Pham et al., 2018)
(Pan et al., 2019)
(Pan et al., 2019)
(Tang and Wang, 2018b, 2018a)
Wan and Wang (2016a, 2016b, 2016c)
(Liu et al., 2018b)
(Pan et al., 2019)
(Niu et al., 2011)
(Zheng et al., 2016)
(Pan et al., 2019)
(Parmar et al., 2017)
(Zhang et al., 2019)
Ampicillin
Azithromycin
Cloxacillin
Ciprofloxacin
Clarithromycin
Chlorpheniramine
Doxycycline
Lincomycin
Metacycline
Metronidazole
Nofloxacin
Ofloxacin
Oxytetracycline
Sulfamethazine
Sulfamethoxazole
Sulfadiazine
Sulfathiazole
Tetracycline
100
95.1
95.32
100
100
100
100
100
100
100
100
99
100
100
100
100
100
84
93
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
2HO2 ! O2 + H2 O2
ð9Þ
Therefore catalytic ozonation process can be used to enhance
the degradation efficiency of organic pollutants, including homogeneous and heterogeneous catalytic ozonation.
In homogeneous catalytic ozonation process, liquid catalysts,
especially transition metal ions are used, such as Fe2+, Mn2+, Ni2+,
Co2+, Cd2+, Cu2+, Ag+, Cr3+, Zn2+ 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 (KasprzykHordern et al., 2003).
3.1. Ozone concentration
The ozone concentration has important influence on the degradation 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 antibiotic 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 degradation 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, azithromycin, clarithromycin, erythromycin, ofloxacin, sulfamethoxazole,
tetracycline, trimethoprim. Paucar et al. (2019) studied the degradation 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 condition, 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 min1 to 2.5219 min1 when pH values
increased from 3.0 to 11.0. In alkaline condition, more O3 was
transformed to OH, and indirect oxidation of OH could be more
beneficial in decomposition of flumequine than direct oxidation
of O3. Moreover, the species of flumequine under different pH also
influenced the result.
Wang et al. (2012) explored the chloramphenicol (CAP) degradation 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 conditions. 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.
5
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 generate OH occurred in alkaline condition. On the other hand, nonprotonated organic amine species (–NH2) was more reactive
toward ozone molecules than the mono-protonated form (–NH3)
(Hoigne & Bader, 1983). At pH 9, non-protonated amine (–NH2)
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 (CO2–
3 ) and bicarbonate
(HCO–3) formed during antibiotic decomposition process are hydroxyl radical scavengers, which can inhibit the removal of antibiotics.
On the other hand, solution pH decreased with the ozonation reaction proceeding, which is adverse for the generation of hydroxyl
radicals (OH). Uslu and Balcioglu (2008) observed that mineralization 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 ozonation, 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 BOD5/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 intermediate 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 BOD5/COD ratio at 9 increased
constantly from 0 to 0.41 after 120 min of ozonation, enhancing
the biodegradability and biological treatability of ampicillincontaining wastewater.
Dantas et al. (2008) reported that the biodegradability
increased from 0 to 0.3 during sulfamethoxazole ozonation, indicating that the antibiotic was conversed to biodegradable intermediate product.
Uslu and Balcioglu (2008) observed that the BOD5/COD ratio of
synthetic oxytetracycline wastewater increased from 0.05 to 0.3
due to the formation of biodegradable intermediate products during ozonation process.
3.5. Antibiotics removal by ozone oxidation
The degradation of various antibiotics by ozonation was summarized in Table 2.
4. Photocatalytic oxidation
Photocatalytic oxidation has been extensively studied for the
degradation of organic pollutants. Semi-conductor materials, such
as TiO2, ZnS, WO3 and SnO2 are used as photo-catalyst. When
6
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
Table 2
Antibiotics removal by ozone oxidation.
Antibiotics
O3 dosage or flow rate
Removal efficiency (%)
References
Amoxicillin
[O3]0 = 18 mg/L
[O3]0 = 0–80 mg/L
[O3]0 = 5 mg/min
[O3]0 = 14–42 mg/L
O3 dosage of 57 mg/h
[O3]0 = 0.125–0.75 gO3/gDOC
[O3]0 = 14–42 mg/L
[O3]0 = 0.4–1.16 g/g DOC
[O3]0 = 3 mg/L
[O3]0 = 660–3680 ppm
[O3]0 = 0.31–0.45 g O3/g TS
[O3]0 = 2 mg/L
[O3]0 = 0–80 mg/L
[O3]0 = 2.96 g/h
O3 flow rate of 0.5–1.0 L/h
[O3]0 = 3 mg/L
O3 flow rate of 0.5 L/min
[O3]0 = 3 mg/L
[O3]0 = 0.125–0.75 gO3/gDOC
[O3]0 = 5–7 mg/L
[O3]0 = 0.1–0.5 mg/L
[O3]0 = 140.5 mg/L
[O3]0 = 3 mg/L
[O3]0 = 14–42 mg/L
O3 flow rate of = 100 mg/h
[O3]0 = 14–42 mg/L
[O3]0 = 5–7 mg/L
[O3]0 = 11.2, 32.7 g/m3
[O3]0 = 0.31–0.45 g O3/g TS
O3 flow rate of 0.5 L/min
[O3]0 = 0.125–0.75 gO3/gDOC
[O3]0 = 3 mg/L
[O3]0 = 0–80 mg/L
[O3]0 = 50 g/m3
[O3]0 = 0–1.6 g/L
[O3]0 = 14–42 mg/L
[O3]0 = 0.4–1.16 g/g DOC
O3 flow rate: 0.4 mL/min
[O3]0 = 3 g/h
O3 flow rate: 5.52 ± 0.32 mL/min
[O3]0 = 2 mg/L
[O3]0 = 3–7 mg/L
99
70–98
(Marcelino et al., 2017)
(Alsager et al., 2018)
(Jung et al., 2012)
(Paucar et al., 2019)
(Wang et al., 2012)
(Iakovides et al., 2019)
(Paucar et al., 2019)
(Hollender et al., 2009)
(El-taliawy et al., 2017)
(De Witte et al., 2009)
(Oncu and Balcioglu, 2013)
(Lu et al., 2019)
(Alsager et al., 2018)
(Balcioglu and Otker, 2003)
(Norte et al., 2018)
(El-taliawy et al., 2017)
(Wang et al., 2018)
(El-taliawy et al., 2017)
(Iakovides et al., 2019)
(Ostman et al., 2019)
(Michael-Kordatou et al., 2017)
(Feng et al., 2016a)
(El-taliawy et al., 2017)
(Paucar et al., 2019)
(Chen et al., 2017)
(Paucar et al., 2019)
(Ostman et al., 2019)
(Uslu and Balcioglu, 2008)
(Oncu and Balcioglu, 2013)
(Wang et al., 2018)
(Iakovides et al., 2019)
(El-taliawy et al., 2017)
(Alsager et al., 2018)
(Goncalves et al., 2012)
(Dantas et al., 2008)
(Paucar et al., 2019)
(Hollender et al., 2009)
(Yin et al., 2017)
(Guo et al., 2015b)
(Urbano et al., 2017)
(Lu et al., 2019)
(Oh et al., 2016)
Ampicillin
Chloramphenicol
Clarithromycin
Ciprofloxacin
Ceftriaxone
Ceftriaxone
Clindamycin
Doxycycline
Erythromycin
Fumequine
Isoproturon
Levofloxacin
Nalidixic acid
Metronidazole
Oxytetracycline
Ofloxacin
Roxithromycin,
Sulphadiazine
Sulfamethoxazole
Sulfaquinoxaline
Trimethoprim
100
97
>70
100
15–99
>70
95
98
85.4
70–98
>95
>70
86.4–93.6
>70
>70
43–100
100
100
>70
100
>90
100
43–100
100
88
86.4–93.6
>70
>70
70–98
100
100
100
15–99
99
>85
>99
70
100
photo-catalysts absorb energy, they excites to generate electrons
(e) with high reducing ability and holes (h+) with high oxidizing
ability. O2 can be reduced to form superoxide radical (O
2 ) by the
excited electrons (e). While holes (h+) migrate to the surface of
the photo-catalysts, H2O will be oxidized to generate hydroxyl radical (OH). Then, the organic pollutants could be decomposed by
superoxide radical (O
2 ) or hydroxyl radical ( OH). TiO2 is the most
commonly used catalyst for its high catalytic efficiency, stability
and no secondary pollution. Its mechanisms of photocatalytic oxidation are as following Eqs. (10)–(16) (Liu et al., 2018a; Saadati
et al., 2016).
TiO2 + hc ! TiO2 + hþ + e
ð10Þ
hþ + OH— ! OH
ð11Þ
hþ + H2 O + O2 ! OH + Hþ + O2 ð12Þ
O2 + e !O2 —
ð13Þ
O2 — + Hþ !HO2 —
ð14Þ
2HO2 — !O2 + H2 O2
ð15Þ
H2 O2 + O2 ! OH + OH— + O2
ð16Þ
4.1. Photocatalytic materials
Photocatalytic reaction means a photochemical reaction and
redox process occurring between a photo-catalysts and its surface
substrates such as H2O2, O2 and target pollutants under light irradiation. Photo-catalysts are very important.
To enhance the efficiency of photo-degradation process, various
photo-catalysts, including metal oxides (TiO2, ZnO), metal sulfides
(such as CdS); precious metal semiconductors (Ag3O4, BiOBr, BiOCl,
BiVO4, GdVO4, SmVO4); non-metallic semiconductors (g-C3N4)
have been tested in photochemical oxidation (Li et al., 2019;
Shandilya et al., 2019, 2018; Sivakumar et al., 2018; Tilley, 2019;
Malathi et al., 2018; Qi et al., 2017; Priya et al., 2016; Wen et al.,
2015; Zangeneh et al., 2015).
Owing to high photocatalytic activity, non-toxicity and high
photo-stability, titanium dioxide (TiO2) photo-catalysts have been
extensively applied in environment remediation, especially for the
degradation of toxic organic pollutants, such as antibiotics (Wen
et al., 2015; Kanakaraju et al., 2014), such as levofloxacin (Kansal
et al., 2014), oxytetracycline (Espindola et al., 2019), tetracycline
(Lyu et al., 2019), ciprofloxacin (Zeng et al., 2019), sulfaquinoxaline
(Sandikly et al., 2019).
BiVO4 showed excellent photocatalytic activity in degradation
of pollutants because it has low band gap, good dispersibility,
non-toxicity, resistance to corrosion (Malathi et al., 2018). It has
been applied for decomposition of antibiotics, such as ciprofloxacin
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
(Chen et al., 2018; Yan et al., 2013), oxytetracycline (Ye et al.,
2019), norfloxacin (Du et al., 2019), penicillin (Liu et al., 2019a),
tetracycline (Wang et al., 2019a).
Zinc oxide (ZnO) has also been explored for degradation of
antibiotics, such as ciprofloxacin (Sarkhosh et al., 2019), norfloxacin (Mamba et al., 2018), sulfamethoxazole (Mirzaei et al.,
2018), cefixime trihydrate (Shooshtari and Ghazi, 2017), tetracycline hydrochloride (Ji et al., 2018) because of its low cost, high
redox potential, non-toxicity, and environmentally-friendly
properties.
Various new carbon materials, such as graphene (Sun and
Chang, 2014), carbon nanotubes (Yan et al., 2015), fullerene (Ge
et al., 2019), carbon quantum dots (Yi et al., 2018) has been investigated for photocatalytic degradation of organic pollutant. They
exhibited high catalytic activity in decomposition of antibiotics,
such as sulfamethazine (Liu et al., 2019b), oxytetracycline (Wang
et al., 2019b), tetracycline (Yuan et al., 2019), levofloxacin (Kaur
et al., 2019), and cefixime (Sheydaei et al., 2018).
4.2. pH value
Solution pH is important for photocatalytic oxidation. When pH
is lower or higher than the potential of zero charge of catalysts, catalyst surface has different charges. Similarly, when pH is lower or
higher than the pKa of substrates, substrates show different
charged form (Mehrjouei et al., 2015).
Dimitrakopoulou et al. (2012) applied UV-A/TiO2 photo-catalyst
to degrade amoxicillin and found that amoxicillin degradation
showed no apparent change at pH 5.0 and 7.5. The mineralization
of amoxicillin decreased from 95% to 75% when pH increased from
5 to 7.5. This may be associated with the ionization states of both
the catalyst and pollutants.
Leon et al. (2017) investigated the degradation of cefotaxime
under sunlight radiation usingTiO2 and ZnO in aqueous solutions.
During the cefotaxime degradation using TiO2 as catalyst, cefotaxime removal rate increased when pH increased from 4 to 6.2,
then it declined when pH further increased from 6.2 to 7.6. Similar
results were also observed during the cefotaxime degradation
using ZnO as catalyst.
Palominos et al. (2008) investigated the photocatalytic oxidation of flumequine (FQ) using TiO2 as catalyst. The maximum
degradation rate of flumequine was observed at medium pH value.
At lower pH, catalyst and flumequine (pKa 6.35) are positively
charged, inhibiting the affinity between them, while at pH higher
than pKa, both catalyst and flumequine are negatively charged.
4.3. Catalysts dosage
Photo-catalyst dosage is important in photocatalytic oxidation
process. Increasing photo-catalyst loads would increase the reactive sites and then improve the oxidation and mineralization efficiencies of pollutants. However, excessively added catalysts
would block the penetration of the photons and cause the loss of
light energy through shielding, reflection and scattering of light
by solid particles (Gong and Chu, 2015; Nezamzadeh-Ejhieh and
Shams-Ghahfarokhi, 2013).
Ahmadi et al. (2017) observed the effect of MWCNT/TiO2 dosage
on photocatalytic degradation of tetracycline (TC). It was found
that tetracycline removal increased when the catalyst dosage
increased from 0.1 g/L to 0.2 g/L. While it further increased to
0.4 g/L, tetracycline degradation efficiency did not further increase.
Zhang et al. (2018) found that TiO2 dosage strongly influenced
the photocatalytic degradation of chloramphenicol (CAP). When
TiO2 dosage increased from 0 to 1 g/L, the kinetic rate constant
increased from 0.00534 to and 0.03160 min1. The excess TiO2
may result in light screening effects and decrease light intensity.
7
Lofrano et al. (2014) observed that the photocatalytic degradation rate constant of vancomycin B was 0.013 and 0.036 min1,
respectively at TiO2 dosage of 0.1 g/L and 0.2 g/L.
4.4. Catalysts stability
The photo-catalytic efficiency is related to the stability of catalysts. In practical application, the stability of catalysts is an important factor considering economic costs. Zhu et al. (2018)
investigated the stability of CdS/Fe3O4/g-C3N4 during the cyclic
photocatalytic degradation of tetracycline, and found that the catalytic activity kept almost unchanged for five cycle experiments.
Wu et al. (2016) evaluated the stability of photo-catalyst CdS/
SrTiO3 heterojunction during five consecutive cycles of photocatalytic degradation of ciprofloxacin, and found that it was stable
and photo-corrosion resistant. Xue et al. (2015) studied the degradation of tetracycline using Au/Pt/g-C3N4 as photo-catalyst, they
found that the photocatalytic efficiency declined 8.7% after four
cycles, indicating that the photo-catalyst was stable. Liu et al.
(2018a) conducted the photo-catalyst recycling experiments for
tetracycline degradation to evaluate the reliability of photocatalyst BGC1.
4.5. Mineralization of antibiotics
The mineralization efficiency is usually lower than degradation
efficiency because there are some transient organic intermediates
formed during the photocatalytic process. Liu et al. (2016) found
that the TOC removal was very slow, it was only 9.5% for oxytetracycline degradation under UV irradiation for 10 h. Ahmadi et al.
(2017) found that TOC removal of tetracycline reached to 83% after
UV irradiation for 300 min with the addition of 0.2 g/L MWCNT/
TiO2. Palominos et al. (2008) observed that the mineralization ratio
of flumequine reached 74% after UV irradiation for 15 min, it
remained almost unchanged after 60 min even a longer irradiation
period. After longer irradiation times, some intermediates such as
aromatic ring remained intact.
4.6. Antibiotics removal by photocatalytic oxidation
Table 3 summarized the antibiotics removal by photocatalytic
oxidation using various catalyst materials.
5. Electrochemical oxidation
Electrochemical oxidation is a process in which organic substance are oxidized and converted or decomposed into non-toxic
and harmless substance under the action of an electric current
(Comninellis, 1994; Martinez-Huitle and Brillas, 2009). The electrochemical oxidation technology includes direct oxidation and
indirect oxidation, they generally exist simultaneously (MartinezHuitle and Ferro, 2006; Simond et al., 1997; Comninellis, 1994).
During direct oxidation process, organic matters in water could
directly react with anode and lose electrons to form small molecular compounds (Comninellis, 1994; Kirk et al., 1985). As for indirect
oxidation process, anions in the water react with anode to produce
intermediate products with strong oxidizing ability, and these
intermediate products further oxidize and decompose organic substances (Do and Yeh, 1996; Chiang et al., 1995). This process is
related to electrolytes.
Different electrolytes will produce different strong oxidative
products and result in different degradation efficiency (Feng
et al., 2016b). Hydroxyl radical (OH) is a kind of intermediate oxidant generated by indirect electrochemical oxidation, which is
adsorbed onto the anode surface. Organic matters also could be
8
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
Table 3
Antibiotics removal by photocatalytic oxidation.
Antibiotics
Photo-catalysts
Removal efficiency (%)
References
Amoxicillin
Cefixime
Cefixime trihydrate
Cefotaxime
Chloramphenicol
Ciprofloxacin
UV-A/TiO2
N-TiO2/GO
ZnO/a-Fe2O3
TiO2/ZnO
TiO2
TiO2/carbon dots
RGO-BiVO4
Ag/AgBr/BiVO
ZnO
CdS/SrTiO3
g-C3N4, Fe3O4/g-C3N4
CdS/SrTiO3
g-C3N4, Fe3O4/g-C3N4
TiO2
g-C3N4, Fe3O4/g-C3N4
TiO2
rGO-CdS
BiVO4/WO3
ZnO/CuOx
TiO2
MWCNT/BiVO4
TiO2@GO
GQDs/g-CNNR
100
80
99.1
84.2
85
91.1
68.2
91.40
100
93.7
100
93.7
100
100
100
90
82.7
70
>80
100
88.8
100
80
100
93.7
100
96.1
95.76
100
100
100
(Dimitrakopoulou et al., 2012)
(Sheydaei et al., 2018)
(Shooshtari and Ghazi, 2017)
(Leon et al., 2017)
(Zhang et al., 2018)
(Zeng et al., 2019)
(Yan et al., 2013)
(Chen et al., 2018)
(Sarkhosh et al., 2019)
(Wu et al., 2016)
(Zhu et al., 2018)
(Wu et al., 2016)
(Zhu et al., 2018)
(Palominos et al., 2008)
(Zhu et al., 2018)
(Kansal et al., 2014)
(Kaur et al., 2019)
(Du et al., 2019)
(Mamba et al., 2018)
(Espindola et al., 2019)
(Ye et al., 2019)
(Wang et al., 2019b)
(Yuan et al., 2019)
(Liu et al., 2016)
(Wu et al., 2016)
(Liu et al., 2019a)
(Liu et al., 2019b)
(Mirzaei et al., 2018)
(Sandikly et al., 2019)
(Zhu et al., 2018)
(Lyu et al., 2019)
(Xue et al., 2015)
(Ji et al., 2018)
(Ahmadi et al., 2017)
(Wang et al., 2019a)
(Lofrano et al., 2014)
Enrofloxacin
FLumequine
Gatifloxacin
Levofloxacin
Norfloxacin
Oxytetracycline
Penicillin
Sulfamethazine
Sulfamethoxazole
Sulfaquinoxaline
Tetracycline
Vancomycin
CdS/SrTiO3
BiVO4
G/A/TNS
UVC lamp (10 W)
TiO2
g-C3N4, Fe3O4/g-C3N4
Mesoporous TiO2
Au/Pt/g-C3N4
ZnO@ZnS
MWCNT/TiO2
BiVO4/Bi2Ti2O7/Fe3O4
TiO2
oxidized by hydroxyl radicals (OH) into small molecular compounds and carbon dioxide (Garcia-Segura et al., 2018).
5.1. Electrode materials
The earliest anode applied in electrochemical oxidation is a
metal electrode, which is a bare electrode without oxide film on
its surface. Such anodes are highly conductive, however, they are
prone to dissolution during electrolysis process, resulting in anode
loss and solution contamination by new impurities.
To avoid its disadvantage and improve oxidation efficiency,
plenty of new anode materials are studied, including graphite
(Liu and Jiang, 2005), glassy carbon (Brimecombe and Limson,
2006), conductive-diamond (Canizares et al., 2006), activated
carbon-steel (Canizares et al., 1999), Pt (Rao and Dube, 1996),
TiO2 (Zhang et al., 2014), nanostructured TiO2 (Tian et al., 2008),
b-PbO2 (Wu and Zhou, 2001), IrO2/Ti (Bonin et al., 2004), Ti/TiO2
-RuO2 -IrO2 (Rajkumar and Palanivelu, 2003), Ti/Pt (Vlyssides
et al., 2004), TiO2/Ti/Ta2O5 - IrO2 (Asmussen et al., 2009), Sb
doped- SnO2 (Zhao et al., 2009), BiOx - TiO2/Ti (Park et al., 2009).
The property of anode is related to the preparation method. The
composition ratio, particle size, surface structure, specific surface
area and bonding force, all affect the performance of the anode
(Feng et al., 2016b). Electrochemical oxidation has been applied
for the degradation of antibiotics, such as chlortetracycline
(Kitazono et al., 2017), cefazolin (Kitazono et al., 2017), tetracycline
(Liu et al., 2015; Miyata et al., 2011), ofloxacin (Jara et al., 2007),
lincomycin (Jara et al., 2007), sulfamethoxazole (Eleoterio et al.,
2013), trimethoprim (Eleoterio et al., 2013), nitrofurazone (Kong
et al., 2015), metronidazole (Kong et al., 2015), ceftriaxone (Li
et al., 2018).
80.9
100
97.14
95
5.2. Current density
The current density affects the driving force of the electrochemical oxidation reaction, thus it affects the electrochemical oxidation
efficiency (Moreira et al., 2017). Kitazono et al. (2017) studied the
electrochemical oxidation of chlortetracycline using Ti/PbO2 as
anode, they found that the degradation of chlortetracycline followed the pseudo first-order kinetics, and the reaction rate constant increased with increase of applied current density, due to
the higher OH yield rate under higher current density. Eleoterio
et al. (2013) investigated the effect of current density on the
removal rate of COD of the wastewater contained antibiotics, such
as sulfamethoxazole and trimethoprim. They observed that the
COD removal efficiency increased when the current density
increased from 10 to 100 mA/cm2. Dirany et al. (2010) studied
the effect of current density on the degradation of sulfamethoxazole. Haidar et al. (2013) studied the degradation of sulfachloropyridazine using a boron-doped diamond (BDD) anode, and found
that the time required for the complete antibiotic decomposition
decreased with the increase of current density. Moreover, the mineralization rate was 76%, 84%, 89%, 93% and 95% when the current
density was 100, 200, 300, 350 and 400 mA, respectively. Moreira
et al. (2014) observed that trimethoprim was degraded rapidly at
high current density, and the kinetic rate constant of trimethoprim
degradation increased with increase of current density.
5.3. pH value
Solution pH influences the performance of electrochemical oxidation. The improvement or inhibition effect is related to water
composition, reaction system (Moreira et al., 2017, 2014;
9
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
Almeida et al., 2012). Moreira et al. (2014) investigated the effect of
pH on trimethoprim degradation using electrochemical oxidation,
they found that at pH 3.0, the presence of HSO-4 in solution may
scavenge hydroxyl radical (OH) and reduce the degradation efficiency of trimethoprim; at pH 4.5, the removal rate decreased
due to iron precipitation.
El-Ghenymy et al. (2013) explored the degradation of sulfamethazine using BDD anode. The mineralization of sulfamethazine reached 90%, which was highest at pH 3.0. DOC was
only reduced by 84%, 84%, 80% and 65% at pH of 2.0, 4.0, 5.0
and 6.0, respectively. At pH 2.0, H2O2 is easily reacts with H+
to form peroxonium ion (H2O+2), which makes H2O2 more electrophilic and decreases its reactivity with Fe2+ in Fenton reaction. When pH was over 4.0, the gradual precipitation of Fe3+
reduced the formation of hydroxyl radical (OH) and inhibited
the sulfamethazine removal.
Wang et al. (2016c) studied the electrochemical oxidation of
ciprofloxacin using SnO2-Sb/Ti electrode. They found that the
removal of ciprofloxacin and COD was slightly higher at higher
pH. The kinetic rate constant and average current efficiency were
maximal at pH 3, higher than that at pH 5, 7, 9 and 11.
6. Ionizing radiation
The ionizing radiation (including gamma ray and electron
beam) is an emerging technology for the degradation of organic
pollutants, either through indirect way or direct way (Fig. 2).
During water radiolytic process, various active species are
formed as Eq. (17).
H2 O ! OH (2.7) + eaq (2.6) + H (0.55) + H2 (0.45)
þ H2 O2 ð0:71Þ þ H3 Oþ ð2:6Þ
ð17Þ
The numbers in brackets are chemical yield (G-value), presenting the number of species formed when absorbed 100 eV energy at
a pH range of 6.0–8.5.
Hydroxyl radicals (OH) can oxidize the organic pollutants, and
solvated electrons (e
aq) can reduce the organic pollutants (Wang
and Chu, 2016; Wang and Wang, 2007). The degradation of antibiotics by ionizing radiation is influenced by various factors, such as
the absorbed dose, initial pH, organic matters and water matrix (Yu
et al., 2010a; 2010b; Hu and Wang, 2007).
6.1. Absorbed dose
5.4. Antibiotics removal by electrochemical oxidation
Antibiotics removal by electrochemical oxidation was summarized in Table 4.
The absorbed dose considerably affects the degradation rate of
antibiotics. Generally, the antibiotics degradation increases with
increase of absorbed dose (Zhuan and Wang, 2019a, 2019b). The
Table 4
Antibiotics removal by electrochemical oxidation.
Type of antibiotics
Anode material
Removal efficiency (%)
References
Cefazolin
Ceftriaxone sodium
Chlortetracycline
Ti/PbO2
RuO2 -TiO2 /Nano-G
Ti/PbO2
Ti/IrO2, Ti/PbO2
SnO2-Sb/Ti
Ti/IrO2, Ti/PbO2
Ti/IrO2, Ti/PbO2
Boron-doped diamond (BDD)
BDD/carbon
Pt/carbon, BDD/Carbon
Boron-doped diamond
Ti/IrO2, Ti/PbO2
Carbon nanotube
100
>97.3
100
>99
99.5
>99
>99
(Kitazono et al., 2017)
(Li et al., 2018)
(Kitazono et al., 2017)
(Miyata et al., 2011)
(Wang et al., 2016c)
(Miyata et al., 2011)
(Miyata et al., 2011)
El-Ghenymy et al. (2013)
Haidar et al. (2013)
Dirany et al. (2010)
(Moreira et al., 2014)
(Miyata et al., 2011)
(Liu et al., 2015)
Ciprofloxacin
Doxycycline
Oxytetracycline
Sulfamethazine
Sulfachloropyridazine
Sulfamethoxazole
Tetracycline
100
100
100
>99
96.3
Fig. 2. Principles of ionizing radiation for the decomposition of organic pollutants.
10
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
radiation-induced degradation of antibiotics follows the pseudo
first-order kinetics model (Eqs. (18) and (19)).
C ¼ C 0 ekD
ln
ð18Þ
C
¼ kD
C0
ð19Þ
where C0 and C are pollutant concentration before and after radiation (mg/L); D is absorbed dose (kGy); and k is dose constant
(kGy1).
The degradation kinetics of various antibiotics followed to
pseudo first-order kinetic model, such as cefaclor (Yu et al.,
2008), diclofenac (Zhuan and Wang, 2020a; He et al., 2014), sulfamethazine (Chu et al., 2015) and sulfamethoxazole (Zhuan and
Wang, 2020b).
OH + HCO3 ! CO3 +OH (k = 3.9 10
8
L/(mol s))
H + HCO3 !H2 + CO3 (k = 4.0 10
4
eaq + HCO3 ! CO3 3 (k = 3.9 10
L/(mol s))
5
L/(mol s))
ð27Þ
ð28Þ
ð29Þ
OH + NO2 ! NO2 + HO (k = 6.0 109 L/(mol s))
ð30Þ
H + NO2 !NO + HO (k = 7.1 108 L/(mol s))
ð31Þ
eaq + NO2 ! NO2 2 (k = 3.5 109 L/(mol s))
ð32Þ
Hþ + NO3 !HNO3 (k = (4.4—6.0) 108 L/(mol s))
ð33Þ
OH + HNO3 ! H2 O + NO3 (k = (0.88—1.2) 108 L/(mol s))
ð34Þ
6.2. pH value
The pH value has significant influence on the degradation of
antibiotics by ionizing radiation (Wang and Wang, 2019c, 2018c).
Solution pH can affect the reactive radical composition by Eqs.
(20)–(22).
H + HNO3 ! H2 + NO3 (k <= 1.0 107 L/(mol s))
ð35Þ
NO3 + H2 O ! HNO3 + HO (k = 3.0 102 L/(mol s))
ð36Þ
eaq + Hþ ! H (k = 2.3 1010 L/(mol s))
ð20Þ
eaq + NO3 ! NO3 2 (k = 9.7 109 L/(mol s))
ð37Þ
eaq + OH ! OH
ð21Þ
NO3 2 + Hþ !NO2 + HO (k = 4.5 1010 L/(mol s))
ð38Þ
OH + OH ! H2 O + O (k = 1.3 1010 L/(mol s))
ð22Þ
H + NO3 ! NO2 + HO (k = 4.4 106 L/(mol s))
ð39Þ
At acidic condition, H concentration was higher than OH concentration, which can combine with e
aq at a rate constant of
2.3 1010 L/(mol s), and inhibit the reaction between e
aq and
OH. As a consequence, more OH would react with antibiotic molecules (Guo et al., 2012).
At alkaline condition, OH– concentration was higher than H+
concentration, which can react with OH at a rate constant of
1.3 1010 L/(mol s) and form weak oxidative O and H2O, reducing
OH concentration and resulting in lower degradation efficiency
(Basfar et al., 2005).
The acidic condition was more effective than alkaline condition
for the degradation of sulphadiazine (Guo et al., 2012), ciprofloxacin (Guo et al., 2015b), metronidazole, norfloxacin.
For the antibiotics with zwitter ion character, solution pH can
affect the distribution of their molecular and ionic forms, as well
as the surface charge property (De Bel et al., 2009), which will generate attraction or repulsion force between different antibiotic
forms, finally affecting the degradation efficiency. Zhuan and
Wang (2019a) found that when pH was higher than pKa2 (5.7), sulfamethoxazole was mainly in negative charged forms, which
would produce repulsion force, decreasing the reaction rate.
NO2 + H ! HNO2 (k = 1.0 1010 L/(mol s))
ð40Þ
OH + NO3 ! HONO3 (k = 1.0 1010 L/(mol s))
ð41Þ
H + NO3 ! HNO3 (k = 1.0 1010 L/(mol s))
ð42Þ
Cl + OH ! ClHO (k = 4.3 109 L/(mol s))
ð43Þ
ClHO + Hþ !Cl + 2H2 O (k = 2.1 1010 L/(mol s))
ð44Þ
+
–
6.3. Inorganic anions, organic matters and matrix
Practical waters are complex matrices which usually include
–
anions (such as Cl-, CO2–
3 , HCO3, NO3 , NO2 ) and organic matters
(such as humic acid). These compouds may interfere with the radiolytic degradation of antibiotic by reacting with the radical species
as Eqs. (23)–(44) (Buxton et al., 1988).
ClHO + eaq !Cl + HO (k = 1.0 1010 L/(mol s))
ð23Þ
ClHO !Cl + OH (k = 6.1 109 L/(mol s))
ð24Þ
OH + CO3
2
! CO3 + H2 O (k = 8.5 106 L/(mol s))
eaq + CO3 2 !HCO3 2 (k = 6.0 10
5
L/(mol s))
ð25Þ
ð26Þ
The presence of CO2
3 reduced the removal rate of sulphadiazine
(Guo et al., 2012), ciprofloxacin, norfloxacin, amoxicillin. The presence of HCO
3 had inhibitory effect on the degradation of antibiotics, such as amoxicillin, sulfamethoxazole (Zhuan & Wang,
2019b), ofloxacin, norfloxacin (Sayed et al., 2016), cefradine. The
presence of NO
3 and NO2 decreased the decomposition of antibiotics, such as ciprofloxacin, norfloxacin (Sayed et al., 2016), sulfamethoxazole (Zhuan and Wang, 2019b).
Wang and Wang (2018d) studied the radiation-induced degradation of sulfamethoxazole in the presence of various inorganic
anions, including chloride, bicarbonate, carbonate, nitrate, sulfate
and phosphate. The results showed that inorganic anions had obvious influence on SMX degradation, which was dependent on their
initial concentrations, suggesting that the effect of inorganic anions
on the radiation-induced degradation of sulfonamides antibiotics
should be considered when radiation technology is used for the
treatment of industrial wastewater.
The existence of humic acid could decrease the degradation efficiency of various types of antibiotics, such as ciprofloxacin, amoxicillin, sulfamethoxazole (Zhuan and Wang, 2019b), ofloxacin,
cefradine, fluoroquinolone (Tegze et al., 2018).
7. Concluding remarks and perspectives
Antibiotics are becoming emerging contaminants, which have
received increasing attention in recent years because they are
ubiquitous in the natural environment. Moreover, antibiotics can
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
transfer and accumulate through food chain. The long-term existence of antibiotics in the environment will cause the generation
of antibiotic resistance genes (ARGs) and antibiotic resistant bacteria (ARBs), posing potential threat to ecosystem and human health.
Due to high degradation efficiency and rate, advances oxidation
processes are promising for degradation of antibiotics in water
and wastewater.
The advantages and disadvantages of different advanced oxidation processes for antibiotics removal were analyzed, summarized
and compared in supporting information (Table S1).
At present, research on antibiotics removal by advanced oxidation processes have made some progress, and the future research
should be focused on the following aspects.
(1) Advanced oxidation processes need to be optimized to
improve their adaptability and practicability, such as
enhancing the efficiency of the catalysts, and the utilization
efficiency of ozone or H2O2.
(2) Degradation effect of antibiotics by advanced oxidation processes has been investigated. The generation mechanism of
free radicals and the degradation mechanism of pollutants
are not yet clear. More attention should be paid to the mechanism study.
(3) Advanced oxidation processes can effectively degrade
antibiotics in water and wastewater, their potential for the
removal of ARGs and ARBs has not been studied, which
needs further investigation.
(4) Actual wastewater is complicated, usually containing multiple antibiotics and other organic pollutants, as well as inorganic compounds, which may decrease the degradation
efficiency of antibiotics compared with single antibiotic in
aqueous solution. Thus, more studies are needed to pay
attention to the practical wastewater and finally fulfil the
industrial application.
(5) It is difficult to efficiently treat the complicated antibiotic
wastewater by only advanced oxidation processes. Combining AOPs with biological treatment methods could be one
way to resolve this problem, especially to enhance the mineralization of pollutants. The integrated treatment methods
can reduce the operational costs and improve the processing
efficiency.
(6) The operational cost is crucial for the practical applications,
how to reduce the treatment cost of AOPs is also important,
and their cost-effect analysis should be considered in future
studies.
Declaration of Competing Interest
None.
Acknowledgements
This study was supported by National Natural Science Foundation of China (51978368) and the Program for Changjiang Scholars
and Innovative Research Team in University (IRT-13026).
Appendix A. Supplementary material
Supplementary data to this article can be found online at
https://doi.org/10.1016/j.scitotenv.2019.135023.
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