Telechargé par Mibeibrahimyeo16

1-s2.0-S0048969719350156-main

publicité
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
[email protected]
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
[email protected] (0.5–1 g/L); pH = 3–9
[email protected] (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
[email protected] 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
[email protected]
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
[email protected]
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.
References
Ahmadi, M., Motlagh, H.R., Jaafarzadeh, N., Mostoufi, A., Saeedi, R., Barzegar, G., Jorfi,
S., 2017. Enhanced photocatalytic degradation of tetracycline and real
11
pharmaceutical wastewater using MWCNT/TiO2 nano-composite. J. Environ.
Manage. 186, 55–63.
Almeida, L.C., Garcia-Segura, S., Arias, C., Bocchi, N., Brillas, E., 2012. Electrochemical
mineralization of the azo dye Acid Red 29 (Chromotrope 2R) by photoelectroFenton process. Chemosphere 89, 751–758.
Alsager, O.A., Alnajrani, M.N., Abuelizz, H.A., Aldaghmani, I.A., 2018. Removal of
antibiotics from water and waste milk by ozonation: kinetics, byproducts, and
antimicrobial activity. Ecotoxicol. Environ. Saf. 158, 114–122.
Asmussen, R.M., Tian, M., Chen, A.C., 2009. A new approach to wastewater
remediation based on bifunctional electrodes. Environ. Sci. Technol. 43, 5100–
5105.
Balcioglu, I.A., Otker, M., 2003. Treatment of pharmaceutical wastewater containing
antibiotics by O3 and O3/H2O2 processes. Chemosphere 50, 85–95.
Barreiro, J.C., Capelato, M.D., Martin-Neto, L., Hansen, H.C.B., 2007. Oxidative
decomposition of atrazine by a Fenton-like reaction in a H2O2/ferrihydrite
system. Water Res. 41, 55–62.
Basfar, A.A., Khan, H.M., Al-Shahrani, A.A., 2005. Trihalomethane treatment using
gamma irradiation: kinetic modeling of single solute and mixtures. Radiat.
Phys. Chem. 72, 555–563.
Bonin, P.M.L., Bejan, D., Schutt, L., Hawari, J., Bunce, N.J., 2004. Electrochemical
reduction of hexahydro-1,3,5-trinitro-1,3,5-triazine in aqueous solutions.
Environ. Sci. Technol. 38, 1595–1599.
Brimecombe, R.D., Limson, J.L., 2006. Electrochemical investigation of the effect of
pH and solvent on amitraz stability. J. Agri. Food Chem. 54, 8139–8143.
Buxton, G.V., Greenstock, C.L., Helman, W.P., Ross, A.B., 1988. Critical-review of rate
constants for reactions of hydrated electrons, hydrogen-atoms and hydroxyl
radicals (OH/O-) in aqueous-solution. J. Phys. Chem. Ref. Data 17, 513–886.
Canizares, P., Dominguez, J.A., Rodrigo, M.A., Villasenor, J., Rodriguez, J., 1999. Effect
of the current intensity in the electrochemical oxidation of aqueous phenol
wastes at an activated carbon and steel anode. Ind. Eng. Chem. Res. 38 (10),
3779–3785.
Canizares, P., Paz, R., Lobato, J., Saez, C., Rodrigo, M.A., 2006. Electrochemical
treatment of the effluent of a fine chemical manufacturing plant. J. Hazard.
Mater. 138, 173–181.
Cerqueira, F., Matamoros, V., Bayona, J., Elsinga, G., Hornstra, L.M., Pina, B., 2019.
Distribution of antibiotic resistance genes in soils and crops. A field study in
legume plants (Vicia faba L.) grown under different watering regimes. Environ.
Res. 170, 16–25.
Chen, F., Yang, Q., Wang, Y.L., Yao, F.B., Ma, Y.H., Huang, X.D., Li, X.M., Wang, D.B.,
Zeng, G.M., Yu, H.Q., 2018. Efficient construction of bismuth vanadate-based Zscheme photocatalyst for simultaneous Cr(VI) reduction and ciprofloxacin
oxidation under visible light: kinetics, degradation pathways and mechanism.
Chem. Eng. J. 348, 157–170.
Chen, K., Zhou, J.L., 2014. Occurrence and behavior of antibiotics in water and
sediments from the Huangpu River, Shanghai, China. Chemosphere 95, 604–
612.
Chen, W.R., Li, X.K., Pan, Z.Q., Ma, S.S., Li, L.S., 2017. Synthesis of MnOX/SBA-15 for
Norfloxacin degradation by catalytic ozonation. Sep. Purif. Technol. 173, 99–
104.
Chiang, L.C., Chang, J.E., Wen, T.C., 1995. Indirect oxidation effect in electrochemical
oxidation treatment of landfill leachate. Water Res. 29, 671–678.
Chu, L.B., Wang, J.L., Liu, Y.K., 2015. Degradation of sulfamethazine in sewage sludge
mixture by gamma irradiation. Radiat. Phys. Chem. 108, 102–105.
Comninellis,
C.,
1994.
Electrocatalysis
in
the
electrochemical
conversion/combustion of organic pollutants for waste-water treatment.
Electrochim. Acta 39, 1857–1862.
Danner, M.C., Robertson, A., Behrends, V., Reiss, J., 2019. Antibiotic pollution in
surface fresh waters: occurrence and effects. Sci. Total Environ. 664, 793–804.
Dantas, R.F., Contreras, S., Sans, C., Esplugas, S., 2008. Sulfamethoxazole abatement
by means of ozonation. J. Hazard. Mater. 150, 790–794.
Daud, N.K., Hameed, B.H., 2010. Decolorization of Acid Red 1 by Fenton-like process
using rice husk ash-based catalyst. J. Hazard. Mater. 176, 938–944.
De Bel, E., Dewulf, J., De Witte, B., Van Langenhove, H., Janssen, C., 2009. Influence of
pH on the sonolysis of ciprofloxacin: biodegradability, ecotoxicity and antibiotic
activity of its degradation products. Chemosphere 77, 291–295.
De Witte, B., Dewulf, J., Demeestere, K., Van Langenhove, H., 2009. Ozonation and
advanced oxidation by the peroxone process of ciprofloxacin in water. J. Hazard.
Mater. 161, 701–708.
Dimitrakopoulou, D., Rethemiotaki, I., Frontistis, Z., Xekoukoulotakis, N.P., Venieri,
D., Mantzavinos, D., 2012. Degradation, mineralization and antibiotic
inactivation of amoxicillin by UV-A/TiO2 photocatalysis. J. Environ. Manage.
98, 168–174.
Dirany, A., Sires, I., Oturan, N., Oturan, M.A., 2010. Electrochemical abatement of the
antibiotic sulfamethoxazole from water. Chemosphere 81, 594–602.
Djeffal, L., Abderrahmane, S., Benzina, M., Fourmentin, M., Siffert, S., Fourmentin, S.,
2014. Efficient degradation of phenol using natural clay as heterogeneous
Fenton-like catalyst. Environ. Sci. Pollut. Res. 21, 3331–3338.
Do, J.S., Yeh, W.C., 1996. Paired electrooxidative degradation of phenol with in situ
electrogenerated hydrogen peroxide and hypochlorite. J. Appl. Electrochem. 26,
673–678.
Du, H., Pu, W.H., Wang, Y.Y., Yan, K., Feng, J., Zhang, J.D., Yang, C.Z., Gong, J.Y., 2019.
Synthesis of BiVO4/WO3 composite film for highly efficient visible light induced
photoelectrocatalytic oxidation of norfloxacin. J. Alloys Compd. 787, 284–294.
Eleoterio, I.C., Forti, J.C., de Andrade, A.R., 2013. Electrochemical treatment of
wastewater of veterinary industry containing antibiotics. Electrocatalysis 4,
283–289.
12
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
El-Ghenymy, A., Rodriguez, R.M., Arias, C., Centellas, F., Garrido, J.A., Cabot, P.L.,
Brillas, E., 2013. Electro-Fenton and photoelectro-Fenton degradation of the
antimicrobial sulfamethazine using a boron-doped diamond anode and an airdiffusion cathode. J. Electroanal. Chem. 701, 7–13.
Elmolla, E., Chaudhuri, M., 2009. Optimization of Fenton process for treatment of
amoxicillin, ampicillin and cloxacillin antibiotics in aqueous solution. J. Hazard.
Mater. 170, 666–672.
Elmolla, E.S., Chaudhuri, M., Eltoukhy, M.M., 2010. The use of artificial neural
network (ANN) for modeling of COD removal from antibiotic aqueous solution
by the Fenton process. J. Hazard. Mater. 179, 127–134.
El-taliawy, H., Ekblad, M., Nilsson, F., Hagman, M., Paxeus, N., Jonsson, K., Cimbritz,
M., Jansen, J.L., Bester, K., 2017. Ozonation efficiency in removing organic micro
pollutants from wastewater with respect to hydraulic loading rates and
different wastewaters. Chem. Eng. J. 325, 310–321.
Espindola, J.C., Cristovao, R.O., Santos, S.G.S., Boaventura, R.A.R., Dias, M.M., Lopes, J.
C.B., Vilar, V.J.P., 2019. Intensification of heterogeneous TiO2 photocatalysis
using the NETmix mili-photoreactor under microscale illumination for
oxytetracycline oxidation. Sci. Total Environ. 681, 467–474.
Etaiw, S.E.H., El-bendary, M.M., 2012. Degradation of methylene blue by catalytic
and photo-catalytic processes catalyzed by the organotin-polymer
3
1[(Me3Sn)4Fe(CN)6]. Appl. Catal. B-Environ. 126, 326–333.
Feng, M.B., Yan, L.Q., Zhang, X.L., Sun, P., Yang, S.G., Wang, L.S., Wang, Z.Y., 2016a.
Fast removal of the antibiotic flumequine from aqueous solution by ozonation:
Influencing factors, reaction pathways, and toxicity evaluation. Sci. Total
Environ. 541, 167–175.
Feng, Y.J., Yang, L.S., Liu, J.F., Logan, B.E., 2016b. Electrochemical technologies for
wastewater treatment and resource reclamation. Environ. Sci.-Water Res.
Technol. 2, 800–831.
Fenton, H.J.H., 1894. LXXIII.—Oxidation of tartaric acid in presence of iron. J. Chem.
Soc. Trans. 65, 899–910.
Focazio, M.J., Kolpin, D.W., Barnes, K.K., Furlong, E.T., Meyer, M.T., Zaugg, S.D.,
Barber, L.B., Thurman, M.E., 2008. A national reconnaissance for
pharmaceuticals and other organic wastewater contaminants in the United
States - II) Untreated drinking water sources. Sci. Total Environ. 402, 201–
216.
Fukuchi, S., Nishimoto, R., Fukushima, M., Zhu, Q.Q., 2014. Effects of reducing agents
on the degradation of 2,4,6-tribromophenol in a heterogeneous Fenton-like
system with an iron-loaded natural zeolite. Appl. Catal. B-Environ. 147, 411–
419.
Garcia-Segura, S., Ocon, J.D., Chong, M.N., 2018. Electrochemical oxidation
remediation of real wastewater effluents - a review. Process Saf. Environ.
Prot. 113, 48–67.
Ge, J., Zhang, Y., Park, S.J., 2019. Recent advances in carbonaceous photocatalysts
with enhanced photocatalytic performances: a mini review. Materials 12, 1916.
Ghosh, P., Kumar, C., Samanta, A.N., Ray, S., 2012. Comparison of a new immobilized
Fe3+ catalyst with homogeneous Fe3+-H2O2 system for degradation of 2,4dinitrophenol. J. Chem. Technol. Biotechnol. 87, 914–923.
Goncalves, A.G., Orfao, J.J.M., Pereira, M.F.R., 2012. Catalytic ozonation of
sulphamethoxazole in the presence of carbon materials: Catalytic
performance and reaction pathways. J. Hazard. Mater. 239, 167–174.
Gong, H., Chu, W., 2015. Photodegradation of sulfamethoxazole with a recyclable
catalyst. Ind. Eng. Chem. Res. 54, 12763–12769.
Guo, R.X., Xie, X.D., Chen, J.Q., 2015a. The degradation of antibiotic amoxicillin in the
Fenton-activated sludge combined system. Environ. Technol. 36, 844–851.
Guo, W.Q., Yin, R.L., Zhou, X.J., Du, J.S., Cao, H.O., Yang, S.S., Ren, N.Q., 2015b.
Sulfamethoxazole degradation by ultrasound/ozone oxidation process in water:
kinetics, mechanisms, and pathways. Ultrason. Sonochem. 22, 182–187.
Guo, Z., Fei, Z., Zhao, Y., Zhang, C., Liu, F., Bao, C., Lin, M., 2012. Gamma irradiationinduced sulfadiazine degradation and its removal mechanisms. Chem. Eng. J.
191, 256–262.
Haidar, M., Dirany, A., Sires, I., Oturan, N., Oturan, M.A., 2013. Electrochemical
degradation of the antibiotic sulfachloropyridazine by hydroxyl radicals
generated at a BDD anode. Chemosphere 91, 1304–1309.
Hassani, A., Karaca, M., Karaca, S., Khataee, A., Acisli, O., Yilmaz, B., 2018. Preparation
of magnetite nanoparticles by high-energy planetary ball mill and its
application for ciprofloxacin degradation through heterogeneous Fenton
process. J. Environ. Manage. 211, 53–62.
He, J., Yang, X.F., Men, B., Wang, D.S., 2016. Interfacial mechanisms of
heterogeneous Fenton reactions catalyzed by iron-based materials: a review.
J. Environ. Sci. 39, 97–109.
He, S.J., Wang, J.L., Ye, L.F., Zhang, Y.X., Yu, J., 2014. Removal of diclofenac from
surface water by electron beam irradiation combined with a biological aerated
filter. Radiat. Phys. Chem. 105, 104–108.
Hernandez, R., Zappi, M., Colucci, J., Jones, R., 2002. Comparing the performance of
various advanced oxidation processes for treatment of acetone contaminated
water. J. Hazard. Mater. 92, 33–50.
Hoigne, J., Bader, H., 1983. Rate constants of reactions of ozone with organic and
inorganic compounds in water – II: dissociating organic-compounds. Water Res.
17, 173–183.
Hollender, J., Zimmermann, S.G., Koepke, S., Krauss, M., McArdell, C.S., Ort, C., Singer,
H., von Gunten, U., Siegrist, H., 2009. Elimination of organic micropollutants in a
municipal wastewater treatment plant upgraded with a full-scale postozonation followed by sand filtration. Environ. Sci. Technol. 43, 7862–7869.
Hou, J., Chen, Z.Y., Gao, J., Xie, Y.L., Li, L.Y., Qin, S.Y., Wang, Q., Mao, D.Q., Luo, Y.,
2019. Simultaneous removal of antibiotics and antibiotic resistance genes from
pharmaceutical wastewater using the combinations of up-flow anaerobic
sludge bed, anoxic-oxic tank, and advanced oxidation technologies. Water
Res. 159, 511–520.
Hu, J., Wang, J.L., 2007. Degradation of chlorophenols in aqueous solution by
gamma-radiation. Radiat. Phys. Chem. 76, 1489–1492.
Huang, Y.H., Liu, Y., Du, P.P., Zeng, L.J., Mo, C.H., Li, Y.W., Lu, H.X., Cai, Q.Y., 2019.
Occurrence and distribution of antibiotics and antibiotic resistant genes in
water and sediments of urban rivers with black-odor water in Guangzhou,
South China. Sci. Total Environ. 670, 170–180.
Iakovides, I.C., Michael-Kordatou, I., Moreira, N.F.F., Ribeiro, A.R., Fernandes, T.,
Pereira, M.F.R., Nunes, O.C., Manaia, C.M., Silva, A.M.T., Fatta-Kassinos, D., 2019.
Continuous ozonation of urban wastewater: removal of antibiotics, antibioticresistant Escherichia coli and antibiotic resistance genes and phytotoxicity.
Water Res. 159, 333–347.
Ikehata, K., Naghashkar, N.J., Ei-Din, M.G., 2006. Degradation of aqueous
pharmaceuticals by ozonation and advanced oxidation processes: a review.
Ozone-Sci. Eng. 28, 353–414.
Jara, C.C., Fino, D., Specchia, V., Saracco, G., Spinelli, R., 2007. Electrochemical
removal of antibiotics from wastewaters. Appl. Catal. B-Environ. 70, 479–487.
Ji, B., Zhang, J.X., Zhang, C., Li, N., Zhao, T.T., Chen, F., Hu, L.H., Zhang, S.D., Wang, Z.Y.,
2018. Vertically aligned [email protected] nanorod chip with improved photocatalytic
activity for antibiotics degradation. ACS Appl. Nano Mater. 1, 793–799.
Jung, Y.J., Kim, W.G., Yoon, Y., Hwang, T.M., Kang, J.W., 2012. pH effect on ozonation
of ampicillin: kinetic study and toxicity assessment. Ozone-Sci. Eng. 34, 156–
162.
Kanakaraju, D., Glass, B.D., Oelgemoller, M., 2014. Titanium dioxide photocatalysis
for pharmaceutical wastewater treatment. Environ. Chem. Lett. 12, 27–47.
Kansal, S.K., Kundu, P., Sood, S., Lamba, R., Umar, A., Mehta, S.K., 2014.
Photocatalytic degradation of the antibiotic levofloxacin using highly
crystalline TiO2 nanoparticles. New J. Chem. 38, 3220–3226.
Kasprzyk-Hordern, B., Ziolek, M., Nawrocki, J., 2003. Catalytic ozonation and
methods of enhancing molecular ozone reactions in water treatment. Appl.
Catal. B-Environ. 46, 639–669.
Kaur, M., Umar, A., Mehta, S.K., Kansal, S.K., 2019. Reduced graphene oxide-CdS
heterostructure: An efficient fluorescent probe for the sensing of Ag(I) and
sunset yellow and a visible-light responsive photocatalyst for the degradation
of levofloxacin drug in aqueous phase. Appl. Catal. B-Environ. 245, 143–158.
Kirk, D.W., Sharifian, H., Foulkes, F.R., 1985. Anodic-oxidation of aniline for wastewater treatment. J. Appl. Electrochem. 15, 285–292.
Kitazono, Y., Ihara, I., Toyoda, K., Umetsu, K., 2017. Antibiotic removal from waste
milk by electrochemical process: degradation characteristics in concentrated
organic solution. J. Mater. Cycles Waste Manage. 19, 1261–1269.
Kong, D.Y., Liang, B., Yun, H., Cheng, H.Y., Ma, J.C., Cui, M.H., Wang, A.J., Ren, N.Q.,
2015. Cathodic degradation of antibiotics: characterization and pathway
analysis. Water Res. 72, 281–292.
Kornmuller, A., Wiesmann, U., 2003. Ozonation of polycyclic aromatic hydrocarbons
in oil/water-emulsions: mass transfer and reaction kinetics. Water Res. 37,
1023–1032.
Kuang, J.M., Huang, J., Wang, B., Cao, Q.M., Deng, S.B., Yu, G., 2013. Ozonation of
trimethoprim in aqueous solution: Identification of reaction products and their
toxicity. Water Res. 47, 2863–2872.
Kummerer, K., 2009. Antibiotics in the aquatic environment - a review - Part II.
Chemosphere 75, 435–441.
Kummerer, K., Al-Ahmad, A., Mersch-Sundermann, V., 2000. Biodegradability of
some antibiotics, elimination of the genotoxicity and affection of wastewater
bacteria in a simple test. Chemosphere 40, 701–710.
Lee, J., Farha, O.K., Roberts, J., Scheidt, K.A., Nguyen, S.T., Hupp, J.T., 2009. Metalorganic framework materials as catalysts. Chem. Soc. Rev. 38, 1450–1459.
Leon, D.E., Zuniga-Benitez, H., Penuela, G.A., Mansilla, H.D., 2017. Photocatalytic
removal of the antibiotic cefotaxime on TiO2 and ZnO suspensions under
simulated sunlight radiation. Water Air Soil Pollut. 228, 361.
Li, D., Guo, X.L., Song, H.R., Sun, T.Y., Wan, J.F., 2018. Preparation of RuO2-TiO2/Nanographite composite anode for electrochemical degradation of ceftriaxone
sodium. J. Hazard. Mater. 351, 250–259.
Li, M.F., Liu, Y.G., Zeng, G.M., Liu, N., Liu, S.B., 2019. Graphene and graphene-based
nanocomposites used for antibiotics removal in water treatment: a review.
Chemosphere 226, 360–380.
Liu, Q.Q., Shen, J.Y., Yang, X.F., Zhang, T.R., Tang, H., 2018a. 3D reduced graphene
oxide aerogel-mediated Z-scheme photocatalytic system for highly efficient
solar-driven water oxidation and removal of antibiotics. Appl. Catal. B-Environ.
232, 562–573.
Liu, S.Q., Feng, L.R., Xu, N., Chen, Z.G., Wang, X.M., 2012. Magnetic nickel ferrite as a
heterogeneous photo-Fenton catalyst for the degradation of rhodamine B in the
presence of oxalic acid. Chem. Eng. J. 203, 432–439.
Liu, X.L., Guo, Z., Zhou, L.B., Yang, J., Cao, H.B., Xiong, M., Xie, Y.B., Jia, G.R., 2019a.
Hierarchical biomimetic BiVO4 for the treatment of pharmaceutical wastewater
in visible-light photocatalytic ozonation. Chemosphere 222, 38–45.
Liu, X.N., Ji, H.D., Li, S., Liu, W., 2019b. Graphene modified anatase/titanate
nanosheets with enhanced photocatalytic activity for efficient degradation of
sulfamethazine under simulated solar light. Chemosphere 233, 198–206.
Liu, Y., Fan, Q., Wang, J.L., 2018b. Zn-Fe-CNTs catalytic in situ generation of H2O2 for
Fenton-like degradation of sulfamethoxazole. J. Hazard. Mater. 342, 166–176.
Liu, Y.B., Liu, H., Zhou, Z., Wang, T.R., Ong, C.N., Vecitis, C.D., 2015. Degradation of
the common aqueous antibiotic tetracycline using a carbon nanotube
electrochemical filter. Environ. Sci. Technol. 49, 7974–7980.
Liu, Y.J., Jiang, X.Z., 2005. Phenol degradation by a nonpulsed diaphragm glow
discharge in an aqueous solution. Environ. Sci. Technol. 39, 8512–8517.
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
Liu, Y.Q., He, X.X., Fu, Y.S., Dionysiou, D.D., 2016. Degradation kinetics and
mechanism of oxytetracycline by hydroxyl radical-based advanced oxidation
processes. Chem. Eng. J. 284, 1317–1327.
Lofrano, G., Carotenuto, M., Uyguner-Demirel, C.S., Vitagliano, A., Siciliano, A., Guida,
M., 2014. An integrated chemical and ecotoxicological assessment for the
photocatalytic degradation of vancomycin. Environ. Technol. 35, 1234–1242.
Lu, J., Sun, J.X., Chen, X.X., Tian, S.H., Chen, D.S., He, C., Xiong, Y., 2019. Efficient
mineralization of aqueous antibiotics by simultaneous catalytic ozonation and
photocatalysis using MgMnO3 as a bifunctional catalyst. Chem. Eng. J. 358, 48–
57.
Lyu, J.Z., Zhou, Z., Wang, Y.H., Li, J., Li, Q.Y., Zhang, Y.K., Ma, X.F., Guan, J.Y., Wei, X.,
2019. Platinum-enhanced amorphous TiO2-filled mesoporous TiO2 crystals for
the photocatalytic mineralization of tetracycline hydrochloride. J. Hazard.
Mater. 373, 278–284.
Ma, J., Yang, M.X., Yu, F., Chen, J.H., 2015. Easy solid-phase synthesis of pHinsensitive heterogeneous CNTs/FeS Fenton-like catalyst for the removal of
antibiotics from aqueous solution. J. Colloid Interface Sci. 444, 24–32.
Mackul’ak, T., Nagyova, K., Faberova, M., Grabic, R., Koba, O., Gal, M., Birosova, L.,
2015. Utilization of Fenton-like reaction for antibiotics and resistant bacteria
elimination in different parts of WWTP. Environ. Toxicol. Pharmacol. 40, 492–
497.
Malathi, A., Madhavan, J., Ashokkumar, M., Arunachalam, P., 2018. A review on
BiVO4 photocatalyst: activity enhancement methods for solar photocatalytic
applications. Appl. Catal. A-Gen. 555, 47–74.
Mamba, G., Kiwi, J., Pulgarin, C., Sanjines, R., Giannakis, S., Rtimi, S., 2018. Evidence
for the degradation of an emerging pollutant by a mechanism involving isoenergetic charge transfer under visible light. Appl. Catal. B-Environ. 233, 175–
183.
Manzetti, S., Ghisi, R., 2014. The environmental release and fate of antibiotics. Mar.
Pollut. Bull. 79, 7–15.
Marcelino, R.B.P., Leao, M.M.D., Lago, R.M., Amorim, C.C., 2017. Multistage ozone
and biological treatment system for real wastewater containing antibiotics. J.
Environ. Manage. 195, 110–116.
Martinez, F., Calleja, G., Melero, J.A., Molina, R., 2007. Iron species incorporated over
different silica supports for the heterogeneous photo-Fenton oxidation of
phenol. Appl. Catal. B-Environ. 70, 452–460.
Martinez-Huitle, C.A., Brillas, E., 2009. Decontamination of wastewaters containing
synthetic organic dyes by electrochemical methods: a general review. Appl.
Catal. B-Environ. 87, 105–145.
Martinez-Huitle, C.A., Ferro, S., 2006. Electrochemical oxidation of organic
pollutants for the wastewater treatment: direct and indirect processes. Chem.
Soc. Rev. 35, 1324–1340.
Mehrjouei, M., Muller, S., Moller, D., 2015. A review on photocatalytic ozonation
used for the treatment of water and wastewater. Chem. Eng. J. 263, 209–219.
Michael-Kordatou, I., Andreou, R., Iacovou, M., Frontistis, Z., Hapeshi, E., Michael, C.,
Fatta-Kassinos, D., 2017. On the capacity of ozonation to remove antimicrobial
compounds, resistant bacteria and toxicity from urban wastewater effluents. J.
Hazard. Mater. 323, 414–425.
Mirzaei, A., Yerushalmi, L., Chen, Z., Haghighat, F., Guo, J.B., 2018. Enhanced
photocatalytic degradation of sulfamethoxazole by zinc oxide photocatalyst in
the presence of fluoride ions: optimization of parameters and toxicological
evaluation. Water Res. 132, 241–251.
Miyata, M., Ihara, I., Yoshid, G., Toyod, K., Umetsu, K., 2011. Electrochemical
oxidation of tetracycline antibiotics using a Ti/IrO2 anode for wastewater
treatment of animal husbandry. Water Sci. Technol. 63, 456–461.
Moreira, F.C., Boaventura, R.A.R., Brillas, E., Vilar, V.J.P., 2017. Electrochemical
advanced oxidation processes: a review on their application to synthetic and
real wastewaters. Appl. Catal. B-Environ. 202, 217–261.
Moreira, F.C., Garcia-Segura, S., Boaventura, R.A.R., Brillas, E., Vilar, V.J.P., 2014.
Degradation of the antibiotic trimethoprim by electrochemical advanced
oxidation processes using a carbon-PTFE air-diffusion cathode and a borondoped diamond or platinum anode. Appl. Catal. B-Environ. 160, 492–505.
Nasseh, N., Taghavi, L., Barikbin, B., Nasseri, M.A., Allahresani, A., 2019. FeNi3/SiO2
magnetic nanocomposite as an efficient and recyclable heterogeneous fentonlike catalyst for the oxidation of metronidazole in neutral environments:
adsorption and degradation studies. Compos. Part B-Eng. 166, 328–340.
Nezamzadeh-Ejhieh, A., Shams-Ghahfarokhi, Z., 2013. Photodegradation of methyl
green by nickel-dimethylglyoxime/ZDM-5 zeolite as a heterogeneous catalyst. J.
Chem. 104093. https://doi.org/10.1155/2013/104093.
Nidheesh, P.V., 2015. Heterogeneous Fenton catalysts for the abatement of organic
pollutants from aqueous solution: A review. RSC Adv. 5, 40552–40577.
Niu, H.Y., Zhang, D., Zhang, S.X., Zhang, X.L., Meng, Z.F., Cai, Y.Q., 2011. Humic acid
coated Fe3O4 magnetic nanoparticles as highly efficient Fenton-like catalyst for
complete mineralization of sulfathiazole. J. Hazard. Mater. 90, 559–565.
Norte, T.H.O., Marcelino, R.B.P., Medeiros, F.H.A., Moreira, R.P.L., Amorim, C.C., Lago,
R.M., 2018. Ozone oxidation of beta-lactam antibiotic molecules and toxicity
decrease in aqueous solution and industrial wastewaters heavily contaminated.
Ozone-Sci. Eng. 40, 385–391.
Oh, J., Medriano, C.A., Kim, S., 2016. The effect of tetracycline in the antibiotic
resistance gene transfer before and after ozone disinfection. Desalinat. Water
Treat. 57, 646–650.
Oncu, N.B., Balcioglu, I.A., 2013. Degradation of ciprofloxacin and oxytetracycline
antibiotics in waste sewage sludge by ozonation. J. Adv. Oxid. Technol. 16, 107–
116.
Ostman, M., Bjorlenius, B., Fick, J., Tysklind, M., 2019. Effect of full-scale ozonation
and pilot-scale granular activated carbon on the removal of biocides,
13
antimycotics and antibiotics in a sewage treatment plant. Sci. Total Environ.
649, 1117–1123.
Ouyang, Q., Kou, F.Y., Tsang, P.E., Lian, J.T., Xian, J.Y., Fang, J.Z., Fang, Z.Q., 2019.
Green synthesis of Fe-based material using tea polyphenols and its application
as a heterogeneous Fenton-like catalyst for the degradation of lincomycin. J.
Clean. Prod. 232, 1492–1498.
Palominos, R., Freer, J., Mondaca, M.A., Mansilla, H.D., 2008. Evidence for hole
participation during the photocatalytic oxidation of the antibiotic flumequine. J.
Photochem. Photobiol. A-Chem. 193, 139–145.
Pan, Y.W., Zhang, Y., Zhou, M.H., Cai, J.J., Tian, Y.S., 2019. Enhanced removal of
antibiotics from secondary wastewater effluents by novel UV/pre-magnetized
Fe-0/H2O2 process. Water Res. 153, 144–159.
Park, H., Vecitis, C.D., Hoffmann, M.R., 2009. Electrochemical water splitting coupled
with organic compound oxidation: the role of active chlorine species. J. Phys.
Chem. C 113, 7935–7945.
Parmar, J., Villa, K., Vilela, D., Sanchez, S., 2017. Platinum-free cobalt ferrite based
micromotors for antibiotic removal. Appl. Mater. Today 9, 605–611.
Paucar, N.E., Kim, I., Tanaka, H., Sato, C., 2019. Ozone treatment process for the
removal of pharmaceuticals and personal care products in wastewater. OzoneSci. Eng. 41, 3–16.
Pham, V.L., Kim, D.G., Ko, S.O., 2018. [email protected] core-shell nanoparticle-catalyzed
oxidative degradation of the antibiotic oxytetracycline in pre-treated landfill
leachate. Chemosphere 191, 639–650.
Prado, N., Ochoa, J., Amrane, A., 2009. Biodegradation and biosorption of
tetracycline and tylosin antibiotics in activated sludge system. Process
Biochem. 44, 1302–1306.
Priya, B., Shandilya, P., Raizada, P., Thakur, P., Singh, N., Singh, P., 2016.
Photocatalytic mineralization and degradation kinetics of ampicillin and
oxytetracycline antibiotics using graphene sand composite and chitosan
supported BiOCl. J. Mol. Catal. A: Chem. 423, 400–413.
Qi, K.Z., Cheng, B., Yu, J.G., Ho, W.K., 2017. Review on the improvement of the
photocatalytic and antibacterial activities of ZnO. J. Alloys Compd. 727, 792–
820.
Qi, Y., Mei, Y.Q., Li, J.Q., Yao, T.J., Yang, Y., Jia, W.J., Tong, X., Wu, J., Xin, B.F., 2019.
Highly efficient microwave-assisted Fenton degradation of metacycline using
pine-needle-like CuCo2O4 nanocatalyst. Chem. Eng. J. 373, 1158–1167.
Rajkumar, D., Palanivelu, K., 2003. Electrochemical degradation of cresols for
wastewater treatment. Ind. Eng. Chem. Res. 42, 1833–1839.
Rao, N.N., Dube, S., 1996. Photoelectrochemical generation of hydrogen using
organic pollutants in water as sacrificial electron donors. Int. J. Hydrogen Energy
21, 95–98.
Rozas, O., Contreras, D., Mondaca, M.A., Perez-Moya, M., Mansilla, H.D., 2010.
Experimental design of Fenton and photo-Fenton reactions for the treatment of
ampicillin solutions. J. Hazard. Mater. 177, 1025–1030.
Saadati, F., Keramati, N., Ghazi, M.M., 2016. Influence of parameters on the
photocatalytic degradation of tetracycline in wastewater: a review. Crit. Rev.
Environ. Sci. Technol. 46, 757–782.
Sandikly, N., Kassir, M., El Jamal, M., Takache, H., Arnoux, P., Mokh, S., AlIskandarani, M., Roques-Carmes, T., 2019. Comparison of the toxicity of
waters containing initially sulfaquinoxaline after photocatalytic treatment by
TiO2 and polyaniline/TiO2. Environ. Technol. DOI: 10.1080/09593330.
2019.1630485.
Sanganyado, E., Gwenzi, W., 2019. Antibiotic resistance in drinking water systems:
Occurrence, removal, and human health risks. Sci. Total Environ. 669, 785–797.
Santos, L.V.D., Meireles, A.M., Lange, L.C., 2015. Degradation of antibiotics
norfloxacin by Fenton, UV and UV/H2O2. J. Environ. Manage. 154, 8–12.
Saputra, E., Muhammad, S., Sun, H.Q., Ang, H.M., Tade, M.O., Wang, S.B., 2013.
Different crystallographic one-dimensional MnO2 nanomaterials and their
superior performance in catalytic phenol degradation. Environ. Sci. Technol.
47, 5882–5887.
Sarkhosh, M., Sadani, M., Abtahi, M., Mohseni, S.M., Sheikhmohammadi, A.,
Azarpira, H., Najafpoor, A.A., Atafar, Z., Rezaei, S., Alli, R., Bay, A., 2019.
Enhancing photo-degradation of ciprofloxacin using simultaneous usage of e-aq
and OH over UV/ZnO/I- process: efficiency, kinetics, pathways, and
mechanisms. J. Hazard. Mater. 377, 418–426.
Sayed, M., Khan, J.A., Shah, L.A., Shah, N.S., Khan, H.M., Rehman, F., Khan, A.R., Khan,
A.M., 2016. Degradation of quinolone antibiotic, norfloxacin, in aqueous
solution using gamma-ray irradiation. Environ. Sci. Pollut. Res. 23, 13155–
13168.
Sekaran, G., Karthikeyan, S., Ramani, K., Ravindran, B., Gnanamani, A., Mandal, A.B.,
2011. Heterogeneous Fenton oxidation of dissolved organics in salt-laden
wastewater from leather industry without sludge production. Environ. Chem.
Lett. 9, 499–504.
Shandilya, P., Mittal, D., Soni, M., Raizada, P., Hosseini-Bandegharaeie, A., Saini, A.K.,
Singh, P., 2018. Fabrication of fluorine doped graphene and SmVO4 based
dispersed and adsorptive photocatalyst for abatement of phenolic compounds
from water and bacterial disinfectionAuthor links open overlay panel. J. Clean.
Prod. 203, 386–399.
Shandilya, P., Mittal, D., Sudhaik, A., Soni, M., Raizada, P., Saini, A.K., Singh, P., 2019.
GdVO4 modified fluorine doped graphene nanosheets as dispersed
photocatalyst for mitigation of phenolic compounds in aqueous environment
and bacterial disinfectionAuthor links open overlay panel. Sep. Purif. Technol.
210, 804–816.
Sheydaei, M., Shiadeh, H.R.K., Ayoubi-Feiz, B., Ezzati, R., 2018. Preparation of nano
N-TiO2/graphene oxide/titan grid sheets for visible light assisted photocatalytic
ozonation of cefixime. Chem. Eng. J. 353, 138–146.
14
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
Shooshtari, N.M., Ghazi, M.M., 2017. An investigation of the photocatalytic activity
of nano a-Fe2O3/ZnO on the photodegradation of cefixime trihydrate. Chem.
Eng. J. 315, 527–536.
Simond, O., Schaller, V., Comninellis, C., 1997. Theoretical model for the anodic
oxidation of organics on metal oxide electrodes. Electrochim. Acta 42, 2009–
2012.
Sivakumar, P., Lee, M., Kim, Y.S., Shim, M.S., 2018. Photo-triggered antibacterial and
anticancer activities of zinc oxide nanoparticles. J. Mater. Chem. B 6, 4852–
4871.
Soon, A.N., Hameed, B.H., 2011. Heterogeneous catalytic treatment of synthetic dyes
in aqueous media using Fenton and photo-assisted Fenton process. Desalination
269, 1–16.
Stockinger, H., Heinzle, E., Kut, O.M., 1995. Removal of chloro and nitro aromatic
waste-water pollutants by ozonation and biotreatment. Environ. Sci. Technol.
29, 2016–2022.
Sun, Z.H., Chang, H.X., 2014. Graphene and graphene-like two-dimensional
materials in photodetection: mechanisms and methodology. ACS Nano 8,
4133–4156.
Szekeres, E., Chiriac, C.M., Baricz, A., Szoke-Nagy, T., Lung, I., Soran, M.L., Rudi, K.,
Dragos, N., Coman, C., 2018. Investigating antibiotics, antibiotic resistance
genes, and microbial contaminants in groundwater in relation to the proximity
of urban areas. Environ. Pollut. 236, 734–744.
Tang, J.T., Wang, J.L., 2018a. Fenton-like degradation of sulfamethoxazole using Febased magnetic nanoparticles embedded into mesoporous carbon hybrid as an
efficient catalyst. Chem. Eng. J. 351, 1085–1094.
Tang, J.T., Wang, J.L., 2018b. Metal organic framework with coordinatively
unsaturated sites as efficient Fenton-like catalyst for enhanced degradation of
sulfamethazine. Environ. Sci. Technol. 52, 5367–5377.
Tegze, A., Sági, G., Kovács, K., Homlok, R., Tóth, T., Mohácsi-Farkas, C., Wojnárovits,
L., Takács, E., 2018. Degradation of fluoroquinolone antibiotics during ionizing
radiation treatment and assessment of antibacterial activity, toxicity and
biodegradability of the products. Radiat. Phys. Chem. 147, 101–105.
Tian, M., Wu, G.S., Adams, B., Wen, J.L., Chen, A.C., 2008. Kinetics of
photoelectrocatalytic degradation of nitrophenols on nanostructured TiO2
electrodes. J. Phys. Chem. C 112, 825–831.
Tian, X.K., Jin, H., Nie, Y.L., Zhou, Z.X., Yang, C., Li, Y., Wang, Y.X., 2017.
Heterogeneous Fenton-like degradation of ofloxacin over a wide pH range of
3.6-10.0 over modified mesoporous iron oxide. Chem. Eng. J. 328, 397–405.
Tilley, S.D., 2019. Recent advances and emerging trends in photo-electrochemical
solar energy conversion. Adv. Energy Mater. 9, 1802877.
Titouhi, H., Belgaied, J.E., 2016a. Heterogeneous Fenton oxidation of ofloxacin drug
by iron alginate support. Environ. Technol. 37, 2003–2015.
Titouhi, H., Belgaied, J.E., 2016b. Removal of ofloxacin antibiotic using
heterogeneous Fenton process over modified alginate beads. J. Environ. Sci.
45, 84–93.
Urbano, V.R., Maniero, M.G., Perez-Moya, M., Guimaraes, J.R., 2017. Influence of pH
and ozone dose on sulfaquinoxaline ozonation. J. Environ. Manage. 195, 224–231.
Uslu, M.O., Balcioglu, I.A., 2008. Ozonation of animal wastes containing
oxytetracycline. Ozone-Sci. Eng. 30, 290–299.
Vlyssides, A., Barampouti, E.M., Mai, S., Arapoglou, D., Kotronarou, A., 2004.
Degradation of methylparathion in aqueous solution by electrochemical
oxidation. Environ. Sci. Technol. 38, 6125–6131.
Wan, Z., Wang, J.L., 2016a. Ce-doped zero-valent iron nanoparticles as a Fenton-like
catalyst for degradation of sulfamethazine. RSC Adv. 6, 103523–103531.
Wan, Z., Wang, J.L., 2016b. Ce-Fe-reduced graphene oxide nanocomposite as an
efficient catalyst for sulfamethazine degradation in aqueous solution. Environ.
Sci. Pollut. Res. 23, 18542–18551.
Wan, Z., Wang, J.L., 2016c. Removal of sulfonamide antibiotics from wastewater by
gamma irradiation in presence of iron ions. Nucl. Sci. Tech. 27, 104.
Wang, J.L., Zhuang, S.T., 2017. Removal of various pollutants from water and
wastewater by modified chitosan adsorbents. Crit. Rev. Environ. Sci. Technol.
47, 2331–2386.
Wang, D.D., Li, J., Xu, Z.F., Zhu, Y.R., Chen, G.X., 2019a. Preparation of novel flowerlike BiVO4/Bi2Ti2O7/Fe3O4 for simultaneous removal of tetracycline and Cu2+:
adsorption and photocatalytic mechanisms. J. Colloid Interface Sci. 533, 344–
357.
Wang, H.H., Zhang, M., He, X.Z., Du, T.T., Wang, Y.Y., Li, Y., Hao, T.W., 2019b. Facile
prepared ball-like [email protected] composites for oxytetracycline removal under solar
and visible lights. Water Res. 160, 197–205.
Wang, J., Sun, W., Xu, C., Liu, W., 2012. Ozone degradation of chloramphenicol:
efficacy, products and toxicity. Int. J. Environ. Technol. Manage. 15, 180–192.
Wang, J.L., Bai, Z.Y., 2017. Fe-Based Catalysts for heterogeneous catalytic ozonation
of emerging contaminants in water and wastewater. Chem. Eng. J. 312, 79–98.
Wang, J.L., Chu, L.B., 2016. Irradiation treatment of pharmaceutical and personal
care products (PPCPs) in water and wastewater: an overview. Radiat. Phys.
Chem. 125, 56–64.
Wang, J.L., Wang, J.Z., 2007. Application of radiation technology to sewage sludge
processing: A review. J. Hazard. Mater. 143, 2–7.
Wang, J.L., Wang, S.Z., 2016. Removal of pharmaceuticals and personal care
products (PPCPs) from wastewater: a review. J. Environ. Manage. 182, 620–640.
Wang, J.L., Wang, S.Z., 2018a. Activation of persulfate (PS) and peroxymonosulfate
(PMS) and application for the degradation of emerging contaminants. Chem.
Eng. J. 334, 1502–1517.
Wang, J.L., Wang, S.Z., 2018b. Microbial degradation of sulfamethoxazole in the
environment. Appl. Microbiol. Biotechnol. 102, 3573–3582.
Wang, J.L., Wang, S.Z., 2019a. Preparation, modification and environmental
application of biochar: a review. J. Clean. Prod. 227, 1002–1022.
Wang, J.L., Xu, L.J., 2012. Advanced oxidation processes for wastewater treatment:
formation of hydroxyl radical and application. Crit. Rev. Environ. Sci. Technol.
42, 251–325.
Wang, J.L., Zhuan, R., Chu, L.B., 2019c. The occurrence, distribution and degradation
of antibiotics by ionizing radiation: an overview. Sci. Total Environ. 646, 1385–
1397.
Wang, J.L., Zhuang, S.T., 2019. Covalent organic frameworks (COFs) for
environmental applications. Coord. Chem. Rev. 400, 213046.
Wang, L., Ben, W.W., Li, Y.G., Liu, C., Qiang, Z.M., 2018. Behavior of tetracycline and
macrolide antibiotics in activated sludge process and their subsequent removal
during sludge reduction by ozone. Chemosphere 206, 184–191.
Wang, L., Yang, J., Li, Y.M., Lv, J., Zou, J.T., 2016a. Removal of chlorpheniramine in a
nanoscale zero-valent iron induced heterogeneous Fenton system: influencing
factors and degradation intermediates. Chem. Eng. J. 284, 1058–1067.
Wang, N.N., Zheng, T., Zhang, G.S., Wang, P., 2016b. A review on Fenton-like
processes for organic wastewater treatment. J. Environ. Chem. Eng. 4, 762–787.
Wang, S.Z., Wang, J.L., 2019b. Activation of peroxymonosulfate by sludge-derived
biochar for the degradation of triclosan in water and wastewater. Chem. Eng. J.
356, 350–358.
Wang, S.Z., Wang, J.L., 2019c. Oxidative removal of carbamazepine by
peroxymonosulfate (PMS) combined to ionizing radiation: degradation,
mineralization and biological toxicity. Sci. Total Environ. 658, 1367–1374.
Wang, S.Z., Wang, J.L., 2018c. Degradation of carbamazepine by radiation-induced
activation of peroxymonosulfate. Chem. Eng. J. 336, 595–601.
Wang, S.Z., Wang, J.L., 2018d. Radiation-induced degradation of sulfamethoxazole
in the presence of various inorganic anions. Chem. Eng. J. 351, 688–696.
Wang, S.Z., Wang, J.L., 2018e. Trimethoprim degradation by Fenton and Fe(II)activated persulfate processes. Chemosphere 191, 97–105.
Wang, Y., Gao, Y.W., Chen, L., Zhang, H., 2015. Goethite as an efficient heterogeneous
Fenton catalyst for the degradation of methyl orange. Catal. Today 252, 107–
112.
Wang, Y., Shen, C.C., Zhang, M.M., Zhang, B.T., Yu, Y.G., 2016c. The electrochemical
degradation of ciprofloxacin using a SnO2-Sb/Ti anode: influencing factors,
reaction pathways and energy demand. Chem. Eng. J. 296, 79–89.
Wen, J.Q., Li, X., Liu, W., Fang, Y.P., Xie, J., Xu, Y.H., 2015. Photocatalysis
fundamentals and surface modification of TiO2 nanomaterials. Chin. J. Catal.
36, 2049–2070.
Wu, G.L., Xiao, L.S., Gu, W., Shi, W.D., Jiang, D.Y., Liu, C.B., 2016. Fabrication and
excellent visible-light-driven photodegradation activity for antibiotics of SrTiO3
nanocube coated CdS microsphere heterojunctions. RSC Adv. 6, 19878–19886.
Wu, Z.C., Zhou, M.H., 2001. Partial degradation of phenol by advanced
electrochemical oxidation process. Environ. Sci. Technol. 35, 2698–2703.
Xu, H.Y., Qi, S.Y., Li, Y., Zhao, Y., Li, J.W., 2013. Heterogeneous Fenton-like
discoloration of Rhodamine B using natural schorl as catalyst: optimization
by response surface methodology. Environ. Sci. Pollut. Res. 20, 5764–5772.
Xu, L.J., Wang, J.L., 2011. A heterogeneous Fenton-like system with nanoparticulate
zero-valent iron for removal of 4-chloro-3-methyl phenol. J. Hazard. Mater. 186,
256–264.
Xu, L.J., Wang, J.L., 2012. Fenton-like degradation of 2,4-dichlorophenol using Fe3O4
magnetic nanoparticles. Appl. Catal. B-Environ. 123, 117–126.
Xue, J.J., Ma, S.S., Zhou, Y.M., Zhang, Z.W., He, M., 2015. Facile photochemical
synthesis of Au/Pt/g-C3N4 with plasmon-enhanced photocatalytic activity for
antibiotic degradation. ACS Appl. Mater. Interfaces 7, 9630–9637.
Yan, Y., Sun, S.F., Song, Y., Yan, X., Guan, W.S., Liu, X.L., Shi, W.D., 2013. Microwaveassisted in situ synthesis of reduced graphene oxide-BiVO4 composite
photocatalysts and their enhanced photocatalytic performance for the
degradation of ciprofloxacin. J. Hazard. Mater. 250, 106–114.
Yan, Y.B., Miao, J.W., Yang, Z.H., Xiao, F.X., Yang, H.B., Liu, B., Yang, Y.H., 2015. Carbon
nanotube catalysts: recent advances in synthesis, characterization and
applications. Chem. Soc. Rev. 44, 3295–3346.
Yargeau, V., Leclair, C., 2008. Impact of operating conditions on decomposition of
antibiotics during ozonation: a review. Ozone-Sci. Eng. 30, 175–188.
Ye, S.J., Zhou, X., Xu, Y.B., Lai, W.K., Yan, K., Huang, L., Ling, J.Y., Zheng, L., 2019.
Photocatalytic performance of multi-walled carbon nanotube/BiVO4
synthesized by electro-spinning process and its degradation mechanisms on
oxytetracycline. Chem. Eng. J. 373, 880–890.
Yi, H., Huang, D.L., Qin, L., Zeng, G.M., Lai, C., Cheng, M., Ye, S.J., Song, B., Ren, X.Y.,
Guo, X.Y., 2018. Selective prepared carbon nanomaterials for advanced
photocatalytic application in environmental pollutant treatment and
hydrogen production. Appl. Catal. B-Environ. 239, 408–424.
Yin, R.L., Guo, W.Q., Du, J.S., Zhou, X.J., Zheng, H.S., Wu, Q.L., Chang, J.S., Ren, N.Q.,
2017. Heteroatoms doped graphene for catalytic ozonation of sulfamethoxazole
by metal-free catalysis: performances and mechanisms. Chem. Eng. J. 317, 632–
639.
Yu, S.H., Lee, B.J., Lee, M.J., Cho, I.H., Chang, S.W., 2008. Decomposition and
mineralization of cefaclor by ionizing radiation: Kinetics and effects of the
radical scavengers. Chemosphere 71, 2106–2112.
Yu, S.Q., Hu, J., Wang, J.L., 2010a. Gamma radiation-induced degradation of pnitrophenol (PNP) in the presence of hydrogen peroxide (H2O2) in aqueous
solution. J. Hazard. Mater. 177, 1061–1067.
Yu, S.Q., Hu, J., Wang, J.L., 2010b. Radiation-induced catalytic degradation of pnitrophenol (PNP) in the presence of TiO2 nanoparticles. Radiat. Phys. Chem. 79,
1039–1046.
J. Wang, R. Zhuan / Science of the Total Environment 701 (2020) 135023
Yuan, A.L., Lei, H., Xi, F.N., Liu, J.Y., Qin, L.S., Chen, Z., Dong, X.P., 2019. Graphene
quantum dots decorated graphitic carbon nitride nanorods for photocatalytic
removal of antibiotics. J. Colloid Interface Sci. 548, 56–65.
Yuan, S.Y., Fan, Y., Zhang, Y.C., Tong, M., Liao, P., 2011. Pd-catalytic in situ generation
of H2O2 from H2 and O2 produced by water electrolysis for the efficient electrofenton degradation of Rhodamine B. Environ. Sci. Technol. 45, 8514–8520.
Zangeneh, H., Zinatizadeh, A.A.L., Habibi, M., Akia, M., Isa, M.H., 2015. Photocatalytic
oxidation of organic dyes and pollutants in wastewater using different modified
titanium dioxides: a comparative review. J. Ind. Eng. Chem. 26, 1–36.
Zeng, Y.Q., Chen, D.N., Chen, T.S., Cai, M.X., Zhang, Q.X., Xie, Z.J., Li, R.B., Xiao, Z.J., Liu,
G.G., Lv, W.Y., 2019. Study on heterogeneous photocatalytic ozonation
degradation of ciprofloxacin by TiO2/carbon dots: Kinetic, mechanism and
pathway investigation. Chemosphere 227, 198–206.
Zha, S.X., Cheng, Y., Gao, Y., Chen, Z.L., Megharaj, M., Naidu, R., 2014. Nanoscale zerovalent iron as a catalyst for heterogeneous Fenton oxidation of amoxicillin.
Chem. Eng. J. 255, 141–148.
Zhang, A.Y., Lin, T., He, Y.Y., Mou, Y.X., 2016a. Heterogeneous activation of H2O2 by
defect-engineered TiO2-X single crystals for refractory pollutants degradation: a
Fenton-like mechanism. J. Hazard. Mater. 311, 81–90.
Zhang, A.Y., Long, L.L., Liu, C., Li, W.W., Yu, H.Q., 2014. Electrochemical degradation
of refractory pollutants using TiO2 single crystals exposed by high-energy 001
facets. Water Res. 66, 273–282.
Zhang, N.Q., Chen, J.Y., Fang, Z.Q., Tsang, E.P., 2019. Ceria accelerated nanoscale
zerovalent iron assisted heterogenous Fenton oxidation of tetracycline. Chem.
Eng. J. 369, 588–599.
Zhang, X., Bai, B., Li Puma, G., Wang, H.L., Suo, Y.R., 2016b. Novel sea buckthorn
biocarbon [email protected] composites: Efficient removal of doxycycline in
aqueous solution in a fixed-bed through synergistic adsorption and
heterogeneous Fenton-like reaction. Chem. Eng. J. 284, 698–707.
Zhang, Y.S., Shao, Y.S., Gao, N.Y., Gao, Y.Q., Chu, W.H., Li, S., Wang, Y., Xu, S.X., 2018.
Kinetics and by-products formation of chloramphenicol (CAP) using
chlorination and photocatalytic oxidation. Chem. Eng. J. 333, 85–91.
Zhao, G.H., Cui, X., Liu, M.C., Li, P.Q., Zhang, Y.G., Cao, T.C., Li, H.X., Lei, Y.Z., Liu, L., Li,
D.M., 2009. Electrochemical degradation of refractory pollutant using a novel
microstructured TiO2 nanotubes/Sb-doped SnO2 electrode. Environ. Sci.
Technol. 43, 1480–1486.
15
Zhao, W.R., Wu, Z.B., Wang, D.H., 2006. Ozone direct oxidation kinetics of cationic
red X-GRL in aqueous solution. J. Hazard. Mater. 137, 1859–1865.
Zheng, C.M., Yang, C.W., Cheng, X.Z., Xu, S.C., Fan, Z.P., Wang, G.H., Wang, S.B., Guan,
X.F., Sun, X.H., 2017. Specifically enhancement of heterogeneous Fenton-like
degradation activities for ofloxacin with synergetic effects of bimetallic Fe-Cu
on ordered mesoporous silicon. Sep. Purif. Technol. 189, 357–365.
Zheng, P., Bai, B., Guan, W.S., Wang, H.L., Suo, Y.R., 2016. Degradation of tetracycline
hydrochloride by heterogeneous Fenton-like reaction using [email protected] subtilis.
RSC Adv. 6, 4101–4107.
Zhou, T., Li, Y.Z., Ji, J., Wong, F.S., Lu, X.H., 2008. Oxidation of 4-chlorophenol in a
heterogeneous zero valent iron/H2O2 Fenton-like system: kinetic, pathway and
effect factors. Sep. Purif. Technol. 62, 551–558.
Zhu, Z., Huo, P.W., Lu, Z.Y., Yan, Y.S., Liu, Z., Shi, W.D., Li, C.X., Dong, H.J., 2018.
Fabrication of magnetically recoverable photocatalysts using g-C3N4 for
effective separation of charge carriers through like-Z-scheme mechanism
with Fe3O4 mediator. Chem. Eng. J. 331, 615–625.
Zhuan, R., Wang, J.L., 2019a. Degradation of sulfamethoxazole by ionizing radiation:
kinetics and implications of additives. Sci. Total Environ. 668, 67–73.
Zhuan, R., Wang, J.L., 2019. Enhanced mineralization of sulfamethoxazole by
gamma radiation in the presence of Fe3O4 as Fenton-like catalyst, Sci. Pollut.
Res Environ. doi: 10.1007/s11356-019-05925-1.
Zhuan, R., Wang, J.L., 2020.a. Degradation of diclofenac in aqueous solution by
ionizing radiation in the presence of humic acid. Sep. Purif. Technol. 234,
116079.
Zhuan, R., Wang, J.L., 2020.b. Enhanced degradation and mineralization of
sulfamethoxazole by integrating gamma radiation with Fenton-like processes.
Radiat. Phys. Chem. 166, 108457.
Zhuang, S.T., Cheng, R., Wang, J.L., 2019a. Adsorption of diclofenac from aqueous
solution using UiO-66-type metal-organic frameworks. Chem. Eng. J. 359, 354–
362.
Zhuang, S.T., Liu, Y., Wang, J.L., 2019b. Mechanistic insight into the adsorption of
diclofenac by MIL-100: experiments and theoretical calculations. Environ.
Pollut. 253, 616–624.
Zhuang, S.T., Liu, Y., Wang, J.L., 2020.. Covalent organic frameworks as efficient
adsorbent for sulfamethazine removal from aqueous solution. J. Hazard. Mater.
383, 121126.
Téléchargement