Ciprofloxacin degradation in UV/chlorine advanced oxidation process: Influencing factors, mechanisms and degradation pathways

Article abstract of DOI:10.1016/j.jphotochem.2018.10.043

Ciprofloxacin (CIP) is a widely used third generation fluoroquinolone antibiotics, and has been often detected in wastewater treatment plants. Finding an effective way to remove them from wastewater is of great concern. Ultraviolet (UV)/chlorine advanced oxidation process (AOP) has many advantages in micropollutant removal. In this study, CIP degradation in UV/chlorine process was investigated. Only 41.2% of CIP was degraded by UV photolysis and 30.5% by dark chlorination in 30 min, while 98.5% of CIP was degraded by UV/chlorine process in 9 min. HCO₃ ^? had markedly inhibition, NO₃ ^? and SO₄ ^2- had slight inhibition, and Cl^? had a marginal inhibition on CIP degradation in UV/chlorine system. The degradation of CIP in UV/chlorine process was mainly attributed to the attack of reactive species. The relative contributions of hydrated electrons (e_aq ^[rad]), hydroxyl radicals (HO^[rad]), chlorine atoms (Cl^[rad]), and UV photolysis were investigated. Under neutral condition in aqueous solution, CIP degradation had highest pseudo first-order reaction rate constant, in which e_aq ^[rad] had the highest contribution, followed by Cl^[rad], HO^[rad], and UV photolysis. The intermediates and byproducts were identified and the degradation pathway was proposed. The total organic chlorine (TOCl) and biotoxicity were further assessed. CIP and natural organic matters (NOMs) were removed efficiently in real water. UV/chlorine showed the potential for the wastewater treatment containing CIP.

Full text of DOI:10.1016/j.jphotochem.2018.10.043

Accepted Manuscript
Title: Ciprofloxacin degradation in UV/chlorine advanced
oxidation process: Influencing factors, mechanisms and
degradation pathways
Authors: Jia Deng, Guangxue Wu, Shoujun Yuan, Xinmin
Zhan, Wei Wang, Zhen-Hu Hu
PII:
S1010-6030(18)31131-6
DOI:
Reference:
https://doi.org/10.1016/j.jphotochem.2018.10.043
JPC 11559
To appear in:
Journal of Photochemistry and Photobiology A: Chemistry
Received date:
Revised date:
Accepted date:
11 August 2018
22 October 2018
23 October 2018
Please cite this article as: Jia D, Guangxue W, Shoujun Y, Xinmin Z, Wei W, Zhen-Hu
H, Ciprofloxacin degradation in UV/chlorine advanced oxidation process: Influencing
factors, mechanisms and degradation pathways, Journal of Photochemistry and amp;
Photobiology, A: Chemistry (2018), https://doi.org/10.1016/j.jphotochem.2018.10.043
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Ciprofloxacin degradation in UV/chlorine advanced oxidation process: Influencing
factors, mechanisms and degradation pathways
a
b
a
c
a,
a,*
Jia Deng , Guangxue Wu , Shoujun Yuan , Xinmin Zhan , Wei Wang *, Zhen-Hu Hu
a
School of Civil Engineering, Hefei University of Technology, Hefei 230009, China
b
Key Laboratory of Microorganism Application and Risk Control (MARC) of
Shenzhen, Graduate School at Shenzhen, Tsinghua University, Shenzhen 518055, China
c
College of Engineering and Informatics, National University of Ireland, Galway,
Ireland
*Corresponding author
Prof. Zhen-Hu Hu, Dr. Wei Wang
School of Civil Engineering,
Hefei University of Technology,
Hefei, 230009
China
Tel:86-551-62904144
Fax: 86-551-62902066
E-mail: zhhu@hfut.edu.cn; dwhit@126.com
Graphical Abstract
1

Highlights
Ciprofloxacin (CIP) is quickly degraded in UV/chlorine process
CIP degradation in UV/chlorine process is affected by many anions and humic acid
.
.
Hydrated electron, Cl , HO , and UV photolysis play key role in CIP degradation
Degradation pathway of CIP in UV/chlorine process is proposed
Abstract
Ciprofloxacin (CIP) is a widely used third generation fluoroquinolone antibiotics,
and has been often detected in wastewater treatment plants. Finding an effective way to
remove them from wastewater is of great concern. Ultraviolet (UV)/chlorine advanced
oxidation process (AOP) has many advantages in micropollutant removal. In this study,
CIP degradation in UV/chlorine process was investigated. Only 41.2% of CIP was
degraded by UV photolysis and 30.5% by dark chlorination in 30 min, while 98.5% of
CIP was degraded by UV/chlorine process in 9 min. HCO -
had markedly inhibition,
3
-
2-
-
NO
3
and SO
4
had slight inhibition, and Cl had a marginal inhibition on the CIP
2

degradation in UV/chlorine system. The degradation of CIP in UV/chlorine process was
mainly attributed to the attack of reactive species. The relative contributions of hydrated
.
.
.
electrons (eaq ), hydroxyl radicals (HO ), chlorine atoms (Cl ), and UV photolysis were
investigated. Under neutral condition in aqueous solution, CIP degradation had highest
.
pseudo first-order reaction rate constants, in which eaq had the highest contribution,
.
.
followed by Cl , HO , and UV photolysis. The intermediates and byproducts were
identified and the degradation pathway was proposed. The total organic chlorine (TOCl)
and biotoxicity were further assessed. CIP and natural organic matters (NOMs) were
removed efficiently in real water. UV/chlorine showed the potential for the wastewater
treatment containing CIP.
Keywords: Ciprofloxacin; UV/chlorine process; advance oxidation process; reactive
species; UV photolysis; wastewater
3

1
. Introduction
Fluoroquinolones are typical antibiotics that are widely used around the world as
human and veterinary medicines. They have been frequently detected in surface water
and wastewater over the last few decades [1]. But, due to the persistence and biotoxicity
of fluoroquinolones, these antibiotics could not be well degraded in conventional water
treatment processes [2]. Ciprofloxacin (CIP) is a widely used third generation
fluoroquinolone, which has been often detected in wastewater treatment plants effluents
(WWTPE) water, surface water, groundwater and soils [3,4]. The presence of low
concentration of CIP may induce selective pressure for microbial populations leading to
fluoroquinolones resistant pathogens [5,6]. Thus, finding an effective way to remove
them from wastewater is of great concern.
Advanced oxidation processes (AOPs) are efficient for the degradation of
recalcitrant pharmaceuticals and personal care products (PPCPs) [7]. Many AOPs
technologies have been conducted for the destruction of fluoroquinolones, such as UV
photolysis, UV/H
2
O
2
and ozonation [8-10]. Aqueous chlorine oxidation and
electrochemical oxidation were also investigated [11,12].
UV and chlorine are widely used for the disinfection in water treatment plants [13].
UV/chlorine AOP has been reported to produce hydrated electrons (eaq.), hydroxyl
.
.
radicals (HO ), chlorine atoms (Cl ) and other oxidation radicals at wavelengths less than
.
.-
.
4
00 nm [14,15]. HO , O and Cl are generated due to the UV photolysis of HOCl and
−
OCl , as listed in Eq. (1).


HClO/ClO hv HO /O Cl
(1)
.
.
HO is a nonselective oxidant with oxidation potential approximately 2.70 V [16]. Cl is
a selective oxidant and can react with electron-rich moieties through one-electron
4

oxidation, H-abstraction and unsaturated carbon-carbon bonds with oxidation potential
.47 V [17]. The reactive species have synergism effect and enhance the degradation of
2
persistent contaminants such as chloramphenicol, ibuprofen, benzotriazole, tolytriazole,
iopamidole, ronidazole and atrazine [18-22]. However, the degradation of CIP in
UV/chlorine process and the relative contributions of reactive species have not been
reported before.
The aims of this study were to (1) compare the degradation of CIP by UV, dark
chlorine and UV/chlorine; (2) analyze the factors affecting CIP degradation in
UV/chlorine process; (3) investigate the relative contributions of the reactive species on
CIP degradation; and (4) propose CIP degradation pathways in UV/chlorine process.
2
. Materials and Methods
2
.1. Materials
Ciprofloxacin (purity > 98.0%) was purchased from TCI (Shanghai) Development
Co. Ltd., China. Sodium hypochlorite (NaClO, 6.0-14.0% - free chlorine w/w) and
tert-butanol (t-BuOH) were purchased from Aladdin Industrial Co. Ltd., China. Sodium
thiosulfate, sodium bicarbonate, potassium phosphate monobasic and sodium hydroxide
were purchased from Sinopharm Chemical Reagent Co. Ltd., China. Methanol,
acetonitrile, and formic acid were purchased from Sigma-Aldrich (Shanghai) Co. LLC.,
China. All chemicals were of analytical grade and used without further purification.
Aqueous solutions were prepared using ultrapure water and all reagents were stored in
the dark at 4.0 °C. The NaClO stock solution was prepared weekly and standardized
[
23].
5

The surface water was collected from Hubing pond in the campus of our university
Hefei, China). The waste water treatment plant effluent (WWTPE) water was collected
(
from Zhuzhuanjing municipal wastewater treatment plant, Hefei, China. The
characteristics of the surface water and WWTPE water were listed in Table 1.
2
.2. Experimental design
The CIP degradation by UV, dark chlorination and the UV/chlorine were
performed in a light-proof tank. The schematic diagram of the apparatus was shown in
Fig. S1. A low pressure UV lamp (16-20 W, ZW40S24Y, Sanyi Co., China) with
monochromatic emissions at 254 nm was installed at the top of tank. Two 300 mL glass
beakers were used as the reactor with working volume of 250 mL. Magnetic stirrer was
used for the mixing. The average UV fluence rate was determined with 0.4 mW.cm-2
[
24]. The UV lamp was warmed up for 30 min before experiments. Control test of CIP
degradation by UV photolysis or dark chlorination was also conducted in a similar way
but only by UV light or chlorine.
In the investigation of the impacts of environmental conditions (e.g., anions, humic
acid, and pH) and application in surface water and WWTPE water, the CIP
concentration was set at 10.0 μM, and 20.0 mM phosphate buffer was used to adjust the
solution pH. Samples (1.0 mL) were collected at different time intervals, and
immediately quenched with 10.0 mM Na
2
S
2
3
O . To identify the intermediate products
and analyze total organic carbon (TOC) and total organic chlorine (TOCl), 100.0 μM of
CIP and 1.0 mM of chlorine were used and the reaction time was extended to 240 min.
To assess the toxicity of intermediates and byproducts, biotoxicity analysis experiments
were conducted in 240 min and samples in different reaction time were collected and
6

quenched by ascorbic acid. All experiments were carried out in triplicate and the
average values with error bar were presented.
2
.3. Analytical Methods
The concentration of CIP in the solution was measured using high-performance
liquid chromatography (HPLC, 1260 Infinity, Agilent, USA) equipped with a C18
column (4.6 mm × 250 mm, 5 μm, Shimadzu Ltd., Japan) and detected by a diode array
detector at 254 nm. The mobile phase was composed of acetonitrile, methanol, water
-1
and formic acid (20: 15: 64.5: 0.5, v/v/v/v) at a flow rate of 0.8 mL.min . The ions in
the degradation solution were analyzed using an ion chromatograph equipped with a
conductivity detector (ICS-90, Dionex, USA). An anion-exchange column (AS9-HC)
was used to separate the ions with an eluent of 9.0 mM Na
2
CO
mL.min . TOC of the solution was measured by total organic carbon analyzer
TOC-VCPN, Shimadzu Ltd., Japan). The TOCl was determined using a precombustion
3
at a flow rate of 1.0
-1
(
method (USEPA, 1988) with an ion chromatograph (IC) system.
The decomposition intermediates of CIP in UV/chlorine process were analyzed
using liquid chromatography/electrospray ionization quadrupole time-of-flight mass
spectrometry (LC/Q-TOFMS, Acquity UP, Waters, USA). A C18 column (4.6 mm × 150
mm, 5 μm, GL Sciences Inc., Japan) was used to separate the relevant components at a
flow rate of 0.4 mL/min. The mobile phase was composed of acetonitrile (solvent A),
methanol (solvent B) and water (solvent C) (20: 20: 60, v/v/v). The scanning mode was
set to m/z 100–1000. The temperature of the desolvation and the ion source block was
-1
set to 350 °C and 110 °C, respectively. Nitrogen was used as the nebulizer (40.0 L.h )
-1
and as the desolvation gas (400.0 L.h ). Byproducts were identified using gas
7

chromatography mass spectrometry (GC-MS, Elite-1, Perkin Elmer, USA) and
equipped with a HP-5MS fused silica capillary column (30 m × 0.25 mm, 0.25 μm,
Perkin Elmer, USA). The details of pretreatment procedure and parameters for
byproducts analysis were shown in Text S1.
The biotoxicity was evaluated using Salmonella typhimurium (Sal94) and
luminescent bacteria photobacterium phosphoreum (Synergy HT, BioTek Ltd., USA).
The details of operation procedure for biotoxicity analysis were shown in Text S2.
NOMs in solution were detected using a three-dimensional excitation-emission matrix
spectrofluorometer (Varian, Palo Alto Co., USA), which used a Xe excitation source.
The parameters of excitation-emission matrix for NOMs analysis were shown in Text
S3.
3
. Results and Discussion
3
.1. CIP degradation in UV, Dark Chlorine and UV/chlorine processes
Fig. 1 showed the CIP degradation by UV photolysis, dark chlorination and
UV/chlorine at the same oxidant molar dosage and pH 7.0. Only 41.2% of CIP was
degraded by UV photolysis and 30.5% by dark chlorination in 30 min. It was interesting
that 98.5% of CIP was degraded by UV/chlorine in 9 min, indicating the strong
oxidative capacity of UV/chlorine process.
3
3
.2. CIP degradation in UV/chlorine process
.2.1. Effects of anions
8

-
-
-, and SO
2-
Generally, HCO
3
, Cl , NO
3
4
widely exist in wastewater and may affect
the oxidation process. Thus, the effects of these ions on CIP degradation in UV/chlorine
process were investigated. Principal reactions (2-16) in UV/chlorine system were listed
in Table 2. Fig. 2a showed that the degradation was obviously inhibited when HCO -
concentration was more than 5.0 mM. It has been reported that HCO -
3
3
could scavenge
.
-
.
.
HO based on reactions 2 and 3, as listed in Table 2. HCO
3
could react with HO and Cl
-.
-
.
to generate CO
3
, which had a low oxidation potential [35]. Cl could react with HO
.
and Cl to form the weak oxidizing capability radicals. Meanwhile, these weak oxidizing
.
.
capability radicals could be decomposed into HO and Cl again with a high rate constant,
-
as listed in reactions 4-7 in Table 2. So, the presence of Cl had a marginal effect on CIP
degradation, as shown in Fig. 2b. Fig. 2c showed the degradation of CIP was also
inhibited by NO
3
-. A possible explanation was that NO -
3
was reduced to NO -
which
2
.
consumed HO to form low oxidizing capability radicals, as listed in reactions 8-10 in
Table 2, which resulted in the inhibition effect. Few studies reported SO 2-
could impact
4
.
.
the HO and Cl . But, Fig. 2d showed CIP degradation was slightly inhibited in the
2
-
which was similar to NO -
and HCO -
. So, high concentration of
presence of SO
4
3
3
SO 2
-
might influence the radicals, too.
4
3
.2.2. Effects of humic acid
Natural organic matters (NOMs) widely exist in natural water bodies, and humic
acid is a typical NOM. Fig. 3 showed the influence of humic acid on CIP degradation.
-1
0
.4 mg.L of humic acid showed significant inhibition on the degradation, and the
degradation efficiency decreased from 98.5% to 62.2%. With the continuous increase of
9

-
1
humic acid to 4.0 mg.L , the degradation efficiency further decreased to 43.2%. The
inhibition might be due to the following reasons. Firstly, NOMs absorbed UV light at
2
54 nm and underwent photolysis which reduced the UV absorption of chlorine and CIP.
So, NOMs acted as an inner filter to impede CIP photodegradation and decrease the rate
at which chlorine photolysis produced radicals. Secondly, in the same system, NOMs
might compete radicals with CIP which decreased the degradation efficiency of CIP [9].
Therefore, NOMs had a negative effect on the organic degradation in UV/chlorine
process.
3
.3. Kinetics analysis and relative contribution of reactive species
Kinetics analysis could describe the reaction process of CIP degradation. In this
study, a pseudo-first-order reaction kinetics model was used to describe the CIP
degradation by UV/chlorine in different pH conditions as Eq. (17):

ln(C/C0 )
k 
(17)
t
where C and C are the CIP concentration at initial time and time t, and k is the
0
pseudo-first-order rate constants. The values of k in different pH conditions were
-3
-1
-3 -1
-3
showed in the inset of Fig. 4a. These values were 3.3 × 10 s , 3.8 × 10 s , 7.4 × 10
-1
-3 -1
-3 -1
s , 7.1 × 10 s and 6.7 × 10 s at pH 5.0, 6.0, 7.0, 8.0 and 9.0, respectively. In this
work, CIP degradation showed higher degradation efficiency in neutral and alkaline
conditions than acidic conditions. It might be attributed to the relative contributions of
reactive species in different pH conditions.
Actually, CIP degradation in UV/chlorine process was effected by many factors,
including UV photolysis and the attack of reactive species. The reactive species
10

included Cl , HO , eaq., Cl
.
.
-.
, O , ClO and ClOH . Due to the low oxidation-reduction
-.
.
-.
2
potential, the reactive species of Cl -
.
, O , ClO and ClOH were negligible [17]. Taking
-.
.
-.
2
the contribution of main reactive species into consideration, the CIP degradation
kinetics could be described using Eq. (18):

ln(C/C )  k t  k
C t  k
C t  k
C t
(18)
0
UV
CIP-HO
HO
CIP-eaq
e
CIP-Cl
Cl
aq
where kUV represents the first-order reaction rate constant of CIP degradation by UV
photolysis. kCIP-HO., kCIP-eaq. and kCIP-Cl. are the second-order reaction rate constants of
.
.
.
CIP degradation by the attack of HO , eaq and Cl , respectively. CHO., Ceaq. and CCl. are
.
.
.
the steady state concentration of HO , eaq and Cl , respectively.
UV was obtained experimentally from CIP degradation by UV photolysis, and was
calculated using Eq. (19):
k
'
'
0

ln(C /C )
kUV
where C’
experiment. Fig. S2 showed the value of kUV in different pH conditions in UV photolysis.

(19)
t
0
and C’ are the CIP concentration at initial time and time t in UV photolysis
kUV obtained in this study was similar to the values which had been reported before [9].
.
The relative contribution of HO was investigated using 10.0 μM nitrobenzene (NB)
.
as the probe compound in UV/chlorine system which only reacted with HO . The
reaction rate constant of NB with HO (kNB-HO.) had been reported (kNB-HO = 3.9 × 109
.
.
-1 -1
M s ) [36]. According to the steady-state assumption, CHO. was calculated using Eq.
(20):

ln(C /C
)
NB0
NB
CHO

(20)
(21)
kNB-HO
t
'
k  k
CHO
1
CIP-HO
where CNB0 and CNB are the concentration of NB at initial time and time t. The result of
11

NB degradation was shown in Fig. S3a. The reaction rate constant of CIP with HO.
CIP-HO.) has been reported (kCIP-HO = 6.2 × 10 M s ) [9]. k
’ was defined as the
.
9
-1 -1
(k
1
.
pseudo-first order rate constant of CIP with HO and the value was obtained in Eq. (21)
-4
-1
-4 -1
-4 -1
as 4.1 × 10 s , 5.4 × 10 s and 5.3 × 10 s at pH 5.0, 7.0 and 9.0, respectively.
To investigate the relative contribution of eaq., excessive tertiary butyl alcohol
(t-BuOH) was added in UV/chlorine system as radicals quencher because of its high
.
.
.
8
-1 -1
.
reaction rate constants with HO and Cl (kt-BuOH-HO = 6.0 × 10 M s , and kt-BuOH-Cl
=
8
-1 -1
.
3
.0 × 10 M s ) [37]. In this system, 100.0 mM t-BuOH was added to scavenge HO
.
.
and Cl Thus, eaq could be remained to react with CIP, and the kinetics was listed in Eq.
22):
(
'
''

ln(C /C )
'
2
k  k
Ceaq
CIP-eaq

(22)
t
where C’’ and C’’ are the concentration of CIP at initial time and time t in t-BuOH
0
scavenging experiment. The CIP degradation kinetics was shown in Fig. S3b. k
2
’ was
defined as the pseudo-first-order rate constant of CIP with eaq.. These values were 1.0 ×
-3
-1
-3 -1
-3 -1
1
0 s , 2.0 × 10 s and 2.9 × 10 s at pH 5.0, 7.0 and 9.0, respectively. Therefore, the
.
only unknown value was the relative contribution of Cl , which could be described in Eq.
(23):
'
3
k  k
C  k (k  k
CHO  kCIP-eaq C )
(23)
CIP-Cl
Cl
UV
CIP-HO
e
aq
.
where k
3
’ was defined as the pseudo-first order rate constant of CIP with Cl . Eq.7 can
be simplified into Eq. (24):
'
3
'
1
'
2
k  k (k  k  k )
(24)
’ was calculated as 1.7 × 10 s , 1.8 × 10 s and 2.3 × 10 s at pH 5.0, 7.0 and 9.0,
respectively.
Fig. 4b showed the relative contributions of UV, HO , Cl and eaq in UV/chlorine
UV
-4
-1
-3 -1
-5 -1
k
3
.
.
.
12

process at different pH conditions. It could be found that the contribution of HO.
remained relatively stable. With the increase of pH value, contributions of UV and eaq.
.
increased. The contribution of Cl was extremely low in non-neutral conditions. Under
+
.
.
.
-
acid condition, H combined Cl and eaq to form low reduction H and Cl according to
-
.
.-
the reaction 15. The reaction of OH with Cl to form the low oxidation ClOH based on
reaction 16, as listed in Table 2. So, comparing with acidic condition, CIP degradation
was more quickly in neutral and alkaline conditions.
3
.4. CIP degradation pathway in UV/chlorine process
The degradation of CIP could be evaluated by the generation of anions during
AOPs. As shown in Table S1, CIP contains carbon-fluorine bond. Thus, the
-
concentration of F was measured during the degradation to evaluate the broken of CIP
structure. Because the energy of carbon-fluorine bond is much higher than
carbon-carbon bond and carbon-nitrogen bond, the carbon-fluorine bond of CIP is very
difficult to be broken. Some reactive species with high oxidation reduction potential
.
such as sulfate radicals (E
0
= 2.5 V ~ 3.1 V), HO (E
0
= 2.7 V) and eaq. (E = -2.9 V)
0
could attack carbon-fluorine bonds [38]. Fig. 5a showed the defluorination in three
systems. Obviously, dark chlorine was invalid for the defluorination, indicating that free
chlorine could not break carbon-fluorine bond down. UV photolysis had a
defluorination efficiency of 40.2% and 68.8% for UV/chlorine defluorination.
UV/chlorine showed a higher defluorination efficiency when compared to UV
photolysis, indicating the advantages of UV/chlorine process.
The mineralization of CIP in UV/chlorine process was assessed using TOC. As
shown in Fig. 5.b, 22.3% of TOC was removed after 4 hours reaction in UV/chlorine
13

process, while there was no obvious decline in dark chlorine and UV photolysis. The
degradation intermediates and byproducts formed during AOPs might bring potential
hazard to the environment. Thus, it was essential to identify the molecular structures of
these intermediates and byproducts. The intermediates were detected via
LC/Q-TOF-MS analysis. Parts of intermediate products structure could be identified,
such as P1 (m/z=332), P2 (m/z=306), P3 (m/z=263) and P4 (m/z=223). In addition,
some detected compounds could be proposed such as a (m/z=320) and b (m/z=320).
The MS spectrum was shown in Fig. S5. P1 (m/z=332) was detected at initial time. Its
secondary amines on the piperazyl ring were attacked firstly. Subsequently,
Carbon-nitrogen bonds were broken down. Then, there were two possible pathways.
The carbon-fluorine bond was broken down and substituted by hydroxyl. In another
pathway, the piperazinyl ring was dropped out and the whole cyclopropyl was dropped
from benzene ring. Some byproducts such as dichloroacetonitrile (DCAN) and
trichloroacetaldehyde (chloral) were detected in CIP degradation process. The MS
spectrum of byproducts was shown in Fig. S6. These intermediates and byproducts
-
could be further degraded into CO
2
, H
2
O, F and other low molecular compounds. The
proposed degradation pathway was shown in Fig. 6.
3
.5. TOCl and biotoxicity analysis
For assessing the potential risk of intermediates, the generation of TOCl during
CIP degradation in dark chlorine and UV/chlorine process was investigated, as shown in
Fig. 7. In dark chlorine process, 130 μM of TOCl was formed at initial 10 minutes, then
decreased to around 70 μM until the end of the process. In UV/chlorine process, the
TOCl concentration continuously increased from initial 0 to about 90 μM at 120 min,
14

and then decreased to 30 μM after 240 min. Initially, dark chlorine and UV/chlorine all
had chlorine substitution process which resulted in TOCl increasing quickly. Comparing
with dark chlorine, UV/chlorine process had a low rate of chlorine substitution before
1
0 min. For dark chlorine, substitution was attributed to chlorine direct oxidation.
However, for UV/chlorine, substitution was attributed to the effect of radicals which
lead to the constant reaction of chlorine substitution. After 120 min, chlorine
substitution rate was lower than decomposing rate which resulted in the decrease of
TOCl.
The biotoxicity of the CIP degradation solution was evaluated. Fig. S7 showed the
change of biototoxicity during CIP degradation in UV/chlorine process. At the initial
time, the inhibition of CIP leaded to the very low bioluminescent. The value of
bioluminescence reached highest at 60 min and decreased subsequently. It represented
the biotoxicity was the lowest on 60 min. After 60 min, with the reaction continued,
byproducts were generated such as DCAN and chloral. So, controlling the reaction time
in UV/chlorine process is very important.
3
.6. CIP degradation in surface water and WWTPE
Fig. 8 showed UV/chlorine process degraded CIP in surface water and WWTPE.
After 9 min, CIP degradation efficiency reached 80.3% and 72.7% in surface water and
WWTPE water, respectively. The values were lower than that in ultrapure water.
According to the investigation results of anions concentration in Table 1, mixing water
-1
+
-1
-
-1
-
-1
2-
containing 0.5 mg.L NH
4
, 7.0 mg.L NO
3
, 30.0 mg.L Cl and 68.0 mg.L SO
4
was used to assess the effect of ions on UV/chlorine process. Interestingly, Fig. 8b
showed a negligible effect of ions on CIP degradation in mixing water. The result might
15

be attributed to the NOMs in real water which consumed the reactive species in solution.
From three-dimensional fluorescence spectra, humic acid-like organics as typical NOMs
were detected in surface water and WWTPE [39]. Fig. 8 c-f showed NOMs could be
efficiently removed in UV/chlorine process. Although, the degradation of CIP in real
water was lower than it in ultrapure water, but the rate constants still maintained a high
value due to the synergism of multiple reactive species [40]. These results showed the
potential of UV/chlorine process for the wastewater treatment containing CIP.
4
. Conclusions
In this study, the degradation of CIP in aqueous solution by UV/chlorine AOP was
investigated. Only 41.2% of CIP was degraded by UV photolysis and 30.5% by dark
chlorination in 30 min, and 98.5% of CIP was degraded by UV/chlorine AOP in 9 min.
-
-
The effect of HCO
3
on CIP degradation in UV/chlorine process depended on HCO
3
concentration, while of NO -
and SO
2-
was slight inhibitory, but of Cl wasn’t obvious.
-
3
4
The degradation of CIP in UV/chlorine process was mainly contributed to the attack of
reactive oxygen species and UV photolysis. Under neutral condition in aqueous solution,
.
CIP degradation had highest first-order reaction rate constants, in which eaq had the
.
.
highest contribution, followed by Cl , HO , and UV photolysis. Based on the
intermediates and byproducts identified, the degradation pathway of CIP in UV/chlorine
process was proposed. The biotoxicity of intermediates in CIP degradation process was
lowest at about 60 min. CIP could be degraded efficiently by UV/chlorine in real water.
The results showed the potential of UV/chlorine process for the wastewater treatment
containing CIP.
16

Acknowledgements
This research was partially supported by the National Natural Science Foundation
of China (51578205, 51538012, and 51728801), and the Project of Science and
Technology in Anhui Province (1501041130).
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19

Figure captions
Fig. 1. CIP degradation in ultrapure water by UV, dark chlorination and UV/chlorine.
Experimental conditions: pH = 7.0, [CIP]
0
= 10 μM, [NaOCl] = 100 μM, and UV
0
-2
fluence rate = 0.4 mW.cm .
Fig. 2. Effect of anions on CIP degradation in UV/chlorine process. Influence of (a)
-
-
-
HCO
-10.0 mM), and (d) SO
CIP] = 10.0 μM, [NaOCl]
Fig. 3. Effect of humic acid on CIP degradation in UV/chlorine process. Experimental
conditions: [humic acid] = 0-4.0 mg/L, [CIP] = 10.0 μM, [NaOCl] = 100.0 μM, and
UV fluence rate = 0.4 mW.cm .
3
([NaHCO
3
] = 0-10.0 mM), (b) Cl ([NaCl] = 0-10.0 mM), (c) NO
3
([NaNO
3
] =
2
-
0
4
([Na SO
2 4
] = 0-10.0 mM). Experimental conditions: pH = 7.0,
-2
[
0
0
= 100.0 μM, and UV fluence rate = 0.4 mW.cm .
0
0
0
-2
Fig. 4. Kinetic analysis and relative contribution of reactive species in CIP degradation
in UV/chlorine process. (a) Effect of pH and reaction rate constants analysis, and (b) the
.
.
relative contribution of HO , Cl , hydrated electrons and UV in the UV/chlorine process.
Experimental conditions: pH = 5.0 - 9.0, [CIP] = 10.0 μM, [chlorine] = 100.0 μM, and
0
0
-2
UV fluence rate = 0.4 mW.cm .
-
-, and TOC in time-dependent CIP degradation by
Fig. 5. The generation of F , NO
3
-
-, and (c) TOC. Experimental
different ways. (a) The generation of F , (b) NO
3
conditions: [CIP]
0
= 100.0 μM, [NaOCl] = 1.0 mM, pH = 7.0, and UV fluence rate =
0
-2
0
.4 mW.cm .
Fig. 6. Proposed pathways of CIP decomposition in UV/chlorine process.
Fig. 7. The generation of total organic chlorine during CIP degradation by dark chlorine
and UV/chlorine. Experimental conditions: [CIP]
0
= 100.0 μM, [NaOCl] = 1.0 mM, and
0
-2
UV fluence rate = 0.4 mW.cm .
Fig. 8. (a) CIP degradation by UV/chlorine process in surface water and WWTPE. (b)
CIP degradation by UV/chlorine process in ultrapure water and mixing water. (c-f)
Fluorescence spectra for NOMs in surface water and WWTPE water. Experimental
-2
conditions: [CIP]
0
= 10.0 μM, [NaOCl] = 100.0 μM, UV fluence rate = 0.4 mW.cm .
0
20

1
0
0
0
0
0
.0
.8
.6
.4
.2
.0
UV
Dark Chlorine
UV Chlorine
0
5
10
15
20
25
30
Time (min)
Fig. 1
21

a
b
1.0
0.8
0.6
0.4
0.2
0.0
1.0
0.8
0.6
0.4
0.2
0.0
UV/Cl+NaHCO3 0μM
UV/Cl+NaHCO3 50μM
UV/Cl+NaHCO3 200μM
UV/Cl+NaHCO3 500μM
UV/Cl+NaHCO3 1mM
UV/Cl+NaHCO3 5mM
UV/Cl+NaHCO3 10mM
UV/Cl+NaCl 0μM
UV/Cl+NaCl 50μM
UV/Cl+NaCl 200μM
UV/Cl+NaCl 500μM
UV/Cl+NaCl 1mM
UV/Cl+NaCl 5mM
UV/Cl+NaCl 10mM
0
2
4
6
8
10
0
2
4
6
8
10
Time (min)
Time (min)
c
d
1.0
0.8
0.6
0.4
0.2
0.0
1.0
0.8
0.6
0.4
0.2
0.0
UV/Cl+Na SO4 0μM
2
UV/Cl+NaNO3 0μM
UV/Cl+NaNO3 50μM
UV/Cl+NaNO3 200μM
UV/Cl+NaNO3 500μM
UV/Cl+NaNO3 1mM
UV/Cl+NaNO3 5mM
UV/Cl+NaNO3 10mM
UV/Cl+Na SO4 50μM
2
UV/Cl+Na SO4 200μM
2
UV/Cl+Na SO4 500μM
2
UV/Cl+Na SO 1mM
2 4
UV/Cl+Na SO4 5mM
2
UV/Cl+Na SO4 10mM
2
0
2
4
6
8
10
0
2
4
6
8
10
Time (min)
Time (min)
Fig. 2
22

1
0
0
0
0
0
.0
.8
.6
.4
.2
.0
0
0
4
mg/L humic acid
.4 mg/L humic acid
mg/L humic acid
0
2
4
6
8
10
Time (min)
Fig. 3
23

6
0
a
0.008
b
1
0
0
0
0
0
.0
.8
.6
.4
.2
.0
hydrated electrons
0
0
0
.006
.004
.002
50
.
HO
.
Cl
4
0
UV
4
5
6
7
8
9 10
30
pH 5
pH 6
pH 7
pH 8
pH 9
pH
2
1
0
0
0
0
2
4
6
8
10
pH 5.0
pH 7.0
pH 9.0
Time (min)
Fig. 4
24

a
1.0
b 22
UV/chlorine
UV
0
0
0
0
0
.8
.6
.4
.2
.0
Dark chlorine
2
0
18
UV/chlorine
UV
Dark chlorine
1
6
0
50
100
150
200
250
0
50
100
150
200
250
Time (min)
Time (min)
Fig. 5
25

Fig. 6
26

1
1
60
20
TOCl(Dark chlorine)
8
4
0
0
0
TOCl(UV/chlorine)
0
50
100
150
200
250
Time (min)
Fig. 7
27

a
1.0
.8
.6
.4
.2
.0
b
1.0
0.8
0.6
0.4
0.2
0.0
UP Water
Surface Water
WWTP Effluent
UP water
mixing water
0
0
0
0
0
0
2
4
6
8
10
0
2
4
6
8
10
Time (min)
Time (min)
c
d
4
00
400
350
300
250
200
before degradation
after degradation
0.000
100.0
350
2
00.0
00.0
00.0
00.0
00.0
00.0
300
3
4
250
5
6
200
7
300
350
400
450
500
550
300
350
400
450
500
550
Emission wavelength (nm)
Emission wavelength (nm)
4
50
00
50
00
50
00
450
400
350
300
250
200
e
f
before degradation
after degradation
0.000
00.0
400.0
00.0
00.0
1000
200
400
4
2
3
6
3
8
2
1
2
1
300
350
400
450
500
550
300
350
400
450
500
550
Emission wavelength (nm)
Emission wavelength (nm)
Fig. 8
28

Table 1. Characteristics of surface water and WWTPE water (mg/L)
2-
Parameters
pH COD NH +
4
-N NO -
3
-N NO -
2
-N TP
ND
Cl-
SO
4
Surface water
WWTP effluent
7.2 42.5 0.3
6.9 89.3 0.5
0.3
7.1
ND
ND
12.9
30.3
23.7
68.2
0.2
Note: (1) the unit is mg/L excluding pH, (2) COD chemical oxygen demand, TP total
phosphorus, ND not detected.
29