Rapid oxidation of histamine H2-receptor antagonists by peroxymonosulfate during water treatment: Kinetics, products, and toxicity evaluation

Article abstract of DOI:10.1016/j.watres.2020.116278

Peroxymonosulfate (PMS) is an appealing oxidant for organic contaminant destruction relying on radical generation after activation. Herein, we report PMS-promoted rapid degradation of histamine H₂-receptor antagonists (HRAs) through non-radical process for the first time. Five commonly used HRAs, i.e., ranitidine (RNTD), cimetidine (CMTD), famotidine (FMTD), nizatidine (NZTD) and roxatidine (RXTD), were examined their reactivity towards PMS. Results show that HRAs (except RXTD) exhibit high reactivity towards PMS, with apparent second-order rate constants from 403 to 872 M^?1s^?1 at pH 7.0. Radical scavenging experiments excluded the contribution of radicals to PMS-promoted degradation of HRAs, and this non-radical process was unaffected by the real water matrices. Structure-activity assessment and theoretical calculation indicated that the thioether sulfur in HRAs (except RXTD) was the main reactive site for PMS oxidation. Transformation product analysis further elucidated oxidation of the thioether sulfur to sulfoxide product through an oxygen atom transfer process. Moreover, the thioether sulfur on the straight chain was more susceptible to oxygen transfer with PMS than that on the thiazole ring of HRAs. Toxicity evaluation indicated the ecotoxicity of HRAs could be remarkably reduced after PMS oxidation. Hence, this work provides a promising strategy to rapidly remove HRAs and significantly reduce their toxicity in water treatment.

Full text of DOI:10.1016/j.watres.2020.116278

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Rapid Oxidation of Histamine H2-Receptor Antagonists by
Peroxymonosulfate during Water Treatment: Kinetics, Products, and
Toxicity Evaluation
Yajie Qian , Jinjing Huang , Xiang Liu , Tongcai Liu , Gang Xue ,
Pin Gao , Xuefei Zhou , Yalei Zhang , Jiabin Chen
PII:
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https://doi.org/10.1016/j.watres.2020.116278
WR 116278
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6 August 2020
Please cite this article as: Yajie Qian , Jinjing Huang , Xiang Liu , Tongcai Liu , Gang Xue ,
Pin Gao , Xuefei Zhou , Yalei Zhang , Jiabin Chen , Rapid Oxidation of Histamine H2-Receptor An-
tagonists by Peroxymonosulfate during Water Treatment: Kinetics, Products, and Toxicity Evaluation,
Water Research (2020), doi: https://doi.org/10.1016/j.watres.2020.116278
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Highlights
 PMS exhibits specific and high reactivity towards HRAs
 HRAs oxidation by PMS proceeds via a non-radical process
 PMS-promoted HRAs degradation is unaffected by water matrices
 Thioether sulfur in HRAs (except RXTD) is oxidized to sulfoxide product
 Ecotoxicity of HRAs is remarkably reduced after PMS oxidation
1

Rapid Oxidation of Histamine H2-Receptor Antagonists by Peroxymonosulfate during
Water Treatment: Kinetics, Products, and Toxicity Evaluation
Yajie Qiana, Jinjing Huanga, Xiang Liua, Tongcai Liub, Gang Xuea, Pin Gaoa, Xuefei Zhoub,
Yalei Zhangb,c, Jiabin Chenb,c,*
a College of Environmental Science and Engineering, Donghua University, Shanghai,
201620, P. R. China
b State Key Laboratory of Pollution Control and Resources Reuse, College of Environmental
Science and Engineering, Tongji University, Shanghai 200092, China
c Shanghai Institute of Pollution Control and Ecological Security, Shanghai, 200092, P.R.
China
* Corresponding Author.
Phone: 86-21-65985811; Fax: 86-21-65986960.
Email: chenjiabincn@163.com (Jiabin Chen).
Manuscript submitted to
Water Research
2

Abstract
Peroxymonosulfate (PMS) is an appealing oxidant for organic contaminant destruction
relying on radical generation after activation. Herein, we report PMS-promoted rapid
degradation of histamine H2-receptor antagonists (HRAs) through non-radical process for the
first time. Five commonly used HRAs, i.e., ranitidine (RNTD), cimetidine (CMTD),
famotidine (FMTD), nizatidine (NZTD) and roxatidine (RXTD), were examined their
reactivity towards PMS. Results show that HRAs (except RXTD) exhibit high reactivity
towards PMS, with apparent second-order rate constants from 403 to 872 M-1s-1 at pH 7.0.
Radical scavenging experiments excluded the contribution of radicals to PMS-promoted
degradation of HRAs, and this non-radical process was unaffected by the real water matrices.
Structure-activity assessment and theoretical calculation indicated that the thioether sulfur in
HRAs (except RXTD) was the main reactive site for PMS oxidation. Transformation product
analysis further elucidated oxidation of the thioether sulfur to sulfoxide product through an
oxygen atom transfer process. Moreover, the thioether sulfur on the straight chain was more
susceptible to oxygen transfer with PMS than that on the thiazole ring of HRAs. Toxicity
evaluation indicated the ecotoxicity of HRAs could be remarkably reduced after PMS
oxidation. Hence, this work provides a promising strategy to rapidly remove HRAs and
significantly reduce their toxicity in water treatment.
Keywords: Histamine H2-receptor antagonists; Peroxymonosulfate; Non-radical process;
Thioether sulfur.
3

1. Introduction
Histamine H2-receptor antagonists (HRAs) are gastrointestinal drugs extensively used to
inhibit gastric acid secretion through blockade of histamine H2-receptors in the gastric mucosa
(Ganellin et al. 1976, Ikehata et al. 2006a). The commonly used HRAs include ranitidine
(RNTD), cimetidine (CMTD), famotidine (FMTD), nizatidine (NZTD) and roxatidine
(RXTD) (Figure 1). After administration, a substantial portion of HRAs are excreted
unchanged, and finally discharged into the environment, and thus they have been widely
detected in the wastewater, surface water and even sediments (Andreozzi et al. 2003, Buth et
al. 2007, Choi et al. 2008, Kolpin et al. 2002, Radjenović et al. 2009, Zuccato et al. 2000). For
example, FMTD, RNTD and CMTD were detected in the wastewater effluent up to 5.38 μg/L
(Choi et al. 2008, Radjenović et al. 2009). RNTD and CMTD were found to occur in the
surface water at concentrations ranging from 10 to 580 ng/L (Kolpin et al. 2002, Zuccato et
al. 2000), and RNTD has been regarded as one of the highest risk pharmaceuticals in the
surface water (Besse and Garric 2008). The occurrence of HRAs in the environment may
exert biological effects on non-target organisms, and thus raises potential risk to the
ecosystem and public health.
Generally, biological treatment processes were reported to be ineffective to remove
HRAs in wastewater treatment plants (Radjenović et al. 2009); while oxidative water
treatment is an important method to degrade HRAs. Disinfection unit is a significant sink for
HRAs before HRAs discharging to the environment. HRAs, e.g., CMTD, are susceptible to
4

rapid oxidation by free chlorine and chloramines, generating some unexpected products with
higher toxicity (Buth et al. 2007). Ozonation was found to be effective in oxidizing nitro-
HRAs, e.g., RNTD and NZTD, but the oxidized products had higher potential to form
halonitromethanes (TCNM), which show high cytotoxicity and genotoxicity and thus have
been regulated (Wang et al. 2016). Although nitro-HRAs are photo-reactive under UV
irradiation, UV intensity currently used in disinfection could not fully eliminate HRAs and
their TCNM formation potential (Dong et al. 2017). Hence, more efficient and clean
techniques are required to remove these HRAs from wastewater effectively.
Persulfate (PS), including peroxymonosulfate (PMS) and peroxydisulfate (PDS), have
become emerging oxidants in wastewater treatment (Zhang et al. 2015a). The homogeneous
cleavage of peroxide bond in PS can generate reactive radical species, e.g., SO4•- (Chen et al.
2019b). Activation of PS to generate SO4•- is generally considered to be crucial for
contaminant destruction in wastewater treatment (Qian et al. 2018). Various strategies have
been used to activate PS for SO4•- generation, including electron transfer reaction (e.g.,
activation with transition metal (Anipsitakis and Dionysiou 2004) or base (Furman et al.
2010)) and external energy stimulation (e.g., UV (Zhang et al. 2015b, Zhou et al. 2018a) and
heating (Chen et al. 2016a, Gao et al. 2016)). SO4•- is a strong oxidant with high redox
potential (E0 = 2.5-3.1 V), and thus can degrade a broad range of contaminants (Chen et al.
2016b). However, SO4•- is non-selective, and thus also shows high reactivity towards water
-
matrices, e.g. , Cl-, HCO3 , and natural organic matters (Qian et al. 2016), which might reduce
the treatment efficiency in the real water matrices. Recently, the non-radical process with PS
has received considerable concerns owing to its high selectivity towards target contaminants
5

(Duan et al. 2018). PMS was reported to effectively oxidize some organic contaminants
without activation, and such reaction is significantly dependent on the contaminant structure
(Yang et al. 2018). For example, PMS shows specific and high reactivity towards β-lactam
antibiotics (Chen et al. 2018), fluoroquinolone (Zhou et al. 2018b), sulfonamides (Ji et al.
2018, Yin et al. 2018), and thioether sulfur, N4 amine on piperazine ring, and amino group on
the benzene were regarded as the effective sites for PMS oxidation, respectively. HRAs seem
to be reactive towards various oxidants due to the presence of multiple reactive
moieties/groups, such as furan ring, amine and thioether sulfur (Ikehata et al. 2006b).
However, it is still unclear whether PMS can degrade HRAs.
In this work, we found PMS exhibited specific and high reactivity towards some HRAs
without external energy or activators. This new discovery motivated us to conduct an in-depth
study to address two major research questions: (1) How PMS reacts with HRAs, and (2) What
characteristics render HRAs highly reactive towards PMS. To our best knowledge, this study
is among the first to systematically reveal the specific and rapid reaction between PMS and
HRAs. The new findings of this study provide a promising and facile strategy to eliminate
HRAs contamination in wastewater treatment.
2. Materials and methods
2.1 Chemicals
The standard chemicals of FMTD, RNTD, CMTD, NZTD and RXTD were purchased from
Sigma-Aldrich or ANPEL laboratory Technologies (Shanghai) Inc. at the highest purity. PMS
6

and PDS were obtained from Sigma-Aldrich at > 99% purity. Methanol (MeOH), acetonitrile
and tert-butyl alcohol (TBA), formic acid, and ammonium acetate were also obtained from
Sigma-Aldrich at HPLC grade. Hydrogen peroxide (HP, H2O2), sodium thiosulfate (Na2S2O3),
sodium dihydrogen phosphate (NaH2PO4), disodium hydrogen phosphate (Na2HPO4), sodium
chloride (NaCl), sodium bicarbonate (NaHCO3), sodium hydroxide (NaOH), humic acid, and
sulfuric acid (H2SO4) were obtained from Sinopharm Chemical Reagent Co., Ltd at analytical
grade. Deionized (DI) reagent water was produced from a Millipore Milli-Q Ultrapure
Gradient A10 purification system.
2.2 Real water samples
Surface water (SW) from a river and municipal wastewater (MWW) were collected at
locations in the southeast region of China. Samples were filtered through 0.45-μm glass fiber
filters immediately upon collection and stored at 5 °C before use. The characteristics of the
samples are shown in Tables S1. A synthetic hospital wastewater (HWW) was prepared by
adopting recipes (Serna-Galvis et al. 2017), and also used to evaluate the efficiency of HRAs
degradation. The composition of HWW is shown in Table S2.
2.3 Experimental procedures
Experiments were conducted in amber glass serum bottles at room temperature (22 °C).
Reaction pH was maintained with 10 mM phosphate buffer. Reaction solutions were prepared
with HRAs and buffer, and then PMS was added to initiate the reaction. Samples were
collected at the predetermined time and quenched with 100 μL of 1 M sodium thiosulfate
(Na2S2O3), which could completely consume the remaining PMS. The initial concentrations
7

used for HRAs and PMS were 10 and 100 μM, respectively, for most experiments. 10-500
μM of PMS were added to the HRAs solution to evaluate the impact of PMS concentration.
Radical quenchers (i.e., MeOH and TBA) were added to investigate whether the radicals
-
contributed to the degradation of HRAs. Cl-, HCO3 , and humic acid were spiked into the
reaction solution to evaluate the impact of water matrices. The above reactions were also
conducted in the real water matrices, i.e., SW, MWW, and HWW. All the experiments were
conducted in duplicate or more.
2.4 Analytical methods
HRAs were analyzed by a high performance liquid chromatograph (HPLC, 1260, Agilent
Technology, USA) equipped with a UV detector. Samples with 15 μL injection was separated
on a Kromasil-C18 column (4.6 × 250 mm, 5 μm). The details are provided in Table S3.
Transformation products were determined on a HPLC system (Utimate 3000, Dionex,
USA) connected to a triple quadrupole mass spectrometer (TSQ Quantum Ultra EMR,
Thermo Fisher Scientific, USA) with electrospray ionization (ESI). Samples with 40 μL
injection were separated on a Zorbax SB-C18 column (2.1 × 150 mm, 5 μm). The detailed
conditions are summarized in Text S1.
PMS was analyzed by a spectrophotometric method proposed by Liang et al. (Liang et
al. 2008), as detailed in Text S2.
8

2.5 Theoretical calculation
The frontier electron density (FED) calculations were conducted by the Gaussian 09
program on a basis set of B3LYP/6-31 G (d, p). The FEDs of the highest occupied molecular
orbital (HOMO) were obtained and visualized. Condensed Fukui function based on the
density functional theory (DFT) was used to predict the regioselectivity of different reactive
species on the HRAs molecules.
2.6 Ecotoxicity assessment
The ecotoxicity of HRAs and the oxidized products were evaluated using the Ecological
Structure-Activity Relationship Model (ECOSAR) program (Liu et al. 2018). This program
has been successfully used to predict the ecotoxicity to screen contaminants. The ecotoxicity
was assessed for three trophic levels of aquatic organisms, i.e., fish, daphnid and green algae.
Half Lethal Concentration (LC50) values for fish (96 h), daphnid (48 h), and Half Effective
Concentration (EC50) values for algae (96 h) were used to assess the acute toxicity. Chronic
Value (ChV) was used to demonstrate chronic toxicity.
3. Results and discussion
3.1. PMS-promoted degradation of HRAs
Advanced oxidation processes (AOPs) are considered as efficient techniques to degrade a
wide range of contaminants owing to the generation of highly reactive radicals, e.g., SO4•- and
HO•. Indeed, RNTD was susceptible to significant degradation in the UV/PDS and UV/HP
9

systems (Figure 2). In the UV/PMS system, RNTD degradation was much faster than that
observed in UV/PDS or UV/HP systems, with complete degradation occurred after 2 min.
Interestingly, PMS alone could also promote rapid degradation of RNTD; while the presence
of PDS or HP alone exhibited negligible effect on RNTD degradation (Figure 2). PMS
oxidation, reactive radical oxidation, and UV photodegradation could contribute to RNTD
degradation in the UV/PMS system. The similar trend of RNTD degradation in PMS and
UV/PMS system indicated that the contribution of PMS oxidation overwhelms other two
degradation processes in UV/PMS system (Figure 2). Indeed, the decomposition of PMS was
only slightly increased after introduction of UV irradiation (Figure S1). Hence, the rapid
degradation of RNTD in UV/PMS system was most likely contributed from the PMS-
promoted degradation. We further conducted radical quenching experiments to investigate
whether radical species involved in the PMS-induced oxidation of RNTD. Generally, MeOH
and TBA are always selected as the radical probes to distinguish SO4•- and HO•. MeOH reacts
with SO4•- and HO• at a comparable rate (9.7 × 108 M-1s-1, 3.2 × 106 M-1s-1 for SO4•- and HO•,
respectively) ; while TBA reacts with SO4•- ((4.0-9.1) × 105 M-1s-1) much slower than that
with HO• ((3.8-7.6) × 108 M-1s-1) (Liang and Su 2009). As shown in Figure S2, the addition of
alcohols did not affect the degradation of RNTD by PMS, even when the concentration of
alcohols increased to 1000 mM. This result indicated that radical species did not involve in
the PMS-promoted degradation of RNTD, and this reaction most likely proceeded via the
non-radical process. It is noted that singlet oxygen (1O2) can be generated from the self-
decomposition of PMS. Furfuryl alcohol (FFA, a scavenger for 1O2 (Yang et al. 2018)) was
added in PMS/RNTD solution, and the results showed that both RNTD degradation and PMS
10

decomposition were not affected after addition of FFA after 2 min (Figure S3), and thus 1O2
did not contribute to the PMS-promoted degradation of RNTD. Hence, PMS likely exhibited
specific reactivity towards RNTD.
We further investigated the degradation of other four HRAs by PMS to evaluate whether
PMS shows specific reactivity towards all HRAs. As shown in Figure 3, PMS promoted rapid
degradation of CMTD, FMTD and NZTD, but was inert for the degradation of RXTD. Hence,
PMS-promoted degradation was dependent on HRAs’ structure. HRAs possess several
electron-rich moieties/groups, e.g., amine, which are always regarded as the potential reactive
sites for oxidants. On the other hand, multiple amine moieties/groups are present in the
investigated HRAs; while thioether sulfur occurs in CMTD, RNTD, FMTD and NZTD, but
not in RXTD (Figure 1). It is thus suggested that the different reactivity of HRAs towards
PMS likely related to the presence/absence of the thioether sulfur, which will be further
elucidated in the following sections.
3.2 Degradation kinetics
The impact of HRAs concentration on the PMS-promoted degradation was investigated
at neutral pH condition. As shown in Figure 4A, PMS-promoted HRAs degradation increased
with PMS concentration increasing from 0.05 mM to 0.5 mM. Moreover, HRAs degradation
followed pseudo-first-order kinetics at various PMS concentrations (equation 1), with the
calculated pseudo-first-order constants (kobs) demonstrated in Figure 4B. Log(kobs) linearly
increased with log(CPMS), and the slope was close to unity. Hence, kobs showed a linear
relationship with respect to PMS concentration, suggesting the overall reaction could be
11

described by a second-order kinetics (equation 2). Apparent second-order reaction rate
constants (k2,app) could be obtained by dividing kobs by PMS concentration. As shown in
Figure 4C, k2,app was calculated to be 481.6 ± (16.7), 872 ± (57.9), 531.6 ± (36.2), 403 ±
(21.5) M-1s-1 for RNTD, FMTD, CMTD and NZTD, respectively. The calculated k2,app values
for HRAs were almost one order of magnitude higher than those for the reaction between
PMS and β-lactam antibiotics (Chen et al. 2018) and tetracyclines (Chen et al. 2019a).
(1)
(2)
3.3 pH impact
The oxidation of organic contaminants by PMS was previously reported to depend on
solution pH (Nihemaiti et al. 2020). Hence, pH impact was further examined for the
degradation of HRAs by PMS at pH 3.0-9.0. As illustrated in Figure 5, PMS-promoted
degradation of HRAs were pH-dependent. Generally, HRAs degradation by PMS was almost
unaffected by the solution pH from 3.0 to 5.0. Afterwards, the values of k2,app slightly
decreased with pH further increased to pH 7.0, and more sharply decreased when pH finally
increased to pH 9.0. PMS possesses two pKa values (pKa1 < 0, pKa2 = 9.4) (Zhang et al.
-
2-
2013), and thus mainly exhibiting HSO5 and SO5 species at the investigated pHs (
, pKa2 = 9.4) (Figure S4). In addition, the pKa values for the amine groups in
HRAs were at around pH 7.0 (i.e., pKa = 8.2 (Latch et al. 2003), 7.1 (Latch et al. 2003), 6.7
(Echizen and Ishizaki 1991), and 6.8 (Wu et al. 2001) for RNTD, CMTD, FMTD and NZTD,
respectively). Hence, HRAs exhibit the protonated state at the acidic pHs (pH 3.0 and 5.0),
12

but deprotonated state at pH 9.0 ( ). The species-specific reaction
between HRAs and PMS was used to explain the pH dependency of k2,app of HRAs (equation
3 and 4).
∑
Hence,
∑
where [HRAs]T and [PMS]T represent the total concentration of HRAs and PMS,
respectively; ki,j represents the species-specific second-order rate constant for the reaction
between species i of HRAs and species j of PMS; and αi and βj represent the distribution
coefficients of HRAs and PMS species, respectively. The ki,j were obtained by least-squares
regression of the experimental data of k2,app-HRAs to equation (4). Because of the small fraction
2-
-
of HRAs0-SO5 (Figure S5), and lower oxidation capacity of SO52- than HSO5 , the specific
reaction between HRAs0 and SO52- was not considered. As shown in Figure 5, the
experimental data of k2,app-HRAs at various pHs could be well explained by the species-specific
reactions. The values of ki,j are summarized in Table S4, and their contribution to the overall
k2,app is described in Figure S6. Results show that the contribution of SO52- could be neglected
at pH 7 and below (Figure S6). Indeed, SO52- showed relatively low reactivity towards HRAs
-
(Table S4), and SO52- were reported to possess lower oxidation capacity than HSO5 (Lei et al.
2016). Furthermore, HRAs0 was more susceptible to oxidation with PMS than HRAs-, which
13

is different from the fact that the deprotonated species with richer electrons normally react
faster with oxidants. It was thus implied that the reaction site on HRAs towards PMS
oxidation is not the deprotonated functional group, i.e., the amine groups on HRAs, which
will be further discussed in the following sections. The higher reactivity of HRAs0 towards
PMS might be explained by electrostatic interactions. The negatively charged PMS could get
close to the protonated HRAs more easily than the deprotonated HRAs, thus exhibited higher
reactivity towards the protonated HRAs.
3.4 Impact of water matrices
We further evaluated the impact of water matrices on PMS-promoted degradation of
HRAs with the results shown in Figure 6. Unlike the AOPs with radical generation, the
-
addition of common anions, e.g., Cl- (0-300 mM) and HCO3 (0-100 mM) did not affect the
degradation of HRAs by PMS at neutral pH (Figure 6A and B), indicating these anions in
water matrices exhibited negligible effect on PMS-promoted degradation of HRAs. Indeed,
-
Cl- reacts with HSO5 and SO52- at the rate constants of 2.06 × 10-3 M-1s-1 and 3.8 × 10-4 M-1s-1,
respectively (Lente et al., 2009), which are much lower than the rate constants between HRAs
with PMS (403-872 M-1s-1). Hence, the impact of Cl- on PMS-induced oxidation of HRAs
could be negligible. Moreover, the degradation of HRAs was not affected in the presence of
HA (0-50 mg/L) (Figure 6C), suggesting that the natural organic matter also showed
neglectable effect on HRAs degradation. Organic matter and Cl- are always known to affect
SO4•- , and thus the water matrices can compete with the target contaminants towards the
radical consumption in AOPs, thus reducing the treatment efficiency (Chen et al. 2019b). The
14

negligible impact of water matrices on PMS-promoted HRAs degradation also implied that
HRAs degradation was not primarily contributed from radicals.
We further investigated PMS-promoted degradation of HRAs in the real water samples.
As shown in Figure 7, HRAs degradation in DI water was similar to those in the SW, HWW
and MWW, thus the degradation of HRAs by PMS was not affected by the real water
matrices. Moreover, we also evaluated the decomposition of PMS along the HRAs
degradation in the real water matrices, and the results show that PMS decomposition was
initially fast, and finally ceased (Figure S7). The trend of PMS decomposition resembled the
degradation of HRAs, indicating that the decomposition of PMS was induced by the reaction
with HRAs, but not by the water matrices. On the other hand, PMS decomposition was also
not influenced by the real water matrices. These evidences further confirmed that PMS
showed specific and rapid reactivity towards HRAs, which was not affected by the water
matrices. Hence, compared to the AOPs with radical generation, PMS oxidation was a more
appealing alternative for degradation of HRAs in real water matrices.
3.5. Reaction pathway and mechanism
3.5.1 Theoretical prediction of reaction sites
The abovementioned radical scavenging experiments verified the non-radical process for
PMS-promoted HRAs degradation. The oxidation of HRAs most likely proceeded via
electrophilic reaction by PMS. PMS tends to attack the moieties/groups with high electron
density in HRAs. Indeed, the electron-rich moieties/groups in HRAs, such as double bonds,
amine, sulfur were regarded as the potential sites susceptible to oxidation by various oxidants,
15

e.g., O3, KMnO4, Cl2, and ClO2 (Wang et al. 2016). Theoretical calculations were frequently
used to predict the potential reaction sites in electrophilic reaction (Qu et al. 2015). Based on
the frontier molecular orbital theory, the electrophilic reaction probably takes place at the
electron rich (HOMO) regions, and the results were shown in Figure 8. Generally, the HOMO
orbits cannot precisely describe the gain or loss of electron on each HRAs molecule site
during reaction. Therefore, we also calculated the Fukui index on electrophilic attack (f−)
according to natural population analysis (NPA) charge distribution of HRAs molecule (Tables
S5-S9). For the investigated HRAs (except RXTD), the HOMOs and f− were mainly
distributed on the double bond, sulfur, and nitrogen atoms, suggesting these sites might act as
the potential reactive sites susceptible to PMS attack. For RXTD, however, the HOMOs
primarily located on the double bond and the nitrogen atoms. Based on the abovementioned
results that RXTD showed negligible reactivity towards PMS, sulfur atom (absence in RXTD)
was likely the reactive sites for PMS oxidation. To facilitate discerning the reaction
mechanism, the transformation products by PMS oxidation were further analyzed.
3.5.2 Transformation products
Transformation products of HRAs were further investigated by LC/MS/MS. The primary
products with molecular weights (MW) of M+16 and M+32 (M: MW of the parent HRAs)
were observed during PMS-promoted oxidation of HRAs (Figure 9). They were most
probably the oxidation products with one/two oxygen addition on the parent HRAs, which
were also observed in the direct oxidation of β-lactam antibiotics/tetracycline by PMS (Chen
et al. 2018, Chen et al. 2019a). In the PMS-promoted oxidation, however, the product profile
of HRAs was different from those of β-lactam antibiotics/tetracycline. For example, two or
16

more M+16 products were generated and verified as the isomers in PMS-promoted oxidation
of β-lactam antibiotics (Chen et al. 2018); while the sole M+16 product was produced in
HRAs degradation by PMS. On the other hand, M+16 products were generated via oxygen
transfer from PMS to the thioether sulfur on the five (six)-membered ring of β-lactam
antibiotics (Chen et al. 2018), but via hydroxylation process in the B ring of tetracycline
(Chen et al. 2019a). It is still unclear how M+16 were generated in the PMS-promoted
oxidation of HRAs. The structure of transformation products were thus analyzed based on the
mass fragments.
RNTD
Both M+16 and M+32 products were generated in PMS-promoted oxidation of RNTD, and
the intensity of M+16 was much higher than that of M+32 (Figure 9A). Hence, the reaction
between PMS and RNTD primarily proceeded via the addition of one oxygen atom on RNTD.
To evaluate the potential reactive sites on RNTD, we analyzed the bond cleavage of RNTD in
the ESI fragmentation at first. In the mass spectrum of the parent compound RNTD (Figure
S8), the presence of m/z 176/138, and m/z 170/144 indicated the breakage of the –S–C– bond,
hence this site is susceptible to cleave in the reaction. Moreover, the m/z 130 is also observed
in the spectrum of RNTD, especially at the higher fragmentation voltage, suggesting that the
–C–C– near the sulfur atom is also likely to cleave during the reaction, but the required
energy is higher than the cleavage of –S–C–. In addition, the fragment of m/z 270 was
generated via the loss of N,N-dimethylamino moiety from RNTD, and further loss of –NO2 of
m/z 270 produced the fragment of m/z 224.
17

In the M+16 product, the presence of m/z 138 and m/z 110 fragments indicated the intact
furan structure in RNTD (Figure S9). Hence, the oxygen addition site most likely located on
the other side part of RNTD, which was verified by the presence of m/z 192 (192 = 176 +16,
m/z 176 is the S containing structure in RNTD) in the M+16 product. Fragment of m/z 286
was generated from the loss of N,N-dimethylamine (45 Da) from RNTD, and further loss of
H2O (18 Da) generated the fragment ion m/z 268. It is noted that both RNTD and M+16
product contain fragment of m/z 130, suggesting the oxygen addition could not occur at the
double bond on the right side (Figure S9). Hence, the –S–C– is the most likely site for oxygen
addition. According to the theoretical calculation, HOMOs of sulfur atom is much higher than
the neighboring carbon, thus sulfur is vulnerable to electrophilic attack by PMS. Indeed, the
thioether sulfur was previously regarded as preferable site for PMS oxidation (Chen et al.
2018). Hence, M+16 product was supposed to be the sulfoxide product of RNTD.
Both the fragments of m/z 130 and m/z 138 were observed in the mass spectra of RNTD
and its M+32 product (Figure S10), indicating the intact furan structure on the left side and
the double bond on the right side in M+32 product. The –S–C– moiety is also the potential
reactive sites for two oxygen atoms addition. In the RNTD sulfoxide, the oxygen is difficult
to further add on the sulfur to generate sulfone products owing to the steric hindrance effect.
The second oxygen is most likely to add on the carbon neighboring to sulfur and underwent
via the hydroxylation process. This hypothesis was verified by the presence of m/z 192
fragment in the mass spectrum of M+32 product. Because of the extremely low intensity of
M+32 product (Figure 9A), the hydroxylation at carbon neighboring sulfur on RNTD could
be negligible compared to the oxidation at thioether sulfur on RNTD.
18

CMTD
Only M+16 product was observed in the reaction between PMS and CMTD (Figure 9 B).
The prominent fragment of CMTD was m/z 159 (Figure S11), suggesting the –S–C– is liable
to cleave, and thus this moiety is vulnerable to reactions. With the increasing fragmentation
voltage, the intensity of m/z 117 increased along with the decreasing m/z 159 intensity,
indicating the loss of –N–CN in CMTD. In M+16 product (Figure S12), the sulfur atom was
most susceptible to electrophilic attack by PMS based on the theoretical calculation. Indeed,
the presence of m/z 175 (175 = 159 + 16) and m/z 133 (133 = 117 + 16) also indicated that
the oxygen atom likely added on the sulfur to generate CMTD sulfoxide product.
FMTD
The product profile of FMTD was similar to that of RNTD. Both M+16 and M+32 products
were observed along FMTD oxidation (Figure 9C). The mass fragment pattern of FMTD
resembled that of CMTD, that is, few fragments were observed even at high voltage (Figure
S13). The presence of m/z 189 and m/z 155 indicated the cleavage of –S–C– moiety, while
m/z 259 means the loss of –SO2–NH2 fragments. In M+16 product, the most abundant
fragment is m/z 205 (Figure S14), indicating the oxygen addition likely occurred on the m/z
189 fragment in FMTD (205 = 189 + 16). Similarly, the presence of m/z 275 suggests the
oxygen addition on m/z 259 in FMTD (275 = 259 + 16). On the other hand, the fragment of
m/z 155 in M+16 product indicated the presence of intact thiazole ring. The fragments of m/z
275, 205 and 155 were also reported in the product of FMTD generated by the photo-catalytic
oxidation, which was finally identified as sulfoxide of FMTD (Murphy et al. 2012). The
19

above evidences indicated that the oxygen addition was located on the sulfur atom, generating
the FMTD sulfoxide product.
For M+32, the intensity of [M+32+Na]+ is much higher than that of [M+32+H]+, and the
mass spectrum of [M+32+Na]+ was obtained and shown in Figure S15. The presence of m/z
313 means the oxygen addition on m/z 259 fragment (313 = 259 + 32 – H + Na), and the loss
of –SO2–NH2 fragments. Moreover, the appearance of m/z 243 indicated the second oxygen
added on the left side of M+16 (243 = 205 + 16 – H + Na), probably on the sulfur on thiazole
ring based on the theoretical calculation.
NZTD
Both M+16 and M+32 products were observed in PMS-promoted oxidation of NZTD
(Figure 9D). In the mass spectrum of NZTD, the presence of m/z 155 and m/z 187 indicated
the cleavage of –S–C– in the fragmentation (Figure S16). It is noted that NZTD and RNTD
share the same right side in their structure (Figure 1), the cleavage of –S–C– in RNTD
generated the primary fragment of m/z 176, which was not observed in NZTD. However, m/z
131 was observed in the fragments of NZTD, and its intensity increased as the most abundant
fragment when the voltage increased to 30 eV. m/z 131 was supposed as the product of m/z
176 after loss of –NO2. The appearance of m/z 286 also verified –NO2 prone to lose in the
fragmentation. According to the theoretical calculation, the sulfur atom and the C atom next
to –NO2 possess high HOMOs, consistent with the fragmentation patterns of NZTD.
In the M+16 product, the presence of m/z 130 and m/z 144 indicated the intact right side of
NZTD (Figure S17), and thus the oxygen addition could not locate on the double bond next to
20

–NO2. In addition, the appearance of fragments of m/z 205 and m/z 147 suggest the oxygen
added on the sulfur on the straight chain. m/z 147 was supposed to be the addition of oxygen
on m/z 131 fragment of NZTD, and the loss of –NO2 in the right side. According to the
evolution of transformation products, M+16 product initially increased and then gradually
decreased. At the same time, the other product, i.e., M+32, continuously increased (Figure
9D). Hence M+16 product was further transformed to M+32 product, and the second oxygen
likely added on M+16 product. The presence of m/z 147 and m/z 192 indicated an intact right
side of M+16 (Figure S18), hence the second oxygen was likely added on the sulfur atom on
thiazole ring based on the theoretical calculation.
Overall, M+16 and M+32 products were identified as the primary products of HRAs after
reaction with PMS. Although the hydroxylation took place in the PMS-promoted degradation
of RNTD to form M+32, its intensity was much lower than M+16, and thus this reaction
process could be neglected in PMS promoted oxidation of HRAs. Oxidation of thioether
sulfur to generate sulfoxide were supposed as the dominant mechanism for HRAs oxidation.
3.5.3 Reaction mechanism
The conventional PMS activation processes rely on the generation of radicals for
contaminant destruction. In this work, PMS-promoted degradation of HRAs did not rely on
radical generation, but proceeded via a direct oxidation. This mild process was previously
reported to dominate in the oxidation of arsenite (Wang et al. 2014), hydrogen sulfide
(Betterton and Hoffmann 1990), and β-lactam antibiotics by PMS (Chen et al. 2018).
Generally, the direct oxidation process proceeds through two-electron transfer, which
21

involves the heterolytic breakage of the peroxide bond and an oxygen transfer from PMS to
•-
the contaminants (Figure 10). PMS was converted to sulfate ion upon reaction without SO4
generation. Indeed, in the direct oxidation between PMS and HRAs, the symmetry-adapted
configuration between 2p orbital of oxygen and empty 3d orbital in sulfur renders lone pair
electrons of oxygen liable to donate to sulfur, generating a feedback coordination bond (Wang
et al. 2016). This process results in an oxygen atom transfer from PMS to the thioether sulfur
with the formation of HRAs sulfoxide. It is noted that two stereoisomeric sulfoxides were
generated in β-lactam antibiotics oxidized by PMS (Chen et al. 2018). The fusion of β-lactam
ring and five (or six)-membered ring makes the electrons of thioether sulfur on the ring more
exposed outside, thus susceptible to peroxide attack from different directions. Such structure
characteristic is not observed in HRAs, thus only one sulfoxide is generated in PMS-
promoted oxidation of HRAs.
It is noted that multiple thioether sulfur containing moieties/groups are present in the
structures of HRAs, e.g., FMTD and NZTD. Based on the transformation product analysis of
FMTD and NZTD, both the thioether sulfurs on the straight chain and the thiazole ring show
reactivity towards PMS oxidation. Moreover, the intensity of M+16 product was much higher
than that of M+32 product, indicating that the thioether sulfur on the straight chain shows
much higher reactivity towards PMS than that on the thiazole ring of HRAs. On the one hand,
the Fukui index (f−) of thioether sulfur on the straight chain was higher than that on the
thiazole ring (Tables S6 and S7), hence the thioether sulfur on the straight chain was more
susceptible to electrophilic attack by PMS than that on the thiazole ring. On the other hand,
the amine groups adjacent to thioether sulfur on the thiazole ring might hinder the attack of
22

PMS on the thioether sulfur owing to the steric hindrance effect, thus contributing to the
observed lower reactivity to PMS. These two factors induced the thioether sulfur on the
straight chain in HRAs more susceptible to oxygen transfer with PMS.
Compared to the conventional PMS activation with radical generation, two electron transfer
(i.e., oxygen atom transfer) process for HRAs oxidation exhibited high efficiency because the
oxidation capacity of the peroxide were fully utilized. Moreover, the direct oxidation by PMS
was not affected by the background ions and NOM in the water matrices, and thus showed
high treatment efficiency in the real water matrices. Hence, this process could minimize the
adverse impact of coexisting matrix constituent, and maximize the utilization efficiency of
PMS. On the other hand, the oxidation of nitro-HRAs by conventional oxidants, e.g., ozone
and chlorine, could generate the products with high TCNM formation potential (Wang et al.
2016); while the oxygen transfer process only involved the oxidation at thioether sulfur, but
made little contribution to TCNM formation. Overall, direct oxidation by PMS is a facile and
environmental-friendly strategy to alleviate HRAs contamination in the water.
3.6. Ecotoxicity evaluation
QSAR analysis was used to predict the ecotoxicity of contaminants and their products by
ECOSAR program. As shown in Table 1, the acute toxicity was categorized as “not harmful”
(> 100 mg/L) to fish for all the target contaminants, and classified as “harmful” (10-100
mg/L) to daphnid and green algae for CMTD and RNTD. Compared to the parent HRAs, the
acute toxicity of the transformation products, e.g., M+16 and M+32 products belong to the
“not harmful” category, and the LC50 (or EC50) values were ten times or even higher than
23

those of HRAs. Hence, acute toxicity of the HRAs could be significantly reduced after
treatment with PMS.
For the chronic toxicity, CMTD and RNTD are classified as “toxic” (1-10 mg/L)
compounds to daphnid, but “harmful” compounds to fish and algae; while their products have
a significantly lower chronic toxicity with the ChV values on 1 order of magnitude higher
than HRAs. Similarly, transformation products of FMTD and NZTD show much lower
chronic toxicity than their corresponding parent contaminants. Overall, ecotoxicity of HRAs
could be remarkably reduced after PMS treatment, and thus oxidation by PMS is a promising
strategy to degrade HRAs and eliminate their toxicity.
4. Conclusions
PMS exhibits specific and high reactivity towards HRAs except RXTD, and the apparent
second-order rate constants were determined to be in the range from 403 to 872 M-1s-1 at pH
7.0. Radical scavenging experiments confirmed lack of involvement of radical species. This
non-radical process was not influenced by the water matrices, and thus maintained high
efficiency in wastewater treatment. Theoretical calculations and transformation product
analysis indicated that the thioether sulfur in HRAs (except RNTD) was the site susceptible to
PMS oxidation, and was oxidized to the corresponding sulfoxide product. Toxicity evaluation
indicated the ecotoxicity of HRAs was remarkably decreased after PMS oxidation. This work
highlights a promising and facile strategy for alleviating HRA contamination in water
treatment.
24

Acknowledgements
We sincerely thank the National Natural Science Foundation of China (51708097,
51878431), State Key Laboratory of Pollution Control and Resource Reuse Foundation (NO.
PCRRF18007), and Fundamental Research Funds for the Central Universities (NO.
2232018D3-08).
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