Photolysis of sulfamethazine using UV irradiation in an aqueous medium

Article abstract of DOI:10.1039/c7ra09564c

Although many studies have been focused on the photochemistry of antibiotics, the roles of reactive species in photolysis and the effects of dissolved substances on antibiotic photochemical behavior have been poorly examined. The photolytic behaviors of s

Full text of DOI:10.1039/c7ra09564c

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Photolysis of sulfamethazine using UV irradiation in
an aqueous medium
Cite this: RSC Adv., 2018, 8, 1427
a
b
a
a
Zhigang Yi, * Juan Wang, Qiong Tang and Tao Jiang
Although many studies have been focused on the photochemistry of antibiotics, the roles of reactive species in
photolysis and the effects of dissolved substances on antibiotic photochemical behavior have been poorly
examined. The photolytic behaviors of sulfamethazine (SMN) in pure water were investigated via adding
different scavengers to quench the active species. Results showed that decomposition of the triplet-
3
excited state of SMN ( SMN*) by direct photolysis was the main path of SMN photolysis in water. Moreover,
self-sensitized SMN cannot be ignored during SMN photodegradation. The main photoproducts of SMN
were identified by LC-MS/MS, which indicated that SMN could not be mineralized although the photolysis
ꢀ
3ꢀ
under UV was effective. The effects of Cl , NO , and fulvic acid (FA) (common substances in natural
water) on SMN photolytic behaviors were also studied. The triplet-induced halogenation of SMN increases
the ionic strength and reduces the ground state SMN; these are the primary causes of promotion of SMN
Received 29th August 2017
Accepted 18th December 2017
ꢀ
3ꢀ
photolysis by Cl . More cOH produced in the presence of NO could promote SMN photolysis.
Competitive absorption of photons of FA with SMN and ROS scavenged by FA were the main reasons for
the inhibition of SMN photolysis. The research findings are helpful for further studies on the environmental
risks of ACs in natural waters and promoting the development of AC pollution treatment technology.
DOI: 10.1039/c7ra09564c
rsc.li/rsc-advances
ꢀ
1
presence in AC manufacturing effluents might reach mg L
1
. Introduction
12
levels. The bacterial resistance and photo-modied toxicities
In the last few decades, large amounts of veterinary antibiotics of ACs could cause severe negative impact on human health and
1,13–18
(
ACs) have been used in animal husbandry as therapeutic the environment.
medicine and feed additives for growth promotion. Moreover, dangerous pollutants for the environment.
one of the most popular groups of ACs, which are mainly used
The obstinate nature of ACs residues interferes with the
ACs have been classied as particularly
1
1,2
in livestock, are sulfonamides (SNs). In China, AC use in stock removal of these compounds from the environment by tradi-
raising was nearly 92 700 tons, and usage of SNs was nearly 7920 tional biological treatments. Finding effective methods to
3
tons in 2013. In the USA, 16 000 tons of ACs, including 2.3% eliminate the ACs residues is required for environmentally
SNs, were consumed per year. In Europe, the output of SNs sustainable development. Heterogeneous photocatalysis, one of
4
ranged from 11% to 23% of the total production of ACs. SNs the typical advanced oxidation techniques (AOPs), is an efficient
19
may be excreted from the body in their parent form without treatment to promote degradation of organic pollutants.
absorption and metabolism by animals. Because of their high Further, the performance of photocatalytic technology could be
20
stability and solubility, conventional sewage disposal tech- effectively improved via photocatalyst modication, electro-
nology used in livestock farming is not efficient for SN removal. chemical, microwave or ultrasonic-assisted photocatalytic
2
1–23
The average removal ratio of SNs by activated sludge was found technology.
The efficiency of photocatalysis is based on the
1
to be 24%, which meant that a mass of these drugs was formation of reactive species through the irradiation of light
introduced into the biosphere every year. In this case, ACs energy on catalysts. During photocatalytic degradation of
were detected in rivers, lakes, estuarine, and coastal waters organic contaminants, photodegradation may play an impor-
1,5
6–8
many years ago. The highest concentration of sulfamethazine tant role, and photochemical decomposition has also been
(
SMN) in manure of swine farms in South China was up to proven to be a major transformation pathway for organic
ꢀ1
9
24,25
0
.250 mg kg . The concentration of ACs in the aquatic envi- pollutants in surface waters.
Therefore, it is signicant to
ꢀ1
ꢀ1 10,11
ronment was relatively low (ng L to mg L ),
whereas its carry out a study on photodecomposition of pollutants.
Photolysis of ACs is mainly caused by ultraviolet wavelength
irradiation, and ACs can undergo not only direct or indirect
College of Chemistry, Leshan Normal University, Leshan, Sichuan 614004, China. photolysis, but also self-sensitized photo-oxidation. During
2
a
E-mail: yizhigang117@hotmail.com
direct photolysis, photon absorption promotes electrons from
the initial ground state of ACs to produce electronically excited
b
Environmental Monitoring Station of Environmental Protection Bureau of Rizhao
Lanshan, Lanshan, Shandong 276800, China
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3
species in the singlet state (AS*) or triplet state ( AS*). On the
one hand, the excited species can be decomposed into photo-
3
products. On the other hand, by self-sensitization, AS* can
3
transfer the energy to the ground state of O or H O to form
2
2
1
26,27
reactive oxygen species (ROS, e.g., cOH and O
2
).
In addition
to the direct and self-sensitized photolysis, indirect photolysis is
a signicant elimination pathway for many aquatic pollutants;
it is initiated by ROS, formed via optical absorption by photo-
sensitizers (e.g., dissolved organic matter (DOM), Fe species,
28–30
halide ions, and H O ).
However, controversy still exists
2
2
about which path should play the leading role in different
research. Studies indicated that many byproducts were still
retained during the photolysis of antibiotics, the intermediates
showed an increasing toxicity, and the solutions aer photolysis
31–35
still had certain residual antibacterial activity.
Because the
photolytic byproducts will be different due to different photo-
lytic reaction paths of ACs, studies on the main factors for
photolysis and the reaction paths of ACs under UV irradiation
are helpful for a further study on the environmental risks of ACs
in natural waters and promoting the development of AC pollu-
tion treatment technology.
Fig. 1 A schematic of the experimental setup.
2.3 Analytical methods
In this study, we employed SMN as a proxy AC to investigate
photolytic behavior under UV irradiation and detect the role of
triplet-excited state and ROS in the photodegradation of ACs.
Furthermore, the photodecomposition pathway of SMN was
proposed. The inuence of common dissolved substances in
natural water on the photodecomposition of SMN was also
discussed in this study.
The concentration of SMN at different irradiation times was
determined by HPLC (LC-2010HT, Shimadzu) coupled with UV/
VIS detection (LC-UV/VIS). The parameters of the analysis were
2
set according to the research of Batista et al. a C18 5 mm to 100 A˚
column (250 mm ꢂ 4.60 mm). The detection wavelength of
SMN was found to be at 268 nm. The sample injection volume
was 50.0 mL. The eluents were (A) H O + 0.200% acetic acid and
2
(
1
0
B) acetonitrile at 80 : 20 ratio, and the ow rate was set at
ꢀ1
.00 mL min . The detection and quantication limits were
2. Materials and methods
ꢀ
1
ꢀ1
.170 mg L and 0.500 mg L , respectively.
2.1 Chemicals
The products of photolysis were identied by a Shimadzu
High purity standard SMN (99%), used as a target pollutant in LC-20A liquid chromatography system coupled with a Shi-
photochemical experiments and as a standard in high- madzu LCMS-8030 triple quadrupole mass spectrometer (LC-
performance liquid chromatography (HPLC) analysis, was MS/MS). The eluent and column used were the same as those
purchased from Aladdin Industrial Corporation (Shanghai, of LC-UV/VIS, but the ratio of A and B eluents was altered during
China). Acetonitrile and isopropanol were of HPLC grade and the analysis: it started with 10% of B, then it increased to 60%
obtained from Chengdu Best Reagent (Chengdu, China). Sorbic aer 10 min, then to 90% in another 8 min, and nally the
acid (t,t-HDA, 99%), sodium chloride, sodium nitrate, fulvic content of B dropped to 10% in another 10 min and remained at
acid (FA), and other chemical reagents were of analytical grade this ratio until the end of the run. The detection was performed
and obtained from Chengdu Best Reagent. Ultrapure water was using an electrospray ionization (ESI) source operating in the
used to prepare SMN solutions and HPLC eluent.
positive mode. The parameters were set as follows: capillary
ꢁ
voltage 4000 V, drying gas temperature 300 C, drying gas ow
ꢀ1
1
2 mL min , and nebulization gas 35 psi. Fragmentor voltages
2
.2 Experimental set-up
A merry-go-round photochemical reactor with a magnetic stirrer
illustrated in Fig. 1) was employed. The high pressure Hg lamp
were adjusted between 10 and 30 V to obtain precursor ions of
degradation products.
The total organic carbon (TOC) was detected by a TOC
analyzer (TOC4100, Shimadzu).
(
was in a quartz sleeve and located at the center of the reactor. The
ꢁ
temperature of the reaction system was 25 C. The 50 mL SMN
solutions were put in quartz tubes around and equidistant from
the Hg lamp, irradiated under UV over a period of 60 min, and
aliquots were withdrawn for analysis at scheduled time intervals.
To investigate the hydrolytic degree of SMN during photolysis,
non-irradiated experiments were conducted simultaneously. The
3
. Results and discussions
3.1 The determination of initial concentration of SMN and
the light source
SMN solutions were wrapped in an aluminum foil, and other The concentration of SMN in the aquatic environment
ꢀ
1
ꢀ1
conditions were kept the same as those for the unwrapped (ng L to mg L ) is relatively low. A pre-concentration step is
36,37
samples. All experiments were performed in triplicate.
required prior to quantication using HPLC.
Although
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previous research has reported that the solid phase extraction and the degradation ratios were 72%, 74%, and 75% in the
SPE) technique can improve the sensitivity of the analytical presence of 2-propanol, NaN , and p-quinone, respectively. The
(
3
1
ꢀ
methods by HPLC, the use of SPE for quantication of sulfon- results indicated that cOH, O , and O c were formed during
amides in aqueous samples has not been examined in detail.
2
2
36
SMN photolysis in pure water via the photosensitization process
During photodegradation of ACs, different initial concentra- of SMN, and the impact of cOH was more noticeable than that of
1
ꢀ
2 2
O and O c on the photolysis of SMN, which were consistent
tions of ACs caused different amounts of reactant molecules to
be photo-excited; this affected the rate of photolysis, but had no with the early research conclusions that hydroxyl radicals
2,6
impact on the reaction mechanism. Furthermore, Nassar played a critical role in the photochemical transformation of
et al. evaluated the photodecomposition of four different ACs organic pollutants.
38
42–44
However, the ratio and rate constants (k)
under irradiation of sunlight and a single wavelength UV source of SMN molecules had no evident decline aer adding scaven-
1
ꢀ
(
254 nm). The results showed that UV irradiation was effective gers for cOH, O , and O c ; this implied that direct photolysis
2 2
to degrade these drugs, and photolysis of ACs under simulated was the main path for photochemical behavior of SMN in pure
solar irradiation was caused by AC absorbing the UV light from water.
sunlight.
Due to all the abovementioned reasons, the selected initial on SMN photolytic reactions, comparative trials under deoxy-
concentration of SMN and the light source in this study were genation by N -sparging, oxygen lling, and adding sorbic acid
To further understand the effect of self-sensitized behavior
2
ꢀ
1
3
2
0 mg L and a 300 W high pressure Hg lamp, respectively. as a scavenger to quench SMN* were carried out in SMN
Although these two parameters were signicantly different from photolysis. The results are also shown in Fig. 2. The value of k
ꢀ
2
ꢀ1
ꢀ2
ꢀ1
ꢀ1
the antibiotic environment in natural water bodies, the increased to 3.31 ꢂ 10 min from 2.75 ꢂ 10 min aer
ꢀ
2
conclusions of this study could provide theoretical reference for deoxygenation, but decreased to 1.99 ꢂ 10 min and 1.60 ꢂ
ꢀ2
ꢀ1
the photochemical behavior of antibiotic residues in natural 10 min aer oxygen lling and addition of sorbic acid,
3
aquatic environments.
respectively (Fig. 2). Removal of
sparging could lead to a decreased formation rate of O
2 2
O from solution by N -
1
2
as well
3
as lower quenching rate of SMN*, and oxygen lling was an
opposite process to strengthen the self-sensitization. Based on
these results, we could state that the self-sensitized behavior
had an inhibiting effect on SMN photolysis. The reduction of k
aer adding sorbic acid further demonstrated that the steady
3.2 Photolysis of SMN in pure water
To investigate the photolytic process of SMN, we explored SMN
photolysis in pure water under 300 W UV radiation. The
experiments have been carried out at pH 7 at which the neutral
39
form of SMN dominates (>80%). Previous studies found that
photo-excited pharmaceutical compounds, formed by the
photosensitization process, interacted with molecular oxygen to
3
state concentration of SMN* was very important during SMN
3
photolysis, and the direct decomposition of SMN* was the
main pathway of SMN photolysis.
1
ꢀ
40
form cOH, O
we selected 2-propanol, sodium azide (NaN
2
, and O
2
c . During the photolytic experiment,
), and p-quinone as
To investigate the hydrolysis of SMN, non-irradiated experi-
ments were conducted simultaneously. No obvious loss of SMN
was observed in pure water solution of SMN; this implied that
hydrolysis of SMN was negligible during the SMN aquatic
environmental behavior.
3
1
ꢀ
41
scavengers to quench cOH, O , and O c respectively. SMN
2
2
degradation and the logarithm of relative SMN concentration
versus irradiation time are displayed in Fig. 2. SMN photolysis
followed pseudo-rst-order kinetics in pure water. The degra-
dation ratio of SMN achieved a value of 78% aer 60 min irra-
diation, and the degradation rate constant (k) of SMN was
3
.3 Pathways of SMN photolysis in pure water
ꢀ
2
ꢀ1
2
.58 ꢂ 10 min in pure water. The values of k decreased to In our experiment, we found that the main path of formation of
2 ꢀ1 ꢀ2 ꢀ1 3
ꢀ2
2
.19 ꢂ 10ꢀ minꢀ1, 2.33 ꢂ 10 min , and 2.37 ꢂ 10 min
SMN photoproducts in water was SMN* decomposition by
Fig. 2 Degradation ratio (A) and kinetics (B) of SMN photolysis under different conditions.
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3.4 Effect of soluble substances on SMN photolysis
Halide ions, nitrate, and dissolved organic matter (DOM) are
ordinary soluble substances in natural waters, which can be
6,45
converted into free radicals under sunlight.
Indirect photo-
lytic behaviors of SMN caused by soluble substances are
common photodecomposition behavior of SMN. We selected
ꢀ
ꢀ
Cl , NO , and FA as probes to explore the effect of photoactive
3
substances on SMN photolysis in water. In view of SMN having
been frequently detected in both freshwater and seawater, the
ꢀ
ꢀ
concentrations of Cl , NO , and FA were set according to their
3
6
content in natural waters (Table 1).
As the content of Cl is very low in freshwater and exhibits
gradient variation in estuarine waters, we set different Cl
ꢀ
ꢀ
concentrations (100, 300, and 500 mM) relevant to estuarine
6
and sea conditions to explore the effects on the photolysis of
SMN. Experiments showed that the k values of SMN photolysis
ꢀ
increased with an increase in Cl concentration (Fig. 5). In pure
Fig. 3 TOC with the SMN photolysis under UV irradiation.
3
water, the main photolysis pathway is SMN* disintegration to
photoproducts by direct photolysis. Hence, the steady state
concentration of triplet-excited state (Tss) is very important for
degradation. Previous studies indicated the Tss of SAs* was
direct photolysis. Moreover, species of ROS formed by self-
sensitized SMN could degrade SMN, which could not be
ignored during SMN photochemical behaviors. To determine
the mineralization degree of SMN and identify the products of
SMN photolysis, TOC was monitored along with SMN photolysis
under UV irradiation. The solutions obtained aer 60 min of
SMN photodecomposition were analyzed using LC-MS/MS. As
shown in Fig. 3, no signicant removal of TOC was observed
even aer 60 min; this meant that most of the organic carbon
was still retained although the SMN photolysis under UV was
effective.
3
4
6,47
higher at a higher ionic strength.
uate the electron accepting ability of the SAs*.
natural water (pH ¼ 6.5–8.5) as two different species: neutral
T
E was employed to eval-
3
48,49
SMN exists in
0
ꢀ
39
state (SMN ) and anionic state (SMN ). The E values are
T
3
ꢀ
ꢀ
ꢀ
2
2
.45 V and 2.25 V, which let SMN* oxidize Cl [E(Cl c /Cl ) ¼
6
3
ꢀ
2
.0 V]. Thus, the deactivation of SMN* induced by Cl may
ꢀ
lead to the formation of Cl
2
c
and halogenated intermediates
via the pathway demonstrated in Scheme 2. Triplet-induced
halogenation of SMN will increase the ionic strength effect,
and at the same time reduce the ground state SMN; this can
explain why the rate of SMN photolysis is faster in the presence
The specic ion mass spectra of major intermediates are
depicted in Fig. 4. There were three main photoproducts that
could be observed: SMN-1, SMN-2, and SMN-3 with m/z of 124,
ꢀ
of Cl as compared to that in the presence of pure water.
ꢀ
The role of NO3 was paradoxical: either its absorption
215, and 295, respectively. cOH radicals formed during the self-
(
200–400 nm) covering the SMN absorption range (200–330 nm)
resulted in a decrease of photolytic rate of SMN or produced
reactive species (e.g., cOH, NO c) via sensitization effects
eqn (1)–(3)); this could increase the photolytic rate of SMN.
sensitization process of SMN attacked the benzene ring or the
dimethyl pyrimidine group of SMN; this resulted in the
formation of SMN-3. SMN-2 was derived from SO removal from
2
2
(
SMN. SMN-1 was produced from the broken carbon–nitrogen
bond via intermediate photolysis. These results were consistent
with the previous research on photolytic mechanism of
Â
Ã
ꢀ
hv
ƒƒƒ ƒƒ ! NO
ꢀ
NO
3
3
*
(1)
(2)
2,38
sulfonamide antibiotics.
The proposed pathways of SMN
Â
ꢀÃ
H2O
c
2
cꢀ
c
ꢀ
NO
3
*/NO þ O ƒƒƒ ƒƒ ! NO þ cOH þ OH
photolysis in pure water are displayed in Scheme 1.
2
3ꢀ
2
ꢀ
3
+ O( P)
[
NO ]* / NO
(3)
Fig. 4 Major photolytic intermediate ion mass spectra of SMN.
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Scheme 1 Proposed degradation pathways of SMN in pure water.
ꢀ
3ꢀ
ꢀ1
Table 1 Concentrations of Cl , NO , and TOC for freshwater and (3.0, 6.0, and 9.0 mg C L ) were added to pure water to inves-
seawater
tigate the effect of DOM on the photolysis of SMN. As shown in
Fig. 8, photolysis rate constants of SMN decreased with an
increase in FA concentration.
Parameters
Freshwater
Seawater
ꢀ
Cl (mM)
3.4
93.6
4.0
469
29.0
6.1
These conclusions were not consistent with previous studies.
ꢀ
3
NO
3
(mM)
Others found that FA formed the excited state FA* and ROS
ꢀ1
TOC (mg C L
)
under irradiation, which further induced the degradation of
organic compounds and was a signicant promoter of organic
51
compound photolysis. The photolysis pathway of FA in SMN
solution is displayed in eqn (4)–(6).
Our experimental results are shown in Fig. 6. In the
ꢀ
concentration range of NO
3
in natural water, the existence of
ꢀ
hv
NO
3
led to the increase of SMN photolytic rate. To know
3
FA ƒƒƒ ƒƒ ! FA*
(4)
(5)
ꢀ
whether cOH formed by NO
3
sensitization played an important
role during SMN photolysis, we added 2-propanol to the reac-
tion system; the experimental ndings are exhibited in Fig. 7.
The k value of SMN photolysis decreased than that in pure
H2O
3
FA* þ O
2
ƒƒƒ ƒƒ ! FA þ ROS
FA + ROS / photolytic products
(6)
water. The experiment conrmed that more production of cOH
ꢀ
3ꢀ
in the presence of NO
3
as compared to that in pure water was
In our case, the inhibiting effect of FA on SMN photolysis
the main cause of NO promoting the SMN photolysis.
FA is a ubiquitous DOM in natural waters that has strong
photochemical activity. In this study, following the concentra-
might have two factors: one, the steady state concentration of
3
SMN* was very important for SMN photolysis; the direct
3
50
decomposition of SMN* has been proven to be the main
tion of DOM in natural waters, different concentrations of FA
pathway of SMN photolysis. Therefore, the competitive
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ꢀ
Fig. 5 Effects of Cl on degradation ratio (A) and kinetics (B) of SMN photolysis.
3
absorption for photons (200–400 nm) by FA could form FA*,
3
3
which inhibited the production of SMN*. Another, the FA*
was quenched by dissolved oxygen in the water. This process
3
takes FA* back to the ground state (FA). FA not only acts as
52
a ROS scavenger, but also inhibits SMN from absorbing the
photons.
Scheme 2 Reaction pathway of SMN induced by chloride ion.
3ꢀ
Fig. 6 Effects of NO on degradation ratio (A) and kinetics (B) of SMN photolysis.
3ꢀ
Fig. 7 The k values of SMN photolysis after adding NO and 2-propanol.
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Fig. 8 Effects of FA on degradation ratio (A) and kinetics (B) of SMN photolysis.
Scientic Research Foundation of Leshan Science & Technology
Bereau, Sichuan Province, China (Grant No. 15ZDYJ0144).
4
. Conclusion
In this study, we investigated the photolytic behaviors of SMN
via adding different scavengers to quench active species in pure
3
water. The research indicated that SMN* played an important
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3
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ignored during study of SMN photochemical behaviors. The
main photoproducts of SMN were identied by LC-MS/MS,
which showed that SMN could not be mineralized although
photolysis under UV was effective. Via studying the effects of
ordinary soluble substances in natural water on SMN photol-
ysis, we also found that the rate of SMN photolysis was faster in
ꢀ
ꢀ
the presence of Cl and NO . The triplet-induced halogenation
3
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ꢀ
ꢀ
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There are no conicts of interest to declare.
7
8
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