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Photochemical reaction of tricresyl phosphate (TCP) in aqueous solution: Influencing factors and photolysis products
Journal Pre-proof Photochemical reaction of tricresyl phosphate (TCP) in aqueous solution: Influencing factors and photolysis products Shibin Sun, Jingqiu Jiang, Hongxia Zhao, Huihui Wan, Baocheng Qu PII:
S0045-6535(19)32210-6
DOI:
https://doi.org/10.1016/j.chemosphere.2019.124971
Reference:
CHEM 124971
To appear in:
ECSN
Received Date: 9 July 2019 Revised Date:
17 September 2019
Accepted Date: 24 September 2019
Please cite this article as: Sun, S., Jiang, J., Zhao, H., Wan, H., Qu, B., Photochemical reaction of tricresyl phosphate (TCP) in aqueous solution: Influencing factors and photolysis products, Chemosphere (2019), doi: https://doi.org/10.1016/j.chemosphere.2019.124971. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.
Photochemical reaction of tricresyl phosphate (TCP) in aqueous solution: Influencing factors and photolysis products
Shibin Sun, 1 † Jingqiu Jiang, 1, † Hongxia Zhao, *, † Huihui Wan ‡, Baocheng Qu *,§
†
Key Laboratory of Industrial Ecology and Environmental Engineering (Ministry of
Education), School of Environmental Science and Technology, Dalian University of Technology, Dalian, 116024, China ‡
College of Chemical Engineering, Analytical Center, Dalian University of Technoloy,
Dalian 116024, China §
College of Marine Technology and Environment, Dalian Ocean University, Dalian
116024, China
#
Corresponding author phone: +86-411-84707965; Fax: +86-411-84707965;
Email: [email protected] #
Corresponding author phone: +86-411-84740281; Fax: +86-411-84740281;
Email: [email protected] 1
These authors contributed equally to this study and share first authorship
hv
0.0
directt indirectt
products
-0.4
TCP
3O
Fe3+ H2O NO2-/NO3-
2
1O
HA … 2
ln(Ct/C0)
O O O P O
-0.2
-0.6 -0.8 3+
0 µmol/L Fe 3+ 50 µmol/L Fe 3+ 10 µmol/L Fe
-1.0
· OH
-1.2 0
2
4
6 8 time (h)
10
12
1
Abstract
2
Organophosphate triesters (OPEs) have caused great concern as a class of emerging
3
environmental contaminants due to their widespread use and their toxicity to
4
organisms. However, the phototransformation behavior of OPE is still not fully
5
understood, which is important for understanding their environmental fate. In the
6
present study, the photodegradation of tricresyl phosphate (TCP), one of the most
7
widely detected OPEs in aqueous environments, was investigated including the direct
8
photolysis and in the presence of several natural water factors, NO2-, Fe3+ and humic
9
acid. The degradation process followed the pseudo-first-order kinetics, with rate
10
constant increasing slightly with increasing initial TCP concentration. The presence of
11
NO2- and Fe3+ was observed to promote the photochemical loss of TCP, while humic
12
acid played a negative role on TCP transformation. Electron spin resonance (EPR)
13
analysis showed that carbon-centered radical was produced in the photolysis process
14
of TCP, and hydroxyl radical contributed to the promotion of rate constant for Fe3+
15
and
16
HPLC-LTQ-Orbitrap MS analysis, and the possible degradation pathways of TCP
17
were proposed. These findings provide a meaningful reference for the fate and
18
transformation of OPEs in natural water.
NO2-.
Four
photolysis
products
were
tentatively
identified
by
19 20
Keywords: tricresyl phosphate (TCP); environmental factors; kinetics; radical
21
analysis; photolysis products
22 1
23
1. Introduction
24
Organophosphate triesters (OPEs) are a class of emerging environmental pollutants
25
which have been widely used as flame retardants and plasticizers in many industrial
26
and household materials, such as furniture, textiles, building materials and electronics
27
(Brandsma, et al., 2013; van der Veen, et al., 2012; Shi, et al., 2016). Because
28
polybrominated diphenyl ethers have been phased out, OPEs have become one of the
29
most frequently used alternative flame retardants. It has been reported that the annual
30
global production of OPEs currently reaches approximately 200 kt (Greaves, et al.,
31
2014), and nearly more than 70 kt being produced in China (Wei, et al., 2015). As
32
OPEs are not chemically bounded to the final products, they can be released into
33
environment via volatilization, abrasion, and leaching process (Rodriguez, et al., 2016;
34
Marvin, et al., 2016; Li et al., 2018). OPEs have been reported to be present in aquatic
35
environments including river waters, seawaters, ground waters, and even drinking
36
waters, with concentrations ranging from ng/L to ug/L (van der Veen and Boer, 2012;
37
Shi, et al., 2016; Wang et al., 2015; Cristale, et al., 2013). For example, Wang et al.
38
reported that in North China, the total concentrations of OPEs ranged from 9.6 to
39
1549 ng·L-1 in 40 major rivers draining into the Bohai Sea (Wang et al., 2015). In the
40
River Aire (UK), OPEs were detected at concentrations ranging from 6.3 to 26050
41
ng·L-1 (Cristale, et al., 2013). The ubiquitous occurrence and high levels of OPEs in
42
the water environments has sparked considerable interest of researchers.
43
An increasing number of toxicological studies have shown that OPEs have negative
44
effects on human health and are toxic to aquatic organisms (van der Veen and Boer, 2
45
2012). For instance, the chloroalkyl OPEs such as tri(2-chloroethyl) phosphate (TCEP)
46
and tri(dichloropropyl) phosphate (TDCP) have been proved to be carcinogenic and
47
neurotoxic and TCEP was considered to be environmentally persistent (Hou, et al.,
48
2016; Wang et al., 2015; Yang et al., 2014). Du et al. (2015) have reported two aryl
49
OPEs which have heart developmental toxicity by disturbing expressions of
50
transcriptional regulators in zebrafish. In addition, tricresyl phosphate (TCP) has been
51
shown to be potential thyroid hormone disruptors and could cause reproductive
52
toxicity (van der Veen and Boer, 2012; Zhang et al., 2016). Accordingly, it is essential
53
to understand the fate of OPEs in aquatic environments for their ecological risk
54
assessment.
55
Photochemical degradation is considered to be an important transformation pathway
56
of organic pollutants in aquatic environment, which may occur by either direct or
57
indirect photolysis. Direct photolysis is the result of light absorbance by pollutants
58
causing their molecular degradation. Some aryl OPEs have considerable sunlight
59
absorbance in the ultraviolet region (290-400 nm) (Cristale et al., 2017), it is believed
60
that direct photolysis plays a crucial role for the photodegradation of OPEs in aquatic
61
environment. For indirect photolysis, NO2−, Fe3+ and humic substances play crucial
62
roles, which are ubiquitous in surface water and absorb solar radiation to reach an
63
excited state, subsequently generating free radicals (Burbano et al., 2005, Cristale et
64
al., 2017). In recent years, researchers have focused on the photodegradation of OPEs
65
using advanced oxidation processes (AOPs) such as ozonation, UV/H2O2, UV/TiO2,
66
UV/persulfate (PS) for their removal in effluents (Watts and Linden 2009; 3
67
Antonopoulou et al., 2016; Antonopoulou et al., 2017; Liu et al., 2018; Ou et al., 2017;
68
Yuan et al., 2015). Nevertheless, until now, only limited studies have reported the
69
environmental fate of OPEs in natural water. Regnery and Püttmann (2010) reported
70
that upon exposure to sunlight, the concentrations of three alkyl OPEs decreased
71
significantly in lake water while no degradation was observed in ultrapure water.
72
Another study pointed out that humic acid inhibited OPEs photodegradation mediated
73
by singlet oxygen while inorganic constituents in river water enhanced their removel
74
(Cristale et al., 2017). These processes contributed the indirect photolysis of OPEs to
75
constituents in natural water. However, there are still some other constituents whose
76
effects on phototransformation of OPEs are not characterized, such as Fe3+ and NO2-.
77
Also, the photodegradation products should be paid more attention because some
78
products are more toxic than their parent compounds (Su et al., 2014).
79
To fill the knowledge gap mentioned above, photodegradation experiments were
80
performed to figure out the transformation kinetics and mechanisms of OPEs under
81
simulated solar light irradiation. TCP, one of the most widely detected OPEs in
82
aquatic environments (van der Veen and Boer, 2012), was selected as the model
83
compound. Direct photodegradation of TCP under different light intensities and
84
wavelengths were firstly studied. Then the influences of different initial
85
concentrations and various environmental factors, including Fe3+, NO2-, and humic
86
acid (HA) on the photodegradation efficiency of TCP were investigated. In addition,
87
photodegradation products of TCP after irradiation were identified and possible
88
degradation pathways were proposed. Electron paramagnetic resonance spectroscopy 4
89
(EPR) techniques was used to study the mechanism for the effects of constituents on
90
photodegradation processes and to capture the intermediates for deducing the
91
degradation pathways. These findings advance our fundamental understanding of fate
92
of OPEs in aquatic environments and ecological risk of OPEs contamination in sunlit
93
water bodies.
94 95
2. Materials and methods
96
2.1 Chemicals
97
Analytical standard TCP (mixture of isomers, CAS number: 1330-78-5, purity ≥ 99%)
98
was purchased from Aladdin Reagent (Shanghai, China), and its molecular formula
99
and weight are C21H21O4P and 368.1178 g/mol, respectively. Standard solution of TCP
100
was prepared in methanol at the concentration of 200 ppm and then stored at 4 oC in
101
the dark. HPLC grade methanol and dichloromethane (DCM) were obtained from
102
Sigma Aldrich (USA). FeCl3, NaNO2 were of reagent grade. Humic acid (CAS:
103
1415-93-6, FA ≥ 98%) was purchased from the International Humic Substance
104
Society (IHSS). Spin traps were used to capture the reactive oxygen species (ROS).
105
5,5-dimethyl-1-pyrroline-Noxide (DMPO) was obtained from Dojindo Laboratories
106
(Shanghai, China). All the solutions used in the experiments were prepared using
107
Milli-Q water (18 MΩ·cm) obtained from OKP ultrapure water system (Shanghai
108
Lake-core Instrument Co., Shanghai, China).
109 110
2.2 Irradiation 5
111
The photochemical experiments were performed in an XPA-1 photochemical reactor,
112
which the schematic diagram was shown in Figure S1 (Nanjing Xujiang
113
Electromechanical Plant, Nanjing, China). A 500 W Xe lamp, a 1000 W Xe lamp and
114
a 800 W mercury (Hg) lamp coupled with the 290 nm filters were used to obtain
115
different light intensities. The 800 W Hg lamp surrounded by 290 nm and 340 nm
116
cut-off filters were used to simulate the natural sunlight and UV-A irradiation sources,
117
and a 254 nm low-pressure Hg lamp was used as the UV-C irradiation source. The
118
light intensities in the center of the solutions were 4.67 mW/cm2, 15.52 mW/cm2 and
119
61.60 mW/cm2 for 500 W Xe lamp, 1000 W Xe lamp and 800 W Hg lamp evaluated
120
by an optical sensors (RAMSES, TriOS).
121
To keep the temperature inside the reactor stable, the lamps was enclosed in a cooling
122
well via the circulation of cold water. A 30 mL aliquot solution (50 ppb -200 ppb) was
123
placed in a quartz tube in a merry-go-round apparatus inside the reactor for irradiation.
124
The initial concentration of TCP was 100 ppb unless specified, certain concentrations
125
of FeCl3, NO2- and HA were added into the reaction solutions. Dark controls were
126
performed under the same experimental condition. At defined time intervals, samples
127
were transferred to glass vials and stored at -20
128
photochemical experiments were performed in triplicate.
o
C until analysis. All the
129 130
2.3 Analytical methods
131
Matrix matched calibration curves were prepared using Milli-Q water with TCP
132
ranging from 1 to 200 ppb. The concentration of TCP was analyzed by ultra-high 6
133
performance liquid chromatography coupled to a triple quadrupole detector
134
(UPLC-MS/MS) (TSQ Quantum Ultra, Thermo, USA). The chromatographic column
135
was a SB-C18 (2.1 mm×150 mm, particle size 5 µm, Agilent, USA) and the
136
temperature was set at 30 oC. The mobile phase was 0.2% formic acid in UPLC water
137
(15%) and methanol (85%) and the flow rate was set at 0.2 mL·min-1. The injection
138
volume was 20 µL. TCP was analyzed under positive electrospray ionization (ESI+)
139
by multiple reaction monitoring (SRM) mode. The optimized collision energy were
140
29 V for the transition of m/z 369.1 → 165.8 and 28 V for the transition of m/z 369.1
141
→243.0. All data were acquired and processed by Xcalibur software.
142
Screening of photolysis products was carried out using an LTQ Orbitrap Velos mass
143
spectrometer (Thermo Fisher Scientific, Bellefonte, PA, USA) equipped with an ESI
144
source in both positive and negative ionization mode. The MS conditions were
145
optimized as follows: sheath gas/aux gas/sweep gas 30/10/0 arb, capillary temperature
146
350 0C, capillary voltage 39 V, tube lens voltage 0 V, and spray voltage 3.5 kV. The
147
system was operated in the full spectral acquisition mode in the mass range of m/z
148
80-600 with a mass resolution of 100,000.
149
The LC system was equipped with an Ultimate XB-C18 column (2.1 mm×100 mm,
150
5µm particle, Thermo Fisher Scientific, USA) and the mobile phase was methanol (A)
151
and water (B). A gradient run was used as follows: 30% A to 60% A in 5 min, to 80%
152
A in 10 min, to 100% A in 5 min, hold 5 min, to 30% A in 5 min, equilibrate to 30% A
153
during 5 min. The flow rate was 200 µL/min and the injection volume was 5 µL. A
154
constant temperature of 30 oC was kept during analysis. All data were acquired and 7
155
processed by Xcalibur software.
156 157
2.4 Electron paramagnetic resonance (EPR) analysis
158
Experiments were still conducted in the XPA-1 photochemical reactor with the same
159
800 W Hg lamp equipped with the 290 nm cut-off filters. All the solutions were
160
prepared in the dark at room temperature (25 oC) and were loaded into high purity
161
capillary quartz tubes immediately after mixing the solution (Fe3+, NO2-, or HA) with
162
TCP and the spin trap DMPO. The final concentration of DMPO was 160 mM.
163
Samples were collected at specific irradiation time to measure the ROS signal by EPR,
164
the detailed information were shown in Supporting Information. EPR experiments
165
were performed on a Bruker EXM A-200 spectrometer (Bruker, Bermen, Germany)
166
and the operating parameters were as follows: central field, 3398 G; microwave
167
frequency, 9.45 GHz (X-band); microwave power, 1.69 Mw; scanning width, 100 G;
168
and scanning frequency, 100 kHz.Simulation of EPR data was accomplished using
169
WinSim software (NIEHS) and Bruker WinEPR. The Spin Trap Database (NIEHS)
170
was referred to in order to interpret and simulate the EPR spectra (Suh et al., 2009)
171
and the simulated spectra correlated well with the corresponding experimental spectra
172
(correlation coefficient > 0.99).
173 174
3. Results and discussion
175
3.1 UV light irradiation of TCP in milli-Q water
176
In surface waters, the photochemical transformation of pollutants could take place by 8
177
direct or indirect photoreaction. Direct photolysis take place when radiation
178
absorption by a molecule triggers its transformation. Figure S1 showed that TCP had a
179
broad absorption spectrum between 220 nm and 320 nm, and the local maximum
180
absorption wavelength was approximately 268 nm. For direct photolysis of TCP, the
181
influences of different light intensities and wavelength were evaluated. Firstly, a 500
182
W Xe lamp, 1000 W Xe lamp and 800 W Hg lamp were used as the light source, the
183
photolysis of TCP underwent with 290 nm cut-off filters and the corresponding light
184
intensities were 4.67, 15.52, 61.60 W/m2 respectively. As shown is Figure 1, the
185
degradation rate increased with the light intensity. Half-life times of 130.8 and 68.0 h
186
were obtained under 500 W Xe lamp and 1000 W Xe lamp, respectively. Under 800
187
W Hg lamp, the degradation rate was significantly increased and the half-life time
188
was 8.75 h. Dark control samples were analyzed during the same period of time and
189
the results showed no obvious degradation without exposure to UV light. The
190
irradiation spectrum of sunlight is similar with that of 1000 W Xe lamp shown in
191
Figure S3 in supporting information, and the light intensity was 20.60 W/m2 for
192
sunlight which is a litter higher than that of 1000 W Xe lamp. Thus, it is deduced the
193
half-life time of TCP for direct photolysis in natural water was close to 68.0 h,
194
meaning that the direct photolysis of TCP under sunlight plays minor role in the
195
photodegradation process in water environment, which is similar as previous study
196
(Cristale et al., 2017).
9
0.0
ln(Ct/C0)
-0.4 -0.8 -1.2
Dark control 2 4.67 mW/cm 2 15.52 mW/cm 2 61.60 mW/cm
-1.6 -2.0 0
198
5
10
15
20
25
time (h)
197
Figure 1. Photodegradation of TCP at different light intensities (C0 = 200 ppb)
199 200
An 800 W Hg arc lamp with 340 nm cut-off filter was used to simulate the UV-A light
201
and a low-pressure Hg light (254 nm) was used to verify the influence of UV-C on the
202
photodegradation of TCP. Figure 2 presented the photodegradation of TCP under
203
UV-A and UV-C irradiation in Milli-Q water. The 800 W Hg arc lamp with 290 nm
204
cut-off filter contains UV-A and UV-B region and obvious degradation was observed
205
for TCP under this condition. As shown in Figure 2, no degradation of TCP under
206
UV-A range irradiation, which indicated that UV-B plays a decisive role in the
207
photodegradation of TCP in natural environment. For UV-C region, 80% of TCP was
208
significantly removed after 20 min of irradiation. UV-C irradiation proved to be the
209
most effective light, which most probably due to its high absorbance in this region to
210
induce the direct photolysis of TCP. In the study of Cristale et al (2017), UV-C light
211
was also used to study the degradation efficiency of nine OPEs, and three of them
212
with aryl groups showed 100% removal after 10 min, which is similar with the result
213
in our study.
10
214 215
Figure 2. Photodegradation of TCP in Milli-Q water under UV-A and UV-C irradiation (C0 =
216
200 ppb)
217 218
3.2 Effects of initial concentration and typical environmental factors on the
219
phototransformation of TCP
220
The phototransformation of TCP under the simulated sunlight irradiation were further
221
investigated under various conditions with different TCP concentrations, together with
222
different concentration of Fe3+, NO2- and HA. The photodegradation rate constant and
223
half-life was determined using pseudo-first-order kinetics, C = C0e-kt. The reaction
224
rate constant (k) and the corresponding correlation coefficient (R2) were given in the
225
table S1.
11
0.0
0.0
a
-0.2
b
-0.2 -0.4 ln(Ct/C0)
ln(Ct/C0)
-0.4 -0.6 -0.8
0
2
-
0 µmol/L NO2
-
20 µmol/L NO2
-1.2 4
6 8 time (h)
10
-
100 µmol/L NO2
0
12
2
4
6 8 time (h)
10
12
0.0
0.0
c
-0.2
d
-0.2
-0.4
-0.4 ln(Ct/C0)
ln(Ct/C0)
-0.8 -1.0
50 ppb 100 ppb 200 ppb
-1.0
-0.6
-0.6 -0.8
-0.6 -0.8 0 ppm HA 2 ppm HA 10 ppm HA
3+
0 µmol/L Fe 3+ 50 µmol/L Fe 3+ 10 µmol/L Fe
-1.0 -1.2 0
2
4
-1.0 -1.2 6 8 time (h)
10
0
12
226
2
4
6 8 time (h)
10
12
227
Figure 3. The effects of environmental factors on the photodegradation of TCP (a) initial
228
concentration, (b) NO2-, (c) Fe3+, (d) HA; the concentration of TCP for the experiments in (b),
229
(c), and (d) is 100 ppb
230 231
3.2.1 Effect of initial concentration of TCP
232
Pollutant concentration may be an important factor concerning the photoreaction
233
process. As TCP was detected in natural surface waters at various concentrations
234
ranging from ppt to ppb level (Lee et al., 2016; Kim and Kannan 2018), it’s
235
significant to investigate the effect of TCP concentration on its photolysis behavior
236
under solar irradiation. In this work, the effect of initial concentration of TCP on the
237
direct degradation efficiency was investigated with three different initial
238
concentrations of 50 ppb, 100 ppb and 200 ppb. As shown in Figure 3a, the k values 12
239
increased with increasing initial concentrations of TCP, they were 0.0509 h-1 for 50
240
ppb, 0.0627 h-1 for 100 ppb and 0.0789 h-1 for 200 ppb, respectively. The
241
photodegradation half-life (t1/2) of TCP varied from 13.6 to 8.8 h with the
242
concentration ranging from 50 to 200 ppb. When the concentrations of TCP in the
243
natural waters are lower than the present study, it is expected that TCP would have a
244
longer photodegradation half-life than being discussed in this study.
245 246
3.2.2 Effect of NO2-
247
Nitrate and Nitrite are ubiquitous constituents coexisting in natural waters. The effect
248
of nitrate on the photodegradation of pollutants in water is mainly realized by the
249
production of nitrite (Beitz et al., 1999; Arakaki, 1999). When compared to nitrate,
250
nitrite absorb a larger fraction of the sunlight spectrum, making nitrite is an important
251
ROS source (Mack, 1999), which may contribute to the photodegradation of
252
pollutants. As illustrated in Figure 3b, the degradation rate of TCP was obviously
253
enhanced by the addition of NO2-, the k values were 0.072 h-1 for control, 0.0818 h-1
254
for 20 µM (NO2-), and 0.0959 h-1 for 100 µM (NO2-). The transformation rate constant
255
increased with increasing concentration of NO2-.
256
Numerous studies have shown that NO2- can generate •OH and •NO2 (Calza, et al.,
257
2012; Vione et al., 2002) under UV light irradiation through the following equations
258
(1)-(3):
259
NO2- + hv (+H+) → •NO + •O-
260
(1)
•O- + H2O → •OH + OH-
(2) 13
261
•OH + NO2- → •NO2 + OH-
(3)
262
With the increase of the concentrations of NO2-, the concentration of generated •OH
263
and •NO2 in the system also increased. The generated free radicals may attack the
264
phenyl structure (Ahn et al., 2003; Poerschmann et al., 2009), thus affecting the
265
photolysis process of TCP and accelerating the photodegradation rate of TCP in water.
266 267
3.2.3 Effect of Fe3+
268
Iron is another significant component with extensive presence in natural aquatic
269
environment at concentrations ranging from 10-7 to 10-4 M (Zhao et al., 2014).
270
Considerable literatures have proved that iron plays an important role in many
271
photochemical reactions of organic compounds relating to ROS production, such as
272
photocatalytic reaction and Fenton reaction (Voelker et al., 1997; Wang et al., 2017).
273
In this study, we used FeCl3 as the source of Fe3+ to investigate its influences on the
274
photodegradation of TCP. As illustrated in Figure 3c, the degradation rate of TCP was
275
enhanced by adding Fe3+ to the system. The k values for TCP were 0.0675 h-1 for
276
control, 0.097 h-1 for solution containing 10 µM Fe3+ and 0.0822 h-1 for solution
277
containing 50 µM Fe3+, respectively. The existing studies proved that Fe3+ could
278
produce •OH via the following reactions (Zhao et al., 2014; Peng et al., 2016; Neamtu
279
and Frimmel, 2006):
280
Fe3+ + H2O ↔ Fe(OH)2+ + H+
(4)
281
Fe(OH)2+ + hv → Fe2+ + •OH
(5)
14
282
Due to the increased •OH production with higher Fe3+ concentration, a positive role of
283
Fe3+ to the photodegradation of TCP was expected. While under our experimental
284
conditions, it’s interesting to find that although Fe3+ did enhance the degradation rate
285
of TCP at concentrations of 10 µM and 50 µM, but the degradation rate at 50 µM is
286
lower than that at 10 µM. Here, the photodegradation rate increased by 43.7% with
287
increasing Fe3+ concentration from 0 to 10 µM but decreased about 15.3% by
288
increasing Fe3+ concentration from 10 µM to 50 µM. Since the photochemical
289
properties of Fe3+ can be strongly influenced by pH levels in natural waters, the result
290
can be attributed to pH value (6.5 ± 0.2) of the solutions used in our experiments. The
291
study of Chowdhury and co-workers (2011) has demonstrated that Fe(OH)2+ is the
292
predominant photoreactive species among the Fe3+-aquo complexes. Moreover, it was
293
reported that the photoactivity of Fe(OH)2+ in aqueous solutions was restrained at pH >
294
5.0 (Zhao et al., 2014) and at a pH of 7.0 (±0.1), Fe3+ at different concentrations
295
revealed an inhibition effect (Zhou et al., 2010). Thus, that’s may be the reason that
296
TCP has a higher k value at the addition of 10 uM Fe3+ when compared to a higher
297
concentration of Fe3+.
298 299
3.2.4 Effect of HA
300
Numerous studies have demonstrated that HA play a significant role in photochemical
301
degradation of organic pollutants in aqueous systems and the dual behavior of HA as
302
photosensitizer and redox inhibitor has been proved (Yu et al., 2010). In order to
303
evaluate the effect of HA on the photodegradation of TCP, experiments without HA 15
304
and with two alternative HA concentrations (2 ppm and 10 ppm) were carried out
305
under simulated solar light irradiation. The obtained pseudo-first-order kinetic
306
degradation curve is shown in Figure 3d. According to the results, the removal rate of
307
TCP was visibly inhibited by HA, with the k values decreasing from 0.00642 h-1 for
308
control to 0.0594 h-1 for 2 ppm HA and 0.0525 h-1 for 10 ppm HA. It was obvious that
309
he presence of HA decreased the photolysis rate constant of TCP, implying HA served
310
as an inhibitor rather than a photosensitizer. Recent studies have demonstrated that
311
HA may act as a light screening agent and free radical quencher to inhibit the excited
312
triplet-induced oxidation of several organic contaminants (Calza et al., 2014; Wenk et
313
al., 2014; Koumaki et al., 2015).
314
HA/1HA*/3HA*/HA+ + •OH→ oxidized HA
315
Cristale and co-works (2017) also found that HA acted as an inhibitor for aryl
316
phosphates, for different type and concentration of HA was selected, which is in
317
consistent with this study. For aryl OPEs, photodegradation decrease in HA solution
318
can be attributed to some inhibition of excited states of photosensitizer OPEs by
319
reducing moieties of HA, because they found that sunlight absorbing OPEs showed
320
photosensitizing properties through generating singlet oxygen. In the study of
321
Kouras-Hadel et al (2012), the inhibiting effect of HA mainly attributed to reduction
322
of quinonic moieties by reaction intermediate superoxide anions.
323
Moreover, the increasing concentration of HA may increase both the generated ROS
324
level and the photoactivated HA species (1HA*, 3HA*, etc.), but the photoactivated HA
325
species are recommended to be more efficient free radical quenchers than the parent 16
326
HA (Chen et al., 2013). On the other hand, HA could absorb the sunlight effectively
327
(data shown in Figure S4), thus the inhibition could be partially related with the
328
reduction of direct light absorption of TCP due to HA scattering and light absorption
329
competition. The light screening effect and increased free radical quenching
330
efficiency of HA may led to an overall inhibition on the photodegradation of TCP.
331 332
3.2.5 Identification of free radicals
333
In this study, EPR was performed to detect •OH or other possible radicals generated
334
during the irradiation process. As seen in Figure 4, no signal was observed in any
335
solution containing DMPO under dark condition, and phosphate buffered saline (PBS)
336
showed no peaks either after 10 min of irradiation. A characteristic sextet peaks of a
337
1:1:1:1:1:1 intensity can be observed after irradiation for 5 min, which is the typical
338
signal of carbon-centered radical (Zhao et al., 2015a). The observed six-line spectrum
339
had hyperfine spiltting constants of αN = 15.78 G, αH = 23.25 G, and g value of
340
2.00298 and NoH = 0.68, where NoH is the ration of the nitrogen-spiltting constant to
341
the hydrogen-spiltting constant (Li and Chignell, 1991). This similarity of this adduct
342
and previously reported 2-chlorophenyl-DMPO adduct (αN = 15.78 G, αH = 23.25 G,
343
NoH = 0.68) (Motten et al., 1985) suggested that it was formed by the reaction of the
344
spin trap with an aryl radical generated during the photolysis of TCP. It is reported
345
that
346
2,2’,4,4’-tetrabromobiphenyl ether could generate carbon-centered radical after
347
irradiation (Zhao et al., 2015a; Zhao et al., 2015b) and sunlight absorbing aryl
some
aryl
organic
pollutants
17
such
as
bromophenol
and
348
phosphate showed photosensitizing properties (Cristale et al., 2017), thus it is
349
speculated that TCP converted to the excited state after irradiation, C-H bond
350
homogenization occurred from the benzene ring structure of the excited TCP,
351
resulting in the formation of the carbon-centered radical.
352
At the same time, the characteristic quartet peaks of the DMPO/•OH adduct with a
353
1:2:2:1 intensity was detected for TCP solution alone and TCP solutions containing
354
Fe3+, NO2- and HA after irradiation. These results are in accordance with several
355
similar studies for the DMPO/•OH adduct (Stan et al., 2005; Huang et al., 2017),
356
confirming the generation of •OH radical in this study. Take TCP solution which
357
containing100 µM NO2- as an example, simulation of the experimental spectra yielded
358
hyperfine splitting constants were: DMPO/•OH (αN = αH = 15.07 G) (see Figure 5).
359 18
360
Figure 4. EPR spectra produced by DMPO adducts in (a) TCP alone, (b) TCP + NO2- solution,
361
(c) TCP + Fe3+ solution, (d) TCP + HA solution.
362 363
Under UV irradiation, solutions containing Fe3+, NO2- and HA can be activated to
364
generate reactive species which can influence the photodegradation of organic
365
compounds in aqueous solution. As shown in Figure 4, the signal intensity of •OH
366
produced in solutions containing NO2- is higher than that in solutions containing Fe3+.
367
As discussed above, we know that the k value of TCP increased by 21.78% in the
368
presence of 50 uM Fe3+, and increased by 57.40% in the presence of 100 uM NO2-,
369
which is consistent of the •OH production in the system. In addition, the signal
370
intensity of carbon-centered radical is much higher than that in other solutions,
371
consisting with the highest degradation efficiency of TCP in NO2- solutions.
372
For TCP solution containing 10 ppm HA (Figure 4d), the •OH signals increased
373
significantly after irradiation while the photodegradation of TCP was inhibited by HA.
374
Form Figure 4d, we can see that •OH and carbon-centered radical were generated
375
after irradiation for 5 min, and less carbon-centered radicals was observed when
376
irradiation time increased to 10 min. The k value of TCP decreased by 18.22% in the
377
presence of 10 ppm HA. Form here we could conclude that HA act as a quencher to
378
inhibit OH or *HA oxidation of TCP and HA also as a light screening agent. This
379
indicated that the light competition effect of HA is too strong to cover the promotion
380
effect of •OH. These results indicated that Fe3+ and NO2- mainly promoted the
381
photodegradation of TCP by the formation of •OH after irradiation while HA mainly 19
382
competed for a light source and acted as a quencher for the reactive species, thus
383
inhibited the degradation of TCP.
384
a Experimental
b
Simulated
3340 3360 3380 3400 3420 Magnetic Field (Gauss)
385
3440
3460
386
Figure 5. Computer simulations of the EPR spectra of DMPO spin adducts after irradiation
387
for 10 min in TCP solution containing 100 µM NO2- : (a) experimental spectrum, (b)
388
composite computer-simulated spectrum; : •OH adduct; : carbon-centered radical.
389
3.3 Identification of photochemical products
390
In order to advance the knowledge of the photolysis mechanism of TCP under the
391
simulated
392
HPLC-LTQ-Orbitrap MS. The samples of TCP in water were analyzed after 12 h
393
irradiation period. Full scan mode with m/z ranged from 50 to 600 was applied, and
394
none of the degradation products were detected in the control solutions. Due to the
395
lack of standard materials for reference, the structure of the photoproducts were
396
identified based on analyzing of the total ion chromatogram (TIC) and the exact mass
397
spectrum coupled with the molecular structure of the parent compound. A total of four
398
degradation products were detected, the corresponding MS spectra and the exacted
solar
light,
photolysis
products
20
of
TCP
were
analyzed
by
399
ion chromatogram were illustrated in Figure S5 and Figure 6, respectively. The
400
proposed elemental composition and the exact mass values of the identified products
401
during the photodegradation of TCP were summarized in Table 1. All the four
402
products had retention time lower than TCP in the chromatogram (20.42 min),
403
meaning polarity of products are higher than that of the parent compound.
404 405
Figure 6. MS spectra of four photoproducts of TCP detected by HPLC-LTQ-Orbitrap MS
406
after irradiation for 12 h.
407 408
Table 1. HPLC-LTQ-Orbitrap-MS retention times, molecular formula, accurate mass data
409
(m/z) and relative mass error (∆ (ppm)) for TCP and the identified photolysis products
compound
RT (min)
Molecular formula
m/z
∆ (ppm)
TCP
20.42
C21H21O4P
369.1250
2.44
P1
7.17
C14H15O3P
263.0832
0.76
P2
11.93
C14H15O4P
277.0635
-0.72
21
P3
13.2
C21H21O5P
383.1054
3.65
P4
8.56
C21H21O6P
401.1149
1.25
410 411
Figure 7 illustrated the possible photochemical pathways of TCP after irradiation in
412
aqueous solution. In pathway a, TCP absorbing light energy lead to the cleavage of a
413
phenoxy bond from the phosphoric center, forming the intermediate M1. Then two
414
photohydrolysis products, P1 and P2, were formed with H2O acts as attacking
415
nucleophiles. P2 is the diester phosphate acid form of TCP, and the corresponding
416
monoester was not detected in the sample of TCP solution over 12 h irradiation,
417
indicating that dicresyl phosphate was more stable than TCP after irradiation. A recent
418
study reported that the organophosphate diesters may have limited nuclear receptor
419
activity compared with the parent triesters (Kojima et al., 2016), but the toxicity
420
information of the other three photolysis products is still unavailable and need further
421
investigation.
422
After irradiation, H-abstraction by UV attacking benzene ring lead to generation of
423
carbon-centered radical (pathway b), this was confirmed by EPR results in the former
424
part. As shown in Figure 4a, •OH was also detected in the photolysis to TCP alone,
425
thus the peroxy product as a key intermediate was taken into consideration in the
426
presence of oxygen. As illustrated in Figure 7, •OH can be obtained as a decay
427
product of M2 because of its instability (Motten et al., 1985; Li and Chignell, 1987).
428
Thus, the hydroxylation products (P3 and P4) were formed during the photolysis of
429
TCP. 22
430
431 432
Figure 7. Proposed reaction pathways of TCP under simulated solar light.
433 434
Four products were characterized in the direct photolysis process of TCP. The
435
products deserve special concern as they may cause toxicity effects on the organisms
436
in water environments. Thus, for accurate ecological risk assessment of OPEs,
437
different photodegradation pathways that may lead to different photoproducts should
438
also be considered.
439 440
4 Conclusions
441
In the present study, the photodegradation of TCP was investigated in aqueous
442
solution under different UV conditions. A half-life at 67.96 h were obtained under 23
443
simulated sunlight, which means direct photolysis may account little for TCP
444
depletion. The initial concentration of TCP, and photoactive components of the
445
surface water including NO2-, Fe3+ and HA were capable of influencing the
446
phototransformation rate constant of TCP. EPR analysis showed that the
447
environmental factors could affect the photodegradation process by the involvement
448
of reactive radicals under irradiation. Carbon-centered radical and hydroxyl radical
449
were produced during the photolysis process and hydroxyl radical contributed to the
450
degradation of TCP in the presence of NO2- and Fe3+. It is the first time that photolysis
451
products of TCP were identified by HPLC-LTQ-Orbitrap MS and possible
452
degradation pathways were proposed, involving the cleavage of a phenoxy bond, C-H
453
bond homogenization from the benzene ring structure of TCP, photoinduced
454
hydrolysis and hydroxylation. Moreover, the toxicity and possible hazards of the
455
products are quite unknown and should be further studied.
456 457
Acknowledgments
458 459
This research was financially supported by National Natural Science Foundation of China (21677023).
460 461 462
24
463
References:
464
Ahn, Y.B., Rhee, S.K., Fennell, D.E., Kekhof, L.J., Hentschel, U., Häggblom, M.M.,
465
2003. Reductive dehalogenation of brominated phenolic compounds by
466
microorganisms associated with the marine sponge aplysina aerophoba. Appl.
467
Environ. Microb. 69, 4159-4166.
468
Antonopoulou, M., Karagianni, P., Konstantinou, I.K., 2016. Kinetic and mechanistic
469
study of photocatalytic degradation of flameretardant Tris (1-chloro-2-propyl)
470
phosphate (TCPP), Appl. Catal. B: Environ. 192, 152-160.
471
Antonopoulou, M., Giannakas, A., Bairamis, F., Papadaki, M., Konstantinou, I., 2017.
472
Degradation of organophosphorus flame retardant tris (1-chloro-2-propyl)
473
phosphate (TCPP) by visible light N,S-codoped TiO2 photocatalysts. Chem. Eng.
474
J. 318, 231-239.
475
Arakaki, T., 1999. pH dependent photoformation of hydroxyl radical and absorbance
476
of aqueous-phase N (III) (HNO2 and NO2-). Environ. Sci. Technol. 33,
477
2561-2565.
478
Beitz, T., Bechmann, W., Mitzner, R., 1999. Investigation on the photoreactions of
479
nitrate and nitrite ions with selected azaarenes in water. Chemosphere, 38,
480
351-361.
481
Burbano, A.A., Dionysiou, D.D., Suidan, M.T., Richardson, T.L., 2005. Oxidation
482
kinetics and effect of pH on the degradation of MTBE with Fenton reagent,
483
Water Res. 39, 107-118.
484
Brandsma, S.H. de Boer,J. Leonards, P.E. Cofino, W.P. Covaci, A., 2013.
485
Organophosphorus
flame-retardant
and
plasticizer
analysis,
including
486
recommendations from the first worldwide inter laboratory study, TrAC Trends
487
Anal. Chem. 43, 217–228.
488
Calza, P., Vione, D., Minero, C., 2014. The role of humic and fulvic acids in the
489
phototransformation of phenolic compounds in seawater. Sci. Total Environ. 493,
490
411-418.
491
Calza, P., Vione, D., Novelli, A., Pelizzetti, E., Minero, C. 2012. The role of nitrite
492
and nitrate ions as photosensitizers in the phototransformation of phenolic
493
compounds in seawater. Sci. Total Environ. 439, 67-75.
494
Chen, Y., Zhang, K., Zuo, Y., 2013. Direct and indirect photodegradation of estriol in
495
the presence of humic acid, nitrate and iron complexes in water solutions. Sci
496
Total Environ. 463-464, 802-809.
497
Chowdhury, R.R., Charpentier, P.A., Ray, M.B., 2011. Photodegradation of 25
498
17β-estradiol in aquatic solution under solar irradiation: Kinetics and influencing
499
water parameters. J. Photochem. Photobiol. A Chem. 219, 67-75.
500
Cristale, J., Katsoyiannis, A., Sweetman, A.J., Jones, K.C., Lacorte, S., 2013.
501
Occurrence and risk assessment of organophosphorus and brominated flame
502
retardants in the River Aire (UK). Environ. Pollut. 179, 194-200.
503
Cristale, J., Dantas, R.F., Luca, A.D., Sans, C., Esplugas, S., Lacorte, S., 2017. Role
504
of
oxygen
and
DOM
in
sunlight
induced
photodegradation
of
505
organophosphorous flame retardants in river water. J. Hazard. Mater. 323,
506
242-249.
507
Du, Z., Wang, G., Gao, S., Wang, Z., 2015. Aryl organophosphate flame retardants
508
induced cardiotoxicity during zebrafish embryogenesis: By disturbing expression
509
of the transcriptional regulators. Aquat. Toxicol. 161, 25-32.
510
Koumaki, E., Mamais, D., Noutsopoulos, C., Nika, M., Bletsou, A.A., Thomaidis,
511
N.S., Eftaxias, A., Stratogianni, G. 2015. Degradation of emerging contaminants
512
from water under natural sunlight: The effect of season, pH, humic acids and
513
nitrate and identification of photodegradation by-products. Chemosphere, 138,
514
675-681.
515
Greaves, A.K. Letcher, R.J., 2014. Comparative body compartment composition and
516
in ovo transfer of organophosphate flame retardants in North American Great
517
Lakes herring gulls. Environ. Sci. Technol. 48, 7942-7950.
518
Hou, R., Xu, Y., Wang, Z., 2016. Review of OPFRs in animals and humans:
519
Absorption, bioaccumulation, metabolism, and internal exposure research.
520
Chemosphere 153 (2016) 78-90.
521
Huang, J., Li, X., Ma, M., Li, D., 2017. Removal of di-(2-ethylhexyl) phthalate from
522
aqueous solution by UV/peroxymonosulfate: Influencing factors and reaction
523
pathways. Chem. Eng. J. 314, 182-191.
524
Kim, U.J., Kannan, K., 2018. Occurrence and distribution of organophosphate flame
525
retardants/plasticizers in surface waters, tap water, and rainwater – Implications
526
for human exposure. Environ. Sci. Technol. 52 (2018) 5625-5633.
527
Kojima, H., Takeuchi, S., Eede, N.V.D., Covaci, A., 2016. Effects of primary
528
metabolites of organophosphate flame retardants on transcriptional activity via
529
human nuclear receptors. Toxicol. Lett. 245, 31-39.
530
Kouras-Hadel, S., Amine-Khodja, A., Halladja, S., Richard, C., 2012. Influence of
531
humic
substance
on
the
nboflavin
532
2,4,6-trimethyphenol, J. Photochem. Photobiol. A 229, 33-38. 26
photosensitized
trandformation
of
533
Koumaki, E., Mamais, D., Noutsopoulos, C., Nika, M.C., Bletsou, A.A., Thomaidis,
534
N.S., Eftaxias, A., Stratogianni, G., 2015. Degradation of emerging contaminants
535
from water under natural sunlight: The effect of season, pH, humic acids and
536
nitrate and identification of photodegradation by-products. Chemosphere 138,
537
675-681.
538
Li, A.S.W., Chignell, C.F., 1987. Spectroscopic studies of cutaneous photosensitizing
539
agents–xii. Spin trapping study of the free radicals generated during the
540
photolysis of photo allergens bithionol and fentichlor. Photochem. Photobiol. 46,
541
445-452.
542
Li, A.S.W., Chignell, C.F., 1991. The NoH value in EPR spin trapping: a new
543
parameter for the identification of 5,5-dimethyl-1-pyrroline-N-oxide spin adducts.
544
J. Biochem. Bioph. Methods. 22, 83-87.
545
Li, C., Wei, G., J. Chen, Zhao, Y., Zhang, Y., Su, L., Qin, W., 2018. Aqueous OH
546
Radical Reaction Rate Constants for Organophosphorus Flame Retardants and
547
Plasticizers: Experimental and Modeling Studies. Environ. Sci. Technol. 52,
548
2790-2799.
549
Liu, J., Ye, J., Chen, Y., Li, C., Ou, H., 2018. UV-driven hydroxyl radical oxidation of
550
tris(2-chloroethyl) phosphate: Intermediate products and residual toxicity.
551
Chemosphere, 190, 225-233.
552
Lee, S., Jeong, W., Kannan, K., Moon, H., 2016. Occurrence and exposure assessment
553
of organophosphate flame retardants (OPFRs) through the consumption of
554
drinking water in Korea. Water Res. 103, 182-188.
555 556
Mack, J., 1999. Photochemistry of nitrite and nitrate in aqueous solution: A review. J. Photochem. Photobiol.A Chem. 128, 1-13.
557
Marvin, C.H., Sverko, E., Tomy, G.T., 2016. Analysis of Chlorinated and Phosphorus
558
Flame Retardants. In: Alaee M. (eds) Dioxin and Related Compounds. The
559
Handbook of Environmental Chemistry, pp, 411-456.
560
Motten, A.G., Buettner, G.R., Chignell, C.F., 1985. Spectroscopic studies of
561
cutaneous photosensitizing agents-VIII. A spin-trapping study of light induced
562
free radicals from chlorpromazine and promazine. Photochem. Photobiol. 42,
563
9-15.
564
Neamtu, M., Frimmel, F.H., 2006. Photodegradation of endocrine disrupting chemical
565
nonylphenol b y simulated solar UV-irradiation. Sci. Total Environ. 369,
566
295-306.
567
Ou, H., Liu, J., Ye, J., Wang, L., Gao, N., Ke, J., 2017. Degradation of 27
568
tris(2-chloroethyl) phosphate by ultraviolet-persulfate: Kinetics, pathway and
569
intermediate impact on proteome of Escherichia coli. Chem. Eng. J. 308,
570
386-395.
571
Peng, J., Wang, G., Zhang, D., Zhang, D., Li, X., 2016. Photodegradation of
572
nonylphenol in aqueous solution by simulated solar UV-irradiation: The
573
comprehensive effect of nitrate, ferric ion and bicarbonate. J. Photochem.
574
Photobiol. A: Chem. 326, 9-15.
575
Poerschmann, J., Trommler, U., Górecki, T., Kopinke, F.D., 2009. Formation of
576
chlorinated biphenyls, diphenyl ethers and benzofurans as a result of
577
Fenton-driven oxidation of 2-chlorophenol. Chemosphere, 75, 772-780.
578
Regnery J., Püttmann, W. 2010. Occurrence and fate of organophosphorus flame
579
retardants and plasticizers in urban and remote surface waters in Germany. Water
580
Res., 44, 4097-4104.
581
Rodriguez, I., Calvo, F., Quintana, J.B., Rubí, E., Rodil, R., Cela, R., 2016. Suitability
582
of solid-phase microextraction for the determination of organophosphate flame
583
retardants and plasticizers in water samples, J. Chromatogr. A, 1108, 158-165.
584
Shi, Y., Guo, L. Li, W. Wang, Y., Liu, J., Cai, Y., 2016. Occurrence, distribution and
585
seasonal variation of organophosphate flame retardants and plasticizers in urban
586
surface water in Beijing, China. Environ. Pollut. 209, 1-10.
587
Stan, S.D., Woods, J.S., Daeschel, M.A., 2005. Investigation of the presence of OH
588
radicals in electrolyzed NaCl solution by electron spin resonance spectroscopy. J
589
Agric. Food Chem. 53, 4901-4905.
590
Su, G., Crump, D., Letcher, R.J., Kennedy, S.W., 2014. Rapid in vitro metabolism of
591
the flame retardant triphenyl phosphate and effects on cytotoxicity and mRNA
592
expression in chicken embryonic hepatocytes. Environ. Sci. Technol. 48,
593
13511-13519.
594
Suh, Y.W., Buettner, G.R., Venkataraman, S., Treimer, S.E., Robertson, L.W.,
595
Ludewig, G., 2009. UVA/B-Induced formation of free radicals from
596
decabromodiphenyl ether. Environ. Sci. Technol. 43, 2581-2588.
597
van der Veen, I., Boer, J.D., 2012. Phosphorus flame retardants: Properties, production,
598
environmental occurrence, toxicity and analysis, Chemosphere 88, 1119-1153.
599
Vione, D., Maurino, V., Minero, C., Pelizzetti, E., 2002. New processes in the
600
environmental chemistry of nitrite: nitration of phenol upon nitrite photoinduced
601
oxidation. Environ. Sci. Technol. 36, 669-676.
602
Voelker, B.M., Morel, F.M.M., Sulzberger, B., 1997. Iron redox cycling in surface 28
603
waters: effects of humic substances and light. Environ. Sci. Technol. 31,
604
1004-1011.
605
Watts, M.J., Linden, K.G., 2009. Advanced oxidation kinetics of aqueous trialkyl
606
phosphate flame retardants and plasticizers. Environ. Sci. Technol. 43,
607
2937-2942.
608
Wang, Y., Feng, C., Li, Y., Gao, J., Yu, C., 2017. Enhancement of emerging
609
contaminants removal using fenton reaction driven by H2O2-producing microbial
610
fuel cells. Chem. Eng. J. 307, 679-686.
611
Wang, R., Tang, J., Xie, Z., Mi, W., Chen, Y., Wolschke, H., Tian, C., Pan, X., Luo, Y.,
612
Ebinghaus, R., 2015. Occurrence and spatial distribution of organophosphate
613
ester flame retardants and plasticizers in 40 rivers draining into the Bohai Sea,
614
north China. Environ. Pollut. 198, 172-178.
615
Wang, Q., Lai, N.L., Wang, X., Guo, Y., Lam, P.K., Lam, J.C., Zhou, B., 2015.
616
Bioconcentration
and
transfer
of
the
organophorous
flame
retardant
617
1,3-Dichloro-2-propyl phosphate causes thyroid endocrine disruption and
618
developmental neurotoxicity in zebrafish larvae. Environ. Sci. Technol. 49,
619
5123-5132.
620
Wei, G., Li, D., Zhuo, M., Liao, Y., Xie, Z., Guo, T., Li, J., Zhang, S., Liang, Z., 2015.
621
Organophosphorus flame retardants and plasticizers: Sources, occurrence,
622
toxicity and human exposure, Environ. Pollut. 196, 29-46.
623
Wenk, J., Eustis, S.N., McNeill, K., Canonica, S., 2013. Quenching of excited triplet
624
states by dissolved natural organic matter. Environ. Sci. Technol. 47,
625
12802-12810.
626 627
Yang, F., Ding, J., Huang, W., Xie, W., Liu, W., 2014. Particle size-specific
628
distributions and preliminary exposure assessments of organophosphate flame
629
retardants in office air particulate matter. Environ. Sci. Technol. 48, 63-70.
630
Yu, C., Zhang, Y. Quan, X., Chen, S., Han, J., Ou, X., Zhao, J., 2010. Photochemical
631
effect of humic acid components separated using molecular imprinting method
632
applying porphyrin-like substances as templates in aqueous solution. Environ.
633
Sci. Technol. 44, 5812-5817.
634
Yuan, X., Lacorte, S., Cristale, J., Dantas, R.F., Sans, C., Esplugas, S., Qiang, Z., 2015.
635
Removal of organophosphate esters from municipal secondary effluent by ozone
636
and UV/H2O2 treatments. Sep. Purif. Technol. 156, 1028-1034.
637
Zhang, Q., Ji, C., Yin, X., Yan, Lu, L., Zhao, M., 2016. Thyroid hormone-disrupting 29
638
activity and ecological risk assessment of phosphorus-containing flame
639
retardants by in vitro, in vivo and in silico approaches. Environ. Pollut. 210,
640
27-33.
641
Zhao, Q., Zhao, H., Quan, X., Chen, S., Zhang, Y., 2014. Photochemical
642
transformation of 2,2′,4,4′-tetrabromodiphenyl ether (BDE-47) in surface coastal
643
waters: Effects of chloride and ferric ions. Mar. Pollut. Bull. 86, 76-83.
644
Zhou, D., Huang, W., Wu, F., Han, C., Chen, Y., 2010. Photodegradation of
645
chloromycetin in aqueous solutions: kinetics and influencing factors. React.
646
Kinet. Mech. Cat. 100, 45-53.
647
Zhao, H., Jiang, J., Wang, Y., Lehmler, H., Buettner, G.R., Quan, X., Chen, J., 2015a.
648
Monohydroxylated
649
dihydroxylated polybrominated biphenyls (Di-OH-PBBs): Novel photoproducts
650
of 2,6-dibromophenol. Environ. Sci. Technol., 49, 14120-14128.
651
polybrominated
diphenyl
ethers
(OH-PBDEs)
and
Zhao, Q., Zhao, H., Quan, X., He, X., Chen, S., 2015b. Photochemical formation of
652
hydroxylated
polybrominated
diphenyl
653
polybrominated diphenyl ethers (PBDEs) in aqueous solution under simulated
654
solar light irradiation. Environ. Sci. Technol. 49, 9092-9099.
655
30
ethers
(OH-PBDEs)
from
Highlights Direct and indirect photolysis of TCP in water were investigated under irradiation. EPR was used for deducing effects of Fe3+, NO2- and HA on phototransformation of TCP. Carbon-centered radical and ·OH were formed during the photolysis process of TCP. Four photoproducts were identified by HPLC-LTQ-Orbitrap MS analysis.
S0045-6535(19)32210-6
DOI:
https://doi.org/10.1016/j.chemosphere.2019.124971
Reference:
CHEM 124971
To appear in:
ECSN
Received Date: 9 July 2019 Revised Date:
17 September 2019
Accepted Date: 24 September 2019
Please cite this article as: Sun, S., Jiang, J., Zhao, H., Wan, H., Qu, B., Photochemical reaction of tricresyl phosphate (TCP) in aqueous solution: Influencing factors and photolysis products, Chemosphere (2019), doi: https://doi.org/10.1016/j.chemosphere.2019.124971. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2019 Published by Elsevier Ltd.
Photochemical reaction of tricresyl phosphate (TCP) in aqueous solution: Influencing factors and photolysis products
Shibin Sun, 1 † Jingqiu Jiang, 1, † Hongxia Zhao, *, † Huihui Wan ‡, Baocheng Qu *,§
†
Key Laboratory of Industrial Ecology and Environmental Engineering (Ministry of
Education), School of Environmental Science and Technology, Dalian University of Technology, Dalian, 116024, China ‡
College of Chemical Engineering, Analytical Center, Dalian University of Technoloy,
Dalian 116024, China §
College of Marine Technology and Environment, Dalian Ocean University, Dalian
116024, China
#
Corresponding author phone: +86-411-84707965; Fax: +86-411-84707965;
Email: [email protected] #
Corresponding author phone: +86-411-84740281; Fax: +86-411-84740281;
Email: [email protected] 1
These authors contributed equally to this study and share first authorship
hv
0.0
directt indirectt
products
-0.4
TCP
3O
Fe3+ H2O NO2-/NO3-
2
1O
HA … 2
ln(Ct/C0)
O O O P O
-0.2
-0.6 -0.8 3+
0 µmol/L Fe 3+ 50 µmol/L Fe 3+ 10 µmol/L Fe
-1.0
· OH
-1.2 0
2
4
6 8 time (h)
10
12
1
Abstract
2
Organophosphate triesters (OPEs) have caused great concern as a class of emerging
3
environmental contaminants due to their widespread use and their toxicity to
4
organisms. However, the phototransformation behavior of OPE is still not fully
5
understood, which is important for understanding their environmental fate. In the
6
present study, the photodegradation of tricresyl phosphate (TCP), one of the most
7
widely detected OPEs in aqueous environments, was investigated including the direct
8
photolysis and in the presence of several natural water factors, NO2-, Fe3+ and humic
9
acid. The degradation process followed the pseudo-first-order kinetics, with rate
10
constant increasing slightly with increasing initial TCP concentration. The presence of
11
NO2- and Fe3+ was observed to promote the photochemical loss of TCP, while humic
12
acid played a negative role on TCP transformation. Electron spin resonance (EPR)
13
analysis showed that carbon-centered radical was produced in the photolysis process
14
of TCP, and hydroxyl radical contributed to the promotion of rate constant for Fe3+
15
and
16
HPLC-LTQ-Orbitrap MS analysis, and the possible degradation pathways of TCP
17
were proposed. These findings provide a meaningful reference for the fate and
18
transformation of OPEs in natural water.
NO2-.
Four
photolysis
products
were
tentatively
identified
by
19 20
Keywords: tricresyl phosphate (TCP); environmental factors; kinetics; radical
21
analysis; photolysis products
22 1
23
1. Introduction
24
Organophosphate triesters (OPEs) are a class of emerging environmental pollutants
25
which have been widely used as flame retardants and plasticizers in many industrial
26
and household materials, such as furniture, textiles, building materials and electronics
27
(Brandsma, et al., 2013; van der Veen, et al., 2012; Shi, et al., 2016). Because
28
polybrominated diphenyl ethers have been phased out, OPEs have become one of the
29
most frequently used alternative flame retardants. It has been reported that the annual
30
global production of OPEs currently reaches approximately 200 kt (Greaves, et al.,
31
2014), and nearly more than 70 kt being produced in China (Wei, et al., 2015). As
32
OPEs are not chemically bounded to the final products, they can be released into
33
environment via volatilization, abrasion, and leaching process (Rodriguez, et al., 2016;
34
Marvin, et al., 2016; Li et al., 2018). OPEs have been reported to be present in aquatic
35
environments including river waters, seawaters, ground waters, and even drinking
36
waters, with concentrations ranging from ng/L to ug/L (van der Veen and Boer, 2012;
37
Shi, et al., 2016; Wang et al., 2015; Cristale, et al., 2013). For example, Wang et al.
38
reported that in North China, the total concentrations of OPEs ranged from 9.6 to
39
1549 ng·L-1 in 40 major rivers draining into the Bohai Sea (Wang et al., 2015). In the
40
River Aire (UK), OPEs were detected at concentrations ranging from 6.3 to 26050
41
ng·L-1 (Cristale, et al., 2013). The ubiquitous occurrence and high levels of OPEs in
42
the water environments has sparked considerable interest of researchers.
43
An increasing number of toxicological studies have shown that OPEs have negative
44
effects on human health and are toxic to aquatic organisms (van der Veen and Boer, 2
45
2012). For instance, the chloroalkyl OPEs such as tri(2-chloroethyl) phosphate (TCEP)
46
and tri(dichloropropyl) phosphate (TDCP) have been proved to be carcinogenic and
47
neurotoxic and TCEP was considered to be environmentally persistent (Hou, et al.,
48
2016; Wang et al., 2015; Yang et al., 2014). Du et al. (2015) have reported two aryl
49
OPEs which have heart developmental toxicity by disturbing expressions of
50
transcriptional regulators in zebrafish. In addition, tricresyl phosphate (TCP) has been
51
shown to be potential thyroid hormone disruptors and could cause reproductive
52
toxicity (van der Veen and Boer, 2012; Zhang et al., 2016). Accordingly, it is essential
53
to understand the fate of OPEs in aquatic environments for their ecological risk
54
assessment.
55
Photochemical degradation is considered to be an important transformation pathway
56
of organic pollutants in aquatic environment, which may occur by either direct or
57
indirect photolysis. Direct photolysis is the result of light absorbance by pollutants
58
causing their molecular degradation. Some aryl OPEs have considerable sunlight
59
absorbance in the ultraviolet region (290-400 nm) (Cristale et al., 2017), it is believed
60
that direct photolysis plays a crucial role for the photodegradation of OPEs in aquatic
61
environment. For indirect photolysis, NO2−, Fe3+ and humic substances play crucial
62
roles, which are ubiquitous in surface water and absorb solar radiation to reach an
63
excited state, subsequently generating free radicals (Burbano et al., 2005, Cristale et
64
al., 2017). In recent years, researchers have focused on the photodegradation of OPEs
65
using advanced oxidation processes (AOPs) such as ozonation, UV/H2O2, UV/TiO2,
66
UV/persulfate (PS) for their removal in effluents (Watts and Linden 2009; 3
67
Antonopoulou et al., 2016; Antonopoulou et al., 2017; Liu et al., 2018; Ou et al., 2017;
68
Yuan et al., 2015). Nevertheless, until now, only limited studies have reported the
69
environmental fate of OPEs in natural water. Regnery and Püttmann (2010) reported
70
that upon exposure to sunlight, the concentrations of three alkyl OPEs decreased
71
significantly in lake water while no degradation was observed in ultrapure water.
72
Another study pointed out that humic acid inhibited OPEs photodegradation mediated
73
by singlet oxygen while inorganic constituents in river water enhanced their removel
74
(Cristale et al., 2017). These processes contributed the indirect photolysis of OPEs to
75
constituents in natural water. However, there are still some other constituents whose
76
effects on phototransformation of OPEs are not characterized, such as Fe3+ and NO2-.
77
Also, the photodegradation products should be paid more attention because some
78
products are more toxic than their parent compounds (Su et al., 2014).
79
To fill the knowledge gap mentioned above, photodegradation experiments were
80
performed to figure out the transformation kinetics and mechanisms of OPEs under
81
simulated solar light irradiation. TCP, one of the most widely detected OPEs in
82
aquatic environments (van der Veen and Boer, 2012), was selected as the model
83
compound. Direct photodegradation of TCP under different light intensities and
84
wavelengths were firstly studied. Then the influences of different initial
85
concentrations and various environmental factors, including Fe3+, NO2-, and humic
86
acid (HA) on the photodegradation efficiency of TCP were investigated. In addition,
87
photodegradation products of TCP after irradiation were identified and possible
88
degradation pathways were proposed. Electron paramagnetic resonance spectroscopy 4
89
(EPR) techniques was used to study the mechanism for the effects of constituents on
90
photodegradation processes and to capture the intermediates for deducing the
91
degradation pathways. These findings advance our fundamental understanding of fate
92
of OPEs in aquatic environments and ecological risk of OPEs contamination in sunlit
93
water bodies.
94 95
2. Materials and methods
96
2.1 Chemicals
97
Analytical standard TCP (mixture of isomers, CAS number: 1330-78-5, purity ≥ 99%)
98
was purchased from Aladdin Reagent (Shanghai, China), and its molecular formula
99
and weight are C21H21O4P and 368.1178 g/mol, respectively. Standard solution of TCP
100
was prepared in methanol at the concentration of 200 ppm and then stored at 4 oC in
101
the dark. HPLC grade methanol and dichloromethane (DCM) were obtained from
102
Sigma Aldrich (USA). FeCl3, NaNO2 were of reagent grade. Humic acid (CAS:
103
1415-93-6, FA ≥ 98%) was purchased from the International Humic Substance
104
Society (IHSS). Spin traps were used to capture the reactive oxygen species (ROS).
105
5,5-dimethyl-1-pyrroline-Noxide (DMPO) was obtained from Dojindo Laboratories
106
(Shanghai, China). All the solutions used in the experiments were prepared using
107
Milli-Q water (18 MΩ·cm) obtained from OKP ultrapure water system (Shanghai
108
Lake-core Instrument Co., Shanghai, China).
109 110
2.2 Irradiation 5
111
The photochemical experiments were performed in an XPA-1 photochemical reactor,
112
which the schematic diagram was shown in Figure S1 (Nanjing Xujiang
113
Electromechanical Plant, Nanjing, China). A 500 W Xe lamp, a 1000 W Xe lamp and
114
a 800 W mercury (Hg) lamp coupled with the 290 nm filters were used to obtain
115
different light intensities. The 800 W Hg lamp surrounded by 290 nm and 340 nm
116
cut-off filters were used to simulate the natural sunlight and UV-A irradiation sources,
117
and a 254 nm low-pressure Hg lamp was used as the UV-C irradiation source. The
118
light intensities in the center of the solutions were 4.67 mW/cm2, 15.52 mW/cm2 and
119
61.60 mW/cm2 for 500 W Xe lamp, 1000 W Xe lamp and 800 W Hg lamp evaluated
120
by an optical sensors (RAMSES, TriOS).
121
To keep the temperature inside the reactor stable, the lamps was enclosed in a cooling
122
well via the circulation of cold water. A 30 mL aliquot solution (50 ppb -200 ppb) was
123
placed in a quartz tube in a merry-go-round apparatus inside the reactor for irradiation.
124
The initial concentration of TCP was 100 ppb unless specified, certain concentrations
125
of FeCl3, NO2- and HA were added into the reaction solutions. Dark controls were
126
performed under the same experimental condition. At defined time intervals, samples
127
were transferred to glass vials and stored at -20
128
photochemical experiments were performed in triplicate.
o
C until analysis. All the
129 130
2.3 Analytical methods
131
Matrix matched calibration curves were prepared using Milli-Q water with TCP
132
ranging from 1 to 200 ppb. The concentration of TCP was analyzed by ultra-high 6
133
performance liquid chromatography coupled to a triple quadrupole detector
134
(UPLC-MS/MS) (TSQ Quantum Ultra, Thermo, USA). The chromatographic column
135
was a SB-C18 (2.1 mm×150 mm, particle size 5 µm, Agilent, USA) and the
136
temperature was set at 30 oC. The mobile phase was 0.2% formic acid in UPLC water
137
(15%) and methanol (85%) and the flow rate was set at 0.2 mL·min-1. The injection
138
volume was 20 µL. TCP was analyzed under positive electrospray ionization (ESI+)
139
by multiple reaction monitoring (SRM) mode. The optimized collision energy were
140
29 V for the transition of m/z 369.1 → 165.8 and 28 V for the transition of m/z 369.1
141
→243.0. All data were acquired and processed by Xcalibur software.
142
Screening of photolysis products was carried out using an LTQ Orbitrap Velos mass
143
spectrometer (Thermo Fisher Scientific, Bellefonte, PA, USA) equipped with an ESI
144
source in both positive and negative ionization mode. The MS conditions were
145
optimized as follows: sheath gas/aux gas/sweep gas 30/10/0 arb, capillary temperature
146
350 0C, capillary voltage 39 V, tube lens voltage 0 V, and spray voltage 3.5 kV. The
147
system was operated in the full spectral acquisition mode in the mass range of m/z
148
80-600 with a mass resolution of 100,000.
149
The LC system was equipped with an Ultimate XB-C18 column (2.1 mm×100 mm,
150
5µm particle, Thermo Fisher Scientific, USA) and the mobile phase was methanol (A)
151
and water (B). A gradient run was used as follows: 30% A to 60% A in 5 min, to 80%
152
A in 10 min, to 100% A in 5 min, hold 5 min, to 30% A in 5 min, equilibrate to 30% A
153
during 5 min. The flow rate was 200 µL/min and the injection volume was 5 µL. A
154
constant temperature of 30 oC was kept during analysis. All data were acquired and 7
155
processed by Xcalibur software.
156 157
2.4 Electron paramagnetic resonance (EPR) analysis
158
Experiments were still conducted in the XPA-1 photochemical reactor with the same
159
800 W Hg lamp equipped with the 290 nm cut-off filters. All the solutions were
160
prepared in the dark at room temperature (25 oC) and were loaded into high purity
161
capillary quartz tubes immediately after mixing the solution (Fe3+, NO2-, or HA) with
162
TCP and the spin trap DMPO. The final concentration of DMPO was 160 mM.
163
Samples were collected at specific irradiation time to measure the ROS signal by EPR,
164
the detailed information were shown in Supporting Information. EPR experiments
165
were performed on a Bruker EXM A-200 spectrometer (Bruker, Bermen, Germany)
166
and the operating parameters were as follows: central field, 3398 G; microwave
167
frequency, 9.45 GHz (X-band); microwave power, 1.69 Mw; scanning width, 100 G;
168
and scanning frequency, 100 kHz.Simulation of EPR data was accomplished using
169
WinSim software (NIEHS) and Bruker WinEPR. The Spin Trap Database (NIEHS)
170
was referred to in order to interpret and simulate the EPR spectra (Suh et al., 2009)
171
and the simulated spectra correlated well with the corresponding experimental spectra
172
(correlation coefficient > 0.99).
173 174
3. Results and discussion
175
3.1 UV light irradiation of TCP in milli-Q water
176
In surface waters, the photochemical transformation of pollutants could take place by 8
177
direct or indirect photoreaction. Direct photolysis take place when radiation
178
absorption by a molecule triggers its transformation. Figure S1 showed that TCP had a
179
broad absorption spectrum between 220 nm and 320 nm, and the local maximum
180
absorption wavelength was approximately 268 nm. For direct photolysis of TCP, the
181
influences of different light intensities and wavelength were evaluated. Firstly, a 500
182
W Xe lamp, 1000 W Xe lamp and 800 W Hg lamp were used as the light source, the
183
photolysis of TCP underwent with 290 nm cut-off filters and the corresponding light
184
intensities were 4.67, 15.52, 61.60 W/m2 respectively. As shown is Figure 1, the
185
degradation rate increased with the light intensity. Half-life times of 130.8 and 68.0 h
186
were obtained under 500 W Xe lamp and 1000 W Xe lamp, respectively. Under 800
187
W Hg lamp, the degradation rate was significantly increased and the half-life time
188
was 8.75 h. Dark control samples were analyzed during the same period of time and
189
the results showed no obvious degradation without exposure to UV light. The
190
irradiation spectrum of sunlight is similar with that of 1000 W Xe lamp shown in
191
Figure S3 in supporting information, and the light intensity was 20.60 W/m2 for
192
sunlight which is a litter higher than that of 1000 W Xe lamp. Thus, it is deduced the
193
half-life time of TCP for direct photolysis in natural water was close to 68.0 h,
194
meaning that the direct photolysis of TCP under sunlight plays minor role in the
195
photodegradation process in water environment, which is similar as previous study
196
(Cristale et al., 2017).
9
0.0
ln(Ct/C0)
-0.4 -0.8 -1.2
Dark control 2 4.67 mW/cm 2 15.52 mW/cm 2 61.60 mW/cm
-1.6 -2.0 0
198
5
10
15
20
25
time (h)
197
Figure 1. Photodegradation of TCP at different light intensities (C0 = 200 ppb)
199 200
An 800 W Hg arc lamp with 340 nm cut-off filter was used to simulate the UV-A light
201
and a low-pressure Hg light (254 nm) was used to verify the influence of UV-C on the
202
photodegradation of TCP. Figure 2 presented the photodegradation of TCP under
203
UV-A and UV-C irradiation in Milli-Q water. The 800 W Hg arc lamp with 290 nm
204
cut-off filter contains UV-A and UV-B region and obvious degradation was observed
205
for TCP under this condition. As shown in Figure 2, no degradation of TCP under
206
UV-A range irradiation, which indicated that UV-B plays a decisive role in the
207
photodegradation of TCP in natural environment. For UV-C region, 80% of TCP was
208
significantly removed after 20 min of irradiation. UV-C irradiation proved to be the
209
most effective light, which most probably due to its high absorbance in this region to
210
induce the direct photolysis of TCP. In the study of Cristale et al (2017), UV-C light
211
was also used to study the degradation efficiency of nine OPEs, and three of them
212
with aryl groups showed 100% removal after 10 min, which is similar with the result
213
in our study.
10
214 215
Figure 2. Photodegradation of TCP in Milli-Q water under UV-A and UV-C irradiation (C0 =
216
200 ppb)
217 218
3.2 Effects of initial concentration and typical environmental factors on the
219
phototransformation of TCP
220
The phototransformation of TCP under the simulated sunlight irradiation were further
221
investigated under various conditions with different TCP concentrations, together with
222
different concentration of Fe3+, NO2- and HA. The photodegradation rate constant and
223
half-life was determined using pseudo-first-order kinetics, C = C0e-kt. The reaction
224
rate constant (k) and the corresponding correlation coefficient (R2) were given in the
225
table S1.
11
0.0
0.0
a
-0.2
b
-0.2 -0.4 ln(Ct/C0)
ln(Ct/C0)
-0.4 -0.6 -0.8
0
2
-
0 µmol/L NO2
-
20 µmol/L NO2
-1.2 4
6 8 time (h)
10
-
100 µmol/L NO2
0
12
2
4
6 8 time (h)
10
12
0.0
0.0
c
-0.2
d
-0.2
-0.4
-0.4 ln(Ct/C0)
ln(Ct/C0)
-0.8 -1.0
50 ppb 100 ppb 200 ppb
-1.0
-0.6
-0.6 -0.8
-0.6 -0.8 0 ppm HA 2 ppm HA 10 ppm HA
3+
0 µmol/L Fe 3+ 50 µmol/L Fe 3+ 10 µmol/L Fe
-1.0 -1.2 0
2
4
-1.0 -1.2 6 8 time (h)
10
0
12
226
2
4
6 8 time (h)
10
12
227
Figure 3. The effects of environmental factors on the photodegradation of TCP (a) initial
228
concentration, (b) NO2-, (c) Fe3+, (d) HA; the concentration of TCP for the experiments in (b),
229
(c), and (d) is 100 ppb
230 231
3.2.1 Effect of initial concentration of TCP
232
Pollutant concentration may be an important factor concerning the photoreaction
233
process. As TCP was detected in natural surface waters at various concentrations
234
ranging from ppt to ppb level (Lee et al., 2016; Kim and Kannan 2018), it’s
235
significant to investigate the effect of TCP concentration on its photolysis behavior
236
under solar irradiation. In this work, the effect of initial concentration of TCP on the
237
direct degradation efficiency was investigated with three different initial
238
concentrations of 50 ppb, 100 ppb and 200 ppb. As shown in Figure 3a, the k values 12
239
increased with increasing initial concentrations of TCP, they were 0.0509 h-1 for 50
240
ppb, 0.0627 h-1 for 100 ppb and 0.0789 h-1 for 200 ppb, respectively. The
241
photodegradation half-life (t1/2) of TCP varied from 13.6 to 8.8 h with the
242
concentration ranging from 50 to 200 ppb. When the concentrations of TCP in the
243
natural waters are lower than the present study, it is expected that TCP would have a
244
longer photodegradation half-life than being discussed in this study.
245 246
3.2.2 Effect of NO2-
247
Nitrate and Nitrite are ubiquitous constituents coexisting in natural waters. The effect
248
of nitrate on the photodegradation of pollutants in water is mainly realized by the
249
production of nitrite (Beitz et al., 1999; Arakaki, 1999). When compared to nitrate,
250
nitrite absorb a larger fraction of the sunlight spectrum, making nitrite is an important
251
ROS source (Mack, 1999), which may contribute to the photodegradation of
252
pollutants. As illustrated in Figure 3b, the degradation rate of TCP was obviously
253
enhanced by the addition of NO2-, the k values were 0.072 h-1 for control, 0.0818 h-1
254
for 20 µM (NO2-), and 0.0959 h-1 for 100 µM (NO2-). The transformation rate constant
255
increased with increasing concentration of NO2-.
256
Numerous studies have shown that NO2- can generate •OH and •NO2 (Calza, et al.,
257
2012; Vione et al., 2002) under UV light irradiation through the following equations
258
(1)-(3):
259
NO2- + hv (+H+) → •NO + •O-
260
(1)
•O- + H2O → •OH + OH-
(2) 13
261
•OH + NO2- → •NO2 + OH-
(3)
262
With the increase of the concentrations of NO2-, the concentration of generated •OH
263
and •NO2 in the system also increased. The generated free radicals may attack the
264
phenyl structure (Ahn et al., 2003; Poerschmann et al., 2009), thus affecting the
265
photolysis process of TCP and accelerating the photodegradation rate of TCP in water.
266 267
3.2.3 Effect of Fe3+
268
Iron is another significant component with extensive presence in natural aquatic
269
environment at concentrations ranging from 10-7 to 10-4 M (Zhao et al., 2014).
270
Considerable literatures have proved that iron plays an important role in many
271
photochemical reactions of organic compounds relating to ROS production, such as
272
photocatalytic reaction and Fenton reaction (Voelker et al., 1997; Wang et al., 2017).
273
In this study, we used FeCl3 as the source of Fe3+ to investigate its influences on the
274
photodegradation of TCP. As illustrated in Figure 3c, the degradation rate of TCP was
275
enhanced by adding Fe3+ to the system. The k values for TCP were 0.0675 h-1 for
276
control, 0.097 h-1 for solution containing 10 µM Fe3+ and 0.0822 h-1 for solution
277
containing 50 µM Fe3+, respectively. The existing studies proved that Fe3+ could
278
produce •OH via the following reactions (Zhao et al., 2014; Peng et al., 2016; Neamtu
279
and Frimmel, 2006):
280
Fe3+ + H2O ↔ Fe(OH)2+ + H+
(4)
281
Fe(OH)2+ + hv → Fe2+ + •OH
(5)
14
282
Due to the increased •OH production with higher Fe3+ concentration, a positive role of
283
Fe3+ to the photodegradation of TCP was expected. While under our experimental
284
conditions, it’s interesting to find that although Fe3+ did enhance the degradation rate
285
of TCP at concentrations of 10 µM and 50 µM, but the degradation rate at 50 µM is
286
lower than that at 10 µM. Here, the photodegradation rate increased by 43.7% with
287
increasing Fe3+ concentration from 0 to 10 µM but decreased about 15.3% by
288
increasing Fe3+ concentration from 10 µM to 50 µM. Since the photochemical
289
properties of Fe3+ can be strongly influenced by pH levels in natural waters, the result
290
can be attributed to pH value (6.5 ± 0.2) of the solutions used in our experiments. The
291
study of Chowdhury and co-workers (2011) has demonstrated that Fe(OH)2+ is the
292
predominant photoreactive species among the Fe3+-aquo complexes. Moreover, it was
293
reported that the photoactivity of Fe(OH)2+ in aqueous solutions was restrained at pH >
294
5.0 (Zhao et al., 2014) and at a pH of 7.0 (±0.1), Fe3+ at different concentrations
295
revealed an inhibition effect (Zhou et al., 2010). Thus, that’s may be the reason that
296
TCP has a higher k value at the addition of 10 uM Fe3+ when compared to a higher
297
concentration of Fe3+.
298 299
3.2.4 Effect of HA
300
Numerous studies have demonstrated that HA play a significant role in photochemical
301
degradation of organic pollutants in aqueous systems and the dual behavior of HA as
302
photosensitizer and redox inhibitor has been proved (Yu et al., 2010). In order to
303
evaluate the effect of HA on the photodegradation of TCP, experiments without HA 15
304
and with two alternative HA concentrations (2 ppm and 10 ppm) were carried out
305
under simulated solar light irradiation. The obtained pseudo-first-order kinetic
306
degradation curve is shown in Figure 3d. According to the results, the removal rate of
307
TCP was visibly inhibited by HA, with the k values decreasing from 0.00642 h-1 for
308
control to 0.0594 h-1 for 2 ppm HA and 0.0525 h-1 for 10 ppm HA. It was obvious that
309
he presence of HA decreased the photolysis rate constant of TCP, implying HA served
310
as an inhibitor rather than a photosensitizer. Recent studies have demonstrated that
311
HA may act as a light screening agent and free radical quencher to inhibit the excited
312
triplet-induced oxidation of several organic contaminants (Calza et al., 2014; Wenk et
313
al., 2014; Koumaki et al., 2015).
314
HA/1HA*/3HA*/HA+ + •OH→ oxidized HA
315
Cristale and co-works (2017) also found that HA acted as an inhibitor for aryl
316
phosphates, for different type and concentration of HA was selected, which is in
317
consistent with this study. For aryl OPEs, photodegradation decrease in HA solution
318
can be attributed to some inhibition of excited states of photosensitizer OPEs by
319
reducing moieties of HA, because they found that sunlight absorbing OPEs showed
320
photosensitizing properties through generating singlet oxygen. In the study of
321
Kouras-Hadel et al (2012), the inhibiting effect of HA mainly attributed to reduction
322
of quinonic moieties by reaction intermediate superoxide anions.
323
Moreover, the increasing concentration of HA may increase both the generated ROS
324
level and the photoactivated HA species (1HA*, 3HA*, etc.), but the photoactivated HA
325
species are recommended to be more efficient free radical quenchers than the parent 16
326
HA (Chen et al., 2013). On the other hand, HA could absorb the sunlight effectively
327
(data shown in Figure S4), thus the inhibition could be partially related with the
328
reduction of direct light absorption of TCP due to HA scattering and light absorption
329
competition. The light screening effect and increased free radical quenching
330
efficiency of HA may led to an overall inhibition on the photodegradation of TCP.
331 332
3.2.5 Identification of free radicals
333
In this study, EPR was performed to detect •OH or other possible radicals generated
334
during the irradiation process. As seen in Figure 4, no signal was observed in any
335
solution containing DMPO under dark condition, and phosphate buffered saline (PBS)
336
showed no peaks either after 10 min of irradiation. A characteristic sextet peaks of a
337
1:1:1:1:1:1 intensity can be observed after irradiation for 5 min, which is the typical
338
signal of carbon-centered radical (Zhao et al., 2015a). The observed six-line spectrum
339
had hyperfine spiltting constants of αN = 15.78 G, αH = 23.25 G, and g value of
340
2.00298 and NoH = 0.68, where NoH is the ration of the nitrogen-spiltting constant to
341
the hydrogen-spiltting constant (Li and Chignell, 1991). This similarity of this adduct
342
and previously reported 2-chlorophenyl-DMPO adduct (αN = 15.78 G, αH = 23.25 G,
343
NoH = 0.68) (Motten et al., 1985) suggested that it was formed by the reaction of the
344
spin trap with an aryl radical generated during the photolysis of TCP. It is reported
345
that
346
2,2’,4,4’-tetrabromobiphenyl ether could generate carbon-centered radical after
347
irradiation (Zhao et al., 2015a; Zhao et al., 2015b) and sunlight absorbing aryl
some
aryl
organic
pollutants
17
such
as
bromophenol
and
348
phosphate showed photosensitizing properties (Cristale et al., 2017), thus it is
349
speculated that TCP converted to the excited state after irradiation, C-H bond
350
homogenization occurred from the benzene ring structure of the excited TCP,
351
resulting in the formation of the carbon-centered radical.
352
At the same time, the characteristic quartet peaks of the DMPO/•OH adduct with a
353
1:2:2:1 intensity was detected for TCP solution alone and TCP solutions containing
354
Fe3+, NO2- and HA after irradiation. These results are in accordance with several
355
similar studies for the DMPO/•OH adduct (Stan et al., 2005; Huang et al., 2017),
356
confirming the generation of •OH radical in this study. Take TCP solution which
357
containing100 µM NO2- as an example, simulation of the experimental spectra yielded
358
hyperfine splitting constants were: DMPO/•OH (αN = αH = 15.07 G) (see Figure 5).
359 18
360
Figure 4. EPR spectra produced by DMPO adducts in (a) TCP alone, (b) TCP + NO2- solution,
361
(c) TCP + Fe3+ solution, (d) TCP + HA solution.
362 363
Under UV irradiation, solutions containing Fe3+, NO2- and HA can be activated to
364
generate reactive species which can influence the photodegradation of organic
365
compounds in aqueous solution. As shown in Figure 4, the signal intensity of •OH
366
produced in solutions containing NO2- is higher than that in solutions containing Fe3+.
367
As discussed above, we know that the k value of TCP increased by 21.78% in the
368
presence of 50 uM Fe3+, and increased by 57.40% in the presence of 100 uM NO2-,
369
which is consistent of the •OH production in the system. In addition, the signal
370
intensity of carbon-centered radical is much higher than that in other solutions,
371
consisting with the highest degradation efficiency of TCP in NO2- solutions.
372
For TCP solution containing 10 ppm HA (Figure 4d), the •OH signals increased
373
significantly after irradiation while the photodegradation of TCP was inhibited by HA.
374
Form Figure 4d, we can see that •OH and carbon-centered radical were generated
375
after irradiation for 5 min, and less carbon-centered radicals was observed when
376
irradiation time increased to 10 min. The k value of TCP decreased by 18.22% in the
377
presence of 10 ppm HA. Form here we could conclude that HA act as a quencher to
378
inhibit OH or *HA oxidation of TCP and HA also as a light screening agent. This
379
indicated that the light competition effect of HA is too strong to cover the promotion
380
effect of •OH. These results indicated that Fe3+ and NO2- mainly promoted the
381
photodegradation of TCP by the formation of •OH after irradiation while HA mainly 19
382
competed for a light source and acted as a quencher for the reactive species, thus
383
inhibited the degradation of TCP.
384
a Experimental
b
Simulated
3340 3360 3380 3400 3420 Magnetic Field (Gauss)
385
3440
3460
386
Figure 5. Computer simulations of the EPR spectra of DMPO spin adducts after irradiation
387
for 10 min in TCP solution containing 100 µM NO2- : (a) experimental spectrum, (b)
388
composite computer-simulated spectrum; : •OH adduct; : carbon-centered radical.
389
3.3 Identification of photochemical products
390
In order to advance the knowledge of the photolysis mechanism of TCP under the
391
simulated
392
HPLC-LTQ-Orbitrap MS. The samples of TCP in water were analyzed after 12 h
393
irradiation period. Full scan mode with m/z ranged from 50 to 600 was applied, and
394
none of the degradation products were detected in the control solutions. Due to the
395
lack of standard materials for reference, the structure of the photoproducts were
396
identified based on analyzing of the total ion chromatogram (TIC) and the exact mass
397
spectrum coupled with the molecular structure of the parent compound. A total of four
398
degradation products were detected, the corresponding MS spectra and the exacted
solar
light,
photolysis
products
20
of
TCP
were
analyzed
by
399
ion chromatogram were illustrated in Figure S5 and Figure 6, respectively. The
400
proposed elemental composition and the exact mass values of the identified products
401
during the photodegradation of TCP were summarized in Table 1. All the four
402
products had retention time lower than TCP in the chromatogram (20.42 min),
403
meaning polarity of products are higher than that of the parent compound.
404 405
Figure 6. MS spectra of four photoproducts of TCP detected by HPLC-LTQ-Orbitrap MS
406
after irradiation for 12 h.
407 408
Table 1. HPLC-LTQ-Orbitrap-MS retention times, molecular formula, accurate mass data
409
(m/z) and relative mass error (∆ (ppm)) for TCP and the identified photolysis products
compound
RT (min)
Molecular formula
m/z
∆ (ppm)
TCP
20.42
C21H21O4P
369.1250
2.44
P1
7.17
C14H15O3P
263.0832
0.76
P2
11.93
C14H15O4P
277.0635
-0.72
21
P3
13.2
C21H21O5P
383.1054
3.65
P4
8.56
C21H21O6P
401.1149
1.25
410 411
Figure 7 illustrated the possible photochemical pathways of TCP after irradiation in
412
aqueous solution. In pathway a, TCP absorbing light energy lead to the cleavage of a
413
phenoxy bond from the phosphoric center, forming the intermediate M1. Then two
414
photohydrolysis products, P1 and P2, were formed with H2O acts as attacking
415
nucleophiles. P2 is the diester phosphate acid form of TCP, and the corresponding
416
monoester was not detected in the sample of TCP solution over 12 h irradiation,
417
indicating that dicresyl phosphate was more stable than TCP after irradiation. A recent
418
study reported that the organophosphate diesters may have limited nuclear receptor
419
activity compared with the parent triesters (Kojima et al., 2016), but the toxicity
420
information of the other three photolysis products is still unavailable and need further
421
investigation.
422
After irradiation, H-abstraction by UV attacking benzene ring lead to generation of
423
carbon-centered radical (pathway b), this was confirmed by EPR results in the former
424
part. As shown in Figure 4a, •OH was also detected in the photolysis to TCP alone,
425
thus the peroxy product as a key intermediate was taken into consideration in the
426
presence of oxygen. As illustrated in Figure 7, •OH can be obtained as a decay
427
product of M2 because of its instability (Motten et al., 1985; Li and Chignell, 1987).
428
Thus, the hydroxylation products (P3 and P4) were formed during the photolysis of
429
TCP. 22
430
431 432
Figure 7. Proposed reaction pathways of TCP under simulated solar light.
433 434
Four products were characterized in the direct photolysis process of TCP. The
435
products deserve special concern as they may cause toxicity effects on the organisms
436
in water environments. Thus, for accurate ecological risk assessment of OPEs,
437
different photodegradation pathways that may lead to different photoproducts should
438
also be considered.
439 440
4 Conclusions
441
In the present study, the photodegradation of TCP was investigated in aqueous
442
solution under different UV conditions. A half-life at 67.96 h were obtained under 23
443
simulated sunlight, which means direct photolysis may account little for TCP
444
depletion. The initial concentration of TCP, and photoactive components of the
445
surface water including NO2-, Fe3+ and HA were capable of influencing the
446
phototransformation rate constant of TCP. EPR analysis showed that the
447
environmental factors could affect the photodegradation process by the involvement
448
of reactive radicals under irradiation. Carbon-centered radical and hydroxyl radical
449
were produced during the photolysis process and hydroxyl radical contributed to the
450
degradation of TCP in the presence of NO2- and Fe3+. It is the first time that photolysis
451
products of TCP were identified by HPLC-LTQ-Orbitrap MS and possible
452
degradation pathways were proposed, involving the cleavage of a phenoxy bond, C-H
453
bond homogenization from the benzene ring structure of TCP, photoinduced
454
hydrolysis and hydroxylation. Moreover, the toxicity and possible hazards of the
455
products are quite unknown and should be further studied.
456 457
Acknowledgments
458 459
This research was financially supported by National Natural Science Foundation of China (21677023).
460 461 462
24
463
References:
464
Ahn, Y.B., Rhee, S.K., Fennell, D.E., Kekhof, L.J., Hentschel, U., Häggblom, M.M.,
465
2003. Reductive dehalogenation of brominated phenolic compounds by
466
microorganisms associated with the marine sponge aplysina aerophoba. Appl.
467
Environ. Microb. 69, 4159-4166.
468
Antonopoulou, M., Karagianni, P., Konstantinou, I.K., 2016. Kinetic and mechanistic
469
study of photocatalytic degradation of flameretardant Tris (1-chloro-2-propyl)
470
phosphate (TCPP), Appl. Catal. B: Environ. 192, 152-160.
471
Antonopoulou, M., Giannakas, A., Bairamis, F., Papadaki, M., Konstantinou, I., 2017.
472
Degradation of organophosphorus flame retardant tris (1-chloro-2-propyl)
473
phosphate (TCPP) by visible light N,S-codoped TiO2 photocatalysts. Chem. Eng.
474
J. 318, 231-239.
475
Arakaki, T., 1999. pH dependent photoformation of hydroxyl radical and absorbance
476
of aqueous-phase N (III) (HNO2 and NO2-). Environ. Sci. Technol. 33,
477
2561-2565.
478
Beitz, T., Bechmann, W., Mitzner, R., 1999. Investigation on the photoreactions of
479
nitrate and nitrite ions with selected azaarenes in water. Chemosphere, 38,
480
351-361.
481
Burbano, A.A., Dionysiou, D.D., Suidan, M.T., Richardson, T.L., 2005. Oxidation
482
kinetics and effect of pH on the degradation of MTBE with Fenton reagent,
483
Water Res. 39, 107-118.
484
Brandsma, S.H. de Boer,J. Leonards, P.E. Cofino, W.P. Covaci, A., 2013.
485
Organophosphorus
flame-retardant
and
plasticizer
analysis,
including
486
recommendations from the first worldwide inter laboratory study, TrAC Trends
487
Anal. Chem. 43, 217–228.
488
Calza, P., Vione, D., Minero, C., 2014. The role of humic and fulvic acids in the
489
phototransformation of phenolic compounds in seawater. Sci. Total Environ. 493,
490
411-418.
491
Calza, P., Vione, D., Novelli, A., Pelizzetti, E., Minero, C. 2012. The role of nitrite
492
and nitrate ions as photosensitizers in the phototransformation of phenolic
493
compounds in seawater. Sci. Total Environ. 439, 67-75.
494
Chen, Y., Zhang, K., Zuo, Y., 2013. Direct and indirect photodegradation of estriol in
495
the presence of humic acid, nitrate and iron complexes in water solutions. Sci
496
Total Environ. 463-464, 802-809.
497
Chowdhury, R.R., Charpentier, P.A., Ray, M.B., 2011. Photodegradation of 25
498
17β-estradiol in aquatic solution under solar irradiation: Kinetics and influencing
499
water parameters. J. Photochem. Photobiol. A Chem. 219, 67-75.
500
Cristale, J., Katsoyiannis, A., Sweetman, A.J., Jones, K.C., Lacorte, S., 2013.
501
Occurrence and risk assessment of organophosphorus and brominated flame
502
retardants in the River Aire (UK). Environ. Pollut. 179, 194-200.
503
Cristale, J., Dantas, R.F., Luca, A.D., Sans, C., Esplugas, S., Lacorte, S., 2017. Role
504
of
oxygen
and
DOM
in
sunlight
induced
photodegradation
of
505
organophosphorous flame retardants in river water. J. Hazard. Mater. 323,
506
242-249.
507
Du, Z., Wang, G., Gao, S., Wang, Z., 2015. Aryl organophosphate flame retardants
508
induced cardiotoxicity during zebrafish embryogenesis: By disturbing expression
509
of the transcriptional regulators. Aquat. Toxicol. 161, 25-32.
510
Koumaki, E., Mamais, D., Noutsopoulos, C., Nika, M., Bletsou, A.A., Thomaidis,
511
N.S., Eftaxias, A., Stratogianni, G. 2015. Degradation of emerging contaminants
512
from water under natural sunlight: The effect of season, pH, humic acids and
513
nitrate and identification of photodegradation by-products. Chemosphere, 138,
514
675-681.
515
Greaves, A.K. Letcher, R.J., 2014. Comparative body compartment composition and
516
in ovo transfer of organophosphate flame retardants in North American Great
517
Lakes herring gulls. Environ. Sci. Technol. 48, 7942-7950.
518
Hou, R., Xu, Y., Wang, Z., 2016. Review of OPFRs in animals and humans:
519
Absorption, bioaccumulation, metabolism, and internal exposure research.
520
Chemosphere 153 (2016) 78-90.
521
Huang, J., Li, X., Ma, M., Li, D., 2017. Removal of di-(2-ethylhexyl) phthalate from
522
aqueous solution by UV/peroxymonosulfate: Influencing factors and reaction
523
pathways. Chem. Eng. J. 314, 182-191.
524
Kim, U.J., Kannan, K., 2018. Occurrence and distribution of organophosphate flame
525
retardants/plasticizers in surface waters, tap water, and rainwater – Implications
526
for human exposure. Environ. Sci. Technol. 52 (2018) 5625-5633.
527
Kojima, H., Takeuchi, S., Eede, N.V.D., Covaci, A., 2016. Effects of primary
528
metabolites of organophosphate flame retardants on transcriptional activity via
529
human nuclear receptors. Toxicol. Lett. 245, 31-39.
530
Kouras-Hadel, S., Amine-Khodja, A., Halladja, S., Richard, C., 2012. Influence of
531
humic
substance
on
the
nboflavin
532
2,4,6-trimethyphenol, J. Photochem. Photobiol. A 229, 33-38. 26
photosensitized
trandformation
of
533
Koumaki, E., Mamais, D., Noutsopoulos, C., Nika, M.C., Bletsou, A.A., Thomaidis,
534
N.S., Eftaxias, A., Stratogianni, G., 2015. Degradation of emerging contaminants
535
from water under natural sunlight: The effect of season, pH, humic acids and
536
nitrate and identification of photodegradation by-products. Chemosphere 138,
537
675-681.
538
Li, A.S.W., Chignell, C.F., 1987. Spectroscopic studies of cutaneous photosensitizing
539
agents–xii. Spin trapping study of the free radicals generated during the
540
photolysis of photo allergens bithionol and fentichlor. Photochem. Photobiol. 46,
541
445-452.
542
Li, A.S.W., Chignell, C.F., 1991. The NoH value in EPR spin trapping: a new
543
parameter for the identification of 5,5-dimethyl-1-pyrroline-N-oxide spin adducts.
544
J. Biochem. Bioph. Methods. 22, 83-87.
545
Li, C., Wei, G., J. Chen, Zhao, Y., Zhang, Y., Su, L., Qin, W., 2018. Aqueous OH
546
Radical Reaction Rate Constants for Organophosphorus Flame Retardants and
547
Plasticizers: Experimental and Modeling Studies. Environ. Sci. Technol. 52,
548
2790-2799.
549
Liu, J., Ye, J., Chen, Y., Li, C., Ou, H., 2018. UV-driven hydroxyl radical oxidation of
550
tris(2-chloroethyl) phosphate: Intermediate products and residual toxicity.
551
Chemosphere, 190, 225-233.
552
Lee, S., Jeong, W., Kannan, K., Moon, H., 2016. Occurrence and exposure assessment
553
of organophosphate flame retardants (OPFRs) through the consumption of
554
drinking water in Korea. Water Res. 103, 182-188.
555 556
Mack, J., 1999. Photochemistry of nitrite and nitrate in aqueous solution: A review. J. Photochem. Photobiol.A Chem. 128, 1-13.
557
Marvin, C.H., Sverko, E., Tomy, G.T., 2016. Analysis of Chlorinated and Phosphorus
558
Flame Retardants. In: Alaee M. (eds) Dioxin and Related Compounds. The
559
Handbook of Environmental Chemistry, pp, 411-456.
560
Motten, A.G., Buettner, G.R., Chignell, C.F., 1985. Spectroscopic studies of
561
cutaneous photosensitizing agents-VIII. A spin-trapping study of light induced
562
free radicals from chlorpromazine and promazine. Photochem. Photobiol. 42,
563
9-15.
564
Neamtu, M., Frimmel, F.H., 2006. Photodegradation of endocrine disrupting chemical
565
nonylphenol b y simulated solar UV-irradiation. Sci. Total Environ. 369,
566
295-306.
567
Ou, H., Liu, J., Ye, J., Wang, L., Gao, N., Ke, J., 2017. Degradation of 27
568
tris(2-chloroethyl) phosphate by ultraviolet-persulfate: Kinetics, pathway and
569
intermediate impact on proteome of Escherichia coli. Chem. Eng. J. 308,
570
386-395.
571
Peng, J., Wang, G., Zhang, D., Zhang, D., Li, X., 2016. Photodegradation of
572
nonylphenol in aqueous solution by simulated solar UV-irradiation: The
573
comprehensive effect of nitrate, ferric ion and bicarbonate. J. Photochem.
574
Photobiol. A: Chem. 326, 9-15.
575
Poerschmann, J., Trommler, U., Górecki, T., Kopinke, F.D., 2009. Formation of
576
chlorinated biphenyls, diphenyl ethers and benzofurans as a result of
577
Fenton-driven oxidation of 2-chlorophenol. Chemosphere, 75, 772-780.
578
Regnery J., Püttmann, W. 2010. Occurrence and fate of organophosphorus flame
579
retardants and plasticizers in urban and remote surface waters in Germany. Water
580
Res., 44, 4097-4104.
581
Rodriguez, I., Calvo, F., Quintana, J.B., Rubí, E., Rodil, R., Cela, R., 2016. Suitability
582
of solid-phase microextraction for the determination of organophosphate flame
583
retardants and plasticizers in water samples, J. Chromatogr. A, 1108, 158-165.
584
Shi, Y., Guo, L. Li, W. Wang, Y., Liu, J., Cai, Y., 2016. Occurrence, distribution and
585
seasonal variation of organophosphate flame retardants and plasticizers in urban
586
surface water in Beijing, China. Environ. Pollut. 209, 1-10.
587
Stan, S.D., Woods, J.S., Daeschel, M.A., 2005. Investigation of the presence of OH
588
radicals in electrolyzed NaCl solution by electron spin resonance spectroscopy. J
589
Agric. Food Chem. 53, 4901-4905.
590
Su, G., Crump, D., Letcher, R.J., Kennedy, S.W., 2014. Rapid in vitro metabolism of
591
the flame retardant triphenyl phosphate and effects on cytotoxicity and mRNA
592
expression in chicken embryonic hepatocytes. Environ. Sci. Technol. 48,
593
13511-13519.
594
Suh, Y.W., Buettner, G.R., Venkataraman, S., Treimer, S.E., Robertson, L.W.,
595
Ludewig, G., 2009. UVA/B-Induced formation of free radicals from
596
decabromodiphenyl ether. Environ. Sci. Technol. 43, 2581-2588.
597
van der Veen, I., Boer, J.D., 2012. Phosphorus flame retardants: Properties, production,
598
environmental occurrence, toxicity and analysis, Chemosphere 88, 1119-1153.
599
Vione, D., Maurino, V., Minero, C., Pelizzetti, E., 2002. New processes in the
600
environmental chemistry of nitrite: nitration of phenol upon nitrite photoinduced
601
oxidation. Environ. Sci. Technol. 36, 669-676.
602
Voelker, B.M., Morel, F.M.M., Sulzberger, B., 1997. Iron redox cycling in surface 28
603
waters: effects of humic substances and light. Environ. Sci. Technol. 31,
604
1004-1011.
605
Watts, M.J., Linden, K.G., 2009. Advanced oxidation kinetics of aqueous trialkyl
606
phosphate flame retardants and plasticizers. Environ. Sci. Technol. 43,
607
2937-2942.
608
Wang, Y., Feng, C., Li, Y., Gao, J., Yu, C., 2017. Enhancement of emerging
609
contaminants removal using fenton reaction driven by H2O2-producing microbial
610
fuel cells. Chem. Eng. J. 307, 679-686.
611
Wang, R., Tang, J., Xie, Z., Mi, W., Chen, Y., Wolschke, H., Tian, C., Pan, X., Luo, Y.,
612
Ebinghaus, R., 2015. Occurrence and spatial distribution of organophosphate
613
ester flame retardants and plasticizers in 40 rivers draining into the Bohai Sea,
614
north China. Environ. Pollut. 198, 172-178.
615
Wang, Q., Lai, N.L., Wang, X., Guo, Y., Lam, P.K., Lam, J.C., Zhou, B., 2015.
616
Bioconcentration
and
transfer
of
the
organophorous
flame
retardant
617
1,3-Dichloro-2-propyl phosphate causes thyroid endocrine disruption and
618
developmental neurotoxicity in zebrafish larvae. Environ. Sci. Technol. 49,
619
5123-5132.
620
Wei, G., Li, D., Zhuo, M., Liao, Y., Xie, Z., Guo, T., Li, J., Zhang, S., Liang, Z., 2015.
621
Organophosphorus flame retardants and plasticizers: Sources, occurrence,
622
toxicity and human exposure, Environ. Pollut. 196, 29-46.
623
Wenk, J., Eustis, S.N., McNeill, K., Canonica, S., 2013. Quenching of excited triplet
624
states by dissolved natural organic matter. Environ. Sci. Technol. 47,
625
12802-12810.
626 627
Yang, F., Ding, J., Huang, W., Xie, W., Liu, W., 2014. Particle size-specific
628
distributions and preliminary exposure assessments of organophosphate flame
629
retardants in office air particulate matter. Environ. Sci. Technol. 48, 63-70.
630
Yu, C., Zhang, Y. Quan, X., Chen, S., Han, J., Ou, X., Zhao, J., 2010. Photochemical
631
effect of humic acid components separated using molecular imprinting method
632
applying porphyrin-like substances as templates in aqueous solution. Environ.
633
Sci. Technol. 44, 5812-5817.
634
Yuan, X., Lacorte, S., Cristale, J., Dantas, R.F., Sans, C., Esplugas, S., Qiang, Z., 2015.
635
Removal of organophosphate esters from municipal secondary effluent by ozone
636
and UV/H2O2 treatments. Sep. Purif. Technol. 156, 1028-1034.
637
Zhang, Q., Ji, C., Yin, X., Yan, Lu, L., Zhao, M., 2016. Thyroid hormone-disrupting 29
638
activity and ecological risk assessment of phosphorus-containing flame
639
retardants by in vitro, in vivo and in silico approaches. Environ. Pollut. 210,
640
27-33.
641
Zhao, Q., Zhao, H., Quan, X., Chen, S., Zhang, Y., 2014. Photochemical
642
transformation of 2,2′,4,4′-tetrabromodiphenyl ether (BDE-47) in surface coastal
643
waters: Effects of chloride and ferric ions. Mar. Pollut. Bull. 86, 76-83.
644
Zhou, D., Huang, W., Wu, F., Han, C., Chen, Y., 2010. Photodegradation of
645
chloromycetin in aqueous solutions: kinetics and influencing factors. React.
646
Kinet. Mech. Cat. 100, 45-53.
647
Zhao, H., Jiang, J., Wang, Y., Lehmler, H., Buettner, G.R., Quan, X., Chen, J., 2015a.
648
Monohydroxylated
649
dihydroxylated polybrominated biphenyls (Di-OH-PBBs): Novel photoproducts
650
of 2,6-dibromophenol. Environ. Sci. Technol., 49, 14120-14128.
651
polybrominated
diphenyl
ethers
(OH-PBDEs)
and
Zhao, Q., Zhao, H., Quan, X., He, X., Chen, S., 2015b. Photochemical formation of
652
hydroxylated
polybrominated
diphenyl
653
polybrominated diphenyl ethers (PBDEs) in aqueous solution under simulated
654
solar light irradiation. Environ. Sci. Technol. 49, 9092-9099.
655
30
ethers
(OH-PBDEs)
from
Highlights Direct and indirect photolysis of TCP in water were investigated under irradiation. EPR was used for deducing effects of Fe3+, NO2- and HA on phototransformation of TCP. Carbon-centered radical and ·OH were formed during the photolysis process of TCP. Four photoproducts were identified by HPLC-LTQ-Orbitrap MS analysis.