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Pyrite enables persulfate activation for efficient atrazine degradation
Journal Pre-proof Pyrite enables persulfate activation for efficient atrazine degradation Xiaobing Wang, Yueyao Wang, Na Chen, Yanbiao Shi, Lizhi Zhang PII:
S0045-6535(19)32808-5
DOI:
https://doi.org/10.1016/j.chemosphere.2019.125568
Reference:
CHEM 125568
To appear in:
ECSN
Received Date: 30 September 2019 Revised Date:
2 December 2019
Accepted Date: 6 December 2019
Please cite this article as: Wang, X., Wang, Y., Chen, N., Shi, Y., Zhang, L., Pyrite enables persulfate activation for efficient atrazine degradation, Chemosphere (2020), doi: https://doi.org/10.1016/ j.chemosphere.2019.125568. 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.
Graphical Abstract
1
1
Pyrite Enables Persulfate Activation for Efficient Atrazine Degradation
2 3
Xiaobing Wang, Yueyao Wang, Na Chen, Yanbiao Shi, and Lizhi Zhang*
4
Key Laboratory of Pesticide & Chemical Biology of Ministry of Education, Institute of Applied &
5
Environmental Chemistry, College of Chemistry, Central China Normal University, Wuhan 430079, P.
6
R. China
7 8 9 10 11 12 13 14 15 16 17 18 19 20
*
Corresponding author 1
21
Phone/Fax: +86-27-6786 7535. E-mail addresses: [email protected]. (Prof. Lizhi Zhang)
22
Abstract
23
Persulfate (PS) is widely used for environmental remediation, but its organic contaminant removal
24
performance strongly depends on its activation. In this study, we demonstrate that pyrite (FeS2) can
25
more effectively activate PS than the commonly used FeSO4 for atrazine degradation. When 3.0 mM
26
of PS and 4.2 mM of iron salts were used, the atrazine degradation efficiency of FeS2/PS was 1.4
27
times that of FeSO4/PS, while the amount of consumed PS in case of FeS2 was only 53% of that by
28
FeSO4. The better PS activation performance of FeS2 could be attributed to its slow and sustainable
29
release of dissolved Fe(II), inhibiting the quenching reaction between •SO4-/•OH and Fe(II) ions, and
30
thus producing more reactive oxygen species for the atrazine degradation. More importantly, the
31
surface bound Fe(II) of FeS2 could activate molecular oxygen to generate superoxide radical (•O2-),
32
which could further promote the effective decomposition of PS by accelerating the Fe(III)/Fe(II)
33
redox cycle. This study unravels the roles of dissolved Fe(II) and surface bound Fe(II) on the
34
persulfate activation, and provides a promising heterogeneous persulfate activator for pollutant
35
control and environmental remediation.
36 37
Keywords: Persulfate activation; Pyrite; Iron-contained minerals; Fe(III)/Fe(II) cycle; Atrazine
38
degradation
39
2
40
1. Introduction
41
In the last decade, advanced oxidation processes (AOPs) have been widely utilized to remove
42
refractory organic contaminants owing to their superior degradation and mineralization efficiency
43
(Esplugas et al., 2002; Neyens and Baeyens, 2003; Pignatello et al., 2006; Landsman et al., 2007;
44
Ahn et al., 2013; Kim et al., 2018). Among the AOPs, persulfate (PS) oxidation technology is very
45
promising for organic pollutant treatment in view of the high oxidant property, the stability at room
46
temperature and the outstanding mobility in subsurface environments (Ahn et al., 2013; Fang et al.,
47
2013; Matzek and Carter, 2016; Kim et al., 2018). With the activation of PS, sulfate radical (•SO4-)
48
can be generated by the cleavage of O-O bond of PS. Meanwhile, •SO4- could react with water and
49
OH- to produce hydroxyl radicals (•OH) (Furman et al., 2010; Ji et al., 2015; Matzek and Carter,
50
2016 ). These two reactive oxygen species (ROSs) are responsible for the oxidation of contaminants
51
(Matzek and Carter, 2016). Obviously, the organic contaminant removal performance strongly
52
depends on the PS activation process.
53
PS can be activated by various methods, including UV-visible light, heat, base, microwave, carbon
54
materials, and transition metal contained materials (Furman et al., 2010; Yang et al., 2011; Gao et al.,
55
2012; Liu et al., 2012; Qi et al., 2014; Matzek and Carter, 2016). Among them, iron-contained
56
materials, with wide availability and environment friendly properties, attract considerable attentions
57
for the PS activation. For example, FeSO4, the most commonly used iron contained material, can
58
induce the rapid decomposition of PS to generate ROSs for organic contaminant removal in a
59
completely homogenous system (Ji et al., 2014; Rao et al., 2014; Shang et al., 2019; Zhu et al., 2019).
60
However, the large amounts of free Fe(II) ions generated from the dissolution of FeSO4 would 3
61
quickly scavenge the ROSs such as •SO4- and •OH (eqs 1 and 2), resulting in the poor utilization
62
efficiency of PS (Li et al., 2014). To solve this problem, scientists recently turn to develop
63
iron-contained minerals, such as hematite, goethite, ferrihydrite, and magnetite, as the
64
heterogeneous PS activators (Teel et al., 2011; Usman et al., 2012; Liu et al., 2016; Li et al., 2017;
65
Wu et al., 2017; Kermani et al., 2018). Unfortunately, most of these iron-contained materials
66
exhibited poor performance on the PS activation due to their insufficient Fe(II) content, hindering
67
their applications for the PS activation. Therefore, it is of great environmental significance to seek
68
for desirable iron minerals with abundant Fe(II) content for the heterogeneous activation of PS.
69
⋅SO 4 − + Fe 2 + → Fe3+ + SO 4 2 −
(1)
70
⋅OH + Fe2+ → Fe3+ + OH −
(2)
71
Pyrite (FeS2) is one of the most widely distributed Fe(II) contained minerals in the earth’s crust
72
(Hall, 2018). Among 13 naturally occurring minerals, including hematite, ilmenite, and so on,
73
pyrite could induce the most rapid decomposition of PS, which was attributed to its release of
74
sufficient dissolved Fe(II) in mildly acidic condition (Teel et al., 2011; Zhou et al., 2018).
75
Similarly, Zhang et al. demonstrated that the PS-pyrite system was effective for p-chloroaniline
76
degradation because of abundant free Fe(II) ions released from the pyrite (Zhang et al., 2017).
77
Besides dissolved Fe(II), pyrite contains plenty of surface bound Fe(II), which might be involved
78
in the PS activation process, but never be considered previously.
79
In this study, we systematically compared the PS activation in a homogenous system of free
80
Fe(II) ions derived from FeSO4 with that in a heterogeneous system of synthetic pyrite for the
81
degradation of atrazine (ATR), a typical herbicide which has been banned in many countries due 4
82
to its endocrine disruptive nature (Shen et al., 2018). The decomposition behavior of PS, the
83
generation of ROSs, and the variation of Fe species were investigated in detail. On the basis of these
84
these experimental results, the roles of dissolved Fe(II) and surface bound Fe(II) in the PS activation
85
were unraveled, and the ATR degradation pathway in FeS2/PS was also proposed.
86
2. Experimental
87
2.1. Chemicals and materials
88
Ferrous sulfate heptahydrate (FeSO4•7H2O), sodium thiosulfate pentahydrate (Na2S2O3•5H2O),
89
sulphur (S), persulfate sodium (PS, Na2S2O8), anhydrous sodium sulfate (Na2SO4), carbon disulfide
90
(CS2), ethanol, and tert-butyl alcohol (TBA) were bought from Aladdin Chemistry Co., Ltd. China.
91
ATR
92
5-dimethyl-1-pyrroline-N-oxide (DMPO) were supplied by Sigma-Aldrich. HPLC grade methanol,
93
acetonitrile, acetone and dichloromethane were bought from Fisher Scientific. The initial pH values
94
were adjusted by sulfuric acid (0.5 M) and sodium hydroxide (0.5 M) solutions. All the reagents were
95
used with at least analytical grade.
96
2.2. Preparation of FeS2
was
obtained
from
Alfa
Aesar.
Superoxide
dismutase
(SOD)
and 5,
97
FeS2 was synthesized with using a one-step hydrothermal method according to the previous report
98
(Liu et al., 2015). Briefly, FeSO4•7H2O (0.02 mol) was dissolved in 60 mL of Na2S2O3 solution (0.33
99
M). Then S powder (0.02 mol) was added into the solution, and the suspension was stirred for 10 min.
100
After that, the mixture was sealed in an 80 mL Teflon-lined stainless steel autoclave, and maintained
101
at 200 oC for 24 h. When the autoclave was naturally cooled down to room temperature, the
5
102
precipitate was centrifuged and then washed with distilled water, carbon disulfide and ethanol to
103
remove impurities, and finally dried at 50 oC in a vacuum oven overnight.
104
2.3. Materials analysis
105
The powder X-ray diffraction (XRD) pattern was recorded on a Bruker D8 Advance X-ray
106
diffractometer with Cu Kα radiation (λ = 0.15418 nm). The morphology of the as-prepared sample
107
was recorded by a scanning electron microscope (SEM, JEOL 6700-F, Japan). High-resolution
108
X-ray photoelectron spectroscopy (XPS) was performed on an ESCALAB 250Xi X-ray
109
photoelectron spectroscope. The XPS spectra for Fe 2p3/2 were fitted using XPSPEAK 4.0
110
software.
111
2.4. ATR degradation procedure
112
Batch trials were performed in conical glass flask with ATR solution (20 mL, 20 mg L-1) at 25
113
o
114
were successively added into the ATR solution. The degradation solution was sampled at
115
predetermined intervals and filtered using 0.22 µm nylon syringe filters, and then 0.1 mL of
116
ethanol was added immediately into 0.9 mL of the sample to terminate the reaction.
C. Then, FeS2 powder (0.01 g, 4.2 mM) or FeSO4•7H2O (0.0232 g, 4.2 mM) and PS (3.0 mM)
117
The ATR degradation processes with corresponding ROSs scavengers were conducted in
118
parallel for comparison. Ethanol was used as a scavenger of •OH and •SO4- (Hou et al., 2012).
119
Tert-butyl alcohol (TBA) was used as a scavenger of •OH and •SO4- (k•OH = (3.8-7.6) × 108 M-1 s-1,
120
k•SO4- = (4.0-9.1) × 105 M-1 s-1), but mainly a •OH scavenger because of its much lower reaction
121
rate constant with •SO4- (Ji et al., 2015). Superoxide dismutase (SOD) was as a scavenger of •O2-
122
(Chen et al., 2017). 6
123
2.5. Analytic methods
124
The ATR’s concentration was analyzed by high performance liquid chromatography (HPLC,
125
Ultimate 3000, Thermo, U.S.A.) with an Agilent A120-C18 column (5 µm; 4.6 mm × 150 mm). The
126
mobile phase contained 50% acetonitrile and 50% water (flow rate: 1.0 mL min-1). The maximum
127
absorption wavelength (λmax) of ATR was 220 nm. The PS concentration in the presence of iron was
128
determined by the iodometric titration method, with a UV-vis spectrophotometer (UV-2250,
129
Shimadzu, Japan) (λ = 352 nm) (Liang et al., 2008). Electron spin resonance (ESR) spectra were
130
measured using Bruker EPR A300 spectrometer (Bruker, Billerica, MA) at room temperature using
131
DMPO as the spin trapping agent. The dissolved ferrous ions and total dissolved iron ions were
132
quantified by the 1, 10-phenanthroline method with a UV-vis spectrophotometer (UV-2250,
133
Shimadzu, Japan) (Harvey et al., 1955). The analysis of total organic carbon (TOC) was carried out
134
on a Shimadzu TOC-VCPH analyzer to evaluate the mineralization of ATR. The intermediate
135
products were determined by high-performance liquid chromatography-tandem mass spectrometry
136
(LC-MS/MS, TSQ Quantum Access MAX, Thermo, U.S.A.) in positive ESI mode with a Hypersil
137
ODS-C18 column (5 µm; 150 mm × 2.1 mm). For pretreatment, the sample was extracted with
138
dichloromethane for three times. After the residual water completely removed by anhydrous sodium
139
sulfate, dichloromethane was evaporated using a rotary evaporator. Finally, the residual sample was
140
re-dissolved in methanol. The eluent contained 50% acetonitrile and 50% water. The flow rate was
141
maintained at 0.2 mL min-1.
142
The concentration of SO42- was conducted on ion chromatography (IC, Dionex ICS-900, Thermo,
143
U.S.A.) with an AS23 column. The concentration of Cl−, ammonium nitrogen (NH4+-N), 7
144
acetaldehyde and acetone were measured during the ATR degradation process, which were
145
described in the Supplementary Material (Text S1).
146
3. Results and discussion
147
3.1. Characterization of FeS2
148
Fig. 1a shows the XRD pattern of FeS2 at 2θ from 20o to 80o. The pattern illustrated that the
149
product mainly consisted of pyrite FeS2 (JCPDS PDF NO. 6-710) and a tiny amount of marcasite
150
FeS2 (JCPDS PDF NO. 3-799). The intergrowth of pyrite and marcasite usually occurred in
151
natural minerals formation and artificially synthesis in laboratory (Goldhaber et al., 1978; Lowson,
152
1982; Liu et al., 2015). Furthermore, the morphology of the sample was investigated by SEM at
153
different magnifications (Fig. 1b), which revealed the existence of irregular nanoblocks and
154
nanorods in the synthesized sample, in agreement with our previous report (Liu et al., 2015).
155 156
Fig. 1. (a) XRD patterns and (b) SEM images of the synthetic pyrite.
157
3.2. Atrazine degradation
8
158
As shown in Fig. 2a, 70% of ATR was degraded within 10 min and then its concentration did not
159
alter in the FeSO4/PS system. Although the ATR degradation percentage of FeS2/PS was lower than
160
that of FeSO4/PS before 30 min of reaction, the concentration of ATR decreased gradually in the
161
FeS2/PS system until about 100% of ATR was degraded in 45 min. Therefore, the final ATR
162
degradation percentage of FeS2/PS was 1.4 times that of FeSO4/PS, demonstrating the higher organic
163
pollutant degradation efficiency of FeS2/PS. Since pH value strongly affects the oxidation of organic
164
contaminants by PS, we monitored the variations of pH in the two systems (Fig. S1), and found pH
165
decreased from 7.0 to 2.7 in the FeSO4/PS system owing to the reaction between FeSO4 and S2O82-
166
(eq 3). As for the FeS2/PS system, pH also decreased sharply from 7.0 to 2.6, which might be
167
ascribed to the proton-releasing reaction between dissolved Fe(II) and S2O82- (eq 3), and the reaction
168
between FeS2 and O2/S2O82- (eqs 4 and 5) (Kang et al., 2011). To check how pH value affect the
169
ATR degradation, we conducted the ATR degradation at the initial pH 2.6 in the two systems. It was
170
found that near 100% of ATR was degraded within 15 min in the FeS2/PS system, much higher than
171
the ATR degradation percentage (63%) of FeSO4/PS in 90 min (Fig. S2). These results ruled out the
172
contribution of slightly lower pH value to the better ATR degradation performance of FeS2/PS.
173
Fe2++ S2O82-+ 3H2O
Fe(OH)3+•SO4- +SO42-+ 3H+
(3)
174
2FeS2+ 7O2+ 8H2O
2Fe2+ + 4SO42-+ 4H+
(4)
175
2FeS2 + 15S2O82- + 16H2O
176
2Fe3+ + 34SO42- + 32H+
(5)
3.3. Persulfate decomposition in FeSO4/PS and FeS2/PS systems
177
As mentioned previously, the organic contaminant removal performance of PS strongly depends
178
on its activation. Hence, we compared the PS decomposition performance of the two systems. In the 9
179
FeSO4/PS system, 78% of PS was decomposed within 30 min, significantly higher than that (41%)
180
of FeS2/PS (Fig. 2b), indicating that more PS was consumed in the FeSO4/PS system even with
181
lower ATR removal percentage, so pyrite could activate PS more effectively to produce ROSs for
182
the ATR degradation.
183 184
Fig. 2. (a) Time profiles of ATR degradation efficiency in FeS2/PS and FeSO4/PS. (b) Time profiles
185
of PS decomposition in FeS2/PS and FeSO4/PS. The initial concentrations of ATR, FeS2, FeSO4, and
186
PS were 20 mg L-1, 4.2 mM, 4.2 mM, and 3.0 mM, respectively. The initial pH values were 7.0.
187
Error bars represent ± one standard deviation derived from triplicate experiments.
188
3.4. Role of dissolved Fe(II) ions in FeS2/PS system
189
Subsequently, trapping experiments were performed to check the generation of ROSs in the PS
190
activation processes. As shown in Fig. 3a, in the FeSO4/PS system, the degradation efficiency of
191
ATR decreased from 70% to 4% and 20% by adding ethanol and TBA, respectively. As for the
192
FeS2/PS system, the ATR degradation was completely inhibited by the addition of ethanol. Upon
193
the addition of TBA, the degradation percentage decreased from 100% to 44% (Fig. 3b). Therefore,
194
both •OH and •SO4- were the major ROSs for the ATR degradation in the two systems. To further 10
195
clarify the formation of ROSs, ESR technique was employed to investigate the production of •OH
196
and •SO4- with DMPO as the spin trapper. As shown in Fig. 3c and 3d, both •OH (αN = αβ-H = 14.9 G)
197
G) and •SO4- radicals (αN = 13.8 G, αβ-H = 9.5 G, αγ1-H = 1.44 G and αγ2-H = 0.79 G) were detected,
198
further confirming that both •OH and •SO4- were produced for the ATR degradation in the two
199
systems (Fang et al., 2013). More importantly, the intensity of the DMPO-•SO4- and DMPO-•OH
200
adducts were much stronger in the FeS2/PS system, revealing that more •SO4- and •OH were
201
generated by the FeS2 activation process than the FeSO4 one.
202
As reported previously, the second-order rate constants were as high as 4.6 × 109 M-1 s-1 for the
203
reaction between dissolved Fe(II) and •SO4-, and 3.2 × 108 M-1 s-1 for the reaction between dissolved
204
Fe(II) and •OH, so •SO4- and •OH could be scavenged rapidly by the dissolved Fe(II) in this study
205
(eqs 1 and 2) (Neyens and Baeyens, 2003; Qi et al., 2014). For the homogenous FeSO4/PS system,
206
the abundant dissolved Fe(II) would react with •SO4- and •OH immediately (Fig. S3), resulting in
207
their serious quenching. Although more PS was rapidly decomposed by abundant free Fe(II) ions,
208
most of generated ROSs was consumed by the dissolved Fe(II) simultaneously, leading to the poor
209
effective utilization of PS in the FeSO4/PS system. As for the FeS2/PS system, dissolved Fe(II) ions
210
were released more slowly into the aqueous solution and its concentration was significantly lower
211
than that in the homogenous FeSO4/PS system (Fig. S3), inhibiting the reaction between the
212
dissolved Fe(II) ions and •SO4-/•OH. Meanwhile, the sustainable release of dissolved Fe(II) ions (eq
213
4) could sustainably activate PS to produce ROSs (Zhang et al., 2017). Therefore, FeS2 exhibited a
214
better activation performance of PS with higher yield of •SO4- and •OH than FeSO4.
11
215 216
Fig. 3. Effects of scavengers on ATR degradation, (a) in FeSO4/PS, and (b) in FeS2/PS. Electron
217
paramagnetic resonance spectra for (c) •SO4- and •OH in FeS2/PS, (d) •SO4- and •OH in FeSO4/PS.
218
The initial concentrations of ATR, FeS2, FeSO4, PS, ethanol, TBA were 20 mg L-1, 4.2 mM, 4.2 mM,
219
3.0 mM, 200 mM, and 200 mM respectively. The initial pH values were 7.0. Error bars represent ±
220
one standard deviation derived from triplicate experiments. (Black squares represent DMPO-•OH
221
adduct and black circles represent DMPO-•SO4- adduct.).
222
3.5. Role of surface bound Fe(II) ions in FeS2/PS system
223
Regarding the existence of ferrous ions in these PS activation systems, •O2- might also be
224
generated via the molecular oxygen activation process with ferrous ions (eq 6). It was known that
225
•O2- play a significant role in the organic contaminant degradation in our previous study (Ai et al., 12
226
2013; Liu et al., 2014; Liu et al., 2015). Therefore, we checked the generation of •O2- in the two
227
systems with using ESR technique. It was found the intensity of DMPO-•O2- adducts in the FeS2/PS
228
FeS2/PS system was much stronger than that in the FeSO4/PS one (Fig. 4a and 4b), indicating more
229
•O2- was produced in the PS activation process by FeS2. This difference revealed that the surface
230
bound Fe(II) of FeS2 was more efficiently for the generation of •O2- than the dissolved Fe(II) ions.
231
Subsequently, trapping experiment was carried out to check the effect of •O2- on the ATR
232
degradation. The presence of SOD did not obviously affect the ATR degradation in the FeSO4/PS
233
system, revealing a negligible contribution of •O2- to the ATR degradation in the homogenous
234
FeSO4/PS system (Fig. 4c). Differently, the ATR degradation efficiency dramatically decreased from
235
100% to 30% in the FeS2/PS system (Fig. 4d), so •O2- played a crucial role on the ATR degradation
236
in the FeS2 activation process. We therefore employed SOD to check the effect of •O2- on the PS
237
decomposition behavior in the FeS2/PS system, and found the decomposition percentage of PS
238
slightly decreased from 41% to 39% within 90 min (Fig. 4e), indicating that •O2- might enhance the
239
effective utilization of PS by other ways, rather than via promoting its decomposition. Subsequently,
240
we compared the generation of •OH and •SO4- in the FeS2/PS system with and without SOD. As
241
expected, the formation of DMPO-•OH and DMPO-•SO4- severely decreased by adding SOD (Fig.
242
4f), confirming that •O2- was responsible for the effective utilization of PS, which might be arisen
243
from the abundant surface bound Fe(II) of FeS2.
244
Fe2+ + O2
Fe3+ + •O2 -
(6)
13
245 246
Fig. 4. (a) ESR spectra of •O2- in FeSO4/PS. (b) ESR spectra of •O2- in FeS2/PS. (c) Effect of •O2-
247
scavenger on ATR degradation in FeSO4/PS. (d) Effect of •O2- scavenger on ATR degradation in
248
FeS2/PS. (e) Time profiles of PS decomposition in FeS2/PS and FeS2/PS/SOD. (f) ESR spectra for
249
•SO4- and •OH in FeS2/PS and FeS2/PS/SOD. The initial concentrations of ATR, FeS2, FeSO4, PS,
250
and SOD were 20 mg L-1, 4.2 mM, 4.2 mM, 3.0 mM, and 150 U mL-1, respectively. The initial pH
251
values were 7.0. Error bars represent ± one standard deviation derived from triplicate experiments.
252
To further demonstrate the effect of surface bound Fe(II) on the •O2- generation for PS
253
activation, the ATR degradation performances in FeS2/PS under different atmospheric conditions
254
were compared. By bubbling N2 continuously, the ATR degradation efficiency decreased from 100%
255
to 42% within 90 min (Fig. S4a), indicating the indispensable role of molecular oxygen, which
256
could be activated by the surface Fe(II) to produce •O2-, further contributing to the PS activation.
257
Similarly, it was found the decomposition percentage of PS under N2 slightly decreased from 41%
258
to 39% within 90 min (Fig. S4b), which was in good agreement with the PS decomposition 14
259
performance in FeS2/PS/SOD (Fig. 4e), further confirming that the •O2- produced from the surface
260
bound Fe(II) could enhance the effective utilization of PS, rather than via promoting its
261
decomposition amount.
262
•SO4- could be produced by the reaction between •O2- and S2O82- (eq 7), and the reaction between
263
ferrous ions and S2O82- (eq 5), in which ferrous ions could be provided continuously via the reduction
264
of ferric ions by •O2- (eq 8) (Fang et al., 2013; Wu et al., 2016). However, the reaction rate constant
265
of •O2- and S2O82- was less than 105 M-1 s-1, much lower than the value (1.5 × 108 M-1 s-1) of ferric
266
ions and •O2-, indicating the latter is more favorable dynamically (King et al., 1995; Mártire and
267
Gonzalez, 1998; Burns et al., 2010; Fujii et al., 2010). Hence, the enhanced effective utilization of PS
268
might be mainly ascribed to the accelerated iron redox (Fe(III)/Fe(II)) cycle in the FeS2/PS system
269
owing to the reduction ability of •O2-. In the FeS2/PS system, the concentration of dissolved Fe(II)
270
was below 0.017 mM during the degradation process, and the dissolved Fe(III) concentration
271
increased gradually to 0.11 mM in 90 min (Fig. 5a). After the addition of SOD, the maximum
272
concentration (0.0018 mM) of dissolved Fe(II) was just about one-eleventh of that in the FeS2/PS
273
system without SOD (Fig. 5b), confirming the promoted iron redox cycle via the reduction of Fe(III)
274
by •O2-.
275
15
276
Fig. 5. (a) Time profiles of dissolved Fe(II), Fe(III) and total Fe generated in FeS2/PS. (b) Time
277
profiles of dissolved Fe(II), Fe(III) and total Fe generated in FeS2/PS/SOD. (c) Time profiles of SO42-
278
formation in FeS2/PS. The initial concentrations of ATR, FeS2, PS and SOD were 20 mg L-1, 4.2 mM,
279
3.0 mM, and 150 U mL-1, respectively. The initial pH values were 7.0. Error bars represent ± one
280
standard deviation derived from triplicate experiments.
281
During this FeS2 activation process, SO42- were produced via the PS decomposition and the
282
FeS2 oxidation. As the decomposition of one mole PS would generate two moles of SO42- (Gao
283
et al., 2012), and about 41% of PS (corresponding to 1.23 mM) was decomposed in the FeS2/PS
284
system, so it could be calculated that 2.46 mM of SO42- was produced from the PS
285
decomposition. Within 90 min, the accumulated SO42- was found to be 5.72 mM (Fig. 5c).
286
Therefore, SO42- produced from the FeS2 oxidation was calculated as 5.72 mM - 2.46 mM = 3.26
287
mM. However, the maximum concentration of dissolved iron ions was merely 0.127 mM, much
288
lower than the theoretical value (1.63 mM) converted from FeS2, indicating that the iron cycle
289
occurred mainly on the surface of FeS2. Moreover, the high-resolution XPS analysis of Fe2p
290
revealed that the surface bound Fe(II) proportion of FeS2 decreased slightly from 86% to 75%
291
after the reaction with PS (Fig. 6a and 6b), suggesting the regeneration of surface bound Fe(II)
292
on FeS2 by the efficient iron redox cycle. These surface Fe(II) could activate PS to produce
293
ROSs continuously (Zhang et al., 2017). Therefore, the surface bound Fe(II) of FeS2 plays an
294
indispensable role in the effective decomposition of PS. Although •SO4-/•OH might also be
295
quenched by the surface Fe(II) sites, the diffusion of •SO4-/•OH from the solution to the FeS2
296
surface and the subsequent surface complex formation would slow the quenching reaction of 16
297
•SO4-/•OH in the FeS2/PS system (Lin and Gurol, 1998). More importantly, the surface bound Fe(II)
298
concentration of FeS2 was much lower than that of dissolved Fe(II) in the homogenous FeSO4 system,
299
further reducing the ineffective consumption of ROSs in the FeS2/PS system.
300
S2O82-+ •O2-
301
Fe3+ + •O2-
•SO4- + SO42- + O2 Fe2+ +O2
(7) (8)
302 303
Fig. 6. High-resolution XPS spectra of pyrite (a) before and (b) after the activation process.
304
According to the above results, we unraveled the roles of dissolved Fe(II) and surface bound Fe(II)
305
of FeS2 in the PS activation process as follows (Scheme S1). First, dissolved Fe(II) ions was released
306
from pyrite slowly and sustainably. These dissolved Fe(II) and surface bound Fe(II) activated PS to
307
produce •SO4- and •OH. Owing to the slow release of dissolved Fe(II), the quenching reaction
308
between •SO4-/•OH and dissolved Fe(II) ions was significantly inhibited, leaving behind more
309
reactive oxygen species for the ATR degradation. Moreover, the surface bound Fe(II) of FeS2 could
310
activate molecular oxygen to generate •O2-, which could further promote the effective decomposition
311
of PS by accelerating the surface iron redox cycle of FeS2. These reasons accounted for the enhanced
312
PS utilization efficiency of FeS2 and the better ATR degradation performance of FeS2/PS. 17
313
Furthermore, we investigated the TOC removal efficiency of the two systems, and found that the
314
TOC removal efficiency of FeS2/PS was 26% within 7 hours, much higher than that (4%) of
315
FeSO4/PS, further confirming the superiority of FeS2 for the PS activation (Fig. S5).
316
When natural pyrite was used to replace the synthesized pyrite, 96% of ATR could be degraded
317
with 90 min, further confirming the desirable PS activation performance of FeS2 (Fig. S6, Fig. S7
318
and Text S2). Moreover, we conducted the ATR degradation experiments with several
319
environmentally relevant iron minerals, including hematite (Fe2O3), goethite (α-FeOOH),
320
magnetite (Fe3O4) and mackinawite (FeS) (Fig. S8). It was found that ATR could not be removed
321
efficiently by Fe2O3/PS and α-FeOOH/PS, suggesting the poor PS activation performance of Fe(III)
322
dominated iron minerals, which could be ascribed to their insufficient Fe(II) content. In the
323
Fe3O4/PS system, 30% of ATR was removed within 90 min, higher than those in the Fe2O3/PS and
324
α-FeOOH/PS systems. The result revealed that Fe(II) bearing minerals was more favorable for the
325
PS activation than Fe(III) dominated iron minerals. However, the ATR removal efficiency of
326
Fe3O4/PS was just one-third of that in FeS2/PS, which might be ascribed to the lower Fe(II)
327
content and higher surface stability of Fe3O4 in comparison with FeS2 (Elsner et al., 2004).
328
Interestingly, 98% of the ATR could be removed within 90 min in the FeS/PS system. However,
329
the low production of natural mackinawite limits its further application. The unit price of the
330
above mentioned iron-contained materials were provided in the Supplementary Material (Text S3
331
and Table S1). These results further reveal that FeS2 is a promising PS activator for organic
332
pollutant treatment.
333
3.6. Atrazine degradation pathway in the FeS2/PS system 18
334
The degradation intermediates of ATR produced in the FeS2/PS system were identified by
335
HPLC-MS/MS.
336
2-hydroxy-4-acetamindo-6-ethylamino-s-triazine
337
-s-triazine
338
2-hydroxy-4-ethylamino-6-amino-s-triazine
339
(OAAT). All these intermediates with mass spectra, structural formula and abbreviation were
340
summarized in Fig. S9. In addition, 2-chloro-4-isopropylamino-6-amino-s-tria-zine (CIAT) was
341
detected using HPLC with the same method for ATR determination. On the basis of these findings,
342
we tentatively proposed the degradation pathway of ATR in Scheme 1 (Choi et al., 2013; Mu et al.,
343
2019). First, dechlorination-hydroxylation took place via the •SO4-/•OH attack on the C-Cl bond of
344
ATR to produce HIET, and then the isopropyl chains of HIET could be oxidized to generate HAET
345
upon the attack of •SO4- and •OH. Subsequently, OEAT and OAAT were generated via the
346
dealkylation of the lateral chains. Meanwhile, CIAT was produced by the direct dealkylation of ATR,
347
which could be further oxidized via the dechlorination-hydroxylation process to generate OIAT.
348
Subsequently, the deamination-hydroxylation process of OIAT generated OOIT. It was possible that
349
OIAT might be further oxidized to generate ODAT via the dechlorination-hydroxylation process, and
350
then the dealkylation of the lateral chains of ODAT would also generate OAAT. Unfortunately,
351
ODAT was not detected in this study. The inorganic ions and low molecular weight organic
352
molecules, including Cl-, NH4+-N, acetaldehyde and acetone, were detected during the ATR
353
degradation, confirming its mineralization in the FeS2/PS system (Text S1 and Fig. S10).
They
included
(OOIT),
2-hydroxy-4-ethylamino-6-isopropylamino-s-triazine (HAET),
2,
4-dihydroxy-6-isopropylamino
2-hydroxy-4-isopropylamino-6-amino-s-triazine (OEAT),
19
and
(HIET),
2-hydroxy-4,
(OIAT),
6-diaminos-s-triazine
354 355
Scheme 1. Proposed degradation pathways of ATR in FeS2/PS.
356
4. Conclusion
357
Persulfate is widely used for environmental remediation, but its organic contaminant removal
358
performance strongly depends on its activation. Although the homogeneous persulfate activation
359
of FeSO4 could generate ROSs for rapid organic contaminant removal, the large amounts of
360
dissolved Fe(II) ions generated from the dissolution of FeSO4 would quickly scavenge the ROSs
361
such as •SO4- and •OH, lowering the utilization efficiency of persulfate. Therefore, the
362
heterogeneous activation of PS with desirable iron minerals is very promising for practical
363
application of persulfate. In this study, we demonstrated that pyrite could more effectively activate
364
PS than the commonly used FeSO4 for the atrazine degradation. The better PS activation
365
performance of FeS2 could be attributed to its slow and sustainable release of dissolved Fe(II),
366
inhibiting the quenching reaction between •SO4-/•OH and Fe(II) ions, and thus producing more
367
reactive oxygen species for the atrazine degradation. More importantly, the surface bound Fe(II) of 20
368
FeS2 could activate molecular oxygen to generate superoxide radical (•O2-), which could further
369
promote the effective decomposition of PS by accelerating the iron redox cycle. This study unravels
370
unravels the roles of dissolved Fe(II) and surface bound Fe(II) on the persulfate activation, and
371
provides a promising heterogeneous persulfate activator for pollutant control and environmental
372
remediation.
373
Acknowledgements
374
This work was financially supported by the National Key Research and Development Program of
375
China (Grant 2018YFC1800701), National Natural Science Funds for Distinguished Young Scholars
376
(Grant 21425728), the National Science Foundation of China (Grant 21936003 and 21872061), the
377
111 Project (Grant B17019), and the CAS Interdisciplinary Innovation Team of the Chinese
378
Academy of Sciences.
379
Declarations of interest
380
None.
381
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28
Highlights
Pyrite (FeS2) could more effectively activate persulfate (PS) than FeSO4 for atrazine degradation.
The slow and sustainable release of dissolved Fe(II) from FeS2 inhibited the quenching reaction between •SO4-/•OH and Fe(II) ions.
The surface Fe(II) of FeS2 could activate molecular oxygen to generate •O2-, which could promote the effective decomposition of PS by accelerating the Fe(III)/Fe(II) cycle.
The TOC removal efficiency of FeS2/PS was 26% within 7 hours, much higher than that (4%) of FeSO4/PS.
Author Contribution Statement Xiaobing Wang: Conceptualization, Investigation, Writing - Original Draft. Yueyao Wang: Formal analysis, Resources. Na Chen: Visualization. Yanbiao Shi: Software. Lizhi Zhang: Writing- Reviewing and Editing.
1
Declarations of interest: none.
S0045-6535(19)32808-5
DOI:
https://doi.org/10.1016/j.chemosphere.2019.125568
Reference:
CHEM 125568
To appear in:
ECSN
Received Date: 30 September 2019 Revised Date:
2 December 2019
Accepted Date: 6 December 2019
Please cite this article as: Wang, X., Wang, Y., Chen, N., Shi, Y., Zhang, L., Pyrite enables persulfate activation for efficient atrazine degradation, Chemosphere (2020), doi: https://doi.org/10.1016/ j.chemosphere.2019.125568. 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.
Graphical Abstract
1
1
Pyrite Enables Persulfate Activation for Efficient Atrazine Degradation
2 3
Xiaobing Wang, Yueyao Wang, Na Chen, Yanbiao Shi, and Lizhi Zhang*
4
Key Laboratory of Pesticide & Chemical Biology of Ministry of Education, Institute of Applied &
5
Environmental Chemistry, College of Chemistry, Central China Normal University, Wuhan 430079, P.
6
R. China
7 8 9 10 11 12 13 14 15 16 17 18 19 20
*
Corresponding author 1
21
Phone/Fax: +86-27-6786 7535. E-mail addresses: [email protected]. (Prof. Lizhi Zhang)
22
Abstract
23
Persulfate (PS) is widely used for environmental remediation, but its organic contaminant removal
24
performance strongly depends on its activation. In this study, we demonstrate that pyrite (FeS2) can
25
more effectively activate PS than the commonly used FeSO4 for atrazine degradation. When 3.0 mM
26
of PS and 4.2 mM of iron salts were used, the atrazine degradation efficiency of FeS2/PS was 1.4
27
times that of FeSO4/PS, while the amount of consumed PS in case of FeS2 was only 53% of that by
28
FeSO4. The better PS activation performance of FeS2 could be attributed to its slow and sustainable
29
release of dissolved Fe(II), inhibiting the quenching reaction between •SO4-/•OH and Fe(II) ions, and
30
thus producing more reactive oxygen species for the atrazine degradation. More importantly, the
31
surface bound Fe(II) of FeS2 could activate molecular oxygen to generate superoxide radical (•O2-),
32
which could further promote the effective decomposition of PS by accelerating the Fe(III)/Fe(II)
33
redox cycle. This study unravels the roles of dissolved Fe(II) and surface bound Fe(II) on the
34
persulfate activation, and provides a promising heterogeneous persulfate activator for pollutant
35
control and environmental remediation.
36 37
Keywords: Persulfate activation; Pyrite; Iron-contained minerals; Fe(III)/Fe(II) cycle; Atrazine
38
degradation
39
2
40
1. Introduction
41
In the last decade, advanced oxidation processes (AOPs) have been widely utilized to remove
42
refractory organic contaminants owing to their superior degradation and mineralization efficiency
43
(Esplugas et al., 2002; Neyens and Baeyens, 2003; Pignatello et al., 2006; Landsman et al., 2007;
44
Ahn et al., 2013; Kim et al., 2018). Among the AOPs, persulfate (PS) oxidation technology is very
45
promising for organic pollutant treatment in view of the high oxidant property, the stability at room
46
temperature and the outstanding mobility in subsurface environments (Ahn et al., 2013; Fang et al.,
47
2013; Matzek and Carter, 2016; Kim et al., 2018). With the activation of PS, sulfate radical (•SO4-)
48
can be generated by the cleavage of O-O bond of PS. Meanwhile, •SO4- could react with water and
49
OH- to produce hydroxyl radicals (•OH) (Furman et al., 2010; Ji et al., 2015; Matzek and Carter,
50
2016 ). These two reactive oxygen species (ROSs) are responsible for the oxidation of contaminants
51
(Matzek and Carter, 2016). Obviously, the organic contaminant removal performance strongly
52
depends on the PS activation process.
53
PS can be activated by various methods, including UV-visible light, heat, base, microwave, carbon
54
materials, and transition metal contained materials (Furman et al., 2010; Yang et al., 2011; Gao et al.,
55
2012; Liu et al., 2012; Qi et al., 2014; Matzek and Carter, 2016). Among them, iron-contained
56
materials, with wide availability and environment friendly properties, attract considerable attentions
57
for the PS activation. For example, FeSO4, the most commonly used iron contained material, can
58
induce the rapid decomposition of PS to generate ROSs for organic contaminant removal in a
59
completely homogenous system (Ji et al., 2014; Rao et al., 2014; Shang et al., 2019; Zhu et al., 2019).
60
However, the large amounts of free Fe(II) ions generated from the dissolution of FeSO4 would 3
61
quickly scavenge the ROSs such as •SO4- and •OH (eqs 1 and 2), resulting in the poor utilization
62
efficiency of PS (Li et al., 2014). To solve this problem, scientists recently turn to develop
63
iron-contained minerals, such as hematite, goethite, ferrihydrite, and magnetite, as the
64
heterogeneous PS activators (Teel et al., 2011; Usman et al., 2012; Liu et al., 2016; Li et al., 2017;
65
Wu et al., 2017; Kermani et al., 2018). Unfortunately, most of these iron-contained materials
66
exhibited poor performance on the PS activation due to their insufficient Fe(II) content, hindering
67
their applications for the PS activation. Therefore, it is of great environmental significance to seek
68
for desirable iron minerals with abundant Fe(II) content for the heterogeneous activation of PS.
69
⋅SO 4 − + Fe 2 + → Fe3+ + SO 4 2 −
(1)
70
⋅OH + Fe2+ → Fe3+ + OH −
(2)
71
Pyrite (FeS2) is one of the most widely distributed Fe(II) contained minerals in the earth’s crust
72
(Hall, 2018). Among 13 naturally occurring minerals, including hematite, ilmenite, and so on,
73
pyrite could induce the most rapid decomposition of PS, which was attributed to its release of
74
sufficient dissolved Fe(II) in mildly acidic condition (Teel et al., 2011; Zhou et al., 2018).
75
Similarly, Zhang et al. demonstrated that the PS-pyrite system was effective for p-chloroaniline
76
degradation because of abundant free Fe(II) ions released from the pyrite (Zhang et al., 2017).
77
Besides dissolved Fe(II), pyrite contains plenty of surface bound Fe(II), which might be involved
78
in the PS activation process, but never be considered previously.
79
In this study, we systematically compared the PS activation in a homogenous system of free
80
Fe(II) ions derived from FeSO4 with that in a heterogeneous system of synthetic pyrite for the
81
degradation of atrazine (ATR), a typical herbicide which has been banned in many countries due 4
82
to its endocrine disruptive nature (Shen et al., 2018). The decomposition behavior of PS, the
83
generation of ROSs, and the variation of Fe species were investigated in detail. On the basis of these
84
these experimental results, the roles of dissolved Fe(II) and surface bound Fe(II) in the PS activation
85
were unraveled, and the ATR degradation pathway in FeS2/PS was also proposed.
86
2. Experimental
87
2.1. Chemicals and materials
88
Ferrous sulfate heptahydrate (FeSO4•7H2O), sodium thiosulfate pentahydrate (Na2S2O3•5H2O),
89
sulphur (S), persulfate sodium (PS, Na2S2O8), anhydrous sodium sulfate (Na2SO4), carbon disulfide
90
(CS2), ethanol, and tert-butyl alcohol (TBA) were bought from Aladdin Chemistry Co., Ltd. China.
91
ATR
92
5-dimethyl-1-pyrroline-N-oxide (DMPO) were supplied by Sigma-Aldrich. HPLC grade methanol,
93
acetonitrile, acetone and dichloromethane were bought from Fisher Scientific. The initial pH values
94
were adjusted by sulfuric acid (0.5 M) and sodium hydroxide (0.5 M) solutions. All the reagents were
95
used with at least analytical grade.
96
2.2. Preparation of FeS2
was
obtained
from
Alfa
Aesar.
Superoxide
dismutase
(SOD)
and 5,
97
FeS2 was synthesized with using a one-step hydrothermal method according to the previous report
98
(Liu et al., 2015). Briefly, FeSO4•7H2O (0.02 mol) was dissolved in 60 mL of Na2S2O3 solution (0.33
99
M). Then S powder (0.02 mol) was added into the solution, and the suspension was stirred for 10 min.
100
After that, the mixture was sealed in an 80 mL Teflon-lined stainless steel autoclave, and maintained
101
at 200 oC for 24 h. When the autoclave was naturally cooled down to room temperature, the
5
102
precipitate was centrifuged and then washed with distilled water, carbon disulfide and ethanol to
103
remove impurities, and finally dried at 50 oC in a vacuum oven overnight.
104
2.3. Materials analysis
105
The powder X-ray diffraction (XRD) pattern was recorded on a Bruker D8 Advance X-ray
106
diffractometer with Cu Kα radiation (λ = 0.15418 nm). The morphology of the as-prepared sample
107
was recorded by a scanning electron microscope (SEM, JEOL 6700-F, Japan). High-resolution
108
X-ray photoelectron spectroscopy (XPS) was performed on an ESCALAB 250Xi X-ray
109
photoelectron spectroscope. The XPS spectra for Fe 2p3/2 were fitted using XPSPEAK 4.0
110
software.
111
2.4. ATR degradation procedure
112
Batch trials were performed in conical glass flask with ATR solution (20 mL, 20 mg L-1) at 25
113
o
114
were successively added into the ATR solution. The degradation solution was sampled at
115
predetermined intervals and filtered using 0.22 µm nylon syringe filters, and then 0.1 mL of
116
ethanol was added immediately into 0.9 mL of the sample to terminate the reaction.
C. Then, FeS2 powder (0.01 g, 4.2 mM) or FeSO4•7H2O (0.0232 g, 4.2 mM) and PS (3.0 mM)
117
The ATR degradation processes with corresponding ROSs scavengers were conducted in
118
parallel for comparison. Ethanol was used as a scavenger of •OH and •SO4- (Hou et al., 2012).
119
Tert-butyl alcohol (TBA) was used as a scavenger of •OH and •SO4- (k•OH = (3.8-7.6) × 108 M-1 s-1,
120
k•SO4- = (4.0-9.1) × 105 M-1 s-1), but mainly a •OH scavenger because of its much lower reaction
121
rate constant with •SO4- (Ji et al., 2015). Superoxide dismutase (SOD) was as a scavenger of •O2-
122
(Chen et al., 2017). 6
123
2.5. Analytic methods
124
The ATR’s concentration was analyzed by high performance liquid chromatography (HPLC,
125
Ultimate 3000, Thermo, U.S.A.) with an Agilent A120-C18 column (5 µm; 4.6 mm × 150 mm). The
126
mobile phase contained 50% acetonitrile and 50% water (flow rate: 1.0 mL min-1). The maximum
127
absorption wavelength (λmax) of ATR was 220 nm. The PS concentration in the presence of iron was
128
determined by the iodometric titration method, with a UV-vis spectrophotometer (UV-2250,
129
Shimadzu, Japan) (λ = 352 nm) (Liang et al., 2008). Electron spin resonance (ESR) spectra were
130
measured using Bruker EPR A300 spectrometer (Bruker, Billerica, MA) at room temperature using
131
DMPO as the spin trapping agent. The dissolved ferrous ions and total dissolved iron ions were
132
quantified by the 1, 10-phenanthroline method with a UV-vis spectrophotometer (UV-2250,
133
Shimadzu, Japan) (Harvey et al., 1955). The analysis of total organic carbon (TOC) was carried out
134
on a Shimadzu TOC-VCPH analyzer to evaluate the mineralization of ATR. The intermediate
135
products were determined by high-performance liquid chromatography-tandem mass spectrometry
136
(LC-MS/MS, TSQ Quantum Access MAX, Thermo, U.S.A.) in positive ESI mode with a Hypersil
137
ODS-C18 column (5 µm; 150 mm × 2.1 mm). For pretreatment, the sample was extracted with
138
dichloromethane for three times. After the residual water completely removed by anhydrous sodium
139
sulfate, dichloromethane was evaporated using a rotary evaporator. Finally, the residual sample was
140
re-dissolved in methanol. The eluent contained 50% acetonitrile and 50% water. The flow rate was
141
maintained at 0.2 mL min-1.
142
The concentration of SO42- was conducted on ion chromatography (IC, Dionex ICS-900, Thermo,
143
U.S.A.) with an AS23 column. The concentration of Cl−, ammonium nitrogen (NH4+-N), 7
144
acetaldehyde and acetone were measured during the ATR degradation process, which were
145
described in the Supplementary Material (Text S1).
146
3. Results and discussion
147
3.1. Characterization of FeS2
148
Fig. 1a shows the XRD pattern of FeS2 at 2θ from 20o to 80o. The pattern illustrated that the
149
product mainly consisted of pyrite FeS2 (JCPDS PDF NO. 6-710) and a tiny amount of marcasite
150
FeS2 (JCPDS PDF NO. 3-799). The intergrowth of pyrite and marcasite usually occurred in
151
natural minerals formation and artificially synthesis in laboratory (Goldhaber et al., 1978; Lowson,
152
1982; Liu et al., 2015). Furthermore, the morphology of the sample was investigated by SEM at
153
different magnifications (Fig. 1b), which revealed the existence of irregular nanoblocks and
154
nanorods in the synthesized sample, in agreement with our previous report (Liu et al., 2015).
155 156
Fig. 1. (a) XRD patterns and (b) SEM images of the synthetic pyrite.
157
3.2. Atrazine degradation
8
158
As shown in Fig. 2a, 70% of ATR was degraded within 10 min and then its concentration did not
159
alter in the FeSO4/PS system. Although the ATR degradation percentage of FeS2/PS was lower than
160
that of FeSO4/PS before 30 min of reaction, the concentration of ATR decreased gradually in the
161
FeS2/PS system until about 100% of ATR was degraded in 45 min. Therefore, the final ATR
162
degradation percentage of FeS2/PS was 1.4 times that of FeSO4/PS, demonstrating the higher organic
163
pollutant degradation efficiency of FeS2/PS. Since pH value strongly affects the oxidation of organic
164
contaminants by PS, we monitored the variations of pH in the two systems (Fig. S1), and found pH
165
decreased from 7.0 to 2.7 in the FeSO4/PS system owing to the reaction between FeSO4 and S2O82-
166
(eq 3). As for the FeS2/PS system, pH also decreased sharply from 7.0 to 2.6, which might be
167
ascribed to the proton-releasing reaction between dissolved Fe(II) and S2O82- (eq 3), and the reaction
168
between FeS2 and O2/S2O82- (eqs 4 and 5) (Kang et al., 2011). To check how pH value affect the
169
ATR degradation, we conducted the ATR degradation at the initial pH 2.6 in the two systems. It was
170
found that near 100% of ATR was degraded within 15 min in the FeS2/PS system, much higher than
171
the ATR degradation percentage (63%) of FeSO4/PS in 90 min (Fig. S2). These results ruled out the
172
contribution of slightly lower pH value to the better ATR degradation performance of FeS2/PS.
173
Fe2++ S2O82-+ 3H2O
Fe(OH)3+•SO4- +SO42-+ 3H+
(3)
174
2FeS2+ 7O2+ 8H2O
2Fe2+ + 4SO42-+ 4H+
(4)
175
2FeS2 + 15S2O82- + 16H2O
176
2Fe3+ + 34SO42- + 32H+
(5)
3.3. Persulfate decomposition in FeSO4/PS and FeS2/PS systems
177
As mentioned previously, the organic contaminant removal performance of PS strongly depends
178
on its activation. Hence, we compared the PS decomposition performance of the two systems. In the 9
179
FeSO4/PS system, 78% of PS was decomposed within 30 min, significantly higher than that (41%)
180
of FeS2/PS (Fig. 2b), indicating that more PS was consumed in the FeSO4/PS system even with
181
lower ATR removal percentage, so pyrite could activate PS more effectively to produce ROSs for
182
the ATR degradation.
183 184
Fig. 2. (a) Time profiles of ATR degradation efficiency in FeS2/PS and FeSO4/PS. (b) Time profiles
185
of PS decomposition in FeS2/PS and FeSO4/PS. The initial concentrations of ATR, FeS2, FeSO4, and
186
PS were 20 mg L-1, 4.2 mM, 4.2 mM, and 3.0 mM, respectively. The initial pH values were 7.0.
187
Error bars represent ± one standard deviation derived from triplicate experiments.
188
3.4. Role of dissolved Fe(II) ions in FeS2/PS system
189
Subsequently, trapping experiments were performed to check the generation of ROSs in the PS
190
activation processes. As shown in Fig. 3a, in the FeSO4/PS system, the degradation efficiency of
191
ATR decreased from 70% to 4% and 20% by adding ethanol and TBA, respectively. As for the
192
FeS2/PS system, the ATR degradation was completely inhibited by the addition of ethanol. Upon
193
the addition of TBA, the degradation percentage decreased from 100% to 44% (Fig. 3b). Therefore,
194
both •OH and •SO4- were the major ROSs for the ATR degradation in the two systems. To further 10
195
clarify the formation of ROSs, ESR technique was employed to investigate the production of •OH
196
and •SO4- with DMPO as the spin trapper. As shown in Fig. 3c and 3d, both •OH (αN = αβ-H = 14.9 G)
197
G) and •SO4- radicals (αN = 13.8 G, αβ-H = 9.5 G, αγ1-H = 1.44 G and αγ2-H = 0.79 G) were detected,
198
further confirming that both •OH and •SO4- were produced for the ATR degradation in the two
199
systems (Fang et al., 2013). More importantly, the intensity of the DMPO-•SO4- and DMPO-•OH
200
adducts were much stronger in the FeS2/PS system, revealing that more •SO4- and •OH were
201
generated by the FeS2 activation process than the FeSO4 one.
202
As reported previously, the second-order rate constants were as high as 4.6 × 109 M-1 s-1 for the
203
reaction between dissolved Fe(II) and •SO4-, and 3.2 × 108 M-1 s-1 for the reaction between dissolved
204
Fe(II) and •OH, so •SO4- and •OH could be scavenged rapidly by the dissolved Fe(II) in this study
205
(eqs 1 and 2) (Neyens and Baeyens, 2003; Qi et al., 2014). For the homogenous FeSO4/PS system,
206
the abundant dissolved Fe(II) would react with •SO4- and •OH immediately (Fig. S3), resulting in
207
their serious quenching. Although more PS was rapidly decomposed by abundant free Fe(II) ions,
208
most of generated ROSs was consumed by the dissolved Fe(II) simultaneously, leading to the poor
209
effective utilization of PS in the FeSO4/PS system. As for the FeS2/PS system, dissolved Fe(II) ions
210
were released more slowly into the aqueous solution and its concentration was significantly lower
211
than that in the homogenous FeSO4/PS system (Fig. S3), inhibiting the reaction between the
212
dissolved Fe(II) ions and •SO4-/•OH. Meanwhile, the sustainable release of dissolved Fe(II) ions (eq
213
4) could sustainably activate PS to produce ROSs (Zhang et al., 2017). Therefore, FeS2 exhibited a
214
better activation performance of PS with higher yield of •SO4- and •OH than FeSO4.
11
215 216
Fig. 3. Effects of scavengers on ATR degradation, (a) in FeSO4/PS, and (b) in FeS2/PS. Electron
217
paramagnetic resonance spectra for (c) •SO4- and •OH in FeS2/PS, (d) •SO4- and •OH in FeSO4/PS.
218
The initial concentrations of ATR, FeS2, FeSO4, PS, ethanol, TBA were 20 mg L-1, 4.2 mM, 4.2 mM,
219
3.0 mM, 200 mM, and 200 mM respectively. The initial pH values were 7.0. Error bars represent ±
220
one standard deviation derived from triplicate experiments. (Black squares represent DMPO-•OH
221
adduct and black circles represent DMPO-•SO4- adduct.).
222
3.5. Role of surface bound Fe(II) ions in FeS2/PS system
223
Regarding the existence of ferrous ions in these PS activation systems, •O2- might also be
224
generated via the molecular oxygen activation process with ferrous ions (eq 6). It was known that
225
•O2- play a significant role in the organic contaminant degradation in our previous study (Ai et al., 12
226
2013; Liu et al., 2014; Liu et al., 2015). Therefore, we checked the generation of •O2- in the two
227
systems with using ESR technique. It was found the intensity of DMPO-•O2- adducts in the FeS2/PS
228
FeS2/PS system was much stronger than that in the FeSO4/PS one (Fig. 4a and 4b), indicating more
229
•O2- was produced in the PS activation process by FeS2. This difference revealed that the surface
230
bound Fe(II) of FeS2 was more efficiently for the generation of •O2- than the dissolved Fe(II) ions.
231
Subsequently, trapping experiment was carried out to check the effect of •O2- on the ATR
232
degradation. The presence of SOD did not obviously affect the ATR degradation in the FeSO4/PS
233
system, revealing a negligible contribution of •O2- to the ATR degradation in the homogenous
234
FeSO4/PS system (Fig. 4c). Differently, the ATR degradation efficiency dramatically decreased from
235
100% to 30% in the FeS2/PS system (Fig. 4d), so •O2- played a crucial role on the ATR degradation
236
in the FeS2 activation process. We therefore employed SOD to check the effect of •O2- on the PS
237
decomposition behavior in the FeS2/PS system, and found the decomposition percentage of PS
238
slightly decreased from 41% to 39% within 90 min (Fig. 4e), indicating that •O2- might enhance the
239
effective utilization of PS by other ways, rather than via promoting its decomposition. Subsequently,
240
we compared the generation of •OH and •SO4- in the FeS2/PS system with and without SOD. As
241
expected, the formation of DMPO-•OH and DMPO-•SO4- severely decreased by adding SOD (Fig.
242
4f), confirming that •O2- was responsible for the effective utilization of PS, which might be arisen
243
from the abundant surface bound Fe(II) of FeS2.
244
Fe2+ + O2
Fe3+ + •O2 -
(6)
13
245 246
Fig. 4. (a) ESR spectra of •O2- in FeSO4/PS. (b) ESR spectra of •O2- in FeS2/PS. (c) Effect of •O2-
247
scavenger on ATR degradation in FeSO4/PS. (d) Effect of •O2- scavenger on ATR degradation in
248
FeS2/PS. (e) Time profiles of PS decomposition in FeS2/PS and FeS2/PS/SOD. (f) ESR spectra for
249
•SO4- and •OH in FeS2/PS and FeS2/PS/SOD. The initial concentrations of ATR, FeS2, FeSO4, PS,
250
and SOD were 20 mg L-1, 4.2 mM, 4.2 mM, 3.0 mM, and 150 U mL-1, respectively. The initial pH
251
values were 7.0. Error bars represent ± one standard deviation derived from triplicate experiments.
252
To further demonstrate the effect of surface bound Fe(II) on the •O2- generation for PS
253
activation, the ATR degradation performances in FeS2/PS under different atmospheric conditions
254
were compared. By bubbling N2 continuously, the ATR degradation efficiency decreased from 100%
255
to 42% within 90 min (Fig. S4a), indicating the indispensable role of molecular oxygen, which
256
could be activated by the surface Fe(II) to produce •O2-, further contributing to the PS activation.
257
Similarly, it was found the decomposition percentage of PS under N2 slightly decreased from 41%
258
to 39% within 90 min (Fig. S4b), which was in good agreement with the PS decomposition 14
259
performance in FeS2/PS/SOD (Fig. 4e), further confirming that the •O2- produced from the surface
260
bound Fe(II) could enhance the effective utilization of PS, rather than via promoting its
261
decomposition amount.
262
•SO4- could be produced by the reaction between •O2- and S2O82- (eq 7), and the reaction between
263
ferrous ions and S2O82- (eq 5), in which ferrous ions could be provided continuously via the reduction
264
of ferric ions by •O2- (eq 8) (Fang et al., 2013; Wu et al., 2016). However, the reaction rate constant
265
of •O2- and S2O82- was less than 105 M-1 s-1, much lower than the value (1.5 × 108 M-1 s-1) of ferric
266
ions and •O2-, indicating the latter is more favorable dynamically (King et al., 1995; Mártire and
267
Gonzalez, 1998; Burns et al., 2010; Fujii et al., 2010). Hence, the enhanced effective utilization of PS
268
might be mainly ascribed to the accelerated iron redox (Fe(III)/Fe(II)) cycle in the FeS2/PS system
269
owing to the reduction ability of •O2-. In the FeS2/PS system, the concentration of dissolved Fe(II)
270
was below 0.017 mM during the degradation process, and the dissolved Fe(III) concentration
271
increased gradually to 0.11 mM in 90 min (Fig. 5a). After the addition of SOD, the maximum
272
concentration (0.0018 mM) of dissolved Fe(II) was just about one-eleventh of that in the FeS2/PS
273
system without SOD (Fig. 5b), confirming the promoted iron redox cycle via the reduction of Fe(III)
274
by •O2-.
275
15
276
Fig. 5. (a) Time profiles of dissolved Fe(II), Fe(III) and total Fe generated in FeS2/PS. (b) Time
277
profiles of dissolved Fe(II), Fe(III) and total Fe generated in FeS2/PS/SOD. (c) Time profiles of SO42-
278
formation in FeS2/PS. The initial concentrations of ATR, FeS2, PS and SOD were 20 mg L-1, 4.2 mM,
279
3.0 mM, and 150 U mL-1, respectively. The initial pH values were 7.0. Error bars represent ± one
280
standard deviation derived from triplicate experiments.
281
During this FeS2 activation process, SO42- were produced via the PS decomposition and the
282
FeS2 oxidation. As the decomposition of one mole PS would generate two moles of SO42- (Gao
283
et al., 2012), and about 41% of PS (corresponding to 1.23 mM) was decomposed in the FeS2/PS
284
system, so it could be calculated that 2.46 mM of SO42- was produced from the PS
285
decomposition. Within 90 min, the accumulated SO42- was found to be 5.72 mM (Fig. 5c).
286
Therefore, SO42- produced from the FeS2 oxidation was calculated as 5.72 mM - 2.46 mM = 3.26
287
mM. However, the maximum concentration of dissolved iron ions was merely 0.127 mM, much
288
lower than the theoretical value (1.63 mM) converted from FeS2, indicating that the iron cycle
289
occurred mainly on the surface of FeS2. Moreover, the high-resolution XPS analysis of Fe2p
290
revealed that the surface bound Fe(II) proportion of FeS2 decreased slightly from 86% to 75%
291
after the reaction with PS (Fig. 6a and 6b), suggesting the regeneration of surface bound Fe(II)
292
on FeS2 by the efficient iron redox cycle. These surface Fe(II) could activate PS to produce
293
ROSs continuously (Zhang et al., 2017). Therefore, the surface bound Fe(II) of FeS2 plays an
294
indispensable role in the effective decomposition of PS. Although •SO4-/•OH might also be
295
quenched by the surface Fe(II) sites, the diffusion of •SO4-/•OH from the solution to the FeS2
296
surface and the subsequent surface complex formation would slow the quenching reaction of 16
297
•SO4-/•OH in the FeS2/PS system (Lin and Gurol, 1998). More importantly, the surface bound Fe(II)
298
concentration of FeS2 was much lower than that of dissolved Fe(II) in the homogenous FeSO4 system,
299
further reducing the ineffective consumption of ROSs in the FeS2/PS system.
300
S2O82-+ •O2-
301
Fe3+ + •O2-
•SO4- + SO42- + O2 Fe2+ +O2
(7) (8)
302 303
Fig. 6. High-resolution XPS spectra of pyrite (a) before and (b) after the activation process.
304
According to the above results, we unraveled the roles of dissolved Fe(II) and surface bound Fe(II)
305
of FeS2 in the PS activation process as follows (Scheme S1). First, dissolved Fe(II) ions was released
306
from pyrite slowly and sustainably. These dissolved Fe(II) and surface bound Fe(II) activated PS to
307
produce •SO4- and •OH. Owing to the slow release of dissolved Fe(II), the quenching reaction
308
between •SO4-/•OH and dissolved Fe(II) ions was significantly inhibited, leaving behind more
309
reactive oxygen species for the ATR degradation. Moreover, the surface bound Fe(II) of FeS2 could
310
activate molecular oxygen to generate •O2-, which could further promote the effective decomposition
311
of PS by accelerating the surface iron redox cycle of FeS2. These reasons accounted for the enhanced
312
PS utilization efficiency of FeS2 and the better ATR degradation performance of FeS2/PS. 17
313
Furthermore, we investigated the TOC removal efficiency of the two systems, and found that the
314
TOC removal efficiency of FeS2/PS was 26% within 7 hours, much higher than that (4%) of
315
FeSO4/PS, further confirming the superiority of FeS2 for the PS activation (Fig. S5).
316
When natural pyrite was used to replace the synthesized pyrite, 96% of ATR could be degraded
317
with 90 min, further confirming the desirable PS activation performance of FeS2 (Fig. S6, Fig. S7
318
and Text S2). Moreover, we conducted the ATR degradation experiments with several
319
environmentally relevant iron minerals, including hematite (Fe2O3), goethite (α-FeOOH),
320
magnetite (Fe3O4) and mackinawite (FeS) (Fig. S8). It was found that ATR could not be removed
321
efficiently by Fe2O3/PS and α-FeOOH/PS, suggesting the poor PS activation performance of Fe(III)
322
dominated iron minerals, which could be ascribed to their insufficient Fe(II) content. In the
323
Fe3O4/PS system, 30% of ATR was removed within 90 min, higher than those in the Fe2O3/PS and
324
α-FeOOH/PS systems. The result revealed that Fe(II) bearing minerals was more favorable for the
325
PS activation than Fe(III) dominated iron minerals. However, the ATR removal efficiency of
326
Fe3O4/PS was just one-third of that in FeS2/PS, which might be ascribed to the lower Fe(II)
327
content and higher surface stability of Fe3O4 in comparison with FeS2 (Elsner et al., 2004).
328
Interestingly, 98% of the ATR could be removed within 90 min in the FeS/PS system. However,
329
the low production of natural mackinawite limits its further application. The unit price of the
330
above mentioned iron-contained materials were provided in the Supplementary Material (Text S3
331
and Table S1). These results further reveal that FeS2 is a promising PS activator for organic
332
pollutant treatment.
333
3.6. Atrazine degradation pathway in the FeS2/PS system 18
334
The degradation intermediates of ATR produced in the FeS2/PS system were identified by
335
HPLC-MS/MS.
336
2-hydroxy-4-acetamindo-6-ethylamino-s-triazine
337
-s-triazine
338
2-hydroxy-4-ethylamino-6-amino-s-triazine
339
(OAAT). All these intermediates with mass spectra, structural formula and abbreviation were
340
summarized in Fig. S9. In addition, 2-chloro-4-isopropylamino-6-amino-s-tria-zine (CIAT) was
341
detected using HPLC with the same method for ATR determination. On the basis of these findings,
342
we tentatively proposed the degradation pathway of ATR in Scheme 1 (Choi et al., 2013; Mu et al.,
343
2019). First, dechlorination-hydroxylation took place via the •SO4-/•OH attack on the C-Cl bond of
344
ATR to produce HIET, and then the isopropyl chains of HIET could be oxidized to generate HAET
345
upon the attack of •SO4- and •OH. Subsequently, OEAT and OAAT were generated via the
346
dealkylation of the lateral chains. Meanwhile, CIAT was produced by the direct dealkylation of ATR,
347
which could be further oxidized via the dechlorination-hydroxylation process to generate OIAT.
348
Subsequently, the deamination-hydroxylation process of OIAT generated OOIT. It was possible that
349
OIAT might be further oxidized to generate ODAT via the dechlorination-hydroxylation process, and
350
then the dealkylation of the lateral chains of ODAT would also generate OAAT. Unfortunately,
351
ODAT was not detected in this study. The inorganic ions and low molecular weight organic
352
molecules, including Cl-, NH4+-N, acetaldehyde and acetone, were detected during the ATR
353
degradation, confirming its mineralization in the FeS2/PS system (Text S1 and Fig. S10).
They
included
(OOIT),
2-hydroxy-4-ethylamino-6-isopropylamino-s-triazine (HAET),
2,
4-dihydroxy-6-isopropylamino
2-hydroxy-4-isopropylamino-6-amino-s-triazine (OEAT),
19
and
(HIET),
2-hydroxy-4,
(OIAT),
6-diaminos-s-triazine
354 355
Scheme 1. Proposed degradation pathways of ATR in FeS2/PS.
356
4. Conclusion
357
Persulfate is widely used for environmental remediation, but its organic contaminant removal
358
performance strongly depends on its activation. Although the homogeneous persulfate activation
359
of FeSO4 could generate ROSs for rapid organic contaminant removal, the large amounts of
360
dissolved Fe(II) ions generated from the dissolution of FeSO4 would quickly scavenge the ROSs
361
such as •SO4- and •OH, lowering the utilization efficiency of persulfate. Therefore, the
362
heterogeneous activation of PS with desirable iron minerals is very promising for practical
363
application of persulfate. In this study, we demonstrated that pyrite could more effectively activate
364
PS than the commonly used FeSO4 for the atrazine degradation. The better PS activation
365
performance of FeS2 could be attributed to its slow and sustainable release of dissolved Fe(II),
366
inhibiting the quenching reaction between •SO4-/•OH and Fe(II) ions, and thus producing more
367
reactive oxygen species for the atrazine degradation. More importantly, the surface bound Fe(II) of 20
368
FeS2 could activate molecular oxygen to generate superoxide radical (•O2-), which could further
369
promote the effective decomposition of PS by accelerating the iron redox cycle. This study unravels
370
unravels the roles of dissolved Fe(II) and surface bound Fe(II) on the persulfate activation, and
371
provides a promising heterogeneous persulfate activator for pollutant control and environmental
372
remediation.
373
Acknowledgements
374
This work was financially supported by the National Key Research and Development Program of
375
China (Grant 2018YFC1800701), National Natural Science Funds for Distinguished Young Scholars
376
(Grant 21425728), the National Science Foundation of China (Grant 21936003 and 21872061), the
377
111 Project (Grant B17019), and the CAS Interdisciplinary Innovation Team of the Chinese
378
Academy of Sciences.
379
Declarations of interest
380
None.
381
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28
Highlights
Pyrite (FeS2) could more effectively activate persulfate (PS) than FeSO4 for atrazine degradation.
The slow and sustainable release of dissolved Fe(II) from FeS2 inhibited the quenching reaction between •SO4-/•OH and Fe(II) ions.
The surface Fe(II) of FeS2 could activate molecular oxygen to generate •O2-, which could promote the effective decomposition of PS by accelerating the Fe(III)/Fe(II) cycle.
The TOC removal efficiency of FeS2/PS was 26% within 7 hours, much higher than that (4%) of FeSO4/PS.
Author Contribution Statement Xiaobing Wang: Conceptualization, Investigation, Writing - Original Draft. Yueyao Wang: Formal analysis, Resources. Na Chen: Visualization. Yanbiao Shi: Software. Lizhi Zhang: Writing- Reviewing and Editing.
1
Declarations of interest: none.