Visible Light Driven Organic Pollutants Degradation with

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Visible Light Driven Organic Pollutants Degradation with Hydrothermally Carbonized Sewage Sludge and Oxalate Via Molecular Oxygen Activation Na Chen, Huan Shang, Shuangyi Tao, Xiaobing Wang, Guangming Zhan, Hao Li, Zhihui Ai, Jiakuan Yang, and Lizhi Zhang Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b03882 • Publication Date (Web): 09 Oct 2018 Downloaded from http://pubs.acs.org on October 9, 2018

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Environmental Science & Technology

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Visible Light Driven Organic Pollutants Degradation with

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Hydrothermally Carbonized Sewage Sludge and Oxalate Via

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Molecular Oxygen Activation

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Na Chen†, Huan Shang†, Shuangyi Tao‡, Xiaobing Wang †, Guangming Zhan†, Hao Li†,

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Zhihui Ai†, Jiakuan Yang‡, and Lizhi Zhang†,*

7 †Key

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Laboratory of Pesticide & Chemical Biology of Ministry of Education, Institute of

9

Environmental & Applied Chemistry, College of Chemistry, Central China Normal University,

10

Wuhan 430079, People’s Republic of China

11 12

‡School

of Environmental Science and Engineering, Huazhong University of Science and Technology, Wuhan 430074, People’s Republic of China

13 14

* To whom correspondence should be addressed. E-mail: [email protected]. Phone/Fax:

15

+86-27-6786 7535

16 17 18 19 20 21

1 ACS Paragon Plus Environment


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22 23 24

ABSTRACT Converting sewage sludge into functional environmental materials has become an

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attractive sewage sludge disposal route. In this study, we synthesize a sewage sludge-based material via

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a facile one-pot hydrothermal carbonization method, and construct a visible light molecular oxygen

27

activation system with hydrothermally carbonized sewage sludge (HTC-S) and oxalate, to degrade

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various organic pollutants. It was found that iron species of HTC-S could chelate with oxalate to

29

generate H2O2 via molecular oxygen activation under visible light, and also promote the H2O2

30

decomposition to produce •OH for the fast organic pollutants degradation. Taking sulfadimidine as the

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example, the apparent degradation rate of HTC-S/oxalate system was almost 5-20 times that of iron

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oxides/oxalate system. This outstanding degradation performance was attributed to the presence of

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iron-containing clay minerals in HTC-S, as confirmed by X-ray diffraction measurements and

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Mössbauer spectrometry. In the oxalate solution, these iron-containing clay minerals could be excited

35

more easily than common iron oxides under visible light, because the silicon species strongly interacted

36

with iron species in HTC-S to form Fe-O-Si bond, which lowered the excitation energy of Fe-oxalate

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complex. This work provides an alternative sewage sludge conversion pathway, and also sheds light on

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the environmental remediation applications of sewage sludge-based materials.

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Keywords: Sewage sludge; Hydrothermal carbonization; Molecular oxygen activation; Visible light;

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Oxalate

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Introduction

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Sewage sludge, a major byproduct produced from wastewater treatment plants, has been defined as a

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pollutant by the US Environmental Protection Agency.1 With the high-speed industrialization and

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urbanization, the annual production of sewage sludge exceeds 30 million tons in China, and will

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continue to increase in the future.2 Traditional sewage sludge managements such as landfilling, ocean

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discharge, and incineration are no longer recognized as environmentally sustainable techniques owing to

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their risks of secondary pollutions and lacking of materials recovery.3 It is therefore a great challenge to

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explore more eco-friendly and value-added routes to dispose sewage sludge.

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Recently, converting sewage sludge into functional environmental materials by pyrolysis attracts

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more and more attention.1 This pathway could alleviate the burden of sewage sludge disposal, and also

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remove pollutants from other contaminated water environment.3 Pyrolysis is a conventional process to

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heat pre-dried sewage sludge at temperatures of 350−800 oC in an anaerobic atmosphere.4 Besides being

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used as sorbents to remove harmful gases (e.g., SO2 and H2S) or organic pollutants (e.g., dyes and

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chlorinated organics),3, 5-8 pyrolyzed sewage sludge was also explored for some new applications.2, 9-14

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For instance, Dai’s group synthesized a sewage sludge-derived Fe-loading nanocomposite via a one-step

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pyrolysis method, and employed it as a stable heterogeneous photo-Fenton catalyst to remove

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rhodamine B and p-nitrophenol.15 Yuan et al. prepared a composite TiO2 photocatalyst with using

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sewage sludge as the support and the dopant source and found the resulting composite photocatalyst

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could remove p-nitrophenol more efficiently than TiO2 P25 under visible light.16 Despite these advances,

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pyrolysis still suffers from some drawbacks, such as energy-intensive pre-dry treatment and release of

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dioxins and furans and harmful gases (e.g. NOx, N2O and SO2) to atmosphere, to convert sewage sludge 3 ACS Paragon Plus Environment


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into functional environmental materials.17

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Hydrothermal carbonization (HTC) is an emerging thermal-chemical technique to treat feedstock at

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relatively lower temperature (150-350 oC) and self-generated pressure (2-6 MPa) with using water as the

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medium.4, 17 In view of high water content (over 95 wt%) and poor dewaterability of sewage sludge,

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HTC might be more energy-saving and eco-friendly for the sewage sludge conversion with using its

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indigenous water as reaction medium, as it can avoid the drawbacks of traditional pyrolysis method, and

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also reserve most of carbonaceous solid and transition metals in the final products.18,

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advantages have already aroused intensive investigation on recovering energy during hydrothermal

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treatment of sewage sludge.20-22 However, the utilization of hydrothermally carbonized sewage sludge

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as functional environmental materials for pollutants degradation is far less explored than its pyrolyzed

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counterpart, hindering its application in environmental remediation.

19

These

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Molecular oxygen activation with using solar energy can generate various reactive oxygen species

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(ROS) to degrade organic pollutants, and thus receives great attention.23, 24 Typically, the photo-induced

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molecular oxygen activation was mediated by semiconductor via photogenerated electron reduction.25

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Iron oxides, with semiconductor properties, are photoactive under solar light irradiation, but its

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photocatalytic activity was poor because of the fast electron-hole charge recombination.26 It was

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reported that polycarboxylates (e.g. oxalate, malonate, and citrate) could form strong complex with

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surface Fe(III) on iron oxides, which may promote their photocatalytic molecular oxygen activation

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performance under solar light via a ligand-to-metal charge transfer (LMCT) process.26-33 Sewage sludge,

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with abundant organic carbon substances (e.g. microorganisms and extracellular polymeric substances)

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and inorganic minerals (e.g. SiO2 and Al2O3), also contained different amounts of iron species, 4 ACS Paragon Plus Environment

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suggesting the existence of iron oxides in the final hydrothermally carbonized sewage sludge product.1

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Therefore, in view of its low cost and waste recycling, the utilization of hydrothermally carbonized

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sewage sludge and polycarboxylates for photo-induced molecular oxygen activation is very promising

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for pollutant control and environmental remediation, but still remain unexplored.

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Herein, we synthesize a sewage sludge-based functional material via a facile one-pot HTC method,

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and construct a photocatalytic molecular activation system with hydrothermally carbonized sewage

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sludge (HTC-S), oxalate and visible light, to degrade various organic contaminates including dyes

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(rhodamine B and methyl blue), pesticides (alachlor and atrazine) and antibiotics (sulfadimidine and

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oxytetracycline). The ROS generation, sulfadimidine degradation pathway and the reusability of HTC-S

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are investigated in detail. The photocatalytic molecular activation performance of HTC-S is compared

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with those of some commercial iron oxides (Fe2O3, FeOOH and Fe3O4) under visible light.

96 97

Experimental Section

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Chemicals and Materials. The chemicals and materials used in this study were described in supporting

99

information (SI Text S1).

100

Preparation of Hydrothermally Carbonized Sewage Sludge. Sewage sludge used in this study was

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transported with polypropylene containers from secondary sedimentation tank of Tangxun Lake

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municipal wastewater treatment plant (Wuhan city, Hubei Province, China) to the laboratory and stored

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at 4 oC before use. The characteristics of raw sewage sludge obtained from different MWWTPs were

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characterized after screened through a 1.0 mm sieve to eliminate suspended residues and impurities (SI

105

Table S1). To synthesize hydrothermally carbonized sewage sludge, 60 g of raw sludge was transferred 5 ACS Paragon Plus Environment


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into an 80 mL Teflon-lined stainless steel autoclave and heated at 180 oC for 5 hours in an electric oven.

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After reaction, the resulting solid was collected by centrifugation and washed with deionized water. The

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product was finally dried at 60 oC for 24 hours in a vacuum oven and called as HTC-S, which was

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placed in the dark for use. The synthesis procedure was provided in supporting information (SI Scheme

110

S1). For comparison, the sludge samples from other local MMWTPs, including Tangxun Lake,

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Longwangzui, Shahu and Sanjintan, were also collected and hydrothermally carbonized under the same

112

conditions.

113

Characterization of Hydrothermally Carbonized Sewage Sludge. The morphology of HTC-S was

114

recorded by a scanning electron microscope (SEM, TESCAN MIRA 3, Czech) and high-angular annular

115

dark field-scanning transmission electron microscopy (HAADF-STEM, JEM-ARM200F, Japan). The

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pore structure of HTC-S was analyzed by nitrogen adsorption (Micrometics ASAP2020) at 77 K. The

117

elements composition of HTC-S was analyzed by elemental analyzer (EA, vario EL cube, Elementar,

118

Germany) and X-ray Fluorescence spectrometer (XRF, Thermo, USA). The elements concentration of

119

HTC-S was determined by inductively coupled plasma spectrometry (ICP, Agilent 720ES, USA). The

120

crystal phases of HTC-S were characterized by a powder X-ray diffractometer (XRD, D/Max-IIIA, Cu

121

Kα radiation, λ = 0.15418 nm). The Mössbauer spectra were measured using a MA-260 Mössbauer

122

spectrometer (Bench MB-500) equipped with a γ-ray source of 0.925 GB, 57Co/Rh at about 25 oC. Fe

123

K-edge extended X-ray absorption fine structure (EXAFS) spectra was collected at the beamline 1W1B

124

of Beijing Synchrotron Radiation Facility, Institute of High Energy Physics, Chinese Academy of

125

Sciences (SI Text S2). The free radicals were recorded on electron paramagnetic resonance (EPR)

126

spectrometer (Bruker E500, Germany) with using 5,5-dimethyl-1-pyrroline -N-oxide (DMPO) as the 6 ACS Paragon Plus Environment

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spin trapper.34

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Pollutants Degradation Procedure. Batch pollutants degradation was performed in a 100 mL

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container under magnetic stirring at 200 rpm. Briefly, 0.01 g of HTC-S and 500 μL of 0.2 mol/L OA

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stock solution were added into 50 mL of sulfadimidine (SM2) stock solution (1 mg L-1), ensuring the

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final concentrations of HTC-S and oxalate were 0.2 g L-1 and 2 mmol L-1, respectively. Then, the initial

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solution pH value was controlled to 5.0 ± 0.2 at 25 °C by adding diluted H2SO4 and NaOH solutions.

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The degradation was triggered by a xenon lamp (PLS-MW2000, Beijing Perfect Light Co., Ltd., China)

134

equipped with a 420 nm cut filter at a distance of 0.2 m from the top. The output energy was 300 W and

135

the irradiation intensity was 1200 mW/cm2, as measured by an optical power meter (PLS-MW2000,

136

Beijing Perfect Light Co., Ltd., China). The temperature of solution was controlled at 25 ± 0.2 °C (SI

137

Scheme S2). Control experiments without HTC-S and oxalate under visible light or in dark were also

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conducted. The degradation solutions were taken out and filtered through 0.22 μm nylon syringe filter at

139

regular intervals, and then 100 μL of ethanol was added quickly into 900 μL of the degradation solution

140

to quench the reaction for the subsequent analysis.

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Analytical Methods. The concentrations of methyl blue and rhodamine B were monitored by a

142

UV-vis spectrometer (UV-2550, Shimadzu, Japan).15 The concentrations of atrazine, alachlor,

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oxycycline and sulfadimidine were monitored by a high performance liquid chromatographer (HPLC,

144

LC-20AT, Shimadzu) with a SB-C18 reverse phase column. The detailed analysis procedures were

145

provided in supporting information (SI Text S3). The degradation intermediates of SM2 were detected

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by liquid chromatography-mass spectrometry with tandem mass spectrometry (LC-MS/MS, TSQ

147

Quantum MAX, Thermo, U.S.A.). The pre-treatment and analytic methods were described in supporting 7 ACS Paragon Plus Environment


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information (SI Text S4).

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The dissolved oxygen concentration was determined by a Dissolved Oxygen Meter (PreSence, Fibox

150

4, Germany) in a closed reaction system. A modified p-hydroxyphenylacetic acid (POHPAA)

151

fluorescence method was used to quantify the concentration of hydrogen peroxide (H2O2),35 while

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benzoic acid (BA) was selected to quantify the accumulative hydroxyl radical (•OH) generation .36, 37

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The concentration of dissolved iron ions were detected by a modified 1, 10-phenanthroline method.38-40

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All of these detailed analysis procedures were provided in supporting information (SI Text S5-S7). The

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total organic carbon (TOC) content variations were determined with using a Shimadzu TOC-V CPH

156

analyzer, and the oxalate concentration variations were measured by an ion chromatograph (IC, Dionex

157

ICS-900, Thermo).

158 159

Results and Discussions

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Characterization of Hydrothermally Carbonized Sewage Sludge. The morphology of HTC-S was

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first characterized by SEM (Figure 1a and SI Figure S1). It was found that the resulting HTC-S was of

162

irregular nanoparticles with porous structure, which was more likely produced by the carbonization of

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biomacromolecules and extracellular polymeric substance of sewage sludge.15 The porous structure of

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HTC-S was analyzed by N2 adsorption method (Figure 1b and SI Table S2). Its pore size and surface

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area were 15 nm and 77 m2 g-1, respectively, higher than those of typically reported pyrolyzed sewage

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sludge materials and biomass-derived hydrochars.10, 14, 15, 17 We found that HTC-S contained 74.16% of

167

ash and 8.54% of C (SI Table S3), much higher and lower than those of previous reported

168

biomass-derived hydrochars, respectively.17, 41, 42 XRF analysis revealed that the ash in HTC-S mainly 8 ACS Paragon Plus Environment

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contained 27.1% of Si, 12.3% of Al, 4.9% of Fe, 2.9% of Ca, 1.9% of K and 1.2% of Mg (Figure 1c).

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Their concentrations were further determined by ICP analysis (Figure 1d). In comparison with those

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reported pyrolyzed sewage sludge materials, HTC-S possessed more Si and less Fe.10, 14, 15 Meanwhile,

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the main five elements of C, O, Al, Si and Fe were uniformly distributed in HTC-S, as revealed by

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HAADF-STEM and element mapping (Figure 1e-j).

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Sulfadimidine Degradation and Reactive Oxygen Species Generation. Regarding that HTC-S

175

contained considerable amount of iron species, we therefore constructed a visible light molecular

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oxygen activation system with using HTC-S and oxalate (HTC-S/oxalate). SM2 was chosen as a model

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contaminant to check the organic pollutant degradation performance of this HTC-S/oxalate system at

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the initial pH value of 5.0. As shown in Figure 2a, neither oxalate nor HTC-S could efficiently remove

179

SM2 in the dark or under visible light, and the degradation of SM2 with HTC-S in the presence of

180

oxalate was also negligible in dark. Moreover, control experiments revealed that SM2 could not be

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self-degraded either in the dark or under visible light. Interestingly, the combination of HTC-S and

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oxalate as well as visible light could induce 94.7 ( ± 2.3)% degradation of SM2 within 120 min,

183

suggesting that HTC-S/oxalate system was highly efficient to degrade SM2 under visible light. More

184

importantly, we noticed that the SM2 degradation process contained two first-order reaction processes,

185

including the first induction period and the subsequent auto-acceleration period. The apparent rate

186

constant (3.10 (±0.61) × 10-2 min-1) in auto-acceleration period was almost 4 times that (7.71 (±1.46)

187

× 10-3 min-1) in induction period (Figure 2b). 30 min was chosen as the inflecting point according to the

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fitting degree of first-order kinetics in different periods (SI Figure S2). For comparison, the anaerobic

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SM2 degradation in the HTC-S/oxalate system under visible light was also performed by bubbling 9 ACS Paragon Plus Environment


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high-purity Ar (1.5 L/min). The removal of molecular oxygen inhibited the SM2 degradation

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completely, confirming the participation of molecular oxygen in the SM2 degradation in the

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HTC-S/oxalate system under visible light (SI Figure S3). To check whether the HTC-S/oxalate system

193

could degrade different organic pollutants under visible light, the degradation experiments of dyes

194

(methyl blue and rhodamine B), pesticides (alachlor and atrazine) and antibiotics (oxycycline) were

195

conducted under the same conditions (SI Figure S4). As expected, all these selected pollutants could be

196

degraded significantly within 120 min (SI Text S9). These results indicated that HTC-S/oxalate system

197

could activate molecular oxygen to generate ROS for various organic pollutants degradation under

198

visible light.

199

To figure out the contribution of different ROS on SM2 degradation, we conducted a series trapping

200

experiments by adding different kinds of excess scavengers (ethanol for both •OH and Fe(IV), IPA for

201

•OH, CAT for H2O2, and SOD for •O2-).43 As shown in Figure 3a, the addition of ethanol and IPA

202

inhibited the SM2 degradation completely, confirming that •OH mainly accounted for the SM2

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degradation in HTC-S/oxalate system under visible light. The addition of CAT suppressed 81.3 ( ±

204

1.64)% of SM2 degradation, indicating that H2O2 was the intermediate for the •OH generation. The

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SM2 degradation inhibition percentage was 30.7 (±3.39)% in the presence of SOD, suggesting that a

206

successive single-electron transfer route (O2→•O2-→H2O2) may contribute to the H2O2 generation.34

207

From the above results, we concluded that HTC-S/oxalate system could activate molecular oxygen to

208

generate H2O2, and also decompose the in-suit generated H2O2 to produce •OH for various pollutants

209

degradation.

210

The kinetics processes of O2 activation and H2O2 decomposition were then investigated. For the O2 10 ACS Paragon Plus Environment

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activation process, we first monitored the dissolved oxygen concentration variation, and found that its

212

concentration decreased from 8.04 ( ±0.01) mg L-1 to 6.05 ( ±0.23) mg L-1 within 120 min of visible

213

light irradiation in HTC-S/oxalate system (Figure 3b). Interestingly, we noticed that the dissolved

214

oxygen consumption process contained three first-order reaction processes (SI Figure S7b), including an

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induction period (0-30 min), a subsequent acceleration period (30-60 min) and a finally decay period

216

(60-120 min), with the rate constants of 1.71 (±0.59) × 10-3 min-1, 4.57 (±0.61) × 10-3 min-1 and 1.49

217

(±0.27) × 10-3 min-1, respectively. As the O2 activation was responsible for the H2O2 generation, we

218

determined the accumulative H2O2 generation in the HTC-S/oxalate system under visible light and

219

calculated its generation rate (SI Text S10). It was found that the accumulative H2O2 generation in the

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HTC-S/oxalate system could reach 44.6 (±1.3) μmol L−1 within 120 min under visible light, and the

221

variation of H2O2 generation rate in different periods was consistent with that of O2 activation rate (SI

222

Figure S8). For the H2O2 decomposition process, we monitored the real-time H2O2 concentration

223

variations, and found that the real-time H2O2 concentration increased up to 14.6 (±0.6) μmol L-1 in the

224

first 60 min, and then decreased to 8.8 (±0.8) μmol L-1, suggesting the in-situ generated H2O2 could be

225

quickly decomposed to produce •OH (Figure 3c). The difference between the concentrations of

226

accumulative and real-time H2O2 represented for the decomposition of H2O2. Therefore, the H2O2

227

decomposition rate was calculated to be 5.25 (±2.80) × 10-3 μmol min-1 within the initial 30 min of the

228

reaction, and increased to 2.36 (±0.16) × 10-2 μmol min-1 in 30-60 min, then slightly decreased to 1.93

229

(±0.05) × 10-2 μmol min-1 in 60-120 min (SI Text S10 and Figure S9a). As the H2O2 decomposition was

230

responsible for the •OH formation, we also measured the accumulative •OH generation in the

231

HTC-S/oxalate system under visible light and calculated its generation rates (SI Text S10). It was found 11 ACS Paragon Plus Environment


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that the generation of accumulative •OH in the HTC-S/oxalate system increased to 38.3 ( ±2.6) μmol

233

L−1 within 120 min under visible light (Figure 3d), and its generation rate variation in different periods

234

was in good agreement with that of H2O2 decomposition rate (SI Figure S9b). According to these

235

analyses, we concluded that both O2 activation and H2O2 decomposition rates were sluggish in the

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induction period, which slowed down the•OH generation rate, and thus limited the SM2 degradation in

237

this period. However, the H2O2 decomposition rate and •OH generation rate in 60-120 min did not

238

significantly decrease even by lowering O2 activation rate in decay period, which guaranteed the

239

constantly efficient SM2 degradation. Therefore, the whole SM2 degradation process only contained an

240

induction period and a subsequent auto-acceleration period.

241

Subsequently, we detected the SM2 degradation intermediates with LC-MS/MS to propose the

242

degradation pathway of (1) sulfadimidine (m/z = 279). Four degradation intermediates, including (2)

243

4-aminophenol

244

2-aminopyrimidine-4,5,6-triol (m/z = 144), and (5) 2-amino-6- methylpyrimidine-4-carboxylic acid

245

(m/z = 153) were found (SI Figure S10). The mass charge ratios (m/z) of all the detected intermediates

246

fitted well with their reported data (SI Table S6).44, 45 The detailed confirmation of these intermediates

247

was described in supporting information (SI Text S11). On the basis of these detected intermediates, a

248

possible SM2 degradation pathway in the HTC-S/oxalate system under visible light was proposed (SI

249

Scheme S3).

(m/z

=

109),

(3)

4,6-dimethylpyrimidin-2-amine

(m/z

=

124),

(4)

250

The Role of Hydrothermally Carbonized Sewage Sludge on Sulfadimidine Degradation. It was

251

reported that dissolved organic substances released from black carbon could generate various reactive

252

oxygen species (e. g. 1O2, O2−) under sunlight irradiation,46 which might be responsible for organic 12 ACS Paragon Plus Environment

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pollutants degradation. Therefore, the roles of dissolved and undissolved HTC-S on SM2 degradation in

254

HTC-S/oxalate system were investigated. We first checked whether dissolved organic carbon could be

255

released from HTC-S by monitoring the total organic carbon (TOC) content variations in different

256

systems (SI Text S12). It was found that dissolved organic carbon leaching contents of HTC-S in

257

different systems obviously increased in the first 15 min, suggesting that dissolved organic carbon could

258

be quickly released from HTC-S into solution (SI Figure S12). Within 120 min, the concentration of

259

dissolved organic carbon leached from HTC-S reached approximately 2.40 (±0.11) mg L-1 in dark in the

260

absence of oxalate, slightly lower than those (3.05 (±0.13) mg L-1) under visible light irradiation and

261

(3.04 (±0.17) mg L-1) in the presence of oxalate and visible light (SI Figure S12). This result suggested

262

that neither visible light nor oxalate could significantly influence the dissolved organic carbon content

263

leached from HTC-S. Then, we checked the contribution of dissolved HTC-S to the SM2 degradation in

264

the HTC-S/oxalate system under visible light. The detailed experimental procedures were illustrated in

265

supporting information (SI Text S13 and Scheme S4). As shown in Figure S13, the SM2 degradation

266

with dissolved HTC-S in the absence or presence of oxalate was negligible, ruling out the contribution

267

of dissolved HTC-S to the SM2 degradation in HTC-S/oxalate system under visible light.

268

The Roles of Iron Species and Oxalate in the HTC-S/oxalate System. To check the roles of iron

269

species of HTC-S in organic pollutants degradation and ROS generation in the HTC-S/oxalate system

270

under visible light, we first treated HTC-S with 1,10-phenanthroline, a strong complexing agent of iron

271

ions,

272

HTC-S-1,10-phenanthroline (SI Text S14 and Scheme S5). As expected, the iron concentration in

273

HTC-S decreased from 8.08 to 2.13 g/kg after this 1,10-phenanthroline chelation treatment (SI Figure

to

chelate

iron

species

in

HTC-S.

The

resulting

solid

sample

was

called

as



Page 14 of 36

274

S14). We then checked the molecular oxygen activation performance of HTC-S-1,10-phenanthroline in

275

the presence of oxalate under visible light. It was found that both the SM2 degradation (SI Figure S15)

276

and the H2O2 generation (SI Figure S16) were significantly suppressed after the removal of iron species

277

with 1,10-phenanthroline chelation, indicating that iron species in HTC-S was responsible for the H2O2

278

generation.

279

HTC-S-1,10-phenanthroline/oxalate system under visible light even after adding 45 μmol L−1 H2O2,

280

which was the same concentration of accumulative H2O2 generated in the HTC-S/oxalate system within

281

120 min under visible light (SI Figure S15). This result indicated that H2O2 could not be efficiently

282

decomposed to produce •OH for the SM2 degradation after the removal of iron species with

283

1,10-phenanthroline chelation. Obviously, iron species in HTC-S played an indispensable role in both

284

the generation and the decomposition of H2O2 in the HTC-S/oxalate system under visible light.

Interestingly,

only

50.0

( ± 6.4)%

of

SM2

could

be

degraded

in

285

We therefore investigated how the oxalate concentration affected the SM2 degradation in the

286

HTC-S/oxalate system under visible light. The SM2 degradation percentage first gradually increased

287

from 28.6 (±1.7)% to 94.7 (±2.3)% with increasing the initial oxalate concentration from 0.4 to 2.0

288

mmol L-1, and then slightly decreased when the oxalate concentration was further increased to 2.4 mmol

289

L-1 (SI Figure S17). This decrease was possibly attributed to the consumption of adsorption site and/or

290

•OH by excess oxalate. To check this possibility, we thus monitored the oxalate concentration variation

291

in the HTC-S/oxalate system under visible light, and found that 43.5% of oxalate was removed

292

simultaneously during the SM2 degradation (SI Figure S18). This simultaneous removal of oxalate

293

could prevent its complexation with various metal ions from causing adverse environmental

294

consequence. For comparison, we selected three typical polycarboxylates that widely existed in fruits 14 ACS Paragon Plus Environment

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295

and vegetables, including malonate, tartrate and citrate, and conducted the SM2 degradation

296

experiments under the same conditions as those in HTC-S/oxalate system. In different HTC-S/

297

polycarboxylates systems, we found that oxalate showed the best photo-reactivity, while tartrate and

298

citrate exhibited lower ability than oxalate and citrate in SM2 degradation (SI Figure S19).

299

pH Effect on the Sulfadimidine Degradation in the HTC-S/oxalate System. As the distribution of

300

oxalate species (pKa1 = 1.22, pKa2 = 4.19, SI Figure S20) and the surface charge of HTC-S (pHPZC = 3.2,

301

SI Figure S21) are strongly dependent on the solution pH, the effect of pH on the SM2 degradation in

302

HTC-S/oxalate system under visible light was thus investigated. We first monitored the temporal pH

303

values during the SM2 degradation at initial pH value of 5.0, and found that the pH value increased

304

from 5.00 (±0.01) to 6.57 (±0.15) within 120 min (SI Figure S22), indicative of the H+ consumption

305

and the OH− generation during the SM2 degradation in the HTC-S/oxalate system under visible light.

306

Then, we compared the SM2 degradation in the HTC-S/oxalate system under visible light at different

307

initial pH values (SI Figure S23), and found that SM2 could be degraded faster under the acidic

308

conditions (pH = 3.0 and 4.0), while the SM2 degradation percentage significantly decreased from 94.7

309

(±2.3)% to 8.9 (±0.7)% along with increasing the initial pH value from 5.0 to 8.0. Therefore, the pH

310

value strongly influenced the organic pollutant removal performance of HTC-S/oxalate system under

311

visible light.

312

The stability and reusability of catalyst are also important factors for practical application. Although

313

the pH value of 3.0-4.0 was beneficial to the efficient SM2 degradation, such lower pH unavoidably

314

induce more iron ions dissolution from HTC-S during the degradation, which was disadvantage for the

315

catalyst reusability and might cause secondary metal ion pollution. When we increased the initial pH 15 ACS Paragon Plus Environment


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316

value to 5.0, we did not detect any dissolved Fe(III) and Fe(II) in the dark or under visible light in the

317

absence of oxalate, suggesting the stability of HTC-S (SI Figure S24). In the presence of oxalate, only a

318

very small amount (3.6 ( ±0.2) μmol L-1 for Fe(II) and 7.3 ( ±0.3) μmol L-1 for Fe(III)) of iron ion

319

leached out from the catalyst (SI Figure S24), much lower than that in reported Fe-oxalate complex

320

systems (SI Table S7). Meanwhile, we also determined the leaching contents of different toxic heavy

321

metal ions during SM2 degradation in HTC-S/oxalate system under visible light (SI Figure S26). Except

322

for Fe and Mn, the leaching contents of the other toxic heavy metals were all below the EU directives of

323

drinking water quality, implying the safety of HTC-S application for wastewater treatment. (SI Table

324

S8). We then checked the reusability of HTC-S during organic pollutants degradation in HTC-S/oxalate

325

system under visible light at initial pH value of 5.0. Although the SM2 degradation efficiency decreased

326

slightly along with repeated use, over 70% of SM2 could be still removed even after five times of use

327

(SI Figure S27a). In comparison with fresh HTC-S, the morphology and chemical structure of HTC-S

328

did not significantly change after reuse (SI Figure S27b). The reusability of HTC-S could be further

329

enhanced by adjusting operating pH value and oxalate concentration to reduce the iron leaching content,

330

which was described in supporting information (SI Text S15, Figure S28-S30). These results suggested

331

that HTC-S could be reused for organic pollutants degradation with adding oxalate under visible light.

332

Comparison between Hydrothermally Carbonized Sewage Sludge and Iron Oxides. We

333

subsequently compared the visible light oxygen molecular activation performance of HTC-S and three

334

common iron oxides (Fe2O3, FeOOH and Fe3O4) in the presence of oxalate, and interestingly found that

335

the SM2 degradation percentages (19.0 (±1.6)% for Fe2O3, 25.1 (±0.6)% for FeOOH and 37.3 (±

336

2.1)% for Fe3O4) in iron oxides/oxalate systems were far lower than that (94.7 ( ± 2.3)%) in 16 ACS Paragon Plus Environment

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337

HTC-S/oxalate system under visible light (Figure 4a). Similarly, all the degradation processes in the

338

iron oxides/oxalate systems were of induction and auto-acceleration periods. The apparent SM2

339

degradation rate constant (3.10 ( ± 0.61)×10-2 min-1) of auto-acceleration period in HTC-S/oxalate

340

system was about 15 times than that (2.06 (±0.28)×10-3 min-1) in Fe2O3/oxalate system, 10 times than

341

that (2.91 ( ± 0.03)×10-3 min-1) in FeOOH/oxalate system, and 6 times than that (4.89 ( ± 0.55)×10-3

342

min-1) in Fe3O4/oxalate system, respectively (Figure 4b). To check the effects of surface area on the

343

reactivity of HTC-S and iron oxides, we normalized their rate constants with specific surface areas (SI

344

Table S10 and Table S11), and thus ruled out the main contribution of surface area to the high activity

345

of HTC-S.

346

As the SM2 degradation efficiency was highly dependent on the ROS generation, we subsequently

347

monitored H2O2 and •OH generation in different systems under visible light. As expected, the

348

accumulative H2O2 generation in iron oxides/oxalate system (5.9 (±0.7) μmol L-1 for Fe2O3, 7.0 (±0.1)

349

μmol L-1 for FeOOH and 8.5 (±0.3) μmol L-1 for Fe3O4) were much lower than that (44.6 (±1.3) μmol

350

L-1) in HTC-S/oxalate system (Figure 4c). Similar phenomena took place for the accumulative •OH

351

generation (Figure 4d). These results confirmed that HTC-S was more efficient on visible light

352

molecular oxygen activation in the presence of oxalate than common iron oxides, which might be

353

related to the unique chemical structure of iron species in HTC-S, forming special surface iron-oxalate

354

complex to efficiently activate molecular oxygen under visible light.

355

To verify this assumption, we thus carefully identified the iron species in HTC-S. As shown in Figure

356

5a, the crystalline phases of graphite, quartz and aluminum oxide were detected in HTC-S. Although we

357

did not observe any obvious characteristic diffraction peaks of iron oxides, the characteristic diffraction 17 ACS Paragon Plus Environment


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358

peaks of iron-containing clay minerals including chlorite and nontronite were recorded. Subsequently,

359

we employed the

360

57Fe

361

(Figure 5b). The parameters, including isomer shift (IS), electric quadrupole splitting (QS) and full line

362

width at half maximum (LW), were summarized in supporting information (SI Table 12). The doublet

363

with IS = 0.37 mm/s and QS = 0.64 mm/s was attributed to Fe(III), which account for 79.7% of the total

364

spectra area. The doublet with IS = 1.10 mm/s and QS = 2.46 mm/s could be ascribed to Fe(II),

365

accounting for 20.3% of the total spectra area. The superimposition of these two doublets of HTC-S was

366

consistent with the characteristic peaks of reported chlorite.47, 48 These results indicated that different

367

from iron oxides, the iron species in HTC-S were mainly occurred in form of iron-containing clay

368

minerals.

57Fe

Mössbauer spectra to further identify these iron-containing clay minerals. The

Mössbauer spectrum of HTC-S at room temperature contained the superimposition of two doublets

369

Clay minerals, with hydrated aluminum silicates as its main composition, have layered structure

370

formed by linking tetrahedral sheet to an octahedral sheet through sharing apical oxygens. The central

371

atom of tetrahedral sheet is Si(IV), while the octahedral sites are occupied by Al(III) (SI Figure S31a).

372

Generally, the iron species in iron-containing clay minerals replace Al(III) and occupy the octahedral

373

center with the formation of Fe-O-Si bond or Fe-O-Al bond, resulting in its strong interactions with

374

silicon and aluminum species (SI Figure S31b).49 To verify their strong interactions, we collected Fe

375

K-edge extended X-ray absorption fine structure (EXAFS) spectra of HTC-S with using Fe foil, FeO,

376

Fe2O3, α-FeOOH and Fe3O4 as standard samples for reference. It was found that the Fourier

377

transformations of EXAFS signals of all the samples exhibited an obvious peak at ~1.5 Å (no phase

378

correction), which could be ascribed to Fe-O shell, while the peak at ~2.6 Å indicated the backscattering 18 ACS Paragon Plus Environment

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379

from other adjacent atoms, such as Fe, Al and Si, existed beyond the Fe-O shell (Figure 5c). To obtain

380

the exact local coordination environment of Fe species in HTC-S, curve fittings were performed. The

381

fitting results were summarized in Table 1. In the first shells, two Fe-O shells with the bond lengths of

382

1.95 (±0.05) Å (Fe-O bond in Fe-O-Al) and 2.11 (±0.07) Å (Fe-O bond in Fe-O-Si) were obtained for

383

HTC-S, quite different from that of those selected iron oxides, further verifying that the Fe species in

384

HTC-S was not in form of iron oxides (Figure 5d). In the second shells, Fe–Fe shells with different bond

385

lengths were found for different iron oxides samples, while Fe-Al shell and Fe-Si shell with the bond

386

lengths of 2.98 (±0.07) Å and 3.14 (±0.07) Å, were obtained for HTC-S (Figure 5d). These results

387

indicated that the Fe species in HTC-S were in form of iron-containing clay minerals with the formation

388

of Fe-O-Al and Fe-O-Si bonds, resulting in its strong interactions with silicon and aluminum species.

389

To further check the influence of silicon and aluminum species on visible light molecular oxygen

390

activation performance of HTC-S/oxalate system, we increased the contents of Si and Al in HTC-S by

391

pretreating the raw sludge with TEOS and Al2(SO4)3•12H2O (SI Text S16). The resulting samples were

392

denoted as HTC-S-Si and HTC-S-Al, respectively. As expected, the Si and Al contents in the final

393

hydrothermal products significantly increased after this pretreatment procedure (Figure 6a and Figure

394

6b). We then examined the SM2 degradation ability of HTC-S/oxalate system after the pretreatments

395

under visible light. As shown in Figure 6c and Figure 6d, the increase of Si amount significantly

396

promoted the SM2 degradation, but more Al exhibited adverse effect on the SM2 degradation,

397

confirming that the strong interaction between Si and Fe species was contributed to the SM2

398

degradation in the HTC-S/oxalate system under visible light. To check the effect of surface areas on the

399

reactivity of catalysts, we also normalized their rate constants with surface areas and found that the 19 ACS Paragon Plus Environment


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400

enhanced reactivity of HTC-S was not ascribed to higher surface area caused by the addition of more Si

401

(SI Table S10 and Table S11), but other factors.

402

To explore how the strong interactions between Si and Fe influenced the photochemical reactivity of

403

HTC-S, density functional theory (DFT) calculation was employed to analyze the energy separation

404

between the highest occupied molecular orbital (HOMO) and the lowest unoccupied molecular orbital

405

(LUMO) of Fe-oxalate complex in the presence or absence of Si (SI Text S17). The Fe atomic cluster

406

model in the presence of Si was derived from the crystal structure of iron-containing clay mineral, while

407

the Fe atomic cluster model in the absence of Si was created according to crystal parameters of iron

408

oxides, to better simulate the Fe chemical environment in iron-containing mineral or iron oxides (SI

409

Figure S32). The HOMO-LUMO gaps for the two configurations (monodentate and bidentate, SI Figure

410

S33 and Figure S34) of Fe-oxalate complex were calculated (SI Table S13). As for monodentate

411

complex, the presence of Si decreased the gap of Fe-oxalate complex from 4.41 eV to 3.57 eV, lowering

412

the excitation energy of Fe-oxalate complex. Similar trend was also observed in the case of bidentate

413

Fe-oxalate complex. So the DFT calculation results revealed that the presence of Si could lower the

414

excitation energy of Fe-oxalate complex, which accounted for the superior visible light oxygen

415

molecular activation performance HTC-S to those three iron oxides. Meanwhile, the generality of our

416

results was also confirmed by a series of HTC-S samples derived from raw sludge samples obtained

417

from different municipal wastewater treatment plants (SI Text S18, Figure S35-S37 and Table S1-S3).

418

On the basis of the above results, we proposed the oxalate promoted organic pollutants degradation

419

over HTC-S under visible light as follows. First, oxalate was adsorbed on the HTC-S surface to form

420

≡FeIII-oxalate complexes. These complexes were then excited by visible light to generate 20 ACS Paragon Plus Environment

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421

carbon-centered radicals (e.g. •C2O4− and •CO2−), as confirmed by EPR technique (SI Figure S38).

422

Meanwhile, the presence of silicon species in HTC-S notably favored the excitation of ≡FeIII-oxalate

423

complexes under visible light, because the silicon species strongly interacted with iron species in

424

HTC-S to form Fe-O-Si bond, which would then bind with oxalate to lower the excitation energy of

425

≡FeIII-oxalate complex. Subsequently, the generated carbon-centered radicals could transfer electron to

426

molecular oxygen for the •O2- formation. •O2- further reacted with H+ to generate H2O2, which would

427

then be decomposed by ≡FeII-oxalate complexes to form •OH (SI Figure S38) to degrade various

428

organic pollutants.

429 430

Environmental Implications.

431

Converting sewage sludge into environmental remediation materilas could alleviate the burden of

432

sewage sludge disposal, and also realize a “Trash to Treasure” strategy. Compared with pyrolysis,

433

hydrothermal carbonization is a more interesting sewage sludge conversion route of energy saving and

434

environmental benignancy characteristics. In this study, we demonstrated that hydrothermally

435

carbonized sewage sludge could be used as an efficient oxygen molecualr activation catalyst in oxalate

436

solution to degrade various organic pollutants under visible light, revealing its potential in wastewater

437

treatment. More importantly, the reactivity of hydrothermally carbonized sewage sludge was even much

438

higher than common iron oxides under visible light, because the silicon species in hydrothermally

439

carbonized sewage sludge notably promoted the excitation of ≡FeIII-oxalate complexes under visible

440

light. These findings offer an alternative pathway for sewage sludge conversion into envrionmental

441

functional materials, and also provide an attractive sewage sludge-based catalyst for polluted water 21 ACS Paragon Plus Environment


442

Page 22 of 36

treatment with solar energy.

443 444

AUTHOR INFORMATION

445

Corresponding Author

446

*Phone/Fax: +86-27-6786 7535; e-mail: [email protected].

447

Notes

448

The authors declare no competing financial interest.

449 450

ACKNOWLEDGEMENTS

451

This work was supported by Natural Science Funds for Distinguished Young Scholars (Grant

452

21425728), National Science Foundation of China (Grant 51472100), the 111 Project (Grant B17019),

453

Self−Determined Research Funds of CCNU from the Colleges’ Basic Research and Operation of MOE

454

(Grant CCNU14Z01001), and the CAS Interdisciplinary Innovation Team of the Chinese Academy of

455

Sciences.

456 457

ASSOCIATED CONTENT

458

Supporting Information

459

Characteristics of raw sewage sludge; analysis procedures of organic pollutants; pre-treatment method

460

and analysis procedure for LC-MS/MS detection; quantification of H2O2 and •OH; diagram of

461

photo-reaction setup; SEM images of HTC-S; elemental composition of HTC-S; mass spectra and

462

structural formula and m/z values of SM2 and its degradation intermediates; possible sulfadimidine 22 ACS Paragon Plus Environment

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463

photodegradation pathway; BET surface area of different samples; EPR signals of carbon centered

464

radicals and •OH; SM2 degradation in different comparative systems; oxalate removal during SM2

465

degradation; temporal pH values during SM2 degradation; determination of dissolved Fe(II) and Fe(III)

466

concentration; cycle experiments; DFT calculations.

467 468

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Reagent to Degrade Alachlor via Superoxide-Mediated Fe(II)/Fe(III) Cycle. ACS Appl. Mater. Inter.

578

2015, 7, 28534-28544.

579

(44)Yang, B.; Mao, X.; Pi, L.; Wu, Y.; Ding, H.; Zhang, W., Effect of pH on the adsorption and

580

photocatalytic degradation of sulfadimidine in Vis/g-C3N4 progress. Environ. Sic. Pollut. Res. Int. 2017,

581

24, 8658-8670.

582

(45)Zhou, T.; Wu, X.; Zhang, Y.; Li, J.; Lim, T.-T., Synergistic catalytic degradation of antibiotic

583

sulfamethazine in a heterogeneous sonophotolytic goethite/oxalate Fenton-like system. Appl. Catal. B:

584

Environ. 2013, 136-137, 294-301.

585

(46) Fu, H.; Liu, H.; Mao, J.; Chu, W.; Li, Q.; Alvarez, P. J.; Qu, X.; Zhu, D. Photochemistry of

586

Dissolved Black Carbon Released from Biochar: Reactive Oxygen Species Generation and

587

Phototransformation. Environ. Sci. Technol. 2016, 50, 1218-1226

588

(47)Heller-Kallai, L.; Rozenson, I., The Use of Mössbauer Spectroscopy of Iron in Clay Mineralogy. 28 ACS Paragon Plus Environment

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Phys. Chem. Miner. 1981, 7, 223-238.

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(48)Smyth, J. R.; Dyar, M. D.; May, H. M.; Bricker, O. P.; Acker, J. G., Crystal Structure Refinement

591

and Mössbauer Spectroscopy of an Ordered Triclinic Clinochlore. Clay. Clay. Miner. 1997, 45,

592

544-550.

593

(49) Madejová, J., FTIR techniques in clay mineral studies. Vib. Spectrosc. 2003, 31, 1-10.

594 595

Figure Captions

596 597

Figure 1. Characterizations of HTC-S: (a) SEM image; (b) BET nitrogen adsorption-desorption

598

isotherm; (c) ash composition and (d) elements concentrations; (e) HAADF-STEM image of the HTC-S.

599

STEM elements mapping images of the HTC-S: (f) C, (g) O, (h) Al, (i) Si, and (j) Fe.



Page 30 of 36

600 601

Figure 2. (a) Time profiles of SM2 degradation in different systems. (b) Plots of -ln(C/C0) versus time

602

for SM2 degradation in HTC-S/oxalate system under visible light. The initial concentrations of SM2,

603

HTC-S, and oxalate were 1 mg L−1, 0.2 g L-1, and 2 mmol L-1, respectively. The initial pH values of the

604

systems were 5.0. Error bars and uncertainties of rate constants represent ± one standard deviation

605

derived from triplicate experiments. Some standard deviations were smaller than symbols shown.

606


Page 31 of 36


607 608

Figure 3. (a) Time profiles of SM2 degradation in HTC-S/oxalate system under visible light with

609

adding different kinds of scavengers (Ethanol for both •OH and Fe(IV), IPA for •OH, CAT for H2O2

610

and SOD for •O2-); (b) Dissolved oxygen consumption and accumulative H2O2 generation in

611

HTC-S/oxalate system under visible light; (c) Real-time H2O2 concentration detected in different

612

systems; (d) Accumulative •OH generation in different systems. The initial concentrations of SM2,

613

HTC-S, oxalate and BA were 1 mg L−1, 0.2 g L-1, 2 mmol L-1 and 10 mmol L-1, respectively. The initial

614

pH values of the systems were 5.0. Error bars represent ± one standard deviation derived from duplicate

615

experiments. Some standard deviations were smaller than symbols shown.

616 617



Page 32 of 36

618 619

Figure 4. (a) Time profiles of SM2 degradation in different systems; (b) Plots of -ln(C/C0) versus time

620

for SM2 degradation different system under visible light; (c) Real-time and accumulative H2O2

621

concentration in different systems under visible light; (d) Accumulative •OH generation in different

622

systems under visible light; The initial concentrations of SM2, HTC-S or iron oxides and oxalate were 1

623

mg L-1, 0.2 g L-1 and 2 mmol L-1, respectively. The initial pH values of the systems were 5.0. Error bars

624

and uncertainties of rate constants represent ± one standard deviation derived from triplicate

625

experiments. Some standard deviations were smaller than symbols shown.

626


Page 33 of 36


627 628

Figure 5. (a) XRD pattern of HTC-S; (b) 57Fe Mössbauer spectra of HTC-S. (c) Fourier transformations

629

of k2-weighted EXAFS oscillations obtained at the Fe K-edge of the various samples; (d) Fitting of the

630

theoretical curve to the experimental curve of HTC-S.



Page 34 of 36

631 632

Figure 6. (a) Si and Al contents before and after pretreating raw sludge with TEOS and

633

Al2(SO4)3•12H2O; (b) FTIR spectra of different HTC-S samples; (c) Time profiles of SM2 degradation

634

in different system under visible light; (d) Plots of -ln(C/C0) versus time for SM2 degradation in

635

different system under visible light. The initial concentrations of SM2, different HTC-S samples and

636

oxalate were 1 mg L−1, 0.2 g L-1 and 2 mmol L-1, respectively. The initial pH values of the systems were

637

5.0. Error bars and uncertainties of rate constants represent ± one standard deviation derived from

638

duplicate experiments. Some standard deviations were smaller than symbols shown.

639 640 641


Page 35 of 36

642


Table 1. Curve fitting results of Fe K-edge EXAFS for various samples.

643 644

aInteratomic

645

eGoodness-of-fit

646

parameters, representing the errors in the last digit.

distance.

bCoordination

number.

cDebye-Waller

factor.

dThreshold

energy shift.

parameter. The values in parentheses were estimated uncertainties of different

647 648 649



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650 651

TOC Art

652


Visible Light Driven Organic Pollutants Degradation with

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