P-g-C3N4 heterojunction photocatalyst: Excellent photocatalytic performance and photocatalytic mechanism for ephedrine degradation

Journal Pre-proof A novel Z-scheme AgBr/P-g-C3 N4 heterojunction photocatalyst: Excellent photocatalytic performance and photocatalytic mechanism for ephedrine degradation Miao Chen (Investigation) (Writing - original draft), Changsheng Guo (Conceptualization)Writing – review and editing), Song Hou (Methodology) (Formal analysis), Jiapei Lv (Methodology), Yan Zhang (Validation), Heng Zhang (Investigation), Jian Xu (Conceptualization)Writing- review and editing) (Supervision)

PII:

S0926-3373(20)30029-1

DOI:

https://doi.org/10.1016/j.apcatb.2020.118614

Reference:

APCATB 118614

To appear in:

Applied Catalysis B: Environmental

Received Date:

4 October 2019

Revised Date:

23 December 2019

Accepted Date:

7 January 2020

Please cite this article as: Chen M, Guo C, Hou S, Lv J, Zhang Y, Zhang H, Xu J, A novel Z-scheme AgBr/P-g-C3 N4 heterojunction photocatalyst: Excellent photocatalytic performance and photocatalytic mechanism for ephedrine degradation, Applied Catalysis B: Environmental (2020), doi: https://doi.org/10.1016/j.apcatb.2020.118614

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A novel

Z-scheme

AgBr/P-g-C3N4

heterojunction

photocatalyst:

Excellent

photocatalytic performance and photocatalytic mechanism for ephedrine degradation

Miao Chen, Changsheng Guo, Song Hou, Jiapei Lv, Yan Zhang, Heng Zhang, Jian Xu* State Key Laboratory of Environmental Criteria and Risk Assessment, Chinese

Corresponding author, Email: [email protected] (J. Xu)

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Research Academy of Environmental Sciences, Beijing, 100012, China

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Graphical Abstract

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Highlights

A series of Z-scheme AgBr/P-g-C3N4 composites were fabricated for

AB/CN (5:1) showed the best photocatalytic performance on EPH degradation.

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the first time.

EPH degradation by AB/CN was pH-dependent.

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DOM or HCO3- accelerated EPH degradation at low concentrations.

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Holes and ·O2− were the dominant reactive species in the EPH reaction

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system.

Abstract

A series of AgBr/P-g-C3N4 (AB/CN) were fabricated via in-situ growth of AgBr particles on the surface of calcined P-g-C3N4 layered mesoporous materials, and the 2

catalysts were used for the photocatalytic degradation of ephedrine (EHP) under simulated solar irradiation. All photocatalysts demonstrated high photocatalytic performances for EPH degradation, and the AB/CN (5:1) showed the best photocatalytic activity. The holes, superoxide radicals and singlet oxygen were the dominant reactive oxygen species involved in EPH degradation. The enhanced photocatalytic performance of AgBr/P-g-C3N4 was ascribed to the formation of Ag

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nanoparticles and the Z-scheme mechanism. EPH degradation was accelerated by dissolved organic matter or HCO3- at low concentrations, but inhibited at high levels.

The addition of CO32- enhanced EPH degradation, and NO3- suppressed the

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degradation. AB/CN was also able to degrade EPH in surface water and secondary

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effluent, implying the potential of the prepared materials in illicit drugs removal in practical applications.

1 Introduction

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degradation pathways

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Keywords: AgBr/P-g-C3N4; ephedrine; Z-scheme mechanism; intermediates;

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As a prescription drug, ephedrine (EPH) is widely used to treat influenza, asthma or hypotension [1]. It is also known as a precursor substance of methamphetamine

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synthesis. After being consumed, pharmaceuticals such as EPH will end in the wastewater treatment plants (WWTPs). The concentration of EPH was up to 185.7 ng·L-1 in surface water and 8.7-1979.5 ng·L-1 in waste water effluents [2,3]. However, conventional wastewater treatment techniques including biodegradation, physical adsorption and photolysis cannot eliminate these emerging contaminants effectively, 3

resulting in their widespread occurrence in natural water bodies [3-5]. The potential ecological toxicity of these drugs to aquatic organisms has been confirmed [6-8], which necessitates the exploration of new efficient methods to remove these emerging contaminants in water. Semiconductor-based photocatalysts have been employed to decompose recalcitrant organic contaminants, produce energy, and inactivate bacteria because of

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their intrinsic properties [9-11]. Among various semiconductors, graphitic carbon

nitride (g-C3N4) has received great attention recently, due to its nontoxicity,

visible-light-driven property and suitable band energy (2.7 eV). g-C3N4 has been

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widely applied for toxic compounds degradation, H2/O2 evolution and CO2 reduction

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[12, 13]. However, low specific surface area, high recombination rate of photoinduced electrons (e-) and holes (h+) and other shortcomings have restricted the further

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practical application of g-C3N4. Methods including metal and non-mental doping [14],

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heterojunction construction [15, 16] and semiconductors combination [17, 18] have been applied to overcome those drawbacks and subsequently resulted in the higher

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photocatalytic performance. Doping of phosphorous on g-C3N4, for instance, can create a relatively low band gap and large specific areas, which can enhance the

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photocatalytic activity of g-C3N4. Phosphorus doping g-C3N4 (P-g-C3N4) possesses good photocatalytic performance with a band energy of 2.49 eV [19] in H2 production and pollutant degradation under visible light irradiation [20]. However, the photocatalytic performance of P-g-C3N4 is still unsatisfactory in practical application of wastewater treatment. 4

Silver bromide (AgBr) is a high-performance visible-light-driven photocatalyst with narrow band gap, unique light sensitivity and high separation rate of charge carriers [21-24]. However, pure AgBr photocatalyst is unstable under visible light illumination [25]. Coupling AgBr with other semiconductors is a feasible method to overcome the photochemical corrosion of AgBr and could induce a high photocatalytic performance by establishing the Z-scheme system. For instance, AgBr

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coupled with hollow porous carbon nitride can degrade 97% Orange G dye within 10 min [26]. Ag/AgBr/g-C3N4 showed higher degradation efficiency of methyl orange

than Ag/AgBr or g-C3N4 under visible light irradiation [27]. High photocatalytic

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activity of the composite by combination of AgBr and P-g-C3N4 is therefore expected.

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In this study, a series of Z-scheme AgBr/P-g-C3N4 (AB/CN) composites were constructed by thermal polymerization coupling with in-situ precipitation-deposition

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method. The as-prepared samples were used to degrade EPH under simulated solar

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irradiation, and the influence of experimental parameters including initial pH, dissolved organic matter (DOM) and some inorganic salt ions on the degradation were

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evaluated. The degradation intermediates/products were identified, and the possible degradation pathways were also proposed. To the best of our knowledge, it is the first

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time to fabricate the Z-scheme composite of AgBr/P-g-C3N4 and apply it to remove EPH in water.

2 Experimental section 2.1 Materials EPH was obtained from Cerilliant Corporation (Round Rock, TX, USA). Silver 5

nitrate (AgNO3), sodium bromide (NaBr), dibasic sodium phosphate (Na2HPO4), 1,4-benzoquinone

(BQ),

isopropanol

(IPA),

5,5-dimethyl-1-pyrrolineN-oxide

(DMPO), ethylenediaminetetraacetic acid disodium (EDTA-2Na), sodium azide (NaN3), humic acid (HA), sodium hydroxide (NaOH), concentrated sulfuric acid (H2SO4, 98%), sodium carbonate (Na2CO3), sodium bicarbonate (NaHCO3) and sodium nitrate (NaNO3) were of analytical grade and purchased from Sinopharm

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Chemical Reagent Co., Ltd (Beijing, China). Acetonitrile, methanol and formic acid

(HPLC grade) were purchased from Fisher Scientific (Poole, UK). Ultrapure water

was acquired by a Milli Q system (Millipore, MA). All reagents and chemicals were

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used directly without further purification.

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2.2 Preparation of AgBr/P-g-C3N4

P-g-C3N4 was fabricated using a facile thermal polymerization method [14, 28].

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AgBr/P-g-C3N4 composites were prepared with the in-situ deposition-precipitation

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method. In a typical procedure, a given amount of P-g-C3N4 material and 30 mL ultrapure water were added in a beaker, followed with ultrasonic dispersion for one

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hour. Adding AgNO3 into the beaker and the suspension was stirred in darkness for 60 min. NaBr solution (50 mL) was dropwise added to the suspension and further

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magnetically stirred for 3 hours in darkness. Finally, the precipitate in the suspension was collected by centrifugation, and washed 3 times by ultrapure water and ethanol, respectively. The obtained samples were then dried at 50 oC under vacuum overnight, and denoted as AB/CN (X:Y), where X:Y represented the mass ratios of AgBr to P-g-C3N4. The detail information of dosage used in sample preparation was shown in 6

Supplementary Materials (Table S1). 2.3 Characterization High resolution transmission electron microscopy (HRTEM, JEM-2100F), transmission electron microscopy (TEM, JEM-100 CXII) and energy-dispersive X-ray spectroscopy (EDX, Oxford Aztec X-MaxN 80) affiliated to scanning electron microscopy (SEM, Hitachi s-4800) were used to investigate the sample’s particle size,

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microstructure and morphology. An X-ray diffractometer (XRD, Rigaku D/Max-2500) using a Cu Kα radiation (λ = 0.15406 nm) was used to study the XRD patterns of

samples. UV–vis diffuse reflectance spectra (UV–vis DRS) were obtained by an

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UV-vis spectrophotometer (Hitachi U3010) with a reference of BaSO4. Fourier

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transform infrared spectroscopy (FT-IR, Nicolet 5SX-FTIR) was used to measure the FT-IR spectra. Photoluminescence (PL) spectra of samples were tested by the

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photoluminescence detector (Hitachi F-4500). X-ray photoelectron spectroscopy

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(XPS) measurements were conducted on XPS spectrometer (PHI Quantera SX). The Brunauer-Emmett−Teller (BET) specific area and pore size distribution were

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measured by an automatic analyzer (Tristar II 3020M). 2.4 Photocatalytic degradation experiments

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In a typical procedure, 0.02 g synthesized catalyst and 50 mL EPH (0.1 mg·L-1)

were added in the quartz tube, which was placed in a photochemical reactor (XPA-7, Xujiang Machinery Factory, Nanjing China) at ambient temperature. A 500 W xenon lamp (420-1100 nm, Institute of Electric Light Source, Beijing) with the average light intensity of 100 mW·cm2 was used to simulate visible light. The suspension in quartz 7

tube was magnetically stirred in darkness for 30 min to guarantee the equilibrium of adsorption-desorption before irradiation. At the given illumination time, 0.5 mL solution was withdrawn and filtered through 0.22 μm nylon membrane for the purpose of analysis. The stability and reusability of the catalyst was tested by the repeated use of AB/CN (5:1) for three runs under the same conditions. The methods and instrument details for the determination of EPH concentration were shown in Text

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S1. Method for EPH degradation intermediates identification was shown in Text S2. 3 Results and discussion 3.1 Characterization

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3.1.1 XRD analysis

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XRD patterns of the synthesized samples were presented in Figure 1a. For P-g-C3N4, two characteristic peaks at 13.1o and 27.8o were observed, which

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corresponded to the (100) and (002) diffraction planes of g-C3N4 (JCPDS 87-1526),

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respectively [29]. AgBr materials showed the structure of face-centered cubic, and the diffraction peaks at 26.9o, 31.2o, 44.6o, 52.7o, 55.3o, 64.7o and 73.4o were attributed to

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(111), (200), (220), (311), (222), (400) and (420) crystal planes of AgBr (JCPDS 06-0438), respectively [30]. For all AB/CN composites, the characteristic peaks

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intensity of g-C3N4 decreased with the increasing of AgBr contents, while the intensity of AgBr peaks increased. However, there was no signal of metallic Ag peaks, and the XRD results confirmed the successful synthesis of AgBr/P-g-C3N4 composites. 3.1.2 Morphology analysis 8

The morphology and microstructure of the synthesized samples was determined by SEM, TEM and HRTEM. As shown in Figure S1a, P-g-C3N4 displayed the mesoporous structure, which could enhance its photocatalytic property by providing more reactive sites. AgBr presented the smooth spherical structure (Figure S1b). For the AB/CN composites (Figure S1b-c), a majority of AgBr particles was deposited on the surface of P-g-C3N4. AgBr particles decorated on the P-g-C3N4 surface possessed

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sphere like structure with the average particle size of 100 ~ 200 nm. The EDX element mapping of AB/CN (5:1) was shown in Figure S2, verifying that AgBr was uniformly deposited on P-g-C3N4 surface.

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The TEM images of P-g-C3N4, AgBr and AB/CN composites were shown in

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Figure 2. AgBr and P-g-C3N4 were linked tightly (Figure 2c-2e), indicating the heterojunction structure formed in AB/CN composite [31]. The heterojunction

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structure could promote charge transformation in AB/CN composite and enhance its

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photocatalytic activities [31]. Figure 2c illustrated the Ag nanoparticles with a diameter of about 10 ~ 20 nm in AB/CN composites, which was ascribed to the partial

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reduction of Ag+ in the AB/CN synthesis processes. Ag nanoparticles could act as ecapture center, which could restrain the recombination of e- and h+. HRTEM image of

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AB/CN (1:1) (Figure 3) indicated that the lattice fringes at 0.24 nm and 0.29 nm were ascribed to the (111) plane of Ag and (200) plane of AgBr, respectively. The interplanar distance of 0.32 nm was corresponding to (002) crystal plane of g-C3N4. The results also confirmed that AgBr nanoparticles were successfully dispersed on the surface of P-g-C3N4. 9

3.1.3 FT-IR analysis FT-IR spectra of the synthesized samples were given in Figure 1b. For the spectrum of P-g-C3N4, several strong bands in the 1639 ~ 1220 cm-1 regions were corresponding to the typical stretching vibration of C=N and C-N heterocycles [32]. The sharp band at 806 cm-1 was attributed to the characteristic breathing mode of P-g-C3N4 s-triazine unit [33]. The broad absorption peaks at 3100 ~ 3500 cm-1

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belonged to the bending vibrations of N-H and O-H bands [34]. AB/CN composites

showed the similar characteristic peaks with P-g-C3N4 materials, which was due to the

weak absorption peaks of AgBr [35]. After combination with AgBr, P-g-C3N4

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remained its characteristic peaks, indicating that there was no covalent bond among

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them, and the combination cannot change the structure of P-g-C3N4 [36]. 3.1.4 UV-vis DRS and band gap analysis

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The UV-vis DRS spectra of the synthesized samples were present in Figure 1c.

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The absorption edges of P-g-C3N4 and AgBr were about 468 and 516 nm, respectively. The absorption edges of AB/CN composites with different AgBr mass ratios were in

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the range of 468 ~ 516 nm. As the contents of AgBr increased, the absorption wavelength of AB/CN composites increased gradually. It indicated that the absorption

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range of AB/CN composite increased with the loading of AgBr on P-g-C3N4 materials, which was ascribed to the formation of Ag nanoparticles. The band gap energy of the photocatalyst can be calculated by the Equation (1) [37]: αhv = A(hv-Eg)n/2 10

(1)

where α, h, v, A and Eg are the absorption coefficient, Planck constant, light frequency, a constant and band gap energy, respectively. For P-g-C3N4, AgBr and AB/CN, the value of n is 4 [27]. According to the plots of (αhv)1/2 versus hv (Figure 1d), the calculated Eg of P-g-C3N4 and AgBr were 2.49 eV and 2.22 eV, respectively. The valance band (VB) and conduction band (CB) of a semiconductor can be determined

EVB = χ – Ee + 0.5Eg (2) ECB = EVB - Eg

(3)

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by the Equation (2) ~ (3) [38]:

where χ is the absolute electronegativity of a semiconductor, and the value is 5.81 for

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AgBr material [27]. Ee is the energy of free electrons on the hydrogen scale (about 4.5

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eV). The CB and VB of AgBr were estimated to be 0.2 eV and 2.42 eV, and that of

3.1.5 XPS analysis

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P-g-C3N4 were -1.02 eV and 1.47 eV, respectively.

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The XPS survey spectrum of AB/CN composite was shown in Figure 4a. The elements of C, N, O, P, Ag and Br were observed in the spectrum, with the

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high-resolution spectra shown in Figure 4b-4g. C1s spectrum had two peaks at 284.8 eV and 287.7 eV, which could be ascribed to C-C bonds and sp2-hybridized carbon

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atoms (N-C=N) in the aromatic ring, respectively [39]. As shown in Figure 4c, the N 1s spectrum can be deconvolved into three peaks. The peaks at 397.5 eV and 398.4 eV were attributed to tertiary nitrogen (N-(C)3) groups and sp2-hybridized N bound to carbon (C=N-C), respectively [40]. The peak at 400.0 eV was assigned to the structure of amino groups (N-H) [41]. The O 1s peaks in Figure 4d were attributed to 11

the absorption of O element on the surface of composites. In the P 2p spectrum, the peak at 132.8 eV was the structure of P-N coordination, suggesting the C-N bonds may be partially replaced by P-N bonds in the triazine rings [20]. In Figure 4f, the Ag 3d spectrum can be deconvoluted into four peaks. The binding energies at 367.0 eV and 373.0 eV were contributed to Ag0 in the AB/CN composite [42]. The peaks at 367.6 eV and 373.6 eV could be attributed to Ag+ in the AgBr nanoparticles,

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corresponding to the binding energies of Ag3d5/2 and Ag3d3/2, respectively [43]. Two major peaks at 68.2 eV and 69.2 eV (Figure 4g) were assigned to Br 3d5/2 and Br 3d3/2 peaks, respectively [44]. The XPS results indicated that Ag nanoparticles were formed

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by photo-reduction of Ag+ during the synthesis of AB/CN composites.

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3.1.6 PL analysis

The photo-generated carriers separation rate of the synthesized samples was

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evaluated by photoluminescence (PL) spectroscopy, which is an indirect standard to

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evaluate the separation rate of electron-hole pairs. The PL spectra of P-g-C3N4, AgBr and AB/CN composites were shown in Figure S3. The excitation wavelength was 315

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nm, and all samples except AB/CN (5:1) showed emission peaks at about 470 nm. It indicated that combination of AgBr with P-g-C3N4 enhanced the separation rate of

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charge carriers and restrained the recombination of photo-generated electron and holes.

3.1.7 N2 adsorption-desorption isotherms The N2 adsorption-desorption isotherms and the corresponding pore size distribution curves of P-g-C3N4 and AB/CN composites were presented in Figure S4. 12

All photocatalysts showed type Ⅳ hysteresis loops, corresponding to the mesoporous structure [45]. The BET specific surface areas (SBET) of AgBr, P-g-C3N4 and AB/CN (5:1) were 5.22, 31.23 and 13.55 m2·g-1. After combination with AgBr, the specific surface area of P-g-C3N4 decreased significantly, which was due to the growth of AgBr particles on the inner and/or surface of P-g-C3N4. The low SBET of AgBr may not contribute to its improvement of photocatalytic degradation efficiency in EPH

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[37]. 3.2 Degradation of EPH

Degradation of EPH by as-prepared samples under visible light irradiation were

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shown in Figure 5a. The adsorption percentage of EPH by the synthesized samples

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before irradiation was below 10%, illustrating that adsorption of EPH by samples was negligible. EPH was hardly diminished under direct visible light irradiation,

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indicating that photolysis of EPH could be ignored. Photocatalytic degradation of

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EPH by as-prepared samples was in line with the pseudo-first-order kinetics model: -ln(Ct/C0) = kt

(4)

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where k and t are rate constant and time, respectively, C0 and Ct are the initial concentration and concentration of EPH at time t. The pseudo-first-order rate

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constants (k) for EPH degradation by different samples were shown in Figure 5b. P-g-C3N4 and AgBr possessed the relatively low degradation performance with the rate constants of 0.0185 min-1 and 0.0296 min-1 under 90 min irradiation, respectively. The degradation rate of EPH gradually increased with the increasing content of AgBr, but AB/CN (10:1) showed a decreased degradation efficiency compared with AB/CN 13

(5:1). It was likely because the Z-scheme structure was destroyed when excessive amount of AgBr doping on the surface of P-g-C3N4 [26, 46]. The AB/CN (5:1) showed the best degradation performance, and almost all EPH was decomposed in 60 min under visible light irradiation, with a rate constant of 0.1022 min-1. The results indicated that combination of AgBr with P-g-C3N4 could enhance the EPH degradation apparently. The physically mixed material of AgBr and P-g-C3N4 with the

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mass ratio of 5:1 was also prepared, which showed lower EPH degradation rate than

that of AB/CN (5:1) composite and even lower than that of AgBr. The results suggested that efficient charge transfer in the interface of AB/CN composite was

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obtained, indicating the heterojunction formed in the composite. The enhanced

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photocatalytic property of AB/CN composites may be due to the formed Z-scheme structure and the surface plasmon resonance (SPR) effect of Ag. SPR effect and

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Z-scheme mechanism were proved to enhance the separation efficiency of

composite [30].

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photo-induced electrons and holes, leading to the high performance of AN/CN

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The cycling run experiment of EPH degradation by AB/CN (5:1) composite under visible light irradiation was carried out to evaluate the recyclability of

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composite. As presented in Figure S5, after three consecutive cycles, there was no significant loss of photocatalytic performance of AB/CN (5:1) composite, and the EPH degradation rate of 84.2% was obtained within 90 min. AB/CN (5:1) was a superior photocatalytic catalyst with favorable stability, indicating its possibility of practical applications in water pollution remediation. 14

3.3 Effect of pH, DOM and inorganic ions 3.3.1 Role of initial solution pH The solution pH value not only influences the transformation of reactive oxygen species (ROS), but also leads to the charge changes in photocatalysts and contaminants, resulting in the different degradation efficiencies [47]. The degradation of EPH over AB/CN (5:1) composite in different pH solutions was presented in

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Figure 6. The degradation efficiency was low at pH value of 3 and 5, which was likely due to the aggregation of photocatalysts under acidic condition, leading to the

reductive utilization efficiency of photos [48]. In addition, the acidic condition would

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result in the release of Ag+ and increase the recombination rate of e- and h+ [43].

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However, high pH value (pH=9 and 10) increased the degradation rates of EPH, which was ascribed to the abundant production of hydroxyl radical (·OH) at alkaline

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condition. The h+ in the system could oxidize OH- to generate ·OH by Equation (5)

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[49]. With the continual increase of pH value to 11, the degradation rate slightly decreased, which can be attributed to the quenching of ·OH at high pH value and

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charge changes of AB/CN and EPH. The best performance for EPH degradation was found at pH 10. The pKa1 and pKa2 values of EPH are 9.4 and 13.5, respectively, and

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the distribution of EPH different species (protonated, non-protonated and deprotonated) under various pH values were shown in Figure S6. The predominant form of EPH was non-protonated form at pH 10, which was more susceptible to the ROS involved in the degradation process. 3.3.2 Effect of DOM 15

DOM could act as a photosensitizer to produce singlet DOM (1DOM), singlet oxygen (1O2), ·OH and excited triplet states (3DOM*) under visible light irradiation, which could destruct contaminants by oxidation reaction [47], meanwhile, ROS involved in the reaction systems could be quenched by DOM as well [50]. The degradation rates of EPH with addition of different concentrations of DOM were given in Figure 7a and Figure S7a. DOM at low concentrations slightly enhanced the

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EPH degradation. The improved EPH degradation rate in the presence of DOM (1 mg·L-1) was likely because various radicals were generated in the reaction system. However, high DOM would compete for active sites and quench ROS, resulting in the

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decreased degradation rate of EPH.

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3.3.3 Effect of HCO3-

The influence of HCO3- on EPH degradation over AB/CN composite was shown

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in Figure 7b and Figure S7b. Low concentrations of HCO3- (0.5 ~ 4 mM) enhanced

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the EPH degradation. The generation of CO3·- by HCO3- in degradation system may be the major reason for the fast degradation of EPH (Equation 6) [51]. However, the

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degradation of EPH decreased with the addition of high concentrations of HCO3- (10 ~ 20 mM). HCO3- could scavenge holes and ·OH involved in the photocatalytic

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degradation system, as indicated by Equation (6) and (7) [52]. In addition, a portion of HCO3- can be absorbed on the surface of composites, resulting in the decreased degradation efficiency. In conclusion, HCO3- had a dual effect on EPH degradation by AB/CN composite, and the dual role of HCO3- were mainly ascribed to the scavenge of ·OH and generation of CO3·-. 16

h+ +OH- → ·OH

(5)

HCO3- + ·OH → CO3·- + H2O (6) HCO3- + h+ → HCO3·-

(7)

CO32- + ·OH→ CO3·- + OH-

(8)

3.3.4 Effect of CO32The influence of CO32- on the degradation of EPH was presented in Figure 7c

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and Figure S7c. EPH degradation was enhanced with the addition of carbonate ion

(0.5~10 mM). The fastest EPH degradation was achieved with the CO32concentration at 2 mM (with the rate constant of 0.1300 min-1). CO32- has been

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reported to increase the degradation rate of oxytetracycline by UV/H2O2 [53]. The

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addition of CO32- resulted in high pH value which favored the EPH degradation. CO32could scavenge the hydroxyl radical to generate carbonate radical (CO3·-) (Equation 8)

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[54]. Furthermore, the carbonate radical is a selective oxidant which tends to attack

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the electron rich compounds containing nitrogen or sulfur [55], resulting in the fast photocatalytic degradation of EPH in the reaction system. However, the increased

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degradation rate by CO32- addition was lower than that in HCO3- addition experiment, which was mainly ascribed to the faster scavenging reaction with ROS by CO32- [56].

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In addition, extensive carbonate in photocatalytic system may decrease the degradation efficiency of compounds by scavenging ROS and increasing the negative charge in the composite surface [57]. The result was coincidence with the phenomenon that higher CO32- concentration (5 mM ~ 10 mM) caused a slighter improvement of EPH degradation in Figure 7c. 17

3.3.5 Effect of NO3The presence of NO3- inhibited the EPH degradation by AB/CN composite (Figure 7d, Figure S7d). The higher the NO3- concentrations, the lower the EPH degradation efficiency. The result was in accordance with the photocatalytic degradation of disulfoton over TiO2 [58]. The inhibitory effect of NO3- on EPH degradation was related to the excessive nitrate present in the system that could create

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a competitive adsorption on the photocatalysts surface, which resulted in the

decreased degradation efficiency [58]. Besides, the larger size of nitride anion than other anions caused the higher suppression since it could cover a larger area on the

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surface of AB/CN composite. Moreover, h+ or ·OH could also be scavenged by

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reacting with NO3- (Equation 9 and 10) [59], which resulted in the decreased EPH degradation.

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NO3- + h+ → NO3·

(9)

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NO3- + ·OH → NO3· + OH- (10) 3.4 Photocatalytic degradation mechanism

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3.4.1 Identification of ROS

Trapping experiments were carried out to identify the major ROS involved in

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EPH degradation by AB/CN (5:1). In this process, EDTA-2Na, IPA, NaN3 and BQ were acted as the scavengers of h+, ·OH, 1O2 and superoxide radicals (·O2-), respectively. Figure 8 showed that EPH degradation was remarkably suppressed when EDTA-2Na or BQ were added, and the rate constants were as low as 0.0041 min-1 and 0.0023 min-1, respectively, indicating the major roles of h+ and ·O2- in reaction system. 18

A moderate inhibition was observed with the addition of NaN3, confirming the relatively important role of 1O2. In the presence of IPA, the photocatalytic degradation of EPH was inhibited slightly, illustrating the minor role of ·OH in the photocatalytic system. The result indicates that h+ and ·O2- are the predominant ROS involved in the EPH degradation over AB/CN composite. 3.4.2 Detection of ·OH and ·O2-

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Electron-spin resonance (ESR) technique was carried out to further confirm the

generation of ·OH and ·O2- during the degradation process, with the experimental methods provided in Text S3. As shown in Figure 9a-9d, no signals were observed in

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the darkness, indicating no ·OH or ·O2- were produced. Four characteristic peaks with

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the intensity ratio of 1:2:2:1 could be found in Figure 9a-9b, indicating that ·OH was generated by P-g-C3N4 and AB/CN (1:1) under visible light irradiation (10 min). Six

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characteristic peaks of DOMP-·O2- adducts were detected in P-g-C3N4 and AB/CN

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photocatalytic degradation system, illustrating the generation of ·O2- by the two materials under visible light irradiation (Figure 9c-9d). The enhanced intensity of

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DMPO-·OH and DOMP-·O2- adducts peaks illustrated that more ·OH and ·O2- were generated by AB/CN composite than that of P-g-C3N4 under visible light irradiation.

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3.4.3 Photocatalytic degradation mechanisms Figure 10 illustrated a proposed Z-scheme mechanism in the process of EPH

degradation over AB/CN composite. Electrons were produced in the CB of P-g-C3N4 and AgBr, where holes were generated in the VB of those two materials under visible light irradiation. The electrons in the CB of P-g-C3N4 (-1.02 eV) could reduce oxygen 19

to generate ·O2- (-0.33 eV [19]) , and holes in the VB of AgBr (2.42 eV) could react with OH- to produce ·OH (2.38 eV [60]). These ROS and holes involved in the reaction system could react with EPH directly, resulting in its efficient degradation. However, e- and h+ could attract each other spontaneously and lead to its rapid recombination, resulting in the lower photocatalytic performance of AB/CN composite. In a Z-scheme mechanism, Ag acted as a cross-linking bridge between

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AgBr and P-g-C3N4, and e- in the CB of AgBr could transfer to Ag nanoparticles

because the CB of AgBr was more negative than the Fermi level of Ag (0.4 V [61]). Simultaneously, the h+ in the VB of P-g-C3N4 was transferred to Ag nanoparticles and

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integrated with e-. As a result, Ag nanoparticles could serve as the charge transmission

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bridge to promote the separation rate of e-/h+ pairs on the AB/CN composite. Therefore, the holes in the VB of AgBr and the ROS generated in the system would

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effectively destruct the contaminants. The Z-scheme heterojunction efficiently

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enhanced the separation rate of e-/h+ pairs, thus enhanced the photocatalytic performance of AB/CN composite dramatically.

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3.5 Degradation of EPH in real natural water The prepared samples were used to degrade EPH in real natural waters to

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evaluate their practical application in removing water contaminants, and the results were shown in Figure 11. Surface water (SW), tap water (TW), ultrapure water (UW) and a secondary effluent (SE) were chosen in the test, and the detail information of water quality parameters was presented in Table S2. The degradation efficiency of EPH in TW was similar to that in UW, with the rate constants of 0.1197 min-1 and 20

0.1022 min-1, respectively. A degradation rate constant of 0.0409 min-1 was achieved in surface water, and 99.5% EPH was removed within 90 min irradiation. Compared to TW and SW, degradation of EPH in SE was significantly inhibited, but it also showed a degradation efficiency of 91.8% within 90 min. The existence of inorganic anions (chlorides, sulfates and carbonates) and DOM in SW and SE resulted in the decreased degradation of EPH [47]. It showed that AgBr/P-g-C3N4 exhibited a high

possibility of its practical application. 3.6 Reaction intermediates and degradation pathways

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photocatalytic performance on EPH degradation in real natural water, implying the

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The intermediates of EPH degradation over AgBr/P-g-C3N4 composite were

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identified by UPLC-MS/MS. The mass spectra collected during the experiment were used to investigate the EPH intermediates. The structure of intermediates was not

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exactly determined because of the lack of standard substances. Therefore, in this work,

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the structure of EPH intermediates was identified by the fragments and specific molecules instead of analyzing the standards [62]. The chromatogram of samples and

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UPLC-MS/MS spectra was shown in Figure S8-S10, implying that the retention time of the parent peak was 1.48 min and several new peaks were obtained. The details of

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retention time, m/z and formulas of intermediates were present in Table S3. Based on the molecular weight and the methods used in previous reports [63-65], a total of 17 intermediates were identified, and the degradation pathways were shown in Figure 12. EPH was attacked by reactive species and resulted in the formation of amphetamine (P1) via dihydroxylation and demethylation. Amphetamine is a common product 21

involved in numerous destruction processes of illicit drugs [66]. Amphetamine experienced the processes of deamination and hydroxylation to generate the intermediate P7, which was further oxidized to create the product P8. P6 and P9 were generated due to the breakage of C-C bond of amphetamine and EPH, respectively. The dehydration of EPH produced the intermediate P2, and P2 was further transformed to P3 through demethylation. The reduction of P3 produced P4, which

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could further generate P5 by oxidation. Oxidation of EPH by ROS attack generated the intermediate P10, and P10 was further transformed to P11 by the reaction of

demethylation and bond breaking. The reaction of hydroxylated EPH caused the

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formation of intermediate P12, which can be further hydroxylated to produce P13.

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The oxidation of P13 by ROS attack contributed to the generation of P14, and the subsequent substitution reaction yielded the intermediate P15. Adding small group

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(-NHCH3) on P15 resulted in the intermediate P16, which further conduced to the

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generation of P17. The reaction mechanism involved in EPH degradation pathways including bond breaking, oxidation, deamination, demethylation, dehydration,

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hydroxylation, dehydroxylation and reduction. The proposed mechanism revealed all possible intermediates of EPH. Considering the mechanism of organic compounds

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degradation over other photocatalytic system disclosed by former researchers [67, 68], all intermediates of EPH were further converted to small molecular substances such as CO2, NH4+ and H2O by ·OH and ·O2-. 3.7 Toxicity measurements

22

EPH and its degradation intermediates may pose threats to aquatic ecosystem, and the biological toxicity of EPH degradation over AB/CN (5:1) composite at different irradiation times was investigated by using Vibrio Qinghaiensis sp. Q-67 (Q67). The method of toxicity measurements was shown in Text S4 [69,70], and the toxicity (expressed as % inhibition) of EPH and its degradation intermediates was shown in Figure 13. The toxicity of parent EPH expressed an inhibition rate of 16.26%

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(0 min of illumination). The inhibition rate was low at first but increased slightly with

the reaction went on, implying that the toxicity of some intermediates was higher than that of parent compound. However, lower toxicity of intermediates was obtained after

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further irradiation, and only 3.09% luminescence inhibition rate of Q67 was found

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after 90 min irradiation. The result indicated that the ecotoxicity of EPH can be weakened or be vanished after degradation by AB/CN composite.

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4 Conclusions

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In this study, a series of AgBr/P-g-C3N4 composites were fabricated by thermal polymerization and in-situ precipitation-deposition methods. AgBr/P-g-C3N4 (5:1)

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showed the best photocatalytic performance on EPH degradation, which could remove 99.9% EPH within 60 min under visible light irradiation. The ROS

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identification and radical quenching tests indicated that h+, ·O2- and 1O2 were the dominant ROS involved in EPH degradation. Photocatalytic degradation of EPH was pH-dependent, with the best degradation performance achieved at pH 10. Low concentrations of DOM or HCO3- accelerated EPH degradation, but high levels inhibited the degradation. CO32- enhanced EPH degradation, and NO3- suppressed the 23

degradation. Seventeen reaction intermediates/products were identified, and the degradation pathways included bond breaking, oxidation, deamination, demethylation, dehydration, hydroxylation, dehydroxylation and reduction. AgBr/P-g-C3N4 also exhibited a relatively good photocatalytic performance on EPH removal in real nature waters.

Author statement Chen:

Investigation,

Writing-

original

draft;

Changsheng

Guo:

ro of

Miao

Conceptualization, Writing- review& editing; Song Hou: Methodology, Formal

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analysis; Jiapei Lv: Methodology; Yan Zhang: Validation; Heng Zhang:

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Investigation; Jian Xu: Conceptualization, Writing- review& editing, Supervision

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Declaration of interests

The authors declare that they have no known competing financial interests or personal

na

relationships that could have appeared to influence the work reported in this paper.

The authors declare the following financial interests/personal relationships which may

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ur

be considered as potential competing interests:

24

Acknowledgement This work was financially supported by National Water Pollution Control and Treatment Science and Technology Major Project (No. 2017ZX07302001 and 2017ZX07301003) and the National Natural Science Foundation of China (No.

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na

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re

-p

ro of

41673120).

25

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Figure 1. XRD patterns (a), FT-IR spectra (b), UV-vis DRS (c) and band energy of the as-prepared

Jo

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samples.

38

ro of -p re lP na ur Jo Figure 2. TEM images of P-g-C3N4 (a), AB/CN (1:2) (b), AB/CN (3:4) (c), AB/CN (5:1) (d, e), and AgBr (f). 39

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Figure 3. HRTEM image of AgBr/P-g-C3N4 (5:1) composite.

40

ro of -p re lP na ur Jo Figure 4. XPS survey spectra (a) of samples and high-resolution XPS spectra of AB/CN (5:1) composite: C 1s (b), N 1s (c), O 1s (d), P 2p (e), Ag 3d (f) and Br 3d (g). 41

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Figure 5. Photocatalytic degradation of EPH by the synthesized samples (a) and the corresponding

Jo

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pseudo-first-order kinetic curves (b).

42

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Figure 6. Effect of initial pH value on degradation of EPH by AB/CN composite (a) and the

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corresponding pseudo-first-order kinetic curves (b).

43

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Figure 7. Effect of DOM (a), HCO3- (b), CO32- (c) and NO3- (d) on the degradation of EPH over

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AB/CN composite.

44

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Figure 8. Effect of different scavengers on degradation of EPH by AB/CN composite (a) and the

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corresponding pseudo-first-order kinetic curves (b).

45

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Figure 9. ESR signals of DMPO adduct with ·OH (a-b) or ·O2- (c-d).

46

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Figure 10. Schematic diagram of the charge separation and transfer in the process of EPH

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degradation over AB/CN composite.

47

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Figure 11. Degradation of EPH in different natural waters by AB/CN composite (a) and the

Jo

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na

lP

re

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corresponding pseudo-first-order kinetic curves (b).

48

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Figure 12. Proposed degradation pathways of EPH by AB/CN composite.

49

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Figure 13. Inhibition rate of Q67 as a function of the irradiation time for EPH degradation.

50