Degradation of Ofloxacin by Perylene Diimide Supramolecular

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Sustainability Engineering and Green Chemistry

Degradation of ofloxacin by perylene diimide supramolecular nanofiber sunlight-driven photocatalysis Ping Chen, Lee Blaney, Giovanni Cagnetta, Jun Huang, Bin Wang, Yujue Wang, Shubo Deng, and Gang Yu Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b05827 • Publication Date (Web): 03 Jan 2019 Downloaded from http://pubs.acs.org on January 5, 2019

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Environmental Science & Technology

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Degradation of ofloxacin by perylene diimide supramolecular nanofiber sunlight-driven

2

photocatalysis

3 4

Ping Chena#, Lee Blaneya,b#, Giovanni Cagnettaa, Jun Huanga, Bin Wanga,

5

Yujue Wanga, Shubo Denga, Gang Yua*

6 7

a School

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Beijing Key Laboratory for Emerging Organic Contaminants Control, Tsinghua University, Beijing 100084, China

9

b University

of Environment, State Key Joint Laboratory of Environmental Simulation and Pollution Control (SKLESPC),

of Maryland Baltimore County, Department of Chemical, Biochemical, and Environmental Engineering, 1000

10

Hilltop Circle, Engineering 314, Baltimore, Maryland 21250, United States

11

# The first two authors contributed equally to this work.

12 13

* Corresponding Author:

14

Gang YU (Ph.D.)

15

School of Environment, Tsinghua University, Beijing, 100084, China

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Tel: +86-10-62787137; Fax: +86-10-62785687; E-mail: [email protected]

17

ACS Paragon Plus Environment


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Abstract

19 20

This study describes a promising sunlight-driven photocatalyst for the treatment of ofloxacin and other fluoroquinolone

21

antibiotics in water and wastewater. Perylene diimide (PDI) supramolecular nanofibers, which absorb a broad spectrum of

22

sunlight, were prepared via a facile acidification polymerization protocol. Under natural sunlight, the PDI photocatalysts

23

achieved rapid treatment of fluoroquinolone antibiotics, including ciprofloxacin, enrofloxacin, norfloxacin, and ofloxacin.

24

The fastest degradation was observed for ofloxacin, which had a half-life of 2.08 min for the investigated conditions.

25

Various light sources emitting in the UV-visible spectrum were tested, and blue light was found to exhibit the fastest

26

ofloxacin transformation kinetics due to the strong absorption by the PDI catalyst. Reactive species, namely h+, 1O2, and

27

O2•-, comprised the primary photocatalytic mechanisms for ofloxacin degradation. Frontier electron density calculations

28

and mass spectrometry were used to verify the major degradation pathways of ofloxacin by the PDI-sunlight

29

photocatalytic system and identify the transformation products of ofloxacin, respectively. Degradation mainly occurred

30

through demethylation at the piperazine ring, ketone formation at the morpholine moiety, and aldehyde reaction at the

31

piperazinyl group. An overall mechanism was proposed for ofloxacin degradation in the PDI-sunlight photocatalytic

32

system, and the effects of water quality constituents were examined to determine performance in real water/wastewater

33

systems. Ultimately, the aggregate results from this study highlight the suitability of the PDI-sunlight photocatalytic

34

system to treat antibiotics in real water and wastewater systems.

35 36

TOC GRAPHIC

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INTRODUCTION

39 40

Antibiotics are common organic micropollutants that are routinely detected in surface water, wastewater, and drinking

41

water.1,2 Over-prescription and -consumption of antibiotics, along with their incomplete removal in wastewater treatment

42

plants, have increased concerns associated with antibiotic presence in water resources. Ofloxacin is a fluoroquinolone

43

antibiotic that has been frequently detected in surface water at concentrations as high as 31.7 μg L-1. 3,4 Despite the

44

relatively low concentration, ofloxacin occurrence can adversely affect ecological systems by inhibiting the growth of

45

microorganisms, inducing the selection of antibiotic resistant bacteria, and leading to other ecotoxicological effects.5,6 For

46

these reasons, novel treatment processes are required for the selective and efficient removal of ofloxacin and other

47

fluoroquinolone antibiotics from diverse water and wastewater matrices.

48 49

Conventional activated sludge systems and advanced oxidation process are commonly used for wastewater treatment.7

50

However, conventional activated sludge generally demonstrates low removal for organic micropollutants, such as

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ofloxacin.8 The treatment of ofloxacin and/or other fluoroquinolone antibiotics has also been studied in various systems,

52

such as electrochemical oxidation, ultraviolet (UV)-coupled systems (e.g., UV-H2O2), acoustic chemical oxidation, ozone-

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coupled systems (e.g., O3-H2O2), Fenton oxidation, persulfate oxidation, laccase-mediated oxidation, and photocatalytic

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oxidation.9-13 Of these, environmentally friendly photocatalytic materials can utilize solar energy without the need for

55

exogenous chemical reagents or electrical energy.14-18 These advantages make novel sunlight-photocatalysis systems a

56

desirable technology for treatment of micropollutants.

57 58

Traditional semiconductor photocatalysts are characterized by high cost, poor structural adjustability, low activity under

59

solar irradiation, and toxicity concerns from leachable metal content. For these reasons, new photocatalytic materials need

60

to be explored. Perylene diimide (PDI) and its derivatives are considered to be the best n-type organic semiconductors.19

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Recently, PDI has been used in solar cells,20 fluorescent materials,21,22 single-molecule sensors,23 and organic transistors24

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due to its unique photoelectrical properties. Importantly, PDI can be formed into supramolecular nanofibers through

63

acidification polymerization,25 and the resultant nanofibers have strong photothermal stability and notable photo-oxidation

64

and -reduction characteristics. Ongoing research has focused on the synthesis, characterization, and application of PDI-

65

based materials.26 Since 2014, traditional semiconductor photocatalysts have been modified with PDI to split water via ACS Paragon Plus Environment


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composite catalysts, such as PDI/Ni,27 PDI/Ru,28 PDI/TiO2,29 and PDI/C3N4.30 The results of these studies have shown that

67

PDI incorporation improved the activity of traditional semiconductor materials under visible light. However, these

68

composite catalysts were still primarily comprised of traditional semiconductor materials but with a small amount of PDI

69

nanofibers to facilitate the migration of charge carriers.

70 71

In recent years, Zhu and coworkers improved the preparation and synthesis protocols for novel PDI supramolecular

72

nanofibers that catalyze phenol degradation with visible light excitation.31-33 Under identical experimental conditions, the

73

phenol degradation rate with PDI supramolecular nanofibers was 20 and 13 times that of non-supramolecular PDI and g-

74

C3N4, respectively. The PDI photocatalyst possessed unique one-dimensional π conjugation and internal electric field

75

effects, which enabled electrons to migrate to the nanofiber surface. At the surface, electrons reacted with dissolved

76

oxygen to form reactive oxygen species that resulted in strong oxidation activity.34 Because many organic micropollutants

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contain phenol moieties or other functional groups that interact with reactive oxygen species, the application of PDI

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photocatalysis to these emerging contaminants merits further study. Given the high energetic costs of UV light in UV-

79

based advanced oxidation processes,35 sunlight-activated photocatalysts are needed to ensure sustainable treatment. No

80

previous efforts have explored this facet of PDI-based catalysts.

81 82

To date, the synthesis, application, and understanding of PDI supramolecular nanomaterials involve the following

83

shortcomings: (i) the metal modification poses a potential threat to aquatic ecosystems; (ii) applications have been limited

84

to production of hydrogen and degradation of simple chemicals (e.g., phenol); (iii) natural sunlight has not been employed

85

for photocatalytic treatment of contaminants; and, (iv) mechanistic insight into the photocatalytic reactions has not been

86

studied. These shortcomings represent knowledge gaps in (1) production of organic PDI nanofibers without transition

87

metals, (2) application of PDI supramolecular nanomaterials to treat organic micropollutants in water and wastewater, (3)

88

utilization of solar irradiation, and (4) identification of reaction mechanisms, pathways, and transformation products. For

89

structurally-complex environmental contaminants, such as ofloxacin, the transformation products generated by PDI

90

photocatalytic reactions may retain toxicological concerns and, therefore, need to be identified. We note that no previous

91

reports have explored the potential for PDI-based materials to transform micropollutants.

92 93

In this study, PDI supramolecular, sunlight-activated photocatalysts were prepared by acidification polymerization, and ACS Paragon Plus Environment

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the catalytic mechanism of these novel materials was investigated for degradation of the ofloxacin micropollutant. The

95

research content includes the preparation and characterization of PDI, the catalyst activity for different light sources, the

96

adsorption and degradation of ofloxacin on the surface of the catalyst, the generation of reactive species, and the impacts

97

of water quality parameters (e.g., pH, cations, anions, and dissolved organic matter (DOM)) in the PDI-sunlight

98

photocatalytic system. The transformation products of ofloxacin were determined and used to inform the photocatalytic

99

mechanism and ofloxacin degradation pathway. Ultimately, these results highlight the potential application of novel,

100

sustainable, sunlight-driven photocatalysts to address growing concerns with micropollutants in water and wastewater.

101 102

MATERIALS AND METHODS

103 104

Preparation and analysis of PDI supramolecular photocatalyst nanofibers. A brief overview of the reagents and ultrapure

105

water is provided in Text S1 of the Supporting Information (SI). Preparation of the PDI supramolecular photocatalyst

106

generally followed the protocol of Zhu et al.,31 although key improvements and modifications were implemented here. Briefly,

107

6 mM 3,4,9,10-perylenetetracarboxylic diimide, 40 mM β-alanine, and 25 g imidazole were accurately weighed and heated at

108

100 °C for 4 h under argon. After cooling to room temperature, the mixture was dispersed in 180 mL of ethanol and 400 mL of

109

2 M hydrochloric acid (HCl) and stirred for 12 h. The resulting alanine-modified material was filtered through a 0.45-μm

110

membrane (Beijing Ruifeng Tongchuang Analysis Instrument Co. Ltd.; Beijing, China). After washing with ultrapure water and

111

drying at 60 °C, the solid substance was dissolved in 250 mL solution containing 0.6% (v/v) triethylamine. Then, 35 mL of 4 M

112

HCl was added to spontaneously form nanofibers. After two repeated cycles of washing with ultrapure water and drying at

113

60 °C, the PDI supramolecular nanofibers were produced.

114 115

The morphology of PDI supramolecular nanofibers was investigated by transmission electron microscopy (TEM) on a

116

JEM-2100F instrument at an acceleration voltage of 200 kV (Japan Electron Optics Laboratory Company; Japan). X-ray

117

diffraction (XRD) patterns were obtained with an Ultima III diffractometer (Rigaku, Japan). The d-spacing from XRD

118

patterns was calculated using Eq. 1.

119 120

2𝑑sin 𝜃 = 𝑛𝜆

121

(Eq. 1)

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In Eq. 1, 𝜃 is the diffraction angle (recorded from the XRD pattern), 𝜆 is the x-ray wavelength, and 𝑛 was the order of

124

diffraction (first). Other structural, morphological, and absorbance characterization methods are described in Text S2 of

125

the SI.

126 127

Photocatalysis experiments. Experimental solutions with 200 mg L-1 of PDI and 8 mg L-1 of ofloxacin were generated in

128

200 mL beakers and ultrasonically dispersed for 30 min to ensure that the PDI catalyst and ofloxacin contaminant

129

achieved adsorption equilibrium.18 The temperature was 28 ± 3 °C, and a magnetic stirrer was used to ensure well-mixed

130

conditions. Photocatalysis experiments were carried out under actual sunlight. Tests were conducted at 10:00 a.m. to

131

14:00 p.m. on sunny days in May through September in 2017 and 2018 in Beijing, China (N 40.00°, E 116.33°). Light

132

intensity was measured using the CEL-NF2000 light power meter from Beijing Zhongjiao Jinyuan Technology Co.

133

(China).

134 135

To examine the photocatalyst activity under different light sources, six 9 W light-emitting diodes (LEDs; Epileds

136

Technologies, Taiwan) were used to generate UV light (390-400 nm; 3.4 mW cm-2), blue light (455-460 nm; 5.8 mW

137

cm-2), green light (515-530 nm; 5.8 mW cm-2), yellow light (590-595 nm; 3.7 mW cm-2), red light (655-660 nm; 13.0 mW

138

cm-2), and white light (400-840 nm; 9.2 mW cm-2). The LEDs were placed 10 cm away from the experimental solutions.

139 140

Reactive oxygen species were investigated by electron spin resonance (ESR) spectroscopy (JES-FA200, Bruker;

141

Germany). Apparent, time-based ofloxacin rate constants were measured in the presence and absence of quenching agents

142

to inform reaction mechanisms. Detailed information about reactive species determination is provided in Text S3 of the

143

SI. These data were used to calculate the contribution of reactive species (e.g., •OH, 1O2, O2•-, h+) to the photocatalytic

144

degradation of ofloxacin using Eqs. 2-5.

145 146

R•OH =

k•OH 𝑘app

≈

(𝑘app ― 𝑘IPA) 𝑘app

147

(Eq. 2)

148 149

R O2 = 1

k1O2 𝑘app

≈

(𝑘IPA ― 𝑘NaN3) 𝑘app


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150

(Eq. 3)

151 152

kO•2 -

RO•2 - =

𝑘app

≈

(𝑘app ― 𝑘TEMPL) 𝑘app

153

(Eq. 4)

154 155

Rh + =

kh + 𝑘app

≈

(𝑘app ― 𝑘Na2C2O2) 𝑘app

156

(Eq. 5)

157 158

In Eqs. 2-5, R is the fractional contribution of a particular reactive species (e.g., •OH, 1O2, O2-•, h+) to the apparent rate

159

constant (kapp) and ki is the rate constant for the degradation of ofloxacin by a certain reactive species (indicated by the

160

•OH, 1O

161

sodium azide (NaN3), 4-hydroxy-2,2,6,6-tetramethylpiperidinyloxy (TEMPL), and sodium oxalate (Na2C2O4) subscripts).

2,

O2-•, and h+ subscripts) or in the presence of a quenching agent (indicated by the isopropyl alcohol (IPA),

162 163

To investigate pH effects on ofloxacin degradation, the fraction of each protonated/deprotonated species in solution (i.e.,

164

α0, α1, and α2) was calculated using Eqs. 6-9.36

165 166

∑ α [OFL]

[OFL]total = [OFL + ] + [OFL +/ - ] + [OFL - ] =

i

total

i = 0,1,2

167

(Eq. 6)

168 169

(

α0 = 1 +

𝐾a1 [H + ]

+

)

𝐾a1𝐾a2 [H + ]

2

―1

170

(Eq. 7)

171

172

(

𝐾a2 [H + ] +1+ α1 = 𝐾a1 [H + ]

)

―1

173

(Eq. 8) ACS Paragon Plus Environment


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(

2

)

[H + ] [H + ] + +1 α2 = 𝐾a1𝐾a2 𝐾a2

175

―1

176

(Eq. 9)

177 178

In Eq. 6, [OFL] is the molar ofloxacin concentration, α is the ionization factor, the “+”, “+/-”, and “-” superscripts indicate

179

the ofloxacin cation, zwitterion, and anion, respectively, the i subscript represents the ofloxacin species from most (0) to

180

least (2) protonated, and the “total” subscript indicates the total concentration. In Eqs. 7-9, 𝐾a1 and 𝐾a2 are the acid

181

dissociation constants for ofloxacin and [H + ] is the molar hydronium concentration.

182 183

When studying the effects of environmental factors, SO42- (Na2SO4, 10 mg L-1), NO3- (NaNO3, 10 mg L-1), Cu2+ (CuSO4, 5

184

mg L-1), DOM (fulvic acid, 100 mg C L-1), Cl- (NaCl, 10 mg L-1), or HCO3- (NaHCO3, 10 mg L-1) were added to the

185

experimental solution containing 200 mg L-1 of PDI and 8 mg L-1 of ofloxacin. To identify the effects of real water

186

matrices on ofloxacin degradation by the PDI-sunlight photocatalytic system, experimental solutions were generated by

187

adding ofloxacin powder to 0.45-μm membrane-filtered tap water (Beijing, China), river water (Pearl River water;

188

Guangzhou, China), and wastewater effluent (Beijing, China). For all conditions, the solution pH (3.0-8.0) and the optical

189

power density of sunlight (45-55 mW cm-2) were measured during experimentation. Samples were collected at pre-

190

determined times. After the catalyst was filtered out, the ofloxacin concentration was determined by high performance

191

liquid chromatography (HPLC; LC-20AT, Shimadzu, Japan).

192 193

Ofloxacin analysis and identification of transformation products. Ofloxacin concentrations were measured by HPLC

194

with photodiode array (PDA) detection (Shimadzu UV model with SPD-M20A) and the transformation products were

195

analyzed by ultraperformance liquid chromatography with quadrupole time-of-flight mass spectrometry (UPLC-Q-TOF-

196

MS; Agilent Technologies 1290UHPLC/6540QTOF-MS; United States). Detailed information on these techniques is

197

provided in Text S4 and Table S1 of the SI.

198 199

Computational methodology to predict attack sites. The conduction band (CB) value was calculated using Eq. 10.


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201


CB = VB ― [GAP]

202

(Eq. 10)

203 204

In Eq. 10, VB is the valence band value and [GAP] is the band gap width of PDI, which was calculated from the UV-

205

visible/diffuse reflectance spectroscopy (UV-vis/DRS) spectrum.

206 207

The B3LYP density functional theory model and the 6-311+g (d, p) pople basis set were used to process the ofloxacin

208

molecular orbital calculations in Gaussian 09. In particular, these tools were used to calculate the frontier electron

209

densities (FEDs) of the highest occupied molecular orbital (HOMO) and the lowest unoccupied molecular orbital

210

(LUMO), along with the point charges. These data were subsequently used to predict the site of reactive species attack on

211

ofloxacin.

212 213 214

RESULTS AND DISCUSSION

215 216

Structure, morphology, and absorbance spectrum of the PDI supramolecular nanofibers.

217

Following the protocols described above, the PDI supramolecular nanofibers were self-assembled with carboxy-

218

substituted PDI molecules through H-type π-π stacking and hydrogen bonding; additional information is available in

219

Wang et al.31 The incorporation of β-alanine improved the growth of the supramolecular system by providing short

220

carboxylic acid side chains that facilitated columnar stacking of perylene groups and increased solubility.31 The phase

221

structures of the synthesized PDI supramolecular nanofibers were elucidated by XRD analysis (Fig. 1a). The PDI material

222

had a high degree of crystallinity and, according to Eq. 1, the typical d-spacing of π-π stacking was 3.44 Å for the labeled,

223

Pa peak. The small d-spacing of π-π stacking in the PDI nanofibers is expected to benefit the migration of charge

224

carriers.29 Fig. 1b–d depict the morphologies of the prepared PDI materials by TEM and atomic force microscopy (AFM).

225

The TEM images revealed fairly uniform nanofibers with ~20-50 nm diameters and ~100-400 nm lengths. The height

226

profile was determined by AFM and confirmed the ~20 nm diameter of individual nanofibers. Following acidification

227

polymerization, the PDI materials exhibited a small surface area (Fig. S1 in the SI), suggesting a low adsorption capacity

228

that may negatively affect photocatalytic mechanisms. ACS Paragon Plus Environment


(a)

(b)

Intensity (a.u.)

PDI 6.2

14 9.9

5

10

229

15

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Pa, 25.9 20 22.9

20

2 Theta (degree)

25

30

(d)

(c)

~40 ~40 nm nm ~20 nm

~20 nm

230 231

Figure 1. (a) XRD pattern of the prepared PDI photocatalyst; (b) and (c) TEM images of PDI supramolecular nanofibers; and,

232

(d) AFM analysis of an individual PDI nanofiber.

233 234

The functional groups of the prepared PDI supramolecular nanofibers were identified through Fourier-transform infrared

235

(FTIR) spectroscopy (Fig. S2a in the SI). The stretching vibration band at ~3450 cm-1 was assigned to C-OH in carboxylic

236

acid moieties; accordingly, the absorption band at 1700 cm-1 was attributed to the C=O of the same group. Another C=O

237

group at 1650 cm-1 was assigned to keto functionalities in the PDI molecular structure.37 The stretching vibration band at

238

1440-1600 cm-1 was attributed to the C=C groups from the aromatic carbon skeleton. The stretching vibration peaks of the

239

C-H (1350 cm-1), CH2 (2850-3090 cm-1), and σC-H (615-1015 cm-1) correspond to in-plane bending vibrations from the

240

aromatic rings, the aliphatic group, and the out-of-plane bending vibrations in the aromatic rings, respectively. The

241

absorption peaks of -N-C=O at ~1400 cm-1 and -N-CH2 at ~1030-1240 cm-1 indicated the presence of amide groups in the ACS Paragon Plus Environment

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PDI nanofibers and chemical reactions between β-alanine and PDI, respectively.38 Overall, the stretching vibration peaks

243

demonstrated the existence of carboxyl, -N-CH2, and -CH2 groups in the PDI supramolecular nanofibers. Among these

244

groups, the carboxyl and -N-CH2 functionalities may aid in sorption of ofloxacin cations/zwitterions and

245

zwitterions/anions, respectively. The innovative acidification polymerization protocol contributed to these newly

246

discovered groups. X-ray photoelectron spectroscopy (XPS) analysis further confirmed that multiple functional groups

247

were present on the surface of the PDI supramolecular nanofibers (Fig. S2b-e in the SI). These results, along with the

248

FTIR spectroscopy analysis, confirmed that β-alanine was successfully introduced into the PDI supramolecular

249

nanofibers.

250 251

Fig. S2f in the SI displays the UV-vis/DRS spectrum of the PDI nanofibers. The PDI photocatalyst presented a wide

252

absorption edge at ~730 nm, suggesting that the supramolecular nanofibers are capable of being activated by visible light

253

(i.e., 390-760 nm). Compared to traditional photocatalysts, such as TiO2 (~320 nm), BiPO4 (~300 nm), and g-C3N4 (~450

254

nm), the absorption spectrum of the PDI photocatalyst demonstrated an obvious red shift and, therefore, enhancement of

255

visible-light absorption. The band gap energy of the PDI material was calculated to be 1.73 eV (see inset of Fig. S2f in the

256

SI). Overall, the structure, morphology, and absorption spectrum of the PDI supramolecular nanofibers demonstrate

257

successful preparation via the modified acidification polymerization protocol and highlight the potential application of

258

these novel materials under sunlight irradiation.

259 260

Photocatalytic degradation kinetics and mechanisms.

261 262

Degradation of fluoroquinolones by the PDI-sunlight photocatalytic system. The activity of the sunlight-activated

263

PDI nanofibers was evaluated for the degradation of four fluoroquinolone antibiotics. Fig. 2a shows the observed time-

264

based rate constant for ciprofloxacin, enrofloxacin, norfloxacin, and ofloxacin degradation by the PDI-sunlight

265

photocatalysis system. The rate constants for ciprofloxacin, enrofloxacin, norfloxacin, and ofloxacin were (9.4 ± 0.8)×10-

266

2,

267

fluoroquinolones were rapidly transformed by the PDI nanofiber photocatalyst. The fastest degradation was obtained for

268

ofloxacin, and this finding was attributed to the chemical structure as discussed below in the degradation pathways

269

section. As ofloxacin showed the fastest degradation, further testing was completed with this fluoroquinolone antibiotic to

(12.6 ± 0.6)×10-2, (7.0 ± 0.7)×10-2, and (33.4 ± 2.0)×10-2 min-1, respectively. These results indicated that all four



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270

understand the reaction kinetics, mechanisms, and products; however, we note that similar mechanisms are expected for

271

the other compounds in this structurally-similar class of antibiotics.

272

(a)

H N

OFL

(b)

O N

N

F O

O

NOR

C/C0 for OFL

H2N N

N

OH

F O

O

H N

ENR

N

N OH

F O

O

H2N N

CIP

Dark

O

0.2

0.3 -1

kobs (min )

0.4

0.5

Sunlight

0.4

PDI-adsorption Photolysis PDI-photocatalysis

0.0

-30 -1

0

1

2

Absorbance (a.u.)

0.25

4

Time (min)

5

6

7

0.00

(c) 0.35 0.30

3

0.05 Blue light

UV light

Green light

Yellow light

Red light

0.10 0.15

kobs (min-1)

273

0.6

OH

F

0.1

0.8

0.2

N

O

0.0

1.0

OH

0.20

0.20

0.25

0.15

0.30 0.10

0.35

0.05

0.40

0.00

0.45

350

400

450

500

550

Wavelength (nm)

600

650

274 275

Figure 2. (a) Rate constants for photocatalytic degradation of four fluoroquinolones (e.g., OFL, ofloxacin; NOR, norfloxacin;

276

ENR, enrofloxacin; CIP, ciprofloxacin) by sunlight-activated PDI; (b) Degradation of ofloxacin by photolysis, PDI adsorption,

277

and PDI photocatalysis under sunlight irradiation; and, (c) The UV-vis absorption spectra of PDI supramolecular nanofibers

278

overlaid with the rate constants for ofloxacin degradation by the PDI photocatalyst. In all cases, the initial concentration of

279

fluoroquinolone antibiotic was 8 mg L-1, the concentration of PDI nanofibers was 200 mg L-1, the solution pH was 5.6, and error

280

bars indicate one standard deviation.

281 282

The contributions of hydrolysis, direct photolysis, adsorption, and photocatalysis to the apparent degradation of ofloxacin ACS Paragon Plus Environment

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are presented in Fig. 2b. No change in ofloxacin concentration was recorded in the absence of PDI photocatalysts and

284

light. A 6% change in the ofloxacin concentration was observed over the 7-min irradiation period in the absence of PDI

285

nanofibers, suggesting that photolysis was not a major degradation mechanism for the experimental conditions. The

286

adsorption and photocatalytic processes were carried out following the methods described in the photocatalysis

287

experiments section. After 30 min adsorption under dark conditions, the 200 mg L-1 of PDI nanofibers adsorbed

288

approximately 15% of the ofloxacin (C0 = 8 mg L-1), indicating an adsorption capacity of 40 mg g-1 for the experimental

289

conditions. With sunlight irradiation, the PDI supramolecular nanofibers enhanced the ofloxacin degradation to ~91%,

290

verifying the ability of the PDI material to rapidly catalyze ofloxacin degradation using solar energy. Importantly, the

291

observed rate constants for ofloxacin degradation by PDI are one to three orders of magnitude higher than those observed

292

in other photocatalytic systems,39-44 suggesting that this material may have unique advantages over previously reported

293

catalysts for micropollutant treatment.

294 295

To further examine the photocatalytic activity of the PDI supramolecular nanofibers, different light sources were

296

employed. The experimental details were included in the photocatalysis experiments section, and the results are displayed

297

in Fig. 2c and Table S2 of the SI. The UV-visible absorption spectra for the PDI materials demonstrated broad absorbance

298

of light at 350-650 nm, with the highest absorption at 470 nm. These results indicate that the PDI catalyst absorbs not only

299

UV-A light, but also visible light. Given the absorption maximum at ~415-550 nm, the PDI catalyst was expected to

300

demonstrate optimal activity under blue light. This hypothesis was confirmed. In fact, all light sources showed effective

301

ofloxacin degradation. The time-based rate constants for ofloxacin degradation in the PDI-UV light, -blue light, -green

302

light, -yellow light, -red light, -white light, and -sunlight photocatalytic systems were 0.223±0.014, 0.403±0.029,

303

0.271±0.027, 0.174±0.007, 0.150±0.012, 0.295±0.009, and 0.334±0.020 min-1, respectively. From these results, the PDI

304

nanofibers exhibited the highest photocatalytic activity for blue light, followed by sunlight and, then, white light. The

305

photocatalysis results were in general agreement with the UV-visible absorption spectra, highlighting the benefits of the

306

broad absorbance spectrum of the PDI nanofibers prepared through the acidification polymerization protocol. Importantly,

307

the PDI-sunlight photocatalytic system yielded a degradation half-life of 2.08 min for ofloxacin. The stability of the PDI

308

supramolecular nanofibers was tested to ensure that this performance was reproducible over sequential treatment cycles.

309

As shown in Fig. S3 of the SI, four sequential cycles showed similar ofloxacin degradation without any apparent change

310

in performance. In combination, these results highlight the potential advantages for implementation of the PDI-sunlight ACS Paragon Plus Environment


311

Page 14 of 28

photocatalytic system in real systems.

312 313

Role of reactive species in ofloxacin degradation. Reactive species generated from the PDI-sunlight photocatalytic system

314

were probed using the ESR spin-trapping technique. As shown in Fig. S4a-c of the SI, no ESR signals were recorded for

315

the dark condition, while three ESR signals were observed under sunlight irradiation. The intensity ratios for two of these

316

signals were 1:2:2:1 (Fig. S4a in the SI) and 1:1:1:1 (Fig. S4b in the SI), indicating the presence of the dimethyl pyridine

317

N-oxide (DMPO)-•OH and DMPO-O2•- adducts,45-46 respectively. These findings confirm the role of •OH and O2•- in the

318

photocatalytic system. In addition to the DMPO-•OH and DMPO-O2•- adducts, another signal (1:1:1 intensity ratio) was

319

detected with 4-oxo-2,2,6,6-tetramethylpiperidine (TMPO) spin-trapping ESR and attributed to the TMPO-1O2 adduct (Fig.

320

S4c in the SI).47,48 The signal intensity of the three adducts increased with irradiation time, suggesting their continuous

321

production by the PDI supramolecular nanofiber photocatalyst during irradiation.

322 323

Chemical quenchers were applied to the experimental solutions to further elucidate the contribution of specific reactive

324

species to ofloxacin degradation. The quenching agents for h+, •OH, •OH and 1O2, and O2•- were Na2C2O4, IPA, NaN3, and

325

TEMPL, respectively. Fig. S5a of the SI shows that the various quenching agents had different effects on ofloxacin

326

degradation by the sunlight-activated PDI nanofibers. In particular, the extents of ofloxacin degradation after 7 min of

327

irradiation for the blank control and solutions containing IPA (100 mM), NaN3 (10 mM), TEMPL (5 mM), and Na2C2O4

328

(100 mM) were approximately 91%, 88%, 78%, 69%, and 53%, respectively; furthermore, the observed rate constants for

329

ofloxacin degradation under the same conditions were 0.334±0.195, 0.302±0.013, 0.216±0.013, 0.171±0.010, and

330

0.120±0.011 min-1, respectively. These results indicate that h+, O2•-, and 1O2 played important roles in ofloxacin

331

degradation by the PDI-sunlight photocatalytic system. Using Eqs. 2-5, the relative contributions of •OH, 1O2, O2•-, and h+

332

to the overall ofloxacin degradation were calculated to be 9.5 ± 4.0 %, 16.4 ± 3.5 %, 48.8 ± 2.9 %, and 64.1 ± 3.4 %,

333

respectively (Fig. S5b in the SI). Note that the contributions sum to more than 100% due to the convoluted radical

334

chemistry involved in the photocatalytic mechanism.39,49,50 Nevertheless, h+ production by the PDI-sunlight photocatalytic

335

system was found to be the major mechanism of ofloxacin degradation. The detailed mechanism for ofloxacin degradation

336

is discussed further in the photocatalytic mechanism section. Table S3 of the SI reports the contributions of •OH, 1O2, O2•-,

337

and h+ to the degradation of ciprofloxacin, norfloxacin, and enrofloxacin. The results followed the same general trend as

338

observed for ofloxacin, although the contribution of O2•- was slightly lower than that of 1O2 for enrofloxacin. For all ACS Paragon Plus Environment

Page 15 of 28

339


fluoroquinolones, h+ exerted the largest impact on the overall degradation kinetics.

340 341

Degradation pathways of ofloxacin. For the 7-min irradiation experiments, the overall ofloxacin degradation was ~91%

342

but the change in total organic carbon (TOC) was just 20%, indicating a relatively low degree of mineralization for this

343

level of treatment (Fig. S6). Instead, ofloxacin degradation in the PDI-sunlight photocatalytic system produced a variety of

344

transformation products that retained structural characteristics of the ofloxacin molecule. The analytical response of these

345

products tended to, first, increase with irradiation time and, then, decrease upon further irradiation. To elucidate the

346

photocatalytic mechanisms associated with ofloxacin degradation, transformation products were identified by mixing

347

samples collected at 3, 7, and 10 min of irradiation for UPLC-Q-TOF MS analysis, which confirmed 13 major

348

transformation products. The total ion current diagram and MS/MS results are shown in Fig. S7, Table S4, and Fig. S8 of

349

the SI. Importantly, the majority of these products (i.e., P2, P4, P6, P7, P8, P9, P10, P11, P12, and P13) retain the

350

fluoroquinolone pharmacophore (i.e., the basic chemical structure needed to convey antibiotic properties) and, therefore,

351

the ability to inhibit bacterial growth when present at high enough concentrations.9 Further efforts to identify the relative

352

potency of individual transformation products will help to elucidate the extent of treatment needed to optimize operating

353

conditions and ensure removal of antimicrobial activity.

354 355

To examine the mechanisms that produced the 13 identified transformation products during photocatalysis of ofloxacin,

356

the FED method was used to predict the sites of reactive species attack. The results of this analysis are summarized in

357

Table S5 of the SI. According to frontline orbital theory: a high FED2HOMO+FED2LUMO value indicates that an atom is more

358

likely to be attacked by •OH (electrophilic reaction); a high, positive point charge indicates the likelihood of O2•- attack

359

(nucleophilic reaction) and H-abstraction; and, a high 2FED2HOMO value suggests that an atom is likely to be attacked by

360

h+.50,51 The C10 and N12 positions of the ofloxacin molecule displayed high FED2HOMO+FED2LUMO values, demonstrating

361

the potential for •OH addition in the quinolone and N-methylpiperazine rings. The C3, C7, C10, C14, C23, and C29

362

positions exhibited high, positive point charges, and so these positions are expected to undergo preferential attack by O2•-;

363

furthermore, the C10, C14, and C29 atoms may undergo H-abstraction reactions. High 2FED2HOMO values were calculated

364

for the C2, N12, and N19 positions, suggesting that these sites may be directly oxidized by h+. These analyses were used

365

with the identified transformation products to propose a tentative ofloxacin photocatalytic pathway for the PDI-sunlight

366

system (see Fig. 3). ACS Paragon Plus Environment


Page 16 of 28

367 HN

O N

Attack N19 hVB+

Pathway I

-CH2

Attack N12

OFL O

O

O

O

-H2

O N

N

+

O

OH

F

O

P10 O

N O

OH

F

O

P2

O

NH

O

P1

O

PDIs

N

O

H

N

Attack C29

F

N N

O

H

1

O2

Attack C14 (-2H)

O N

O O

P8

O

O

-CO

H2O

OH P11 O

OH

O

+ CO2

N

F

N

F

O P3

O HN

O

O OH

P13 O N

N

N

F

O

Pathway III

OH

O N

+O OH

O

O N

Sunlight

Pathway II

N H

N H

O

O N O

368 369

OH O

P4 HN

F

H2N

N

F

OH P9

OH

O

N

H2N

N

F

O

O HN

-CH2O

O

O N

OH P7

HN

O2·-

F

-C5H9N N

N

-CO

N

O N

HN

OH

F P6

O

O

OH

F P12 O

O

Figure 3. Proposed degradation pathway for ofloxacin transformation in the PDI-sunlight photocatalytic system.

370 371

The MS analysis demonstrated that ofloxacin and a majority of the transformation products (e.g., P2, P4, P6, P7 P8, P9,

372

P10, P11, and P12) contained an MS/MS fragment with m/z 261, suggesting that these compounds retained the quinolone

373

moiety. Further analysis of the degradation products indicated that the N-methylpiperazinyl functionality was the main

374

site of attack for reactive species. In particular, three main pathways were determined (see Fig. 3). Pathway I was the

375

demethylation reaction, which has been widely reported for photolysis of fluoroquinolones.9,52,53 The photogenerated h+

376

attacked N19, releasing the methyl group and producing P7. Then, P7 underwent dehydrogenation to form P2.54 The

377

generation of P4 and P10 was proposed to begin with the direct oxidation of P7, as an electron donor, by h+, a reaction

378

that produces the demethylated cation radical [OFL-CH3]•+.55,56 For heterogeneous photocatalysts, the presence of

379

dissolved oxygen influences the degradation mechanisms and products.57 For instance, dissolved oxygen can react with

380

conduction band e- to produce nucleophilic O2•-. From the FED analysis in Table S5 of the SI, O2•- was predicted to attack

381

the C3, C7, C10, C14, C23, and C29 positions. However, the retained m/z 261 fragment suggests that O2•- mainly attacked

382

the C14 position on the N-methylpiperazinyl group. In addition, O2•- may preferentially interact with [OFL-CH3]•+,

383

forming a zwitterion or a diradical. The detailed mechanism for O2•- reaction with the piperazine ring is shown in Eq. 11.51

384

However, the species produced in the third reaction step was unstable in water and decomposed into acetaldehyde. Similar ACS Paragon Plus Environment

Page 17 of 28


385

ring cleavage products were observed in previous reports.58,59 Under acidic conditions, the acetaldehyde products undergo

386

H+ addition and hydrolysis to generate the P4 and P10 isomers. In a separate reaction, h+ attacked the N12 position on

387

ofloxacin to form P9 by effectively releasing the N-methylpiperazine.

388 O2·-

HN

HN

HN

Open ring

N

N

N

O

389

O

O

H+

HN N

O

O

H2N

Hydrolysis

N

O

HN N

H2O

390

HN

O

P10 O

O

O

H2N

P4

(Eq. 11)

391 392

Pathway II was the ketone formation reaction. Since electron-withdrawing groups can weaken the stability of the

393

transition state, the H-abstraction reaction greatly affects the formation of degradation products.60 According to the FED

394

calculations, the H-abstraction reaction was most likely to occur at C10, C14, and C29. These reactions form unstable

395

molecules that are likely to quickly react with •OH to produce hydroxylated ofloxacin products (Eq. 12). Depending on

396

the atoms adjacent to the hydroxylated site, subsequent dehydrogenation generates a keto-ofloxacin product. These

397

reactions could possibly generate m/z 378 (C10), m/z 376 (C14), or m/z 376 (C29). An m/z 378 product was not identified

398

in the UPLC-Q-TOF MS analysis; therefore, C10 was ruled out as an attack site. Only P13 (m/z 376) exhibited the

399

characteristics of the ketone formation reaction, namely the H-abstraction (-1 H), hydroxyl addition (+1 OH), and

400

dehydrogenation (-2 H) to yield a +14 change in m/z relative to ofloxacin. These results indicate that H-abstraction

401

occurred on C14 or C29. The fragmentation pattern from UPLC-Q-TOF MS analysis did not include the quinolone

402

structure (m/z 261); therefore, the H-abstraction reaction likely occurred at the C29 position according to the mechanism

403

shown in Eq. 12.

404 N

O N

N N

H OH Attack C29

F

405

OFL O

O

N

O



O N

 OH

F O

O

N

O N

2H

N

O

O N

OH

F

406

N

N OH

F P13 O

O

O

(Eq. 12)

407 408

Pathway III was the aldehyde reaction. In this case, 1O2 attacked the C14 position in ofloxacin to form endoperoxide (see

409

Eq. 13), which decomposed to form acetaldehyde groups in a similar manner as that reported by Wahab et al. for pACS Paragon Plus Environment


Page 18 of 28

410

chlorophenol.51 The resulting product is P8 in Fig. 6. As indicated in Eq. 13, the acetaldehyde groups in P8 were prone to

411

H+ attack and subsequent hydrolysis, resulting in formation of the P11 and P12 isomers. Further decomposition of the

412

aldehyde groups generated P6, which exhibited a strong response in the total ion current chromatogram (see Fig. S7 of the

413

SI). These mechanisms align with previous reports of photocatalysis of fluoroquinolone antibiotics by TiO2.54 Ultimately,

414

the identified transformation products from all three pathways were degraded into smaller, oxygen-containing compounds

415

(e.g., P1 and P3) before being mineralized into water and carbon dioxide.

416 N

1

O2

N N

N

417

OFL

O

O

O P8 O

H N

H+

N

Open ring

N

Hydrdysis

N

O

H2O

N H

O

N N

O

O P12

418

HN P11

N H

HN P6

(Eq. 13)

419 420

Photocatalytic mechanism governing the PDI-sunlight photocatalytic system. Based on the above results and discussion,

421

the mechanism shown in Fig. S10 of the SI (also shown in the graphical abstract) was proposed for ofloxacin degradation

422

in the PDI-sunlight photocatalytic system. An organic semiconductor supramolecular structure was created through self-

423

assembly of PDI molecules via acidification polymerization. The terminal carboxyl group aided in the formation of H-type

424

π-π stacking of supramolecular nanofibers. Under sunlight irradiation, the PDI was excited through absorption of UV-A and

425

visible light, causing the formation of photogenerated electrons and holes (see Eq. 14).

426 427

PDI + hv

PDI + h + + e -

428

(Eq. 14)

429 430

The PDI supramolecular structures demonstrate nanocrystalline, high-order H-type π-π stacking effects, and the electric

431

field caused by the carboxyl-rich side facilitated the transport of electrons to the surface, leading to the high separation of

432

the conduction band e- and valence band h+. Since the conduction band potential of PDI (-0.34 eV, calculated by Eq. 10

433

and Fig. S11 in the SI) was lower than the standard reduction potential of O2/O2•- (-0.33 eV), the conduction band e- can

434

be captured by dissolved oxygen to form O2•- (see Eq. 15).


Page 19 of 28


e - + O2

436

O•2 -

437

(Eq. 15)

438 439

Importantly, the long-range electron delocalization in PDI π-π stacking greatly improved the migration efficiency of

440

photogenerated e-, resulting in a high yield of O2•- at the catalyst-solution interface.31 The O2•- reacts with H+ (Eq. 16) to

441

produce 1O2 (Eq. 17) and •OH (Eq. 18).

442 443

O•2 - + H +

HO•2

444

(Eq. 16)

445 446

2HO•2

H 2O 2 + 1O 2

447

(Eq. 17)

448 449

H 2O 2 + e -

•

OH + OH -

450

(Eq. 18)

451 452

The h+ had a valence band potential of 1.39 eV, suggesting minimal reaction with water or HO- to form •OH. In this case,

453

the •OH measured by ESR likely derived from O2•--initiated reactions. Nevertheless, the reactive species quenching

454

experiments demonstrated the negligible role of •OH in the photocatalysis of ofloxacin. The main contribution to

455

ofloxacin transformation was attributed to direct reaction with h+.

456 457

Effects of water quality and interfering substances on ofloxacin photocatalysis. As shown in Fig. 4a, the observed rate

458

constant for ofloxacin degradation decreased from 0.632 ± 0.033 min-1 at pH 3.0 to 0.095 ± 0.008 min-1 at pH 8.0,

459

indicating that ofloxacin degradation was inhibited at higher pH in the PDI-sunlight photocatalytic system. The acid-base

460

speciation reactions for ofloxacin reveal that the cation, zwitterion, and anion dominate at pH less than 5.6, pH between

461

5.6 and 7.9, and pH above 7.9, respectively (Eqs. 6-9). Ofloxacin deprotonation from the cation to the zwitterion increases

462

the density of the electron cloud around the carboxylic acid group, making it more vulnerable to attack by reactive

463

species. This mechanism was confirmed with the g-C3N4-visible light system (see Fig. 4a); note that the preparation of ACS Paragon Plus Environment


Page 20 of 28

464

this material is reported in Text S5 of the SI. However, the PDI-sunlight photocatalytic system showed the opposite trend,

465

suggesting that the stimulatory effects of ofloxacin deprotonation were suppressed by another phenomenon. One possible

466

explanation may involve concurrent deprotonation of PDI nanofibers, which could cause depolymerization of the

467

supramolecular nanofibers with increasing pH and, thus, impede photocatalytic activity. However, the strong hydrogen

468

bond interactions between carboxyl groups of the supramolecular fibers were stable for the tested experimental

469

conditions, as indicated in a previous report.31 For this reason, the mechanism involved with pH effects requires further

470

investigation.

471

(a) 1.2

0.07 0.7 OFL+

OFL-

OFL+/-

0.06 0.6 0.05 0.5

PDI g-C3N4

0.6

0.04 0.4 0.03 0.3

0.4

0.02 0.2

0.2

0.3

kobs (min-1)

0.8

kobs (min-1)

Ionization fraction

1.0

(b) 0.4

0.2

0.1

0.01 0.1

0.0

(c)

Absorbance (a.u.)

472

2

4

6

8

pH

3.5

10

12

0.00 0.0

2.0

DOM (100 mg C L-1) Cl- (10 mg L-1) HCO3- (10 mg L-1)

Visible light absorption

1.5

Milli-Q water Tap water Pearl River water Wastewater effluent

1.0

Cl-

HCO3-

Milli-Q water Tap water Pearl River water Wastewater effluent

0.8 0.6 0.4 0.2

0.5 0.0 200

SO42- NO3 DOM

1.0

NO3- (10 mg L-1)

2.5

Control Cu2+

(d) 1.2

Cu2+ (5 mg L-1) SO42- (10 mg L-1)

3.0

0.0

C/C0 for OFL

0

300

400

500

600

700

0.0

0

1

2

3

4

5

6

7

8

473 474

Wavelength (nm) Time (min) Figure 4. (a) Ionization fractions (primary y-axis) and rate constants for ofloxacin (8 mg L-1) photocatalysis by the g-C3N4-

475

white light system (400-800 nm, 200 mg L-1, secondary y-axis, orange bar) and the PDI-sunlight photocatalytic system (200 mg

476

L-1, secondary y-axis, blue bar) as a function of solution pH; the curves through the rate constant data stem from a species-

477

specific kinetics model similar to that reported by Snowberger et al. (2016)9; (b) rate constants for ofloxacin degradation by

478

sunlight-activated PDI nanofibers in the presence of water quality constituents (e.g., [Cu2+] = 5 mg L-1, [DOM] = 100 mg C L-1,

479

concentrations of all other water quality constituents were 10 mg L-1); (c) UV-vis absorbance spectra of solutions containing the ACS Paragon Plus Environment

Page 21 of 28


480

water quality constituents included in (b); and, (d) ofloxacin (8 mg L-1) degradation in different water and wastewater matrices

481

by 200 mg L-1 of PDI nanofibers for sunlight irradiation at pH 5.6. All error bars indicate one standard deviation.

482 483

The effects of water quality constituents on the rate constant for ofloxacin photocatalysis are shown in Fig. 4b. In the

484

presence of SO42- and NO3-, the rate constant for ofloxacin degradation was not significantly affected; furthermore, only

485

minor differences were observed for Cl- and HCO3-. However, the Cu2+ cation significantly reduced the transformation

486

kinetics of ofloxacin, a finding that was attributed to Cu2+ scavenging of O2•- (𝑘"Cu2 + ,O•2 ― = 8.1×109 M−1s−1) and 1O2

487

(𝑘"Cu2 + ,1O2 = 1.2×109 M-1s-1).61,62 Martínez et al. reported the second-order rate constant for ofloxacin reaction with 1O2 to

488

be 5.6×106 M-1s-1.63 Then, for the experimental conditions used here, Cu2+ reacted with approximately 1.5× more 1O2 than

489

ofloxacin, resulting in slower ofloxacin degradation (kobs = 0.334 ± 0.195 min-1) compared to the control solution (kobs =

490

0.126 ± 0.007 min-1). The relative impacts of other metals on ofloxacin degradation can be investigated by comparing the

491

rate constants for metal ion reaction with 1O2 and O2•-. For example, the second-order rate constant for Fe3+ with O2•- is

492

2.4×106 M−1s−1, and rate constants for reaction of Fe3+-complexes with 1O2 are on the order of 109 M−1s−1.64,65 In this case,

493

competition effects from Fe3+ will be similar to Cu2+ for 1O2-based reactions and less than Cu2+ for O2•-. So, Fe3+ will exert

494

a lower impact than Cu2+ (at the same molar concentration). However, we note that iron can also (i) precipitate on the PDI

495

supramolecular photocatalyst to reduce activity or (ii) enable Fenton reactions in the presence of hydrogen peroxide to

496

increase activity. The impacts of environmentally-relevant metal concentrations are also inherent in the results described

497

below for ofloxacin degradation in real surface water and wastewater effluent. The presence of DOM inhibited the

498

photocatalytic degradation of ofloxacin, presumably due to competition for reaction with the reactive species and light

499

screening effects in the visible region (see Fig. 4c).

500 501

To evaluate the suitability of the PDI-sunlight photocatalytic system to degrade antibiotics under environmental

502

conditions, ofloxacin transformation was studied in different water matrices. The primary chemical properties associated

503

with known impacts on ofloxacin degradation kinetics (from Fig. 4b) are shown in Table S6 of the SI for the four water

504

sources. As indicated in Fig. 4d and Table S7 in the SI, the observed rate constants for ofloxacin degradation in ultrapure

505

water, tap water, Pearl River water, and wastewater effluent were 0.334 ± 0.195, 0.250 ± 0.018, 0.208 ± 0.014, and 0.224

506

± 0.009 min-1, respectively. Unsurprisingly, ultrapure water, which contains negligible background inorganic or organic

507

species, demonstrated the fastest ofloxacin transformation due to the absence of reactive species scavenging or light ACS Paragon Plus Environment


Page 22 of 28

508

screening. Ofloxacin degradation in tap water was slightly inhibited compared to ultrapure water, presumably due to the

509

low TOC, Cl-, and NO3- content that caused some scavenging effects. A greater extent of interference was observed in the

510

Pearl River water and wastewater effluent, for which the ofloxacin transformation decreased by 17.8% and 15.9%,

511

respectively, over the 7-min irradiation period. The TOC concentrations in the Pearl River water and wastewater effluent

512

were 1.75× and 1.09× higher than that of the tap water, increasing the scavenging of reactive species by DOM and

513

inhibiting ofloxacin degradation. The higher DOM content of the Pearl River water and wastewater effluent likely exerted

514

light-screening effects that further inhibited ofloxacin transformation. Nevertheless, the PDI-sunlight photocatalytic

515

system was successfully applied to ofloxacin treatment in tap water, surface water, and wastewater effluent with half-lives

516

of 2.77, 3.33, and 3.09 min, respectively. These results confirm the suitability of this novel system to ameliorate concerns

517

with antibiotics in real drinking water and wastewater matrices.

518 519

ACKNOWLEDGEMENTS.

520 521

This work was supported by the Major Science and Technology Program for Water Pollution Control and Treatment in

522

China (No. 2017ZX07202001), Program for Changjiang Scholars and Innovative Research Team in University, and China

523

Postdoctoral Science Foundation (Grant No. 2017M620796).

524 525

SUPPORTING INFORMATION.

526 527

Five sections of text describe chemical reagents, characterization methods, reactive species determination, ofloxacin analysis

528

and identification of transformation products, and preparation of g-C3N4. Seven tables provide fluoroquinolone analysis

529

parameters, effects of different light sources, contributions of reactive species, transformation product information, FED

530

calculations, water quality, and rate constants for ofloxacin degradation in different wastewater matrices. Eleven figures

531

show the characterization results from BET, FTIR, XPS, and DRS analyses, sequential photocatalysis experiments, ESR

532

spectra, quenching experiments, TOC, total ion current chromatogram, MS/MS fragmentation, atomic numbering,

533

proposed photocatalytic mechanism, and valence band XPS spectra.

534 535

REFERENCES ACS Paragon Plus Environment

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536 537 538 539 540 541

(1) Calamari, D.; Zuccato, E.; Castiglioni, S.; Bagnati, R.; Fanelli, R., Strategic survey of therapeutic drugs in the rivers Po and Lambro in northern Italy. Environ. Sci. Technol. 2003, 37, (7), 1241-1248. (2) He, K.; Blaney, L., Systematic optimization of an SPE with HPLC-FLD method for fluoroquinolone detection in wastewater. J. Hazard. Mater. 2015, 282, 96-105. (3) Fatta-Kassinos, D.; Vasquez, M. I.; Kummerer, K., Transformation products of pharmaceuticals in surface waters and

542

wastewater formed during photolysis and advanced oxidation processes - Degradation, elucidation of byproducts and

543

assessment of their biological potency. Chemosphere 2011, 85, (5), 693-709.

544

(4) Radjenovic, J.; Petrovic, M.; Barcelo, D., Fate and distribution of pharmaceuticals in wastewater and sewage sludge of the

545

conventional activated sludge (CAS) and advanced membrane bioreactor (MBR) treatment. Water Res. 2009, 43, (3), 831-

546

841.

547

(5) He, K.; Soares, A. D.; Adejumo, H.; McDiarmid, M.; Squibb, K.; Blaney, L., Detection of a wide variety of human and

548

veterinary fluoroquinolone antibiotics in municipal wastewater and wastewater-impacted surface water. J. Pharmaceut.

549

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