Photochemical behavior of antibiotics impacted by complexation effects of concomitant metals: a case for ciprofloxacin and Cu(II).

Environmental Science

View Article Online View Journal

Processes & Impacts Accepted Manuscript

This article can be cited before page numbers have been issued, to do this please use: X. Wei, J. Chen, Q. Xie, S. Zhang, Y. Li, Y. Zhang and H. Xie, Environ. Sci.: Processes Impacts, 2015, DOI: 10.1039/C5EM00204D.

This is an Accepted Manuscript, which has been through the Royal Society of Chemistry peer review process and has been accepted for publication. Accepted Manuscripts are published online shortly after acceptance, before technical editing, formatting and proof reading. Using this free service, authors can make their results available to the community, in citable form, before we publish the edited article. We will replace this Accepted Manuscript with the edited and formatted Advance Article as soon as it is available. You can find more information about Accepted Manuscripts in the Information for Authors. Please note that technical editing may introduce minor changes to the text and/or graphics, which may alter content. The journal’s standard Terms & Conditions and the Ethical guidelines still apply. In no event shall the Royal Society of Chemistry be held responsible for any errors or omissions in this Accepted Manuscript or any consequences arising from the use of any information it contains.

rsc.li/process-impacts

Page 1 of 21

Environmental Science: Processes & Impacts View Article Online

DOI: 10.1039/C5EM00204D

Table of contents entry

Cu(II) complexation altered the light absorption, direct photolytic pathways, 1O2 photo-generation ability, and the reactivity of H2CIP+ towards 1O2 by changing its molecular orbitals and atomic charge distribution.

Environmental Science: Processes & Impacts Accepted Manuscript

Published on 09 June 2015. Downloaded by North Dakota State University on 24/06/2015 14:00:24.

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Environmental Science: Processes & Impacts

Page 2 of 21 View Article Online

DOI: 10.1039/C5EM00204D

Environmental impact To support 10 billion people in 2050, aquaculture and livestock breeding could be an important strategy. To ensure the production, thousands of tons of antibiotics and metal trace elements are used as feed additives per year and their combined pollution becomes more and more prominent. The interaction between them may affect the environmental behavior and ecotoxicology of antibiotics, and increase the uncertainty of risk assessment. As many antibiotics are resistant to microbial-degradation, photodegradation becomes pivotal in determining their fate and ecological risk. Here we proved that metal can complex with antibiotics, and then alter their photolytic reactivity and pathways. Additionally, we enlighten a computational approach to predict the photochemical behavior for antibiotics in different aqueous forms and clarify the photolytic mechanisms.



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Page 3 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

1

Photochemical Behavior of Antibiotics Impacted by Complexation Effects

2

of Concomitant Metals: A Case for Ciprofloxacin and Cu(II)†

3 4

Xiaoxuan Wei,a Jingwen Chen,∗a Qing Xie,a Siyu Zhang,ab Yingjie Li,a Yifei Zhang,a and

5

Hongbin Xiea

6

a

7

Environmental Science and Technology, Dalian University of Technology, Dalian 116024,

8

China

9

b

10

Key Laboratory of Industrial Ecology and Environmental Engineering (MOE), School of

Key Laboratory of Pollution Ecology and Environmental Engineering, Institute of Applied

Ecology, Chinese Academy of Science, Shenyang 110016, China

11 12

Abstract

13

Many water bodies, especially those adjacent to aquaculture and livestock breeding areas,

14

are contaminated by both antibiotics and transition metals. However, the effects of the

15

interaction between antibiotics and transition metals on the environmental behavior and the

16

ecotoxicology of antibiotics are largely unknown. We hypothesized that antibiotics may

17

coordinately bind with metal ions, and this complexation may affect the environmental

18

photochemical behavior of antibiotics. We took ciprofloxacin (CIP) and Cu(II) as a case, and

19

employed simulated sunlight experiments and density functional theory calculations to

20

investigate the underlying reaction mechanisms. Results showed that monovalent cationic

21

ciprofloxacin (H2CIP+) that is predominant in the normal pH range (6~9) of surface waters, †

Electronic Supplementary Information (ESI) available: details of the total ion chromatograms and mass spectra of the identified products, the product yields, and the formation and degradation rate constants.

∗

Correspondence to: Jingwen Chen, e-mail: [email protected]; Phone: +86-411-8470 6269

a

Dalian University of Technology

b

Institute of Applied Ecology, Chinese Academy of Science 1


DOI: 10.1039/C5EM00204D




1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

22

can chelate with hydrated Cu(II) to form [Cu(H2CIP)(H2O)4]3+. Compared with H2CIP+,

23

[Cu(H2CIP)(H2O)4]3+ has different molecular orbitals, and atomic charge distribution. As a

24

result, [Cu(H2CIP)(H2O)4]3+ showed dissimilar light absorption properties, slower direct

25

photolytic rates, lower 1O2 generation ability and weaker reactivity towards 1O2. Due to the

26

Cu(II) complexation, the apparent photodegradation of H2CIP+ was inhibited, and the

27

photolytic pathways and product distribution were altered. This study implies that for accurate

28

ecological risk assessment of antibiotics under transition metal co-contamination conditions,

29

the effects of metal complexation should be considered.

30

Keywords: Photochemical behavior, Antibiotics, Metal complexation

31

Introduction

32

Antibiotics are of acute concern as they are pseudo-persistent in the aquatic environment1

33

and many of them can induce bacterial resistance even at environmental concentrations.2 In

34

recent years, the combined pollution of antibiotics and transition metals (Cu, Zn, Fe, etc.)

35

became evident, especially in natural waters adjacent to aquaculture and livestock breeding

36

areas where both trace elements and antibiotics are used as feed additives.3,4 For example, the

37

concentrations of some antibiotics and transition metals were up to 6.8 and 55.0 µg/L,

38

respectively, in coastal waters of the China Bohai bay due to marine aquaculture.5,6 The

39

molecular structures of most antibiotics, such as fluoroquinolones and tetracyclines, contain

40

heteroatoms (e.g., O and N). Theoretically, these antibiotics can coordinately bind with metal

41

ions, leading to the formation of metal complexes.7,8 The complexation may alter the

42

physicochemical properties (e.g., octanol-water partition coefficient) and chemical reactivity

43

of antibiotics.9,10 As a result, the environmental fate and ecotoxicology of antibiotics are

44

altered.11−13 Thus, it is of importance to explore the effects of complexation on the

45

environmental behavior and ecotoxicology of antibiotics, for the purpose of ecological risk

46

assessment of antibiotics under the combined pollution with transition metals. 2


DOI: 10.1039/C5EM00204D

Page 5 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

47

Photochemical degradation has been proven to be a central factor in determining the

48

environmental fate of most antibiotics.14,15 In addition to direct photodegradation, some

49

antibiotics, such as fluoroquinolones and tetracyclines, can also photo-generate reactive

50

oxygen species (ROS, e.g., singlet oxygen 1O2 and hydroxyl radical ·OH), and be oxidized by

51

ROS produced by themselves (i.e. self-sensitized photooxidation) or by other water

52

constituents.16-18 Some previous studies have observed that metal ions could affect the

53

photolytic kinetics of antibiotics. For example, the photolytic rate constants (k) of tetracycline

54

at varied Mg(II) and Ca(II) concentrations relevant to the natural conditions were found to

55

vary by up to one order of magnitude.11 Clarithomycin and roxithromycin cannot directly

56

photodegrade in environmental waters as they do not absorb sunlight (λ > 290 nm),

57

nevertheless their Fe(III)-complexes undergo direct photolysis in sunlit waters.19

58

The mechanisms underlying the impact of metal complexation on the photodegradation of

59

antibiotics and other trace organic pollutants are largely unknown. We hypothesized that metal

60

complexation can have effects on the photochemical reactivity of antibiotics through the

61

following four ways: (1) As new chemical bonds are formed by complexation, the formed

62

metal complexes may have different light absorption characteristics; (2) The formed metal

63

complexes may have direct photolytic pathways that are different from the pathways of the

64

corresponding antibiotics; (3) The complexes and the corresponding antibiotics may have

65

different ROS generation abilities; and (4) The complexes and the corresponding antibiotics

66

may have different reactivity towards ROS.

67

In this study, we adopted ciprofloxacin (CIP, a widely used fluoroquinolone) and Cu(II) (a

68

common transition metal with strong complexing ability) as a case to verify the hypothesis. In

69

the normal pH range (6~9) of surface waters, monovalent cationic ciprofloxacin (H2CIP+) is

70

the predominant dissociation species of CIP.20 It can undergo both direct photolysis and

71

self-sensitized photooxidation via ROS (e.g., 1O2). Previous studies have observed that metals,

3


DOI: 10.1039/C5EM00204D




1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

72

e.g. Cu(II), Zn(II), Fe(III), can affect the photodegradation of fluoroquinolones.21,22 However,

73

the underlying mechanisms are not clear and need clarification.

74

We employed fluorescence spectrophotometric titration and infrared spectrometry to

75

examine the complexation between CIP and Cu(II). Simulated sunlight experiments were

76

performed to explore the effects of Cu(II) complexation on the photochemical behavior of CIP.

77

In addition, density functional theory (DFT) calculations were performed to provide

78

complementary insights into the impact of Cu(II) complexation on the photolysis of CIP.

79

Materials and methods

80

Chemicals and Reagents

81

Ciprofloxacin (CIP) with 98.0% purity was provided by Zhejiang Guobang Pharmaceutical

82

Co., Ltd. Deuterium oxide (D2O, 99.9%) and perinaphthenone (97% purity) were obtained

83

from Sigma-Aldrich. Furfuryl alcohol (FFA, 90% purity) was purchased from Shanghai

84

Jinshan Ting New Chemical Reagent Factory. Acetonitrile and trifluoroacetic acid were of

85

HPLC grade and purchased from Tedia Inc. Other reagents (purity > 99.0%) were purchased

86

from Kermel Chemical Reagent Co., Ltd. Ultrapure water (18 MΩ cm) was obtained from an

87

OKP ultrapure water system produced by Shanghai Lakecore Instrument Co., Ltd.

88

Characterization of the Complex

89

The complexation of CIP with Cu(II) in pH = 7.5 solutions was verified by a fluorescence

90

spectrophotometric titration technique. Fluorescence spectra were recorded using a Hitachi

91

F-4500 fluorescence spectrometer. Fluorescence intensity of the solutions at Ex/Em =

92

280/430 nm was used to calculate the complex ratio and the conditional stability constant (K'f)

93

on the basis of the Lineweaver-Burk equation13 (detailed in Text S1 of the Electronic

94

Supplementary Information, ESI).

95

To identify the complexation sites, the Cu(II)-CIP complex was prepared according to the

96

procedures reported by Zhang et al.13 (Text S2). IR spectra from 1% solid dispersions in KBr 4


DOI: 10.1039/C5EM00204D

Page 7 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

97

were recorded by Fourier transform infrared spectroscopy (IR-Prestige-21, Shimadzu, Japan).

98

Photodegradation Experiments

99

The photolytic experiments were performed with an XPA-7 merry-go-round photoreactor

100

(Xujiang Electromechanical Plant, China). A water-refrigerated 1000 W xenon lamp

101

surrounded by Pyrex filters was used to simulate the sunlight with λ > 290 nm. The irradiance

102

measured by a UV-365 radiometer (Photoelectric Instrument Factory of Beijing Normal

103

University, China) was 294 µW/cm2 in the center of the tubes. All the CIP solutions (5

104

µmol/L) were adjusted to pH = 7.5 with HCl/NaOH only, so as to avoid the possible effects of

105

buffers.23 pH values were measured with a Mettler Toledo S40-K pH-meter equipped with an

106

InLab®Expert Pro combined electrode. The pH values changed slightly (< 0.5 pH units)

107

during the photolytic experiments. The ionic strength was controlled with NaCl (1 mmol/L).

108

UV-vis absorption spectra of CIP (5 µmol/L) were recorded via a Hitachi U-2900

109

spectrophotometer using a 1 cm cuvette. As the absorbance in the effective wavelength range

110

(290~400 nm) is > 0.02, all the observed photolytic rate constants were corrected for

111

light-screening.20

112

Ethylene diamine tetraacetic acid (EDTA), a typical complexing agent (conditional stability

113

constant K'f,Cu(EDTA) = 1.0 × 1016 L/mol, pH = 7.5),24 was employed to assess the role of Cu(II)

114

complexation on the photodegradation of CIP. The direct photodegradation of CIP was

115

investigated in N2 saturated solutions. To examine the 1O2 generating ability of CIP irradiated

116

by simulated sunlight, FFA was used as a probe compound, and D2O was employed to

117

prolong the lifetime of 1O2.25 The steady-state concentration of 1O2 was calculated as follows:

118

[ 1O2 ] =

kFFA k1O2,FFA

(1)

119

where kFFA is the observed degradation rate constant for FFA, k1O2,FFA is the known reaction

120

rate constant for FFA with 1O2 (k1O2,FFA = 8.3 × 107 L mol−1s−1 in D2O),26 and [FFA]0 = 10

121

µmol/L. The reactivity of CIP towards 1O2 was assessed by a second-order reaction rate 5


DOI: 10.1039/C5EM00204D



DOI: 10.1039/C5EM00204D

1

122

constant, k1O2. For the determination of k1O2, perinaphthenone was used as the

123

photosensitizer, and FFA as a reference compound (k1O2,FFA = 1.2 × 108 L mol−1s−1 in H2O).27

124

The initial concentrations of perinaphthenone, FFA and CIP were 20 µmol/L and the

125

concentration of Cu(II) was 40 µmol/L. A 300 W Hg lamp surrounded by 380 nm cut-off

126

filters was employed as the light source, under which the direct photodegradation of CIP and

127

FFA can be neglected, and CIP cannot photogenerate 1O2 as these compounds have no light

128

absorption at these conditions. The k1O2 value of CIP was calculated with the following

129

equation:

130

k1O2,CIP =

ln([CIP]t / [CIP]0 ) k1O2,FFA ln([FFA]t / [FFA]0 )

O2

(2)

131

Analytical Determinations

132

The concentration of CIP was determined by an Agilent 1100 HPLC with a ZORBAX

133

SB-C18 column (2.1 mm × 150 mm, 3.5 µm). CIP was excited at λ = 280 nm and recorded at

134

λ = 445 nm with a fluorescence detector. The mobile phases were acetonitrile (A) and

135

trifluoroacetic acid in water (pH = 2.4) with the following gradient elution: 5% A (0 min) to

136

20% A (3 min) to 40% A (15 min) to 5% A (17−20 min). For FFA, the detection wavelength

137

was 230 nm with a diode array detector, and the mobile phase was made up of 10%

138

acetonitrile and 90% H2O. The flow rate was 0.2 mL/min, injection volume was 5 µL, and

139

column temperature was 30 °C.

140

Photolytic products were separated by an Agilent 1200 HPLC. The gradient eluting

141

condition of HPLC was: 2% A (0 min) to 12% A (3 min) to 20% A (28 min) to 38% A (43 min)

142

to 2% A (44−50 min). The other separation conditions were the same as in case of the

143

analytical method of CIP. The accurate masses of the products were determined by an Agilent

144

6224 TOF mass spectrometer, and the mass spectral fragmentation patterns were examined by

145

an Agilent 6410B triple quadrupole mass spectrometer. The mass spectrometer conditions

146

were: ionization mode: ESI, positive mode; scan range: m/z 50−1000; drying gas flow: 9 6



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

Page 9 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

147

L/min; drying gas temperature: 350 °C; fragmentor: 150 V; nebulizer pressure: 40 psi;

148

capillary voltage: 4000 V; skimmer voltage: 65 V; octapole radio frequency voltage: 750 V.

149

DFT Computation

150

All the calculations were performed with the Gaussian 09 program suite,28 using the M06-2X

151

hybrid meta exchange-correlation functional29 with the 6-31+G(d,p) basis set for C, N, O, F,

152

H, and the LanL2DZ pseudopotential30 and the basis set for Cu. The solvent effect of water

153

was considered by the integral equation formalism of the polarized continuum model

154

(IEFPCM) based on the self-consistent-reaction-field (SCRF) method.31 The possible

155

defluorination reactions were calculated for H2CIP+ (the dominant dissociation species of CIP

156

in pH = 7.5 solution) at the lowest excited triplet state (T),32,33 and for its Cu(II) complex at

157

the excited tripquartet state (4T, where the quartet superscript refers to the total spin of the

158

complex, T refers to the local multiplicity of H2CIP+).34,35 The excited-state geometries of

159

H2CIP+ and its Cu(II) complex were calculated by adjusting the spin multiplicity to 3 and 4,

160

respectively.36,37 Frequency calculations were performed to determine the character of

161

stationary points. Transition states (TS) were characterized with only one imaginary

162

vibrational frequency. Intrinsic reaction coordinate (IRC) analysis38 was executed to verify

163

that each TS uniquely connected the designated reactants with the products. Values of the

164

Gibbs free energy and the enthalpy were corrected by thermal energy at 298 K. Atomic

165

charges were evaluated by Mulliken charge analysis.39

166

Results and discussion

167

Complexation between Ciprofloxacin and Cu(II)

168

The fluorescence spectrophotometric titration results (Fig. S1) show that Cu(II) and CIP can

169

form a 1:1 complex with a conditional stability constant K'f,Cu(CIP) = 1.23 × 106 L/mol in pH =

170

7.5 solutions. As shown in Equation (3), the fraction of Cu(CIP) (αCu(CIP)) depends only on the

171

equilibrium concentration of the free Cu(II) ion ([Cu2+]). According to the previous studies of 7


DOI: 10.1039/C5EM00204D




1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

172

Cu(II) speciation in natural waters, [Cu2+] is in the range of 10−6~10−16 mol/L at the

173

complexation equilibrium with organic matters.40−42 When [Cu2+] = 1 µmol/L, αCu(CIP) is about

174

55% (Fig. S2). These findings indicate that Cu(II) can significantly affect the chemical

175

speciation of CIP in some water bodies.

αCu(CIP) =

176

' K f,Cu(CIP) [Cu 2+ ] [Cu(CIP)] = ' [CIP] + [Cu(CIP)] 1 + K f,Cu(CIP) [Cu 2+ ]

(3)

177

We observed pronounced differences in IR spectra for CIP alone and CIP with the

178

coexistence of Cu(II) (Fig. S3). From the IR spectrum of CIP and the spectrum of the

179

Cu(II)-CIP complex, the C=O stretching vibration of the carboxyl group (~1710 cm−1)

180

disappeared, and a shift to low wavenumbers (from 1610 cm−1 to 1585 cm−1) was observed in

181

the carbonyl group.43,44 These IR spectra indicate the participation of the carboxyl and

182

carbonyl groups in the complexation of CIP with Cu(II). Thus, it is inferred that the ketonic

183

and carboxylate oxygen atoms of CIP bind directly with Cu(II).

184

We also optimized the structure of the Cu(II)-CIP complex by DFT calculation, and found

185

that Cu(II) is bound to six O atoms, with two coming from the carbonyl and the carboxyl

186

group of CIP and four from water molecules (Fig. 1). Thus, H2CIP+ (the dominant

187

dissociation species of CIP in pH = 7.5 solution) can chelate with hydrated Cu(II) to form

188

[Cu(H2CIP)(H2O)4]3+.

189 190

Fig. 1 Complexing reaction between H2CIP+ and [Cu(H2O)6]2+

191 192

Effects of Cu(II) Complexation on the Apparent Photodegradation Kinetics in Aerated 8


DOI: 10.1039/C5EM00204D

Page 11 of 21


193

Solutions

194

In the dark control experiments, the loss of CIP [either in the presence or absence of Cu(II)

195

and/or EDTA] was < 3%. Thus, the degradation by other processes was negligible during the

196

photolytic experiments. Linear regression of ln(Ct/C0) vs time (t) showed that the

197

photodegradation of CIP follows pseudo-first order kinetics (r > 0.99, p < 0.05). The apparent

198

photolytic rate constants (k) of CIP under different conditions are shown in Fig. 2. With the

199

concentration of Cu(II) increasing from 0 to 20 µmol/L, the k values decreased from 1.21 ±

200

0.03 to 0.21 ± 0.02 h−1, indicating that Cu(II) significantly inhibits the photodegradation of

201

H2CIP+. However, with the coexistence of Cu(II) (10 µmol/L) and EDTA (20 µmol/L), the k

202

value is similar to that in pure water. Under this condition, Cu(II) mainly chelates with EDTA

203

(> 99.9%) instead of CIP, as K'f,Cu(EDTA) is about ten orders of magnitude higher than K'f,Cu(CIP).

204

Additionally, the k values are nearly the same for CIP dissolved in pure water and in the

205

solution with coexistence of EDTA, indicating that EDTA does not interfere with the apparent

206

photodegradation of H2CIP+. Thus, we found that Cu(II) affects the apparent

207

photodegradation of H2CIP+ mainly through its complexation with H2CIP+. CIP + Cu(II) CIP + EDTA CIP + Cu(II) + EDTA pH = 7.5

1.6 1.2 -1

k (h )


1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

0.8 0.4 0.0

208

0.0

2.5 5.0 10.0 15.0 20.0 [Cu(II)] (µmol/L)

209

Fig. 2 Effects of Cu(II) and EDTA on the apparent photolytic rate constants (k) of CIP in

210

aerated solutions ([CIP]0 = 5 µmol/L, [EDTA] = 20 µmol/L. The error bars represent the 95%

211

confidence interval, n = 3)

9


DOI: 10.1039/C5EM00204D




1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

212

Effects of Cu(II) Complexation on the Light Absorption Characteristics

213

The UV-vis absorbance spectra of CIP at various Cu(II) concentrations are shown in Fig. S4.

214

In the presence of Cu(II), the short wavelength absorption band (λmax ≈ 275 nm) of CIP

215

showed a red shift with its intensity increasing, and the broad absorption band (around

216

310~335 nm) displayed a blue shift.

217

We employed DFT to calculate the molecular orbital composition for the main absorptions

218

of H2CIP+ and [Cu(H2CIP)(H2O)4]3+ (oscillator strengths > 0.1). As shown in Table S1, the

219

lowest-lying absorption (λ = 322.3 nm) for H2CIP+ corresponds to the transition from the

220

highest occupied molecular orbital (HOMO) to the lowest unoccupied molecular orbital

221

(LUMO). For [Cu(H2CIP)(H2O)4]3+, in contrast, the lowest-lying absorption (λ = 315.5 nm)

222

corresponds to the transitions both from the second highest occupied molecular orbital

223

(HOMO−1) to HOMO and from HOMO to LUMO. The difference in the light absorption

224

between H2CIP+ and [Cu(H2CIP)(H2O)4]3+ can be further rationalized by inspection of their

225

relevant orbital structures. As shown in Fig. S5, the lowest-lying excitation for H2CIP+ is of a

226

π,π* nature. For [Cu(H2CIP)(H2O)4]3+, the excitation involves both charge transfer transition

227

from the ligand to the metal and d,d transition. Thus, the Cu(II) complexation can alter the

228

molecular orbital components of the excitation and orbital structures, which causes

229

[Cu(H2CIP)(H2O)4]3+ to show different light absorption characteristics from H2CIP+.

230

Effects of Cu(II) Complexation on the Direct Photolytic Pathways

231

As shown in Fig. 2 and Fig. 3, the direct photolytic rates (determined in N2-saturated solutions)

232

of CIP with and without Cu(II) are higher than the corresponding apparent photolytic rates

233

(determined in the aerated solutions) due to the removal of dissolved oxygen, an excited

234

triplet quencher. The direct photolytic rate constant (kd) of CIP was decreased in the presence

235

of Cu(II), indicating that Cu(II) can significantly inhibit the direct photodegradation of

236

H2CIP+. We identified two defluorination products (P288 and P330) formed in the direct

10


DOI: 10.1039/C5EM00204D

Page 13 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

237

photodegradation of CIP (Table S2 and Fig. S6). The evolution of the products in the absence

238

and presence of Cu(II) is depicted in Fig. 4. With the co-existence of Cu(II), the formation of

239

P288 was slightly suppressed, and the generation of P330 was significantly inhibited.

240

According to our previous study, P288 was generated by cleavage of the C−F bond, and P330

241

was formed by OH− addition.20 Thus, the Cu(II) complexation inhibited the direct

242

photodegradation of H2CIP+ mainly through lowering its reactivity towards OH−. As the

243

molecular structures of P288 and P330 (Fig. 4) contain carbonyl and carboxyl groups, P288

244

and P330 may also chelate with Cu(II), and the photodegradation of P288 and P330 could be

245

impacted by the concomitant Cu(II) as well.

9

CIP + EDTA CIP + Cu(II) + EDTA

CIP CIP + Cu(II) pH = 7.5

6

3

0

kd -1

(h )

246

1

13

[ O2] × 10 -1 (mol L )

-5

k1O2,CIP × 10 -1

-1

(L mol s )

247

Fig. 3 Effects of Cu(II) and EDTA on the direct photolytic rate constant (kd, determined in

248

N2-saturated solutions), steady concentration of 1O2 and bimolecular rate constants of CIP

249

with 1O2 (For the determination of kd and [1O2], [CIP]0 = 5 µmol/L, [Cu(II)] = [FFA]0 = 10

250

µmol/L, [EDTA] = 20 µmol/L. For the determination of k1O2,CIP, [CIP]0 = [FFA]0 = 20 µmol/L,

251

[Cu(II)] = 40 µmol/L, [EDTA] = 80 µmol/L. The error bars represent the 95% confidence

252

interval, n = 3)

253 254

We further employed the DFT calculation to simulate the C−F bond cleavage and OH−

255

addition reactions for H2CIP+ and [Cu(H2CIP)(H2O)4]3+, for which the results are detailed in

11


DOI: 10.1039/C5EM00204D



256

Text S3 and Table S3. The activation free energies (∆G‡) for the C−F bond cleavage of

257

H2CIP+ and [Cu(H2CIP)(H2O)4]3+ are 27.3 and 32.8 kcal/mol, respectively. As shown in Fig.

258

S7, there is no significant difference in the calculated C−F bond lengths and atomic charges

259

on F between the two species. Thus, the Cu(II) complexation has negligible effects on the

260

C−F bond cleavage of H2CIP+. For OH− addition, ∆G‡ of [Cu(H2CIP)(H2O)4]3+ (10.8 kcal/mol)

261

is about 2 times higher than that of H2CIP+ (Table S3). Since the OH− addition is a

262

nucleophilic attack, we calculated the atomic charges on C12 (qC12). As shown in Fig. S8, qC12

263

of [Cu(H2CIP)(H2O)4]3+ was about half of that of H2CIP+. Thus, we conclude that the lower

264

reactivity of [Cu(H2CIP)(H2O)4]3+ with OH− is to be ascribed to the charge rearrangement of

265

H2CIP+ caused by the complexation.

266 O

O

O

NH2

N H

OH

N

N

N

HN

P288 P288 (CIP + Cu(II))

400

O

HO

OH

500

Peak area


1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

P330 P330 (CIP)

P288 (CIP)

300 pH = 7.5

P288 (CIP + Cu(II))

200 P330 (CIP)

100 P330 (CIP + Cu(II))

0 0.0 267

0.2

0.4

0.6 0.8 (C0-Ct)/C0

1.0

268

Fig. 4 Molecular structures and evolution profiles for the photoproducts from direct

269

photolysis (determined in N2-saturated solutions) of CIP with and without Cu(II), where

270

[CIP]0 = 5 µmol/L, [Cu(II)] = 10 µmol/L, and (C0−Ct)/C0 indicates the proportion of CIP

271

consumed at time t

272 12


DOI: 10.1039/C5EM00204D

Page 15 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

273

Effects of Cu(II) Complexation on the 1O2 Reactions

274

As shown in Fig. 3, with the co-existence of Cu(II), both the steady concentration of 1O2

275

([1O2]) and the bimolecular rate constants of CIP with 1O2 (k1O2,CIP) were lower, implying that

276

[Cu(H2CIP)(H2O)4]3+ has a lower 1O2 generation ability and weaker reactivity towards 1O2. In

277

a previous study on phototoxic potential of norfloxacin, Martinez et al. also observed that the

278

1

279

and P263) were identified in the 1O2 oxidation reactions of CIP with and without Cu(II)

280

(Table S2 and Fig. S6). According to the molecular structures of these products, the 1O2

281

oxidation mainly occurs at the piperazine ring of CIP. As shown in Fig. 5, the presence of

282

Cu(II) enhanced the production of P306, and decreased the generation of P334. Thus, the

283

Cu(II) complexation can not only inhibit the reactivity of CIP towards 1O2, but also alter the

284

oxidation product distribution. It deserves mentioning that as the molecular structures of P306,

285

P334 and P263 (Fig. 5) contain carbonyl and carboxyl groups, these products may also

286

chelate with Cu(II), and the complexes may have different 1O2 reactivity compared with the

287

ligands.

O2 yields were depressed in the presence of Ca(II) and Mg(II).45 Three products (P306, P334

288

According to Leach et al.46 and Gollnick et al.47, 1O2 is an electrophile that can oxidize

289

unsaturated C=C bonds and amino N atoms. The attack of 1O2 is expected to primarily occur

290

at molecular sites with the highest electron density. Our DFT calculation results show that for

291

the two CIP species, the largest negative atomic charges are located at N15 and N18 of the

292

piperazine ring (Table S4). Previous DFT calculation studies also indicated that the N atoms

293

of levofloxacin and norfloxacin bear negative charges.48,49 The calculated atomic charges

294

imply that N15 and N18 are the preferred sites for attack of 1O2, and the identified products also

295

confirmed this preference. For H2CIP+, the charges on N15 and N18 are −0.63e and −0.61e,

296

respectively, indicating that N15 and N18 have similar reactivity towards 1O2. In the case of

297

[Cu(H2CIP)(H2O)4]3+, the charge on N15 (−0.68e) is more negative than that on N18 (−0.58e),

13


DOI: 10.1039/C5EM00204D



298

which implies that 1O2 may attack N15 more favorably than N18. The evolution of the two

299

products, P306 formed from the attack at N15 and P334 formed from the attack at N18 (Fig. 5),

300

supports the results suggested by the atomic charges. Thus, this study also found that the

301

Cu(II) complexation can alter the 1O2 oxidation reaction pathways of CIP due to the

302

rearrangement of atomic charges. O

1000

NH2

O

O

F

OH

N H

N HN

O

O

F

F

OH

N H

O P306 (CIP + Cu(II))

Peak area


1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

N

H2N

O OH

N

P263 (CIP + Cu(II))

P334 (CIP)

750 P306 (CIP + Cu(II)) pH = 7.5

P306 (CIP)

500

P263 (CIP + Cu(II)) P263 (CIP) P334 (CIP)

250

P334 (CIP + Cu(II))

0 0.0 303

0.2

0.4

0.6 0.8 (C0-Ct)/C0

1.0

304

Fig. 5 Molecular structures and evolution profiles for the 1O2 photooxidation products of CIP

305

with and without Cu(II), where [Perinaphthenone]0 = [CIP]0 = 20 µmol/L, [Cu(II)] = 40

306

µmol/L, and (C0−Ct)/C0 indicates the proportion of CIP consumed at time t

307

Implications

308

In natural waters, Cu(II) can not only complex with CIP, but also with the concomitant

309

dissolved organic matter (DOM).50 According to the studies of Ahmed et al.41 and Lofts et

310

al.42 on the copper-speciation in freshwaters over a wide range of metal-organic matter ratios,

311

the concentration of free bivalent copper ion ([Cu2+]) under equilibrium conditions is in the

312

range of 10−6~10−16 mol/L. As shown in Fig. 2, the k value in the solution with the total

14


DOI: 10.1039/C5EM00204D

Page 17 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

313

concentration of Cu(II) ([Cu(II)]) = 2.5 µmol/L, where [Cu2+] is 0.5 µmol/L and is within the

314

range of [Cu2+] in natural waters, is lower than that in pure water. Thus, the Cu(II)

315

complexation can inhibit the apparent photodegradation of H2CIP+ under the environmental

316

conditions.

317

This study also found that Cu(II) complexation can alter the light absorption, direct

318

photolytic pathways, 1O2 photo-generation ability, and the reactivity of H2CIP+ towards 1O2.

319

As the 1O2 photo-generation ability of H2CIP+ was inhibited by the complexation, it can be

320

inferred that the photosensitized toxicity51 of H2CIP+ may also be altered by Cu(II)

321

complexation. As the photolytic pathways and product distribution were also altered by the

322

complexation, it can be assumed that the photomodified toxicity51 of H2CIP+ may be altered

323

too. Certainly, in addition to the photosensitized and photomodified toxicity, the toxicity of

324

Cu(II), H2CIP+ and the photoproducts can also be impacted by the complexation. Zhang et al.

325

investigated the toxicity of CIP and oxytetracycline to Scenedesmus obliquus and Vibrio

326

fischeri, and found that in the presence of the transition metals Cu(II), Zn(II) and Cd(II), the

327

formed complexes are commonly more toxic than the corresponding metals or ligands.13 Thus,

328

the complexation of concomitant Cu(II) can alter the environmental risk of H2CIP+. For

329

accurate risk assessment of antibiotics in the presence of metals, it is of importance to

330

understand the effect of metal complexation on their environmental behavior and toxicology.

331

According to our previous studies,20 CIP may exhibit five dominant dissociation forms in

332

water. The different dissociation species of CIP can chelate with different metal ions to form

333

distinct complexes.52 At different pH conditions, the impacts of Cu(II) on the apparent

334

photolytic rate constants of CIP are dissimilar (Fig. S9), suggesting that the effects of Cu(II)

335

complexation on the photodegradation of CIP in its different dissociation forms are distinct.

336

As shown in Fig. S10 and Fig. S11, due to complexation mainly, Ca(II) and Fe(III) can also

337

inhibit the apparent photolysis of CIP in pH = 7.5 aerated solutions. The order of inhibition

15


DOI: 10.1039/C5EM00204D




1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

338

strength was Cu(II) > Fe(III) > Ca(II). Thus, it can be concluded that one specific metal ion

339

can have different complexation effects on the photochemical behavior of different organic

340

pollutants and their different dissociation species; and different metal ions have different

341

impacts on the photochemical behavior of a specific organic pollutant. More studies are

342

needed to understand the general effects of metal complexation on the photochemical

343

behavior of organic micropollutants. Given the huge and ever-increasing number of organic

344

pollutants,4 it is also necessary to develop computational models that can predict the aquatic

345

photochemical behavior of organic pollutants and their different dissociation and metal

346

complexation forms.

347

Acknowledgements

348

We thank Prof. Dr. Willie Peijnenburg (Leiden University) for improving the English

349

expression and suggestions. The study was supported by the National Basic Research

350

Program (2013CB430403) and National Natural Science Foundation (21137001, 21325729)

351

of China.

352

Notes and references

353

1

354

8211-8217.

355

2

356

Pathog., 2011, 7, e1002158.

357

3

358

Environ. Int., 2008, 34, 1215-1226.

359

4

360

Gunten, B. Wehrli, Science, 2006, 313, 1072-1077.

361

5

G.T. Ankley, B.W. Brooks, D.B. Huggett, J.P. Sumpter, Environ. Sci. Technol., 2007, 41,

E. Gullberg, S. Cao, O.G. Berg, C. Ilback, L. Sandegren, D. Hughes, D.I. Andersson, Plos.

A. Sapkota, A.R. Sapkota, M. Kucharski, J. Burke, S. McKenzie, P. Walker, R. Lawrence,

R.P. Schwarzenbach, B.I. Escher, K. Fenner, T.B. Hofstetter, C.A. Johnson, U. von

S.C. Zou, W.H. Xu, R.J. Zhang, J.H. Tang, Y.J. Chen, G. Zhang, Environ. Pollut., 2011,

16


DOI: 10.1039/C5EM00204D

Page 19 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

362

159, 2913-2920.

363

6

W. Meng, Y.W. Qin, B.H. Zheng, L. Zhang, J. Environ. Sci., 2008, 20, 814-819.

364

7

A. Serafin and A. Stanczak, Russ. J. Coord. Chem., 2009, 35, 81-95.

365

8

M.O. Schmitt and S. Schneider, Phys. Chem. Comm., 2000, 3, 42-55.

366

9

D.L. Ross, S.K. Elkinton, S.R. Knaub, C.M. Riley, Int. J. Pharm., 1993, 93, 131-138.

367

10 W.R. Chen and C.H. Huang, Environ. Sci. Technol., 2009, 43, 401-407.

368

11 J.J. Werner, W.A. Arnold, K. McNeill, Environ. Sci. Technol., 2006, 40, 7236-7241.

369

12 Z. Simon, B. Katja, U. Darko, V. Marjan, K. Albin, J. Pharmaceut. Biomed., 2010, 53,

370

655-659.

371

13 Y. Zhang, X.Y. Cai, X.M. Lang, X.L. Qiao, X.H. Li, J.W. Chen, Environ. Pollut., 2012,

372

166, 48-56.

373

14 Y. Li, J.F. Niu, W.L. Wang, Chemosphere, 2011, 85, 892-897.

374

15 L.K. Ge, J.W. Chen, J. Lin, X.Y. Cai, Environ. Sci. Technol., 2009, 43, 3101-3107.

375

16 S.W. Yan and W.H. Song, Environ. Sci.: Processes Impacts, 2014, 16, 697-720.

376

17 L.K. Ge, J.W. Chen, X.X. Wei, S.Y. Zhang, X.L. Qiao, X.Y. Cai, Q. Xie, Environ. Sci.

377

Technol., 2010, 44, 2400-2405.

378

18 J.K. Challis, M.L. Hanson, K.J. Friesen, C.S. Wong, Environ. Sci.: Processes Impacts,

379

2014, 16, 672-696.

380

19 D. Vione, J. Feitosa-Felizzola, C. Minero, S. Chiron, Water Res., 2009, 43, 1959-1967.

381

20 X.X. Wei, J.W. Chen, Q. Xie, S.Y. Zhang, L.K. Ge, X.L. Qiao, Environ. Sci. Technol.,

382

2013, 47, 4284-4290.

383

21 M.C. Cuquerella, A. Belvedere, A. Catalfo, M.A. Miranda, J.C. Scaiano, G. de Guidi, J.

384

Photoch. Photobio. B, 2010, 101, 295-303.

385

22 U. Hubicka, J. Krzek, B. Zuromska, M. Walczak, M. Zylewski, D. Pawlowski, Photoch.

386

Photobio. Sci., 2012, 11, 351-357.

17


DOI: 10.1039/C5EM00204D




1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

387

23 M. Mella, E. Fasani, A. Albini, Helv. Chim. Acta, 2001, 84, 2508-2519.

388

24 G.D. Christian, Analytical Chemistry, 6th ed, Wiley, New York, 2007.

389

25 F. Wilkinson, W.P. Helman, A.B. Ross, J. Phys. Chem. Ref. Data, 1995, 24, 663-1021.

390

26 D.E. Latch, B.L. Stender, J.L. Packer, W.A. Arnold, K. McNeill, Environ. Sci. Technol.,

391

2003, 37, 3342-3350.

392

27 W.R. Haag, J. Hoigne, E. Gassman, A.M. Braun, Chemosphere, 1984, 13, 631-640.

393

28 M.J. Frisch, G.W. Trucks, H.B. Schlegel, G.E. Scuseria, M.A. Robb, J.R. Cheeseman, G.

394

Scalmani, V. Barone, B. Mennucci, G.A. Petersson, H. Nakatsuji, M. Caricato, X. Li, H.P.

395

Hratchian, A.F. Izmaylov, J. Bloino, G. Zheng, J.L. Sonnenberg, M. Hada, M. Ehara, K.

396

Toyota, R. Fukuda, J. Hasegawa, M. Ishida, T. Nakajima, Y. Honda, O. Kitao, H. Nakai, T.

397

Vreven, J.A. Montgomery, Jr., J.E. Peralta, F. Ogliaro, M. Bearpark, J.J. Heyd, E. Brothers,

398

K.N. Kudin, V.N. Staroverov, R. Kobayashi, J. Normand, K. Raghavachari, A. Rendell, J.C.

399

Burant, S.S. Iyengar, J. Tomasi, M. Cossi, N. Rega, J.M. Millam, M. Klene, J.E. Knox, J.B.

400

Cross, V. Bakken, C. Adamo, J. Jaramillo, R. Gomperts, R.E. Stratmann, O. Yazyev, A.J.

401

Austin, R. Cammi, C. Pomelli, J.W. Ochterski, R.L. Martin, K. Morokuma, V.G. Zakrzewski,

402

G.A. Voth, P. Salvador, J.J. Dannenberg, S. Dapprich, A.D. Daniels, O. Farkas, J.B. Foresman,

403

J.V. Ortiz, J. Cioslowski, D.J. Fox, Gaussian 09, Revision A.02; Gaussian, Inc.: Wallingford,

404

CT (2009).

405

29 Y. Zhao and D.G. Truhlar, Theor. Chem. Acc., 2008, 120, 215-241.

406

30 P.J. Hay and W.R. Wadt, J. Chem. Phys., 1985, 82, 270-283.

407

31 J. Tomasi, B. Mennucci, R. Cammi, Chem. Rev., 2005, 105, 2999-3093.

408

32 S. Monti, S. Sortino, E. Fasani, A. Albini, Chem.-Eur. J., 2001, 7, 2185-2196.

409

33 A. Albini and S. Monti, Chem. Soc. Rev., 2003, 32, 238-250.

410

34 X.W. Yan and D. Holten, J. Phys. Chem.-Us., 1988, 92, 5982-5986.

411

35 M.H. Ha-Thi, N. Shafizadeh, L. Poisson, B. Soep, J. Phys. Chem. A, 2013, 117,

18


DOI: 10.1039/C5EM00204D

Page 21 of 21



1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60

412

8111-8118.

413

36 K. Nozaki, K. Takamori, Y. Nakatsugawa, T. Ohno, Inorg. Chem., 2006, 45, 6161-6178.

414

37 F. Wang and T. Ziegler, J. Chem. Phys., 2005, 122, 074109.

415

38 K. Fukui, Accounts Chem. Res., 1981, 14, 363-368.

416

39 R.S. Mulliken, J. Chem. Phys., 1955, 23, 2338-2342.

417

40 C. Fortin, Y. Couillard, B. Vigneault, P.G.C. Campbell, Aquat. Geochem., 2010, 16,

418

151-172.

419

41 I.A.M. Ahmed, J. Hamilton-Taylor, S. Lofts, J.C.L. Meeussen, C. Lin, H. Zhang, W.

420

Davison, Environ. Sci. Technol., 2013, 47, 1487-1495.

421

42 S. Lofts and E. Tipping, Environ. Chem., 2011, 8, 501-516.

422

43 J. Coates, Interpretation of infrared spectra, a practical approach, John Wiley and Sons,

423

New York, 2000.

424

44 P. Trivedi and D. Vasudevan, Environ. Sci. Technol., 2007, 41, 3153-3158.

425

45 L. Martinez, P. Bilski, C.F. Chignell, Photochem. Photobiol., 1996, 64, 911-917.

426

46 A.G. Leach and K.N. Houk, Chem. Commun., 2002, 1243-1255.

427

47 K. Gollnick and J.H.E. Lindner, Tetrahedron Lett., 1973, 14, 1903-1906.

428

48 B. De Witte, H. Van Langenhove, K. Hemelsoet, K. Demeestere, P. De Wispelaere, V.

429

Van Speybroeck, J. Dewulf, Chemosphere, 2009, 76, 683-689.

430

49 K.A.K. Musa and L.A. Eriksson, J. Phys. Chem. A, 2009, 113, 10803-10810.

431

50 K. Ndungu, Environ. Sci. Technol., 2012, 46, 7644-7652.

432

51 M. Grote, G. Schüürmann, R. Altenburger, Environ. Sci. Technol., 2005, 39, 4141-4149.

433

52 B. Urbaniak and Z.J. Kokot, Anal. Chim. Acta, 2009, 647, 54-59.

19


DOI: 10.1039/C5EM00204D

Degradation of ciprofloxacin and sulfamethoxazole by ferrous-activated persulfate: implications for remediation of groundwater contaminated by antibiotics.

Bioaccumulation of metals and metalloids in medicinal plant Ipomoea pes-caprae from areas impacted by tsunami.

A new method for investigating oxidation behavior of liquid metals.

Case of Impacted Urethral Calculus.

Management of multiple impacted teeth: a case report and review.

Complexation behavior of oppositely charged polyelectrolytes: Effect of charge distribution.

Combined toxic effects of heavy metals and antibiotics on a Pseudomonas fluorescens strain ZY2 isolated from swine wastewater.

A case of impacted central incisor due to dentigerous cyst associated with impacted compound odontome.

A Case of Excision of the Knee-Joint: And of Urethrotomy for Impacted Calculus: Recovery.

The influence of the various central metals on photophysical and photochemical properties of benzothiazole-substituted phthalocyanines.

Photochemical behavior of sethoxydim in the presence of vegetable oils.

Ciprofloxacin does not inhibit mitochondrial functions but other antibiotics do.

Representation of phase equilibrium behavior of antibiotics.

Industrial air pollution by metals monitored by a synthetic fabric: A case study of a community.

Effects of trazodone on night eating behavior: a case report.

Inhibition of multiplication of tobacco mosaic virus in protoplasts by antibiotics and its prevention by divalent metals.

An unusual case of deciduous impacted molar.

Photochemical cis-trans isomerization of furylacryloylpeptides and their different kinetic behavior as substrates for carboxypeptidase Y.

Chemical behavior of Cu, Zn, Cd, and Pb in a eutrophic reservoir: speciation and complexation capacity.

A rare case of impacted maxillary first premolar.

Sources and fates of heavy metals in a mining-impacted stream: temporal variability and the role of iron oxides.

Clinical and economic evaluation of oral ciprofloxacin after an abbreviated course of intravenous antibiotics.

Ameloblastic fibroma and impacted mandibular first molar. A case report.

Cyclic peptide marine metabolites and CuII.