Recyclable Capture and Destruction of Aqueous Micropollutants

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Environmental Science & Technology

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Recyclable capture and destruction of aqueous micropollutants using

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the molecule-specific cavity of cyclodextrin polymer coupled with

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KMnO4 oxidation

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Xiyun Cai,†, * Qingquan Liu,† Chunlong Xia,†, ‡ Danna Shan,† Juan Du,† and Jingwen

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Chen†

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†

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Education), School of Environmental Science and Technology, Dalian University of

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Technology, Dalian 116024, China

9

‡

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Key Laboratory of Industrial Ecology and Environmental Engineering (Ministry of

Current address: Fushun Branch of Liaoning Province Hydrology and Water

Resources Investigation Bureau, Fushun 113008, China

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1

ACS Paragon Plus Environment


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ABSTRACT

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The removal of aqueous micropollutants remains challenging because of the

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interference of natural water constituents that are typically 3-9 orders of magnitude

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more concentrated. Cyclodextrins, which feature molecular recognition and are

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widely applied in separation and catalysis, are promising materials in the development

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of pollutant treatment technologies. Here, we described the facile integration of

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cyclodextrin polymer (CDP) adsorption and KMnO4 oxidation for recyclable capture

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and destruction of aqueous micropollutants (i.e., antibiotics and TBBPA). CDP

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exhibited adsorption efficiencies of 0.81-88% and 0.81-94% toward 14 pollutants at

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50.0 ng/L and 50.0 µg/L, respectively, at a solid-to-liquid ratio of 1:1250. The

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presence of simulated or natural water constituents (e.g., Mg2+, Ca2+, DOC, and a

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combination thereof) did not decrease the adsorption potential of CDP toward these

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pollutants because the pollutants, based on molecular specificity, were entrapped in

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the CD cavity. Subsequent KMnO4 oxidation completely degraded the retained

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pollutants, demonstrating that the pollutants could be broken down in the cavity.

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Pristine CDP was rearranged into the structurally loose composites that featured a

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porous CDP architecture with uniform embedment of δ-MnO2 nanoparticles and

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different adsorption efficiencies. δ-MnO2 loading was a linear function of the number

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of times the integrated procedure was repeated, underlying the accurate control of

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CDP recycling. Thus, this approach may represent a new method for the removal of

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aqueous micropollutants.

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INTRODUCTION

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Micropollutants are ubiquitously detected at low levels (ng/L-µg/L) in aquatic

36

systems, comprising thousands of synthetic and geogenic compounds and their

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transformation products.1,2 Many micropollutants raise considerable concerns, as they

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are continuously released into the environment,3 form problematic products,4-7 or

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cause largely long-term adverse effects.8 Examples of this category include antibiotics

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and bisphenol flame retardants,1,2 which cause the induction and spread of antibiotic

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resistance genes and endocrine disruption effects, respectively.

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A variety of unit processes (e.g., adsorption and oxidation) and combinations of

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these processes have been developed to mitigate organic pollutants (mg/L-g/L) in

44

wastewater.1,2 However, they often fail to remove aqueous micropollutants due to the

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interference of natural water components that are 3 to 9 orders of magnitude more

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concentrated.9-12 In particular, in the adsorption unit, the adsorption capacity of

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adsorbents decreases over time due to the accumulation of natural organic matter, and

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further treatment strategies are needed to regenerate adsorbents and degrade

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concentrated pollutants.1 For commonly used oxidation-based treatments, which are

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powerful and versatile, both the target and non-target compounds can decompose,

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which consumes a large volume of oxidants and may form toxic by-products.13,14

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Worse still, natural organic matter may bind pollutants and alter their reaction

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pathway with the formation of unexpected products.9,15-17 Therefore, specificity-based

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approaches to mitigate aqueous micropollutants are highly desirable.1,2

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The host-guest chemistry of cyclodextrins (CDs) is widely recognized to be specific

56

at the molecular level;18 as a result, CDs are applied in the design of adsorbents and

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catalysts.19-23 CDs, derived from starch, are composed of 6-8 α-1,4-linked glucose

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units and have a toroidal shape with a hydrophobic inner cavity and hydrophilic

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external surface. The cavity reversibly encapsulates size-matched guest compounds

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via coordination of various weak interactions (e.g., hydrophobic interactions,

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hydrogen bonds, and steric effects) and features molecular recognition.24-27 Thus, CDs

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are recognized as typical host compounds. Interestingly, CDs that are commonly

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water soluble in a monomer form can be used as functional units to construct

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insoluble materials by various methods (e.g., polymerization), without loss of

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molecular recognition.20,21 These materials have been proven to efficiently adsorb and

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separate organic pollutants (e.g., POPs,19 pesticides,20 dyes,28 and pharmaceuticals29,30)

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at high levels in wastewater and natural water. In some pioneering studies, CDs have

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been reported to have the unique ability to act as molecular containers that mediate

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oxidation and photochemical reactions of organic compounds.31-34 This behavior is

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attributable to cage effects and cavity encircling effects that influence the reaction

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potential and accessibility of compounds entrapped in the cavity, respectively.

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Considering the profound success of CDs in catalysis and separation, the

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integration of CD-containing materials and chemical technologies may provide an

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opportunity for the removal of aqueous micropollutants. However, the application of

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CDs in this field remains in its infancy.35 Here, we reported on the recyclable capture

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and destruction of aqueous micropollutants (including 13 antibiotics and 1 flame

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retardant, Table S1) by combining cyclodextrin polymer (CDP) adsorption and

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KMnO4 oxidation. KMnO4 is a versatile, environmentally friendly oxidant used in

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water decontamination.36-41 Pollutants were spiked at two environmentally relevant

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levels (i.e., 50.0 ng/L and 50.0 µg/L).42-45 Pollutant removal was investigated in the

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presence of simulated water constituents or in river water. Degradation processes of

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pollutants in the CD cavity were revealed using the flame-retardant TBBPA as a

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model compound, and changes in the structure of CDP were probed.

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Experimental Section

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Reagents and Materials. Tetrabromobisphenol A (TBBPA, 98+% purity) was

86

purchased from Aladdin Reagent Database Inc. (Shanghai, China). Antibiotics (98+%

87

purity) were purchased from Dr. Ehrenstorfer GmbH (Augsburg, Germany), including

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sulfathiazole, sulfapyridine, sulfadiazine, sulfachloropyridazine, sulfadimethoxine,

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sulfamerazine, sulfadimidine, sulfamethoxazole, enoxacin, enrofloxacin, lincomycin,

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penicillin G, and rifampin. β-Cyclodextrin (CD, 99+% purity) and epichlorohydrin

91

(EPI, AR) were obtained from Bodi Chemical Co., Ltd. (Tianjin, China). All organic

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solvents were of HPLC grade (Tedia). Other chemicals were provided by Sinopharm

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Chemical Reagent Co., Ltd. (Shanghai, China). A hydrogel-like cyclodextrin polymer

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(CDP) was prepared via the cross-linking reaction between CD and EPI.20 Raw CDP

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was washed successively with methanol, deionized water, 0.2-mol/L HCl, and

96

deionized water. The polymer was oven-dried at 37 °C and sieved, and 80-120-mesh

97

portions were collected and used in this study.

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Adsorption of aqueous micropollutants by CDP. Batch adsorption experiments

99

were conducted in 40-mL glass centrifuge tubes. Adsorbents (CDP, 20.0 mg or 200.0

100

mg) were added to tubes containing 25 mL of 50.0-µg/L pollutants (pH 7.0) in the

101

absence or presence of water constituents. The pH was adjusted with dilute HCl and

102

NaOH solutions. All tubes were shaken at 180 rpm and at 25 °C for 10 h to ensure

103

adsorption equilibrium. Solutions were withdrawn in 3.0-mL aliquots and filtered

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through a 0.45-µm Millipore membrane. Pollutants in the filtrate were analyzed on an

105

HPLC-MS/MS system (Table S2) to calculate adsorption distribution coefficient and

106

adsorption efficiency (i.e., adsorption removal rate).

107

Destruction of micropollutants by KMnO4. For the homogeneous system,

108

50.0-µg/L pollutants and 20.0-µmol/L KMnO4 were used. As the oxidant was added to

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initiate the reaction, 1.0-mL aliquots of reaction solutions were periodically

110

withdrawn and transferred into 2.0-mL HPLC vials containing 25.0 µL of 20-mmol/L

111

L-ascorbic acid. The vials were immediately vortexed, and the reaction was

112

terminated. Pollutant residuals were measured on the HPLC-MS/MS system. For the

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heterogeneous system, CDP (20.0 or 200.0 mg) was added to tubes containing 25 mL

114

of 50.0-µg/L pollutants (pH 7.0). All tubes were shaken at 180 rpm and at 25 °C for

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10 h. The adsorbents were filtered and transferred into 20 mL of 100-µmol/L KMnO4

116

to initiate the reaction. At certain time intervals, the reaction was terminated with 250

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µL of 200-mmol/L L-ascorbic acid, and then, the adsorbents were filtered and

118

ultrasonically washed three times with methanol. The washing solutions were

119

combined and evaporated to dryness under a gentle stream of nitrogen gas (N2), after

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which the residue was redissolved in 1.0 mL of methanol for the HPLC-MS/MS

121

measurement. The effects of KMnO4 levels (10.0 µmol/L-1.00 mmol/L) were

122

investigated under the same conditions.

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The removal of pollutants (50.0 ng/L) in deionized water and river water was also

124

investigated. In brief, CDP (400 mg) was added to 1000-mL Erlenmeyer flasks

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containing 500 mL of 50.0-ng/L pollutants. The mixture was shaken at 180 rpm at

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25 °C for 10 h. The adsorbents were separated by filtration and treated with 400 mL

127

of 100-µmol/L KMnO4 for 2 h. Then, the adsorbents were then filtered again and

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extracted three times with methanol. The extraction solutions were combined and

129

evaporated to dryness, and the residue was redissolved in 1.0 mL of ultrapure water

130

for the HPLC-MS/MS measurement. In parallel, the filtrates were acidified to pH 2.5

131

with concentrated HCl. Antibiotics and TBBPA in the supernatants were extracted and

132

concentrated using solid phase extraction (SPE) and liquid-liquid extraction (LLE),

133

respectively. SPE was performed using an Auto Trace 280 SPE system (Thermo,

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USA). Oasis HLB cartridges (500 mg) were preconditioned successively with 20 mL

135

of methanol, 6 mL of pure water, and 6 mL of HCl (pH 2.5). Acidified supernatants

136

were passed through cartridges at a flow rate of 10 mL/min. The cartridges were

137

washed with 10 mL of deionized water, and then dried under a flow of N2. Analytes

138

retained by the cartridges were eluted with 12 mL of methanol. The washing solutions

139

were evaporated to dryness, and the residue was redissolved in 1.0 mL of ultrapure

140

water. The LLE procedure was performed three times with mixed dichloromethane

141

and n-hexane (50/50, V/V). The combined extracts were evaporated to dryness, and

142

the residue was redissolved in 1.0 mL of methanol. All final samples were measured

143

on the HPLC-MS/MS system. Recoveries of micropollutants were compiled in Table

144

S3, and all data were corrected accordingly.

145

Destruction of TBBPA in the CD cavity by KMnO4. Experiments investigating the

146

destruction of TBBPA in the presence of CD monomer were conducted at 20±2 °C in

147

100-mL conical flasks containing 50-mL mixtures of TBBPA (1.00 mg/L) and CD (0,

148

2.0, 4.0, 8.0, and 10.0 mmol/L). The mixtures were magnetically stirred at 180 rpm

149

and incubated for 24 h to ensure inclusion equilibrium. After the reaction was initiated

150

using 10.0- or 20.0-µmol/L KMnO4, 1.0-mL aliquots of reaction solutions were

151

periodically withdrawn and transferred into 2.0-mL HPLC vials containing 25 µL of

152

10- or 20-mmol/L L-ascorbic acid. The vials were immediately vortexed to terminate

153

the reaction. TBBPA residual was measured on a Hitachi L2000 HPLC system

154

equipped with a Thermo Hypersil C-18 column (4.6×250 mm, 5 µm). The mobile

155

phase was composed of 60% acetonitrile and 40% water. The flow rate was set at 1.0

156

mL/min. The column oven was maintained at 30 °C. The detection wavelength was

157

set at 210 nm. The destruction of TBBPA retained on CDP was also investigated.

158

Adsorbents (CDP, 20.0 mg) were added to 20 mL of 1.00-mg/L TBBPA and incubated

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for 4 h, after which they were filtered and transferred into 20-mL solutions containing

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100 µmol/L KMnO4. At certain time intervals, 1.0-mL aliquots of reaction solutions

161

were withdrawn and mixed with 250 µL of 200-mmol/L L-ascorbic acid to terminate

162

the reaction. The adsorbents were filtered again and ultrasonically washed three times

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with methanol. The washing solutions were combined and evaporated to dryness, and

164

the residue was redissolved in 1.0 mL of methanol for the HPLC measurement.

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Additionally, the products of TBBPA and CD resulting from KMnO4 oxidation

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were identified. For TBBPA product identification in the homogeneous system, 100

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mL of TBBPA (1.00 mg/L) at pH 7.0 was oxidized in the absence or presence of CD

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(2.0 mmol/L) by 5.0-µmol/L KMnO4. At certain time intervals, the reaction was

169

quenched, and 20.0 mL of quenched solution was withdrawn and extracted with ethyl

170

acetate (20.0 mL×3). The combined extracts were evaporated to dryness, and the

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residue was redissolved in 1.0 mL of methanol. In the heterogeneous oxidation case,

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80.0 mg of CDP was added to 20 mL of 1.00-mg/L TBBPA (pH 7.0). The mixtures

173

were shaken at 180 rpm at 25 °C for 4 h. The adsorbents were filtered and transferred

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into 20 mL of 20-µmol/L KMnO4. At certain time intervals, the reaction was

175

terminated, and then, the adsorbents were filtered again and washed three times with

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methanol. Combined extracts were evaporated to dryness, and the residue was

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redissolved in 1.0 mL of methanol. The final samples from the two treatments were

178

analyzed on an Agilent 6224 Q-TOF LC/MS system, and mass spectra were recorded

179

in both the positive and negative ESI MRM modes. In the case of CD itself, CD (2.0

180

mmol/L) was oxidized by KMnO4 (2.0 mmol/L). The reaction was periodically

181

terminated,

182

α-cyano-4-hydroxycinnamic acid. The samples were analyzed on a Waters Micromass

183

MALDI micro MX-TOF MS system, and mass spectra were recorded in both the

and

1.0

mL

of

quenched

solution

was

mixed

with

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positive and negative ESI MRM modes to identify intermediates of CD.

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Integration of CDP adsorption and KMnO4 oxidation for TBBPA removal.

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Adsorbent (CDP, 2.00 g) was added to a 50-mL solution of 0.50-mg/L TBBPA, and

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the mixed system was shaken at 180 rpm at 25 °C for 4 h. The solution was filtered to

188

measure TBBPA residual. The adsorbents were filtered and transferred into 20 mL of

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0.50 mmol/L KMnO4. The reaction was terminated after 4 h. The adsorbents were

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withdrawn and washed three times with methanol. The combined extracts were

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evaporated to dryness. TBBPA residual was redissolved and measured. This procedure

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was repeated 20 times. After each cycle, both the manganese loading and adsorption

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efficiency of adsorbents and the removal of TBBPA were investigated.

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Surface characterization and measurements of CDP upon KMnO4 oxidation.

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To probe alterations of structure of CDP, we conducted KMnO4 treatments at high

196

levels (i.e., 5.0-200 mmol/L). Briefly, adsorbents (CDP, 5.0 g) were added to 100 mL

197

solutions of KMnO4 at different levels, and the mixed solutions were magnetically

198

stirred at room temperature. After 2 h, the reaction was terminated and insoluble

199

materials were filtered. The as-treated materials, denoted as CDP-x KMnO4 (where x

200

is the concentration of the oxidant in mmol/L), were washed with deionized water and

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oven dried at 70 °C for further surface characterization and measurements. The

202

morphology of materials in aqueous and solid state was recorded by an OLYMPUS

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TH4-2000 fluorescence inverse microscope and a JSM-5600LV scanning electron

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microscope equipped with an IE 300X energy dispersive spectroscopy, respectively.

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TEM images were obtained with a Tecnai G220 S-Twin transmission electron

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microscope that operated at 200 kV and had a dot resolution of 0.248 nm. Mass

207

fractions of C and H elements were obtained with a Vario EL III Element Analyzer.

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The BET surface area, pore volume, and BJH pore size distribution of materials were

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analyzed on a Quantachrome Quadrasorb-SI system by the N2 adsorption-desorption

210

method at 77 K. The particle size distribution of materials in water was measured on a

211

Malvern Mastersize 2000 Particle Analyzer. The manganese content of materials was

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measured on a Perkin Elmer Optima 2000DV ICP-AES with prior digestion of mixed

213

HNO3 and H2O2. The CD content of materials was measured using the phenol-sulfuric

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acid method. Swelling ratio was calculated as the relative weight gain of materials

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immersed in water at 25 °C for 24 h. FTIR analysis was performed on a Bruker

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EQUINOX55 spectrometer. XRD analysis was performed on a PANalytical Empyrean

217

diffractometer with Cu Kα radiation (λ ≈ 1.54 Å). For XPS analysis, survey scans

218

(0-1,400 eV) of materials and high-resolution scans of the C1s, O1s and Mn2p regions

219

were performed on a Thermo ESCALAB 250 system with a monochromatic Al Kα

220

X-ray source (1,486.7 eV) operating at 150 W. XPS Peak 4.0 software was used to

221

deconvolute core-level spectra. The binding energies of element core levels were

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determined relative to C1s (284.8 eV), and elemental concentrations were quantified

223

using instrument-specific sensitivity factors.

224

Data Analysis. All experiments were performed in triplicate. Statistical

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significance was determined using one-way ANOVA in Origin 8.0 (Microcal Software,

226

Inc., USA). Values with non-overlapping 95% confidence intervals were considered

227

significantly different.

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RESULTS AND DISCUSSION

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Selective capture of aqueous micropollutants by CDP. The kinetics of the

230

adsorption of aqueous micropollutants onto CDP was rapid and adsorption

231

equilibrium was typically reached within 4 h, following a pseudo-second order model

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(Fig. S1). The adsorption efficiencies of CDP toward micropollutants (50.0 µg/L) at a

233

solid-to-liquid ratio of 1:1250 demonstrated that 94% of TBBPA was adsorbed by 10


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CDP, whereas less than 30% of the other 13 pollutants were adsorbed (Fig. 1). As the

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solid-to-liquid ratio was increased to 1:125, the adsorption of the weakly adsorbed

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pollutants was enhanced by a factor of 2.4-12.6, and removal rates ranged from 12%

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(penicillin G) to 79% (enrofloxacin). The presence of Ca2+ (50 mg/L), Mg2+ (50

238

mg/L), DOM (10 mg/L), or a combination thereof, representative of typical natural

239

water constituents, did not diminish the adsorption efficiency of CDP (Fig. S1).

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Similar adsorption efficiencies were observed in deionized water and river water, even

241

when these pollutants were present at a low level of 50.0 ng/L (Fig. S2). These results

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indicate that simulated or natural water constituents don’t interfere with the adsorption

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of micropollutants on CDP. Similar phenomena have been reported for many

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pollutants (e.g., dyes, pesticides, and pharmaceuticals) in the mg/L-g/L range in

245

wastewater or natural water.19,20,28,30 Such specificity of CDP is commonly attributable

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to molecular recognition of the CD unit whose complexation with organic compounds

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(e.g., cyanotoxins) is typically regardless of natural organic matter and salinity27 and

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which is considered as the principal adsorption site of the polymer.20,21

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This molecular recognition appears to apply to the micropollutants in this study.

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Adsorbent CDP is enriched in CD units (approximately 72%), ensuring sufficient

251

cavities for micropollutants. A majority of these compounds have great potential to

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enter the cavity of the CD unit and form binary complexes, as indicated by the

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complexes’ stability constants (Table S1). For example, TBBPA, with an adsorption

254

efficiency of 87-95% (Figs. 1, S1 and S2), was tightly trapped in the CD cavity and

255

formed a highly stable complex with the stability constant of 2233 L/mol (Table S1,

256

Text S1). In particular, one of TBBPA’s hydroxyl benzene rings projected into the CD

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cavity from the cavity’s large rim, the other hydroxyl benzene ring leaned against the

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same rim, and the bridge moiety was axially oriented to the cavity (Fig. S3).

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Additionally, the adsorption coefficient of TBBPA on CDP decreased with increasing

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pH, as did the stability constant of the TBBPA-CD monomer complex (Text S1). The

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adsorption coefficient and stability constant yielded a good linear correlation, with an

262

R2 value of 0.94 (Fig. S4), and it further supported the role of cavity entrapment.

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Destruction of micropollutants retained on CDP by KMnO4. Treatment with 20

264

µmol/L KMnO4 led to decomposition of aqueous micropollutants (50.0 µg/L), and the

265

reaction followed a first-order kinetic model (Fig. S5). As the oxidation of organic

266

compounds by KMnO4 is a first-order reaction with respect to both reagents,38,46 the

267

second-order rate constants of micropollutants could be calculated from their

268

respective first-order rate constants (Table S4). As expected, micropollutants

269

underwent decomposition to lesser extent in river water than in deionized water, with

270

the exception of lincomycin, and fitted first-order rate constants were reduced by

271

about 50% (rifampin) to 95% (sulfachloropyridazine, enrofloxacin, and TBBPA).

272

Even when the oxidant increased to 100 µmol/L, 11 of the 14 pollutants still exhibited

273

low reaction rate constants in river water, relative to them in deionized water with

274

20-µmol/L KMnO4 treatment. This reduction of rate constants can be attributable to

275

the well-recognized interference of natural water constituents (particularly DOM).

276

However, a comparison of calculated second-order rate constants shows that these

277

pollutants are susceptible to KMnO4 oxidation, compared to other compounds (e.g.,

278

soil organic carbon, nitro-phenols, PAHs, and chlorinated ethylenes).36,37,47,48

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Micropollutants retained on CDP are degradable (Fig. 2) even though they are

280

entrapped in the cavity of the CD unit in a polymer form. The 2-h decomposition rates

281

of retained pollutants increased with increasing oxidant levels from 10 to 1,000

282

µmol/L. For most of the pollutants, a plateau of degradation was approached at 100

283

µmol/L KMnO4, as highlighted by degradation efficiencies of >91% for all retained

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pollutants. This degradation reaction proceeded rapidly (Fig. S6). For example, in the

285

treatment with 100-µmol/L KMnO4, >90% of the retained pollutants were destroyed

286

within 10 min, except penicillin G (78%), and a degradation plateau was achieved for

287

10 of the 14 pollutants. It is feasible for KMnO4 oxidation to degrade these pollutants

288

when they are retained on CDP, as this oxidant at similar concentrations has

289

successfully remediated soils or waters contaminated by phenols (10-160 µmol/L

290

KMnO4),49 24 particular contaminants of concern (100 µmol/L KMnO4),37 antibiotics

291

(100-800 µmol/L KMnO4),41 and chlorinated ethylenes (1,000 µmol/L KMnO4).36

292

Degradation processes of TBBPA in the CD cavity by KMnO4. The degradation

293

processes of micropollutants in the cavity of the CD unit in soluble and insoluble

294

forms were investigated with TBBPA as a model compound. Aqueous TBBPA (1.00

295

mg/L) was rapidly depleted in 20-µmol/L KMnO4, yielding first- and second-order

296

rate constants of 0.346 min-1 and 288 M-1 s-1, respectively, and no TBBPA residue was

297

detected after 20 min (Fig. S7). The depletion of aqueous TBBPA was inhibited in the

298

presence of soluble CD monomer to different extents, whereas >98% of the

299

compound, in all cases, could be degraded within 30 min (Fig. S7, Table S5). A

300

remarkable decrease in degradation was observed for TBBPA retained on CDP; under

301

the same conditions, TBBPA was degraded by approximately 60% within 120 min

302

and some intermediates accumulated (Fig. S8). Near-complete removal (>99%) of

303

both retained TBBPA and formed intermediates was achieved only in 0.50- and

304

1.0-mmol/L KMnO4 treatments. Despite distinct degradation efficiencies, TBBPA in

305

the free-, CD monomer-entrapped, and CDP-retained forms exhibited similar patterns

306

of transformation products, and all three species were degraded via the β-scission

307

pathway (Fig. S9, Table S6, Text S2). This pathway has been extensively documented

308

for TBBPA in various reaction systems (e.g., fenton and active sludge),50-53 occurring

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in parallel with the sequential debromination reaction of TBBPA. The latter leads to

310

the formation of low brominated bisphenols that are bioaccumulative and cause

311

endocrine disrupting effects.51,53 Thus, the formation of secondary pollutants from

312

TBBPA in association with KMnO4 oxidation is not of concern.54

313

Whereas CD either freely dissolved in water or crosslinked in the polymer lowers

314

TBBPA degradation, this host compound remained nearly intact within 120 min in

315

20-µmol/L KMnO4 treatment and degraded by approximately 10% in 2.0-mmol/L

316

KMnO4 treatment (Fig. S10). CD itself had a second-order rate constant of 0.0095 M-1

317

s-1, indicating that it be 3.0×104-fold more recalcitrant to KMnO4 oxidation than

318

TBBPA. In parallel with the decomposition of CD, one dominant product with m/z of

319

1232 was formed and accumulated, which appeared to be one of hepta-carboxylated

320

derivatives of CD monomer (Fig. S10). Taking in account the finding that KMnO4

321

selectively attacks the 6-OH positions at the small rim of the cavity,55 this product is

322

actually the derivative of complete carboxylation of these groups and consequently,

323

the oxidant attacks the OH groups on the small rim of the cavity. This site is

324

orientated oppositely to the inclusion site of TBBPA with CD (Fig. S3), so CD does

325

not compete with TBBPA for the oxidant. However, once entering the CD cavity,

326

TBBPA must adjust its configuration to match the cavity, likely influencing its own

327

reaction potential. In particular, the bridge of one of hydroxyl benzene rings and the

328

isopropyl group at which the β-scission reaction occurs is encircled by the large rim of

329

the cavity (Figs. S3, S9), limiting the accessibility of this reaction site by the oxidant.

330

CD in either soluble or insoluble form may act as a molecular container in which the

331

reaction between TBBPA and KMnO4 is of low dimensionality and thus reduced.

332

Quantification of the degradability of the CD-entrapped species was attempted (Eqs.

333

1-4) with the preconditions that CD remains intact due to its low reaction potential

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and that inclusion equilibrium is reached rapidly. The apparent rate constant (ka,0) of

335

TBBPA in the presence of CD monomer would be a linear function of the inverse of

336

1+KC*[CD], as CD concentration is the sole variable. Fitting Eq. 4 yielded an

337

intercept of 0.143 min-1 (Fig. S11), corresponding to the first-order rate constant of the

338

entrapped species. The second-order rate constant was calculated as 119 M-1 s-1 and

339

approximated 41% of the value of the free species. Retained TBBPA that is mainly

340

entrapped in the insoluble CD cavity decomposed with rate constants of 0.013 min-1

341

(first-order) and 2.16 M-1 s-1 (second-order) (Fig. S12); this species was 2-3 orders of

342

magnitude more inert than the other two species. It is well recognized that the

343

inclusion potential of the CD unit in a polymer form is greatly enhanced due to cavity

344

modification, compared with that of the CD monomer.21 In this study, the insoluble

345

CD unit is 8.54 times more attractive than free CD, based on the finding that the

346

adsorption coefficient and inclusion potential of TBBPA are linearly correlated with a

347

slope of 9.54 (Fig. S4). Therefore, the high inclusion potential of the insoluble CD

348

unit may account for the low degradability of retained TBBPA by KMnO4 oxidation.

[P − CD ] ←→[P − CD] = K ⋅[P ] ⋅ [CD] C [P] ⋅ [CD ]

(1)

350

[P0 ] = [P ] + [P − CD ] = [P ] + K C ⋅ [P ]⋅ [CD ]

(2)

351

[P0 ] d [P0 ] = ( k a ,1 + k a ,2 ⋅ K C ⋅ [CD ] ) ⋅ dt 1 + K C ⋅ [CD ]

(3)

349

352

KC =

k a ,0 =

k a ,1 + k a ,2 ⋅ K C ⋅ [CD ] k a ,1 − k a ,2 = k a ,2 + 1 + K C ⋅ [CD ] 1 + K C ⋅ [CD ]

(4)

353

where [P0] is the apparent concentration of TBBPA; [P] and [P-CD] are the

354

concentrations of the free and entrapped species, respectively; [CD] is the

355

concentration of CD; ka,0 is the apparent rate constant of TBBPA; ka,1 and ka,2 are the

356

rate constants of the free and entrapped species, respectively; and KC is the stability

357

constant of the complex. 15



Page 16 of 31

358

Repeated use of CDP in the degradation of micropollutants. In the scenario of

359

micropollutants (50.0 ng/L) in river water (Fig. S13), removal efficiencies of these

360

pollutants in the homogeneous oxidation increased with KMnO4 level. Five of the 14

361

pollutants (i.e., sulfathiazole, enoxacin, lincomycin, rifampin, and TBBPA) were

362

degraded by more than 93% within 2 h by 100-µmol/L KMnO4, whereas the rest had

363

degradation efficiencies of 15% (sulfamethoxazole) to 88% (penicillin G). In the

364

adsorption-oxidation integration treatment, however, pollutants retained on CDP were

365

nearly completely degraded within 2 h by 100-µmol/L KMnO4, with the exception of

366

sulfadiazine (78%), enoxacin (65%), and enrofloxacin (51%). This integration

367

treatment yielded similar removal efficiencies for these pollutants in deionized water,

368

indicating that natural water constituents do not interfere with their destruction.

369

Furthermore, TBBPA was used as a model compound to assess the feasibility of

370

repeated use of CDP in this integration treatment, as TBBPA has a high adsorption

371

efficiency. As the integration procedure was repeated up to 20 times, CDP became

372

slightly yellow-to-orange in color, associated with the cumulative loading of

373

manganese (Mn) (Figs. 3, S14). A plot of Mn content versus the number of repetitions

374

yielded a linear relationship with R2 higher than 0.99, implying the uniform loading of

375

manganese. After each cycle, approximately 98% of TBBPA spiked in the water was

376

adsorbed, and retained TBBPA was degraded completely without the accumulation of

377

intermediates. These results show that CDP is recyclable in this integrated treatment.

378

Beside this integration treatment, many advanced treatment methods (i.e.,

379

fenton-based processes, photocatalysis, and ozone oxidation) have been developed to

380

deal with water pollution.2,41,56 Two issues of concerns are raised for most of these

381

advanced treatment methods. One is that a large volume of oxidants may be needed

382

due to the consumption of natural water constituents themselves and thus low removal

16


Page 17 of 31


383

efficiencies of pollutants (Fig. S5); the other is that the parent pollutants and/or

384

intermediates may accumulate and be recharged into the environment because

385

complete mineralization of these structurally diverse compounds is difficult to achieve

386

(Fig. S13). The two concerns appear to be overcome in this integration treatment,

387

where CDP adsorption can selectively concentrate micropollutants from the bulk

388

solution and KMnO4 oxidation can destruct them under controllable process

389

parameters (e.g., KMnO4 level). This integration treatment, therefore, may be

390

complementary to current advanced treatment methods.

391

Rearrangement of CDP upon KMnO4 oxidation. The FT-IR spectra of CDP with

392

and without KMnO4 treatment exhibited characteristics typical of CD monomer, e.g.,

393

pyranoid ring vibrations at 400-1500 cm-1, OH-bending vibrations at 1647 cm-1,

394

C-O-C and C-O stretching vibrations at 1030-1150 cm-1, CH2 stretching vibrations at

395

2928 cm-1, and O-H stretching vibrations at 3368 cm-1 (Fig. S15). Two bands formed

396

at 516 and 576 cm-1 corresponded to Mn-O vibrations and implied Mn loading,

397

consistent with the elementary analysis and XPS characterization of the as-treated

398

materials (Table S7, Fig. S16). The KMnO4 residue in the materials was not

399

considered because of the absence of Mn 2p3/2 (KMnO4) centered at 647 eV (Fig.

400

S16). In the high-resolution XPS spectrum of the Mn region, however, Mn 2p3/2

401

centered at 642.27 eV and Mn 2p1/2 centered at 653.34 eV were detected (Fig. S16),

402

yielding a spin-energy separation of 11.07 eV. Mn 2p3/3 and Mn-O* had a

403

spin-energy separation of 112.76 eV, approximating the value (i.e., 112.37 eV) of the

404

spin-energy separation between Mn 2p3/2 and the O 1s component at the lowest

405

position of MnO2.57 All peaks and spin-energy separations indicate the formation of

406

MnO2 in CDP with KMnO4 treatment,57,58 as does the ratio of OMn-O to Mn, with a

407

value of 1.91 (Fig. S16). These results demonstrate that treating CDP with KMnO4

17



408

Page 18 of 31

yields a composite of the pristine polymer and MnO2.

409

The XRD patterns of the composites exhibited a broad peak at 2θ=5-30°, typically

410

indicative of an amorphous or non-crystalline structure of the pristine polymer (Fig.

411

S17). This peak became weakened with MnO2 loading, whereas upward sloping

412

occurred around 37.6° and 65.7°, corresponding to the (006) and (119) diffraction

413

peaks of δ-MnO2 (JCPDF 18-0802), respectively. Parallel proof for δ-MnO2 formation

414

was provided by the TEM measurement, showing that MnO2 is composed of 2-4-nm

415

nanorod particles with a lattice spacing of 2.49 Å (Fig. S18), characteristic of the (006)

416

plane of δ-MnO2.59 Furthermore, both the SEM and TEM images show that δ-MnO2 is

417

uniformly embedded in the matrix (Fig. S18), supporting the linearly accumulative

418

loading of MnO2 (Fig. 3).

419

Collectively, the composite features a porous CDP architecture with uniform

420

embedment of δ-MnO2 nanoparticles. This material has distinct characteristics from

421

the pristine polymer and more than 92% of particles participate in water (Table 1, Fig.

422

S19). The CD unit is less abundant in the composite with δ-MnO2 loading, although

423

the cavity is modified by KMnO4 oxidation (Fig. S10) and may have enhanced

424

inclusion potential.21 The pore sizes of the composite are enlarged to 3.5-8.0 nm, a

425

size indicative of mesoporous materials (2-50 nm),60,61 and this change makes

426

adsorption sites accessible. Enlarged pores are also observed, suffering from either the

427

carboxylation modification of the CD unit and/or destruction of EPI derivatives

428

plugging pores in the polymer, rather than pore-filling mechanism of MnO2 that

429

brings about remarkable reduction of pore size of porous materials.62-64 Moreover, the

430

swelling ratio and particle size are enhanced to different extents, both of which make

431

the composite structurally loose and hydrophilic65 and favor mass transfer and access

432

to the CD cavity of reagents. Although significant correlations between these changes

18


Page 19 of 31


433

are not observed, they probably have opposite effects on the number and accessibility

434

of adsorption sites (e.g., a decrease in CD content versus enlargement of pore size).

435

Such rearrangement has profound effects on adsorption potential of adsorbents, as

436

indicated by the fact that the composite shows an initial increase and subsequent

437

decrease in the adsorption efficiency with increasing δ-MnO2 loading (Fig. S20). The

438

composite derived from 10-mmol/L KMnO4 treatment had a δ-MnO2 loading of

439

2.242% and exhibited similar or higher adsorption potential toward 6 of the 14

440

micropollutants compared to the pristine polymer. This composite may also be made

441

from the pristine polymer that is used 77 times in the integrated procedure for the

442

removal of TBBPA based on the linearly accumulative loading of δ-MnO2.

443

Environmental Implications. We have communicated evidence for the utilization

444

of CD unit immobilized in the polymer to repeatedly capture and destroy aqueous

445

micropollutants when coupled with KMnO4 oxidation. Although far from being

446

optimized, this approach already achieves good removal efficiencies for these

447

pollutants at environmentally relevant levels, for instance that removal efficiencies

448

range from 12% to 95% for micropollutants (Figs. 1 and 2). This is comparable to

449

other advanced treatment methods with removal efficiencies of 12.5% to 100% for

450

concerned organic pollutants.54,66-68 Furthermore, KMnO4 oxidation, which triggers

451

the degradation of micropollutants entrapped in the CD cavity of CDP, rearranges the

452

pristine polymer to a structurally loose composite of CDP and δ-MnO2 nanoparticles.

453

Accurate control of the number of times that CDP is recycled in this approach can be

454

performed, taking into account the linearly accumulative loading of δ-MnO2

455

nanoparticles and changes in the adsorption efficiency of the composite.

456

Interestingly, MnO2 nanoparticles (including δ-MnO2 nanomaterials) have been

457

proven good catalysts in pollutant degradation.69-71 Further understanding of CDP

19



Page 20 of 31

458

rearrangement, associated with MnO2 nanoparticle loading, may provide strategies for

459

the design of new adsorbents from other host materials and in situ reactors; for the

460

latter, pollutants may be concentrated from soil or water and then catalytically

461

degraded in the matrix. Moreover, the outcome of CD mediation in chemical reactions

462

may be stimulating or inhibiting, depending on reaction types and reactants.31-34

463

Rational integration of CD-related material adsorption and other chemical treatment

464

methods can offer a promising choice for micropollutant removal, taking into account

465

the fact that many chemical treatment methods are intrinsically efficient in pollutant

466

removal but limited by the interference of natural water constituents.1-2

467

ASSOCIATED CONTENT

468

Supporting Information

469

Additional text, tables, and figures. This material is available free of charge via the

470

Internet at http://pubs.acs.org.

471

AUTHOR INFORMATION

472

Corresponding Author

473

* E-mail: [email protected]; Tel./Fax: +86-411-8470-7844.

474

Notes

475

The authors declare no competing financial interest.

476

ACKNOWLEDGMENTS

477

This study was supported by the National Basic Research Program of China (No.

478

2013CB430403), the Special Fund for Agro-scientific Research in the Public Interest

479

(No. 201503107), the National Natural Science Foundation of China (Nos. 21477013

480

and 41171382), and the Fundamental Research Funds for the Central Universities. We

481

acknowledge anonymous reviewers for their valuable comments on this paper. 20


Page 21 of 31


482

REFERENCES

483 484 485 486 487 488 489 490 491 492 493 494 495 496 497 498 499 500 501 502 503 504 505 506 507 508 509 510 511 512 513 514 515 516 517 518 519 520 521 522 523 524

[1] Schwarzenbach, R. P.; Escher, B. I.; Fenner, K.; Hofstetter, T. B.; Johnson, C. A.; von Gunten, U.; Wehrli, B. The challenge of micropollutants in aquatic systems. Science 2006, 313, 1072-1077. [2] Shannon, M. A.; Bohn, P. W.; Elimelech, M.; Georgiadis, J. G.; Marinas, B. J.; Mayes, A. M. Science and technology for water purification in the coming decades. Nature 2008, 452, 301-310. [3] Daughton, C. G. Cradle-to-cradle stewardship of drugs for minimizing their environmental disposition while promoting human health. I. Rationale for and avenues toward a green pharmacy. Environ. Health Perspect. 2003, 111, 757-774. [4] Boxall, A. B. A.; Sinclair, C. J.; Fenner, K.; Kolpin, D.; Maud, S. J. When synthetic chemicals degrade in the environment. Environ. Sci. Technol. 2004, 38, 368A-375A. [5] Cui, N.; Zhang, X. X.; Xie, Q.; Wang, S.; Chen, J. W.; Huang, L. P.; Qiao, X. L.; Li, X. H.; Cai, X. Y. Toxicity profile of labile preservative bronopol in water: the role of more persistent and toxic transformation products. Environ. Pollut. 2011, 159, 609-615. [6] Fenner, K.; Canonica, S.; Wackett, L. P.; Elsner, M. Evaluating pesticide degradation in the environment: blind spots and emerging opportunities. Science 2013, 341, 752-758. [7] Cai, X. Y.; Xiong, W. N.; Xia, T. T.; Chen, J. W. Probing the stereochemistry of successive sulfoxidation of the insecticide fenamiphos in soils. Environ. Sci. Technol. 2014, 48, 11277-11285. [8] Hayes, T. B.; Khoury, V.; Narayan, A.; Nazir, M.; Park, A.; Brown, T.; Adame, L.; Chan, E.; Buchholz, D.; Stueve, T.; Gappipeau, S. Atrazine induces complete feminization and chemical castration in male African clawed frogs (Xenopus laevis). Proc. Natl. Acad. Sci. USA 2010, 107, 4612-4617. [9] Matatov-Meytal, Y. I.; Sheintuch, M. Catalytic abatement of water pollutants. Ind. Eng. Chem. Res. 1998, 37, 309-326. [10] Carballa, M.; Omil, F.; Lema, J. M.; Lompart, M.; Garcia-Jares, C.; Rodriguez, I.; Gomez, M.; Ternes, T. Behavior of pharmaceuticals, cosmetics and hormones in a sewage treatment plant. Water Res. 2004, 38, 2918-2926. [11] Gogate, P. R.; Pandit, A. B. A review of imperative technologies for wastewater treatment II: hybrid methods. Adv. Environ. Res. 2004, 8, 5553-5597. [12] Oller, I.; Malato, S.; Sanchez-Perez, J. A. Combination of advanced oxidation processes and biological treatments for wastewater decontamination-A review. Sci. Total Environ. 2011, 409, 4141-4166. [13] von Gunten, U. Ozonation of drinking water: Part I. Oxidation kinetics and product formation Water Res. 2003, 37, 1443-1467. [14] Richardson, S. D.; Plewa, M. J.; Wagner, E. D.; Schoeny, R.; DeMarini, D. M. Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection by-products in drinking water: A review and roadmap for research. Mutat. Res. 2007, 636, 178-242. [15] Westerhoff, P.; Yoon, Y.; Snyder, S.; Wert, E. Fate of endocrine-disruptor, pharmaceutical, and personal care product chemicals during simulated drinking water treatment processes. Environ. Sci. Technol. 2005, 39, 6649-6663. [16] Lu, J.; Huang, Q. Removal of acetaminophen using enzymemediated oxidative coupling

21



525 526 527 528 529 530 531 532 533 534 535 536 537 538 539 540 541 542 543 544 545 546 547 548 549 550 551 552 553 554 555 556 557 558 559 560 561 562 563 564 565 566 567 568

[17]

[18] [19] [20]

[21] [22] [23]

[24] [25] [26] [27]

[28] [29]

[30]

[31] [32] [33] [34]

Page 22 of 31

processes: II. Cross-coupling with natural organic matter. Environ. Sci. Technol. 2009, 43, 7068−7073. Jiang, J.; Gao, Y.; Pang, S. Y.; Wang, Q.; Huangfu, X. L.; Liu, Y. Z.; Ma, J. Oxidation of bromophenols and formation of brominated polymeric products of concern during water treatment with potassium permanganate. Environ. Sci. Technol. 2014, 48, 10850−10858. Crini G. Review: a history of cyclodextrins. Chem. Rev. 2014, 114, 10940-10975. Sawicki, R.; Mercier, L. Evaluation of mesoporous cyclodextrin-silica nanocomposites for the removal of pesticides from aqueous media. Environ. Sci. Technol 2006, 40, 1978-1983. Liu, H. H.; Cai, X. Y.; Wang, Y.; Chen, J. W. Adsorption mechanism-based screening of cyclodextrin polymers for adsorption and separation of pesticides from water. Water Res. 2011, 45, 3499-3511. Morin-Crini, N.; Crini, G. Environmental applications of water-insoluble β-cyclodextrin-epichlorohydrin polymers. Prog. Polym. Sci. 2013, 38, 344-368. Yoshizawa, M.; Tamura, M.; Fujita, M. Diels-Alder in aqueous molecular hosts: unusual regioselectivity and efficient catalysis. Science 2006, 312, 251-254. Lee, T. C.; Kalenius, E.; Lazar, A. I.; Assaf, K. I.; Kuhnert, N.; Grün, C. H.; Jänis, J.; Scherman, O. A.; Nau, W. M. Chemistry inside molecular containers in the gas phase. Nature Chem. 2013, 5, 376–382. Szejtli, J. Introduction and general overview of cyclodextrin chemistry. Chem. Rev. 1998, 98, 1743-1753. Liu, L.; Guo, Q. X. The driving forces in the inclusion complexation of cyclodextrins. J. Incl. Phenom. Macro. 2002, 42, 1-14. Liu, H. H.; Cai, X. Y.; Chen, J. W. Mathematical model for cyclodextrin alteration of bioavailability of organic pollutants. Environ. Sci. Technol. 2013, 47, 5835-5842. Chen, L.; Dionysiou, D. D.; O’Shea, K. Complexation of microcystins and nodularin by cyclodextrins in aqueous solution, a potential removal strategy. Environ. Sci. Technol. 2011, 45, 2293-2300. Crini G. Studies on adsorption of dyes on β-cyclodextrin polymer. Bioresour. Technol. 2003, 90, 193-198. Nagy, Z. M.; Molnar, M.; Fekete-Kertesz, I.; Molnar-Perl, I.; Fenyvesi, E.; Gruiz, K. Removal of emerging micropollutants from water using cyclodextrin. Sci. Total. Environ. 2014, 485-486, 711-719. Zhang, Y.; Cai, X. Y.; Xiong, W. N.; Jiang, H.; Zhao, H. T.; Yang, X. H.; Li, C.; Fu, Z. Q.; Chen, J. W. Molecular insights into the pH-dependent adsorption and removal of ionizable antibiotic oxytetracycline by adsorbent cyclodextrin polymers. PLoS ONE, 2014, 9(1), e86228. Takahashi, K. Organic reaction mediated by cyclodextrins. Chem. Rev. 1998, 98, 2013-2033. Barr, L.; Dumanski, P. G.; Easton, C. J.; Harper, J. B.; Lee, K.; Lincoln, S. F.; Meyer, A. G.; Simpson, J. S. Cyclodextrin molecular reactors. J. Incl. Phenom. Macro. 2004, 50, 19-24. Easton, C. J. Cyclodextrin-based catalysts and molecular reactors. Pure Appl. Chem. 2005, 77, 1865-1871. Sletten, E. M.; Nakamura, H.; Jewett, J. C.; Bertozzi, C. R. Difluorobenzocyclooctyne: synthesis, reactivity, and stabilization by β-cyclodextrin. J. Am. Chem. Soc. 2010, 132, 11799-11805.

22


Page 23 of 31


569 570 571 572 573 574 575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597 598 599 600 601 602 603 604 605 606 607 608 609 610 611 612

[35] Shen, H.M.; Ji, H.B.; Wu, H.K.; Shi, H.X. Recent advances in the immobilization of β-cyclodextrin and their application. Chin. J. Org. Chem. 2014, 34, 1549-1572. [36] Yan, Y. E.; Schwartz, F. W. Oxidative degradation and kinetics of chlorinated ethylenes by potassium permanganate. J. Contam. Hydrol. 1999, 37, 343-365. [37] Waldemer, R. H.; Tratnyek, P. G. Kinetics of contaminant degradation by permanganate. Environ. Sci. Technol. 2006, 40, 1055-1061. [38] Guan, X. H.; He, D.; Ma, J.; Chen, G. H. Application of permanganate in the oxidation of micropollutants: a mini review. Front. Environ.Sci. En. 2010, 4, 405-413. [39] Hu, L. H.; Martin, H. M.; Strathmann, T. J. Oxidation kinetics of antibiotics during water treatment with potassium permanganate. Environ. Sci. Technol. 2010, 44, 6416-6422. [40] Liu, R. P.; Liu, H. J.; Zhao, X.; Qu, J. H.; Zhang, R. Treatment of dye wastewater with permanganate oxidation and in situ formed manganese dioxides adsorption: cation blue as model pollutant. J. Hazard. Mater. 2010, 176, 926-931. [41] Hu, L. H.; Stemig, A. M.; Wammer, K. H.; Strathmann, T. J. Oxidation of antibiotics during water treatment with potassium permanganate: reaction pathways and deactivation. Environ. Sci. Technol. 2011, 45, 3635-3642. [42] Managaki, S.; Murata, A.; Takada, H.; Tuyen, B. C.; Chiem, N. H. Distribution of macrolides, sulfonamides, and trimethoprim in tropical waters: ubiquitous occurrence of veterinary antibiotics in the Mekong Delta. Environ. Sci. Technol. 2007, 41, 8004–8010. [43] Edwards, M.; Topp, E.; Metcalfe, C. D.; Li, H.; Gottschall, N.; Bolton, P.; Curnoe, W.; Payne, M.; Beck, A.; Kleywegt, S. Pharmaceutical and personal care products in tile drainage following surface spreading and injection of dewatered municipal biosolids to an agricultural field. Sci. Total. Environ. 2009, 407, 4220–4230. [44] Luo, Y.; Xu, L.; Rysz, M.; Wang, Y. Q.; Zhang, H.; Alvarez, P. J. J. Occurrence and transport of tetracycline, sulfonamide, quinolone, and macrolide antibiotics in the Haihe River Basin, China. Environ. Sci. Technol. 2011, 45, 1827–1833. [45] Li, W. H.; Shi, Y. L.; Gao, L. H.; Liu, J. M.; Cai, Y. Q. Occurrence of antibiotics in water, sediments, aquatic plants, and animals from Baiyangdian Lake in North China. Chemosphere 2012, 89, 1307-1315. [46] Ladbury, J. W.; Cullis, C. H. Kinetics and mechanism of oxidation by permanganate. Chem. Rev. 1958, 58, 403-438. [47] Huang, K. C.; Hoag, G. E.; Chheda, P.; Woody, B. A.; Dobbs, G. M. Oxidation of chlorinated ethenes by potassium permanganate: a kinetics study. J. Hazard. Mater. 2001, 87, 155-169. [48] Forsey, S. P.; Thomson, N. R.; Barker, J. F. Oxidation kinetics of polycyclic aromatic hydrocarbons by permanganate. Chemosphere 2010, 79, 628-636. [49] Jiang, J.; Pang, S. Y.; Ma, J.; Liu, H. L. Oxidation of phenolic endocrine disrupting chemicals by potassium permanganate in synthetic and real waters. Environ. Sci. Technol. 2012, 46, 1774-1781. [50] Barontini, F.; Cozzani, V.; Marsanich, K.; Raffa, V.; Petarca, L. An experimental investigation of tetrabromobisphenol A decomposition pathways. J. Anal. Appl. Pyrolysis 2004, 72, 41-53. [51] Brenner, A.; Mukmenew, I.; Abeliovich, A.; Kushmaro, A. Biodegradability of tetrabromo bisphenol A and tribromophenol by activated sludge. Ecotoxicology 2006, 15, 399-402.

23



613 614 615 616 617 618 619 620 621 622 623 624 625 626 627 628 629 630 631 632 633 634 635 636 637 638 639 640 641 642 643 644 645 646 647 648 649 650 651 652 653 654 655 656

Page 24 of 31

[52] Lin, K. D.; Liu, W. P.; Gan, J. Reaction of tetrabromobisphenol A (TBBPA) with manganese dioxide: kinetics, products, and pathways. Environ. Sci. Technol. 2009, 43, 4480-4486. [53] Zhong, Y. H.; Liang, X. L.; Zhong, Y.; Zhu, J. X.; Zhu, S. Y.; Yuan, P.; He, H. P.; Zhang, J. Heterogeneous UV/Fenton degradation of TBBPA catalyzed by titanomagnetite: catalyst characterization, performance and degradation products. Water Res. 2012, 46, 4633-4644. [54] Pang, S. Y.; Jiang, J.; Gao, Y.; Zhou, Y.; Huangfu, X. L.; Liu, Y. Z.; Ma, J. Oxidation of flame retardant tetrabromobisphenol A by aqueous permanganate: reaction kinetics, brominated products, and pathways. Environ. Sci. Technol. 2014, 48, 615-623. [55] Manhas, M. S.; Mohammed, F.; Khan, Z. A kinetic study of oxidation of β-cyclodextrin by permanganate in aqueous media. Colloids Surf., A 2007, 295, 165-171. [56] Ribeiro, A. R.; Nunes, O. C.; Pereira, M. F.; Silva, A. M. An overview on the advanced oxidation processes applied for the treatment of water pollutants defined in the recently launched Directive 2013/39/EU. Environ. Int. 2015, 75, 33-51. [57] Chigane, M.; Ishikawa, M. Manganese oxide thin film preparation by potentiostatic electrolyses and electrochromism. J. Electrochem. Soc. 2000, 147, 2246-2251. [58] Tan, B. J.; Klabunde, K. J.; Sherwood, P. M. A. XPS studies of solvated metal atom dispersed catalysts evidence for layered cobalt-manganese particles on alumina and silica. J. Am. Chem. Soc. 1991, 113, 855-861. [59] Patel, M. N.; Wang, X. Q.; Wilson, B.; Ferrer, D. A.; Dai, S.; Stevenson, K. J.; Johnston, K. P. Hybrid MnO2-disordered mesoporous carbon nanocomposites: synthesis and characterization as electrochemical pseudocapacitor electrodes. J. Mater. Chem. 2010, 20, 390-398. [60] Huq, R.; Mercier, L.; Kooyman, P. J. Incorporation of cyclodextrin into mesostructured silica. Chem. Mater. 2001, 13, 4512-4519. [61] Liu, C. Q.; Lambert, J. B.; Fu, L. A novel family of ordered, mesoporous inorganic/organic hybrid polymers containing covalently and multiply bound microporous organic hosts. J. Am. Chem. Soc. 2003, 125, 6452-6461. [62] Zhang, L. X.; Shi, J. L.; Yu, J.; Hua, Z. L.; Zhao, X. G.; Ruan, M. L. A new in-situ reduction route for the synthesis of pt nanoclusters in the channels of mesoporous silica SBA-15. Adv. Mater. 2002, 14, 1510-1513. [63] Li, L.; Shi, J. L.; Zhang, L. X.; Xiong, L. M.; Yan, J. N. A novel and simple in-situ reduction route for the synthesis of an ultra-thin metal nanocoating in the channels of mesoporous silica materials. Adv. Mater. 2004, 16, 1079-1082. [64] Dong, X. P.; Shen, W. H.; Gu, J. L.; Xiong, L. M.; Zhu, Y. F.; Li, Z.; Shi, J. L. MnO2-Embedded-in- Mesoporous-Carbon-Wall structure for use as electrochemical capacitors. J. Phys.Chem. B 2006, 110, 6015-6019. [65] Peters, O; Ritter., H. Supramolecular Controlled Water Uptake of Macroscopic Materials by a Cyclodextrin-Induced Hydrophobic-to-Hydrophilic Transition. Angew. Chem. Int. Ed. 2013, 52, 8961-8963. [66] Zimmermann, S. G.; Wittenwiler, M.; Hollender, J.; Krauss, M.; Ort, C.; Siegrist, H.; von Gunten, U. Kinetic assessment and modeling of an ozenation step for full-scale municipal wastewater treatment: Micropollutant oxidation, by-product formation and disinfection. Water Res. 2011, 45, 605-617. [67] Margot, J.; Kienle, C.; Magnet, A.; Weil, M.; Rossi, L.; de Alencastro, L. F. ; Abegglen, C.;

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657 658 659 660 661 662 663 664 665 666 667 668 669

[68]

[69] [70]

[71]

Thonney, D.; Chèvre, N.; Schärer, M.; Barry, D. A. Treatment of micropollutants in municipal wastewater: Ozone or powered activated carbon? Sci. Total Environ. 2013, 461-462, 480-498. Luo, Y. L.; Guo, W. S.; Ngo, H. H.; Nghiem, L. D.; Hai, F. I.; Zhang, J.; Liang, S.; Wang, X. C. A review on the occurrence of micropollutants in the aquatic environment and their fate and removal during wastewater treatment. Sci. Total Environ. 2014, 473-474, 619-641. Allarda, S.; von Guntenb, U.; Sahlib, E.; Nicolaua, R.; Gallarda, H. Oxidation of iodide and iodine on birnessite (δ-MnO2) in the pH range 4–8. Water Res. 2009, 43, 3417-3426. Saputra, E.; Muhammad, S.; Sun, H. Q.; Ang, H. M.; Tadé, M. O.; Wang S. B. Different crystallographic one-dimensional MnO2 nanomaterials and their superior performance in catalytic phenol degradation. Environ. Sci. Technol. 2013, 47, 5882-5887. Wang, L. L.; Cheng, H. F. Birnessite (δ-MnO2) mediated degradation of organoarsenic feed additive p-arsanilic acid. Environ. Sci. Technol. 2015, 49, 3473-3481.

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Page 26 of 31

Table 1. Major properties and structure characterization of CDP with and without KMnO4 treatment.

670

Materials

CD content (%)

Swelling ratio (g/g)

Particle Size (µm)

Surface area (m2/g)

Pore size (nm)

Pore volume (10-3cc/g)

MnO2 content (wt%)

β-CDP

71.84

5.844

334.69

1.155

1.927

5.256

0.000

CDP-5-mmol/L KMnO4

63.28

5.635

358.27

0.725

7.946

3.932

0.948

0.480

3.851

4.342

2.242

CDP-10-mmol/L KMnO4

47.45

6.002

354.43

CDP-50-mmol/L KMnO4

39.63

6.931

385.46

1.087

3.858

3.602

9.742

CDP-100-mmol/L KMnO4

33.61

7.562

359.69

0.977

3.909

3.475

11.514

CDP-200-mmol/L KMnO4

21.49

8.035

380.92

0.721

3.487

2.366

19.959

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Page 27 of 31


671

FIGURE CAPTIONS

672

Fig. 1. Adsorption efficiency of CDP toward 14 micropollutants (50.0 µg/L) at

673

solid-to-liquid ratios of 1:1250 and 1:125.

674

Fig. 2. Effects of KMnO4 concentration on degradation of micropollutants retained on

675

CDP.

676

Fig. 3. Effects of repetition times on manganese (Mn) loading of CDP used in the

677

adsorption and oxidation integration procedure.

678

27


Su lfa Su thi lfa azo p le Su yrid lf a in Su d e lfa iaz m i Su er ne l S f az Su ulf adi ine lfa ad mi ch im din lo eth e r Su op oxi lfa yri ne m da et zin ho xa e En zole En ox ro aci fl n Li oxa nc cin o Pe my ni cin cil l R in G ifa m p TB in BP A

Adsorption efficiency (%)


100


Page 28 of 31

679

Fig. 1. Adsorption efficiency of CDP toward 14 micropollutants (50.0 µg/L) at

680

solid-to-liquid ratios of 1:1250 and 1:125.

1:1250 1:125

80

60

40

20

0

681

682

28

Page 29 of 31


683

Fig. 2. Effects of KMnO4 concentration on degradation of micropollutants retained on

684

CDP.

Degradation efficiency (%)

100

10 µmol/L

100 µmol/L

500 µmol/L

1000 µmol/L

90

80

70

Su lfa Su thia lfa zo py le Su rid in lf Su adi e a lf a z m ine Su era S lfa zin Su ulfa dim e id lfa di ch me ine th lo r o Su opy xin lfa rid e a m et zin ho e xa zo En le En oxa ro cin flo Li xac nc i om n Pe yc i ni ci n lli R nG ifa m p TB in BP A

60

685

A

686

29



Page 30 of 31

687

Fig. 3. Effects of repetition times on manganese (Mn) loading of CDP used in the

688

adsorption and oxidation integration procedure.

a, morphology of CDP used in the adsorption and oxidation integration procedure. Microscope images (×100) were recorded using an Olympus IX71inverted microscope. 2

MnO2 y = 0.0291x, R = 0.99

Mn or MnO2 loading, %

0.8

Mn

2

y = 0.0184x, R = 0.99

0.6 0.4 0.2 0.0 0

2

4

6

8

10 12 14 16 18 20

Repetition times b, effects of repetition times on Mn loading. 689 690

30


Page 31 of 31


691

Graphic for manuscript

692

31


Recyclable Capture and Destruction of Aqueous Micropollutants

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