Rapid Destruction of Tetrabromobisphenol A by Iron(III) - Doi.org

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Rapid Destruction of Tetrabromobisphenol A by Iron(III)-Tetraamidomacrocyclic Ligand/Layered Double Hydroxide Composite/H2O2 system Chao Wang, Juan Gao, and Cheng Gu Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b04294 • Publication Date (Web): 02 Dec 2016 Downloaded from http://pubs.acs.org on December 2, 2016

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Environmental Science & Technology

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Rapid Destruction of Tetrabromobisphenol A by

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Iron(III)-Tetraamidomacrocyclic Ligand/Layered Double

3

Hydroxide Composite/H2O2 System Chao Wang1, Juan Gao2, Cheng Gu*1

4 5 1

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State Key Laboratory of Pollution Control and Resource Reuse, School of the Environment, Nanjing University, Nanjing 210023, P.R. China

7 8 9

2

Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science, Chinese Academy of Sciences, Nanjing, Jiangsu 210008, P. R. China

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*To whom correspondence should be addressed

13

Cheng Gu

14

Professor

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School of the Environment

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Nanjing University

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Nanjing, Jiangsu, 210023

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P. R. China

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Phone/Fax: +86-25-89680636

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E-mail: [email protected]

21 22 1

ACS Paragon Plus Environment


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Abstract

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Iron(III)-tetraamidomacrocyclic ligand (Fe(III)-TAML) activators have received

25

widespread attentions for their abilities to activate hydrogen peroxide to oxidize many

26

organic pollutants. In this study, Fe(III)-TAML/layered double hydroxide (LDH)

27

composite was developed by intercalating Fe(III)-TAML into the interlayer of LDH.

28

Electrostatic interaction and hydrogen bonding might account for the adsorption of

29

Fe(III)-TAML on LDH. The newly synthesized Fe(III)-TAML/LDH composite

30

showed

31

tetrabromobisphenol A (TBBPA) in the presence of hydrogen peroxide, which can be

32

fully degraded within 20 s, and the degradation rate increased up to 8 times compared

33

to free Fe(III)-TAML. In addition, the toxicity of the system was significantly reduced

34

after the reaction. The higher reactivity of Fe(III)-TAML/LDH system is attributed to

35

the enhanced adsorption of TBBPA on LDH, which could increase the contact

36

possibility between Fe(III)-TAML and TBBPA. Based on the analysis of reaction

37

intermediates, β-scission at the middle carbon atom and C-Br bond cleavage in phenyl

38

ring of TBBPA were involved in the degradation process. Furthermore, our results

39

demonstrated that the Fe(III)-TAML/LDH composite can be reused several times,

40

which could lower the overall cost for environmental implication and render

41

Fe(III)-TAML/LDH as the potential environmental-friendly catalyst for future

42

wastewater treatment under mild reaction conditions.

43

Keywords:

44

hydroxide, tetrabromobisphenol A, waste remediation, green catalyst

superior

reactivity

as

indicated

iron(III)-tetraamidomacrocyclic

by

ligand

efficient

decomposition

activator,

45 46 47 2


layered

of

double

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48 49 50 51

52

TOC Art:

53 54 55 56 57 58 59 60 61 62 63 64 65 66

3



67

Introduction

68

Tetrabromobisphenol A [4,4’-isopropylidenebis (2,6-dibromophenol), TBBPA],

69

accounting for nearly 60% of the total brominated flame retardant (BFR) market

70

share1,2, is primarily used as the flame retardant in printed circuit boards and also as

71

additives in many consumer products, e.g. textiles and plastics3-5. Recent studies

72

showed that TBBPA has been released into the environment and frequently detected in

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various environmental matrices, including air, soil, water and sediment. Zweidinger et

74

al.6 reported a level of 1.8 µg TBBPA m-3 in the air near the production site. Studies

75

found that the concentrations of TBBPA in contaminated soil and sediment ranged

76

from 0.5 to 140 µg kg-1 (dry weight) and from 2 to 150 µg kg-1 (dry weight),

77

respectively7,8. It was also reported that the TBBPA concentration could reach 4.87 µg

78

L-1 in water from Lake Chaohu in China9. Based on the biodegradation results of

79

TBBPA in different environmental media, including soil, river sediment and water, the

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estimated half-life of TBBPA in the environment was ~2 months3. TBBPA was even

81

detected in serum, adipose tissue and breast milk samples of humans at the level of ng

82

kg-1 and has exerted severe health effect10-12. In vitro studies had confirmed that

83

TBBPA could exhibit biological activity, including the disruption of thyroid

84

hormone13,14 and high lethal toxicity towards cerebellar granule cells15. The research

85

by Sjödin et al.16 showed a close correlation between air concentrations of TBBPA

86

and human serum levels, indicating that occupational exposure might be the important

87

route for human exposure.

88

Considering the recalcitrant nature of TBBPA in the environment and its potential 4


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risk, it is greatly needed to develop efficient techniques to eliminate TBBPA. It is

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reported that TBBPA cannot be effectively degraded by traditional methods, such as

91

δ-MnO217, UV irradiation18 and photosensitized oxidation19, and the toxic byproducts,

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e.g. brominated isopropylphenol derivatives were generated during the degradation

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process. Recently, some advanced strategies were tested to fully degrade TBBPA.

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Results from Ding et al. showed that TBBPA (18.4 µM) could be degraded in 30 min

95

to achieve 99% of TBBPA degradation, 67% of debromination and 56% of TOC

96

removal by sulfate radicals, which were produced from 1.5 mM peroxymonosulfate

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catalyzed by 0.1 g L-1 CuFe2O4 nanoparticles20. However, during the reaction, Cu2+

98

was released and the concentration of Cu2+ could reach 1.3 mg L-1, that would

99

potentially lead to secondary pollution20. Ozonation was also utilized to achieve rapid

100

removal of TBBPA, but the generation of carcinogenic bromate ions retarded its

101

application21. Therefore, processes that can achieve rapid destruction and don’t

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present any toxicity concerns during the degradation of TBBPA have still remained

103

elusive.

104

Iron(III)-tetraamidomacrocyclic ligand (Fe(III)-TAML, structure shown in Figure 1)

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activators have been classified as the new family of green catalysts. Prior studies

106

showed that Fe(III)-TAML/H2O2 can completely mineralize chlorophenols in a few

107

minutes22,23. Further research indicated that Fe(III)-TAML activator/H2O2 systems

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also exhibited high reactivity for degradation of fenitrothion and other

109

organophosphorus pesticides24, dibenzothiophene25, natural and synthetic estrogens26,

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azo- and ruthenium-dyes27,28, pharmaceutical ingredient (e.g., sertraline29), and 5



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explosives (e.g., 2,4,6-trinitrotoluene and 1,3,5-trinitrobenzene30). Although the

112

reactivity of Fe(III)-TAML catalyst is strongly pH dependent, it decreases

113

significantly under neutral or acidic conditions31,32. However, even at pH 7,

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Fe(III)-TAML still showed high reactivity compared to other advanced oxidation

115

techniques33. Previous studies have shown that immobilization of metalloporphyrin

116

molecules onto various supports, including graphene, single-walled carbon nanotubes

117

and natural montmorillonite clay mineral would significantly increase the stability,

118

catalytic activity and reusability34-36. However, only limited studies were conducted

119

for immobilization of Fe(III)-TAML activators. It was reported that the deposition of

120

Fe(III)-TAML on electrode surface would significantly increase its turnover

121

numbers37. So far, there is still lack of information to use immobilized Fe(III)-TAML

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for environmental applications.

123

The layered double hydroxides (LDHs), also called anionic clays, consist of

124

brucite-like layers. Due to isomorphic substitutions in LDH structure, LDHs possess

125

positive charges, which are commonly compensated by exchangeable anions, e.g., Cl-，

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CO32- et al.38,39 With the pronounced versatility and low cost, LDHs have been widely

127

utilized as the supports for synthesis of heterogeneous catalysts40-42 and for drug

128

delivery43,44. Given that Fe(III)-TAML is negatively charged, we hypothesize that

129

Fe(III)-TAML can be intercalated into the interlayer of LDH to form

130

Fe(III)-TAML/LDH composite, which would increase the stability and catalytic

131

reactivity of Fe(III)-TAML.

132

The objective of this study was to develop a novel material by immobilization of 6


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Fe(III)-TAML on the surface of LDH to effectively degrade TBBPA. Our results

134

showed that Fe(III)-TAML can form strong interaction with LDH through

135

electrostatic attraction and hydrogen bonding, which was also supported from our

136

spectroscopic analysis. The newly formed material exhibited extraordinary reactivity

137

as indicated by the degradation of TBBPA, the degradation rate increased up to 8

138

times compared to free Fe(III)-TAML, especially at lower pH. More importantly, after

139

reaction, less toxic byproducts are formed. The experimental results further

140

demonstrated that Fe(III)-TAML/LDH composite can be reused, that would extend its

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potential application. The role of LDH in the reaction, in addition to protecting

142

Fe(III)-TAML from suicidal inactivation, is to provide a unique matrix to increase the

143

contact possibility between Fe(III)-TAML and TBBPA.

144

Materials and methods

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Chemicals. Details of chemicals are listed in Text S1.

146

Preparation and characterization of LDH. Detailed information on preparation

147

and characterization of LDH is provided in Text S2.

148

Preparation and characterization of Fe(III)-TAML/LDH composite material.

149

The details of preparation and characterization of Fe(III)-TAML/LDH are given in

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Text S3.

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Degradation of TBBPA by Fe(III)-TAML and Fe(III)-TAML/LDH composite.

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Degradation experiment was conducted in a 200 mL Erlenmeyer flask. For each

153

experiment, TBBPA stock solution was added to Fe(III)-TAML/LDH suspension with

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the initial concentrations for TBBPA and Fe(III)-TAML of 18.4 and 1.0 µM, 7



155

respectively. To investigate the pH effect on the reactivity of Fe(III)-TAML/LDH, the

156

degradation experiment was conducted at pH 8, 9 and 10, respectively. The reaction

157

was initiated by addition of 30 µL of H2O2 (30%) stock solution with the molar ratio

158

of H2O2/Fe(III)-TAML = 110/1. To avoid any pH fluctuation, both the pHs of

159

TBBPA/Fe(III)-TAML/LDH solution and H2O2 stock solution were pre-adjusted to

160

desired pH by addition of 0.1 M NaOH or 0.1 M HClO4 solution. The pH was stable

161

and the variation was within ±0.1 during the 120 s reaction period. At predetermined

162

time interval, an aliquot of 500 µL sample was withdrawn and mixed with 5 µL

163

concentrated perchloric acid (7 M) to terminate the reaction by demetalation of

164

Fe(III)-TAML45, and 1 mL methanol was added to extract TBBPA adsorbed by

165

Fe(III)-TAML/LDH composite. Then the sample was filtered through a 0.22 µm

166

polytetrafluoroethylene syringe filter and the concentrations of TBBPA were analyzed

167

using a high pressure liquid chromatography (HPLC) (Waters Alliance 2695, Milford,

168

MA). The released Br- concentrations during the reaction were determined by an ion

169

chromatography (IC) (Dionex Co., USA). In addition, the total organic carbon (TOC)

170

contents before and after the reaction were measured using a TOC-5000A analyzer

171

(Shimadzu, Japan) to quantify the mineralization of TBBPA during the reaction. For

172

comparison, similar experiment was conducted for degradation of TBBPA by 1.0 µM

173

free Fe(III)-TAML. The adsorption of TBBPA on Fe(III)-TAML/LDH was also

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studied at different pHs. The adsorption kinetic experiment was conducted in a 0.22

175

µm polytetrafluoroethylene syringe filter to ensure the quick termination of the

176

adsorption process. The control experiments with full recovery (> 98%) demonstrated 8


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that either Fe(III)-TAML/LDH or LDH+H2O2 showed no effect on the degradation of

178

TBBPA (Fig. S3). The information for quantification of TBBPA and bromide ion,

179

mass spectrometric measurement and minimization of time lag in degradation

180

experiment is summarized in Text S4, S5 and S6, respectively.

181 182

Reusability of Fe(III)-TAML/LDH composite. The detailed information for reusability test of Fe(III)-TAML/LDH composite is provided in Text S7.

183

Computational methods. Gaussian 09W program46 with DFT method (B3LYP)47

184

was used to investigate the structures of Fe(III)-TAML and TBBPA, the infrared

185

spectrum of Fe(III)-TAML was also calculated. The 6-311G(d,p) basis set was

186

applied to all atoms of Fe(III)-TAML and TBBPA (i.e. C, H, O, N, Br and Fe). The

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calculated infrared spectrum of Fe(III)-TAML was analyzed using Gview program

188

based on the optimized results. The frontier electron densities (FEDs) of the highest

189

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

190

(LUMO) of TBBPA were determined based on Gaussian output files, and the values

191

of 2FED2HOMO and (FED2HOMO + FED2LUMO) were calculated to predict the possible

192

reaction sites for the redox reactions.

193 194

Toxicity assay. The detailed information on toxicity assay is provided in Text S8.

Results and discussion

195

Characterization of Synthesized LDH. Based on the XRD and TEM

196

measurements (Fig. 2 and Fig. S1), the synthesized nanosheets of LDH show good

197

crystallinity, the crystallite is partially overlapped and shaped in a hexagonal form

198

with lateral size of 60-120 nm (Fig. S1), which is consistent with the results reported 9



199

by Xu et al.48

200

Adsorption of Fe(III)-TAML on LDH. Due to the electrostatic interaction

201

between negatively charged Fe(III)-TAML and positively charged LDH layers, high

202

adsorption of Fe(III)-TAML on LDH was observed with fitted parameters

203

KL=0.00011 L µmol-1 and Qmax=320.72 µmol g-1 for Langmuir isotherm (Fig. S2 and

204

Table S1). The XRD patterns show that as more Fe(III)-TAML molecules are loaded,

205

the basal spacings of LDH increase from 7.6 to 10.9 Å (Fig. 2), confirming the

206

intercalation of Fe(III)-TAML into the hydrotalcite interlayer. Taking into account the

207

thickness of hydroxide layer sheet of ~4.8 Å49, the interlayer distance can be

208

estimated as 6.1 Å when the adsorption reaches the plateau, which is smaller than the

209

length of Fe(III)-TAML molecule (7.7 Å) as indicated by the distance between the

210

two oxygen atoms in the carbonyl groups (22O and 24O, Fig. 1(b)). To demonstrate

211

the conformation change of Fe(III)-TAML molecule in clay interlayer, the orientation

212

of Fe(III)-TAML was simulated at different interlayer expansion of 3.5, 5.4, and 6.1 Å,

213

respectively (Fig. 3). It is clearly shown that at higher intercalation loading, the

214

Fe(III)-TAML molecule tends to arrange into tilted configuration to accommodate

215

more Fe(III)-TAML molecules (Fig. 3), similar results were also observed for

216

biomolecules adsorbed on clay minerals36,50,51.

217

The strong interaction between Fe(III)-TAML and LDH was further supported

218

from FTIR spectroscopic analysis (Fig. 4). To help interpret the IR peak shift,

219

theoretical spectrum of Fe(III)-TAML was calculated and compared with our

220

experimental results (Table S2). In accordance with the calculated IR bands, peak 10


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assignments for Fe(III)-TAML are listed in Table S2. The synchronous vibration of

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both C-N in-plane stretching (14C-11N and 13C-16N) is assigned to the peak at 1393

223

cm-1. The peak at 1455 cm-1 corresponds to the synchronous bending of hydrogen at

224

20C, 21C, 37C and 41C, while the 1475 cm-1 peak for the same bending at 46C and

225

50C. In addition, the peaks at 1575 and 1626 cm-1 are attributed to the C=O

226

(14C=22O and 13C=45O) in-plane stretching and HOH bending of water molecule

227

axially coordinated to Fe atom, respectively. Compared to the IR spectrum of free

228

Fe(III)-TAML, clear peak shifts were observed as Fe(III)-TAML is adsorbed on LDH.

229

The 1562 cm-1 peak results from the red shift of 1575 cm-1 peak, which is attributed to

230

the electrostatic interaction between the negatively charged oxygen atoms of carboxyl

231

groups and the positively charged LDH layers. This electrostatic effect indirectly

232

influences the C-N in-plane stretching, leading to the red shift from 1393 cm-1 in

233

Fe(III)-TAML to 1386 cm-1 in Fe(III)-TAML/LDH. The 1459 and 1477 cm-1 peaks,

234

corresponding to C-H bending, appear in Fe(III)-TAML/LDH with less obvious shifts.

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The 1626 cm-1 peak in Fe(III)-TAML is not observable due to the overlap with the

236

HOH bending band for water molecule associated with LDH. Compared to the 1630

237

cm-1 peak in LDH, a blue shift occurs for the HOH bending of water molecule

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adsorbed on LDH (1641 cm-1), which could be accounted for the interactions of H2O

239

molecule in LDH interlayer with O and N atoms of acylamino groups in

240

Fe(III)-TAML via hydrogen bonding. The occurrence of new peaks and the IR shifts

241

demonstrate the interaction between Fe(III)-TAML anions and positively charged

242

LDH layers, e.g. electrostatic and hydrogen bonding interaction. 11



243

Degradation of TBBPA by Fe(III)-TAML and Fe(III)-TAML/LDH at different

244

pHs. The degradation of TBBPA catalyzed by Fe(III)-TAML and Fe(III)-TAML/LDH

245

shows strong pH dependent behavior (Fig. 5). As listed in Table S3, the fitted

246

pseudo-first-order rate constant is significantly enhanced as pH increases to 9 or 10

247

compared to that at pH 8. The highest degradation rate was observed at pH 10 with

248

complete degradation of TBBPA in less than 20 s, similar results were also observed

249

in the literature22,26,29. The high reactivity of Fe(III)-TAML under alkaline condition

250

can be explained by the increasing proportion of deprotonated species of

251

Fe(III)-TAML at higher pH. During the catalytic process, Fe(III)-TAML is initially

252

oxidized by H2O2 to form Fe(IV)-TAML52 or Fe(V)-TAML53, which can be then

253

quickly reduced by organic pollutants. However, different species of Fe(III)-TAML

254

shows distinct reactivity to form oxidized Fe(IV)-TAML or Fe(V)-TAML. As listed in

255

Scheme S1, the deprotonated species [Fe(III)-TAML(OH)]2- with higher electron

256

density compared to [Fe(III)-TAML(OH2)]- would act as a better electron donor to be

257

oxidized by H2O2 at a much faster reaction rate (k2 ≫ k1)28. With the similar

258

explanation, [Fe(III)-TAML(OH)]2- is more prone to react with H2O2 than the

259

deprotonated HO2- (k4 < k2), in addition, this reaction might also be affected by

260

electrostatic repulsion28. Therefore, the ionization for both the axial water molecule in

261

the coordination sphere of Fe(III)-TAML and H2O2 would play vital roles for the

262

generation of oxidized Fe(III)-TAML and subsequent degradation of pollutants. As

263

reported in prior studies, the dissociation constants (pKa) for water associated with

264

Fe(III)-TAML and H2O2 are ~10 and >11, respectively28,32, which indicates that the 12


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highest reactivity for Fe(III)-TAML would be in a narrow range between pH 10 and

266

11.

267

Compared to Fe(III)-TAML, the pseudo-first-order rate constants for TBBPA

268

degradation increase in the presence of Fe(III)-TAML/LDH, especially at pH 8 (Table

269

S3). To further investigate the underlying mechanism, the adsorption of TBBPA on

270

Fe(III)-TAML/LDH was studied. As illustrated in Figure S4, the TBBPA adsorption

271

on Fe(III)-TAML/LDH can be fitted with Langmuir isotherm model (Table S4), with

272

the maximum adsorption of 76.38, 107.76 and 157.88 µmol g-1 for pH at 8, 9 and 10,

273

respectively. The pH-dependent adsorption is attributed to the extent of ionization

274

forms for TBBPA at different pHs. As simulated by Visual MINTEQ, with the two

275

pKa values of 7.4 and 8.5, the proportion of TBBPA2- increases from 20.96% to 97.35%

276

as increase of pH from 8 to 10, whereas the percentage of TBBPA- decreases from

277

60.49% to 2.65% (Table S5). It indicates that at higher pH, TBBPA exists as a more

278

negatively charged ion, which would be more facile to be adsorbed by LDH through

279

an ion exchange process. The strong adsorption of TBBPA on LDH could explain the

280

enhancement of the degradation rate catalyzed by Fe(III)-TAML/LDH. Furthermore,

281

as shown in Figure S5, the adsorption of TBBPA on LDH is fast, after 4 s, the ratios

282

of adsorption amount (Qt) to the equilibrium adsorption amount (Qe) could reach 22%,

283

29% and 32% for pH 8, 9 and 10, respectively. According to Fick’s law, the diffusate

284

concentration gradient is the driving force for mass transfer process54. Due to the high

285

reaction rate between TBBPA and active Fe(III)-TAML, the concentration of TBBPA

286

in LDH interlayer decreases so fast that a great concentration difference is generated 13



287

between LDH and homogeneous solution, therefore, a high diffusion rate would be

288

expected during the reaction period. To further quantitatively investigate the diffusion

289

process, the total amount of TBBPA remained (including both TBBPA in solution and

290

adsorbed on LDH surface) and the amount of TBBPA in solution were separately

291

measured by addition of methanol to extract the adsorbed TBBPA and by simple

292

centrifugation, respectively. The amount of TBBPA adsorbed on LDH could be

293

obtained by calculating the difference between these two measurements. As shown in

294

Figure S6, the amount of TBBPA adsorbed on LDH maintains in the range of 33-55

295

µmol g-1 in the first 30 s reaction period at pH 8, which is even higher than that (32

296

µmol g-1) after 120 s for adsorption kinetics of TBBPA on Fe(III)-TAML/LDH (inset

297

of Figure S6). The fast TBBPA adsorption on LDH provides direct evidence that the

298

diffusion of TBBPA onto LDH may not limit the reaction between TBBPA and

299

Fe(III)-TAML. In addition, Fe(III)-TAML/LDH with different Fe(III)-TAML loadings

300

of 4.42, 6.96, 8.36, 11.92, and 22.38 µmol g-1 were prepared to degrade the same

301

amount of TBBPA (18.4 µM). As illustrated in Figure S7, the pseudo-first-order

302

reaction rate constants are linearly proportional to the amount of Fe(III)-TAML

303

loaded on LDH, indicating that the amount of Fe(III)-TAML rather than mass transfer

304

dominated the surface reaction. So, it can be concluded that the adsorption of TBBPA

305

on LDH would not be a rate limiting step to slow down the degradation reaction,

306

instead it would enhance the dissipation of TBBPA by increasing the contact

307

possibility between TBBPA and TAML activator.

308

The most significant enhancement for TBBPA degradation efficiency by 14


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Fe(III)-TAML/LDH occurred at pH 8 (Table S3), which is not consistent with the

310

adsorption results (Fig. S4). At pH 10, the predominant species of TBBPA is TBBPA2-,

311

and the increase of degradation rate for Fe(III)-TAML/LDH is 1.58 compared to free

312

Fe(III)-TAML (Table S3), so the contribution of adsorption for the increase of

313

reactivity is < 2. Whereas, when pH increases from 8 to 9, the degradation rate

314

increases 15 times for free Fe(III)-TAML (Table S3). Therefore, the pH effect on the

315

reactivity is more pronounced. It has been reported that the basicity in clay interlayer

316

may increase resulting from the bicarbonate hydrolysis and the presence of Mg-O

317

Lewis base sites on MgAl LDH materials55. In the pH range between 8 and 10,

318

bicarbonate is the predominant species56. The presence of bicarbonate can be

319

evidenced by the strong band at 1370 cm-1 (Fig. 4), which is assigned to the

320

antisymmetric v3 vibration of bicarbonate anion57. The exposure to air during LDH

321

preparation inevitably introduces carbon dioxide into the system. On the other hand,

322

the Mg-O basic sites on MgAl LDH could also enhance the basicity55. Therefore, it

323

can be concluded that the pH in clay interlayer is higher than that in homogeneous

324

solution58. Compared to pH 9 and 10, the contributions from bicarbonate and Mg-O

325

sites are more pronounced at pH 8, since the change of OH- level would be more

326

significant at neutral or near neutral pH, which might account for the higher increase

327

for interlayer pH and higher reactivity for Fe(III)-TAML/LDH system at pH 8. It was

328

summarized by DeNardo et al.59 that the rate constants for both Fe-TAML activation

329

by H2O2 and substrate oxidation by reactive Fe-TAML are positively relevant to

330

solution pH, indicating that both rate constants increase with the increase of pH levels. 15



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331

Therefore, it is likely that the main utility of Fe(III)-TAML/LDH is to increase the

332

activity at lower solution pH by raising the interlayer pH.

333

To investigate the degradation efficiency of TBBPA by Fe(III)-TAML/LDH, both

334

mineralization and debromination during the degradation process were studied. A

335

significant increase of debromination rate was observed for Fe(III)-TAML/LDH

336

system compared to free Fe(III)-TAML (Fig. S8 and Table S6), which can also be

337

explained by the strong interaction between TBBPA and LDH. As reported by Gupta

338

et al.22, Fe(III)-TAML could achieve nearly complete mineralization and

339

dechlorination for chlorophenols. However, only limited debromination was observed

340

in our experiment. As shown in Figure S8, at 120 s, the debromination ratios are

341

15.66, 23.07 and 30.92% for Fe(III)-TAML at pH 8, 9, and 10, respectively, while the

342

respective

343

Fe(III)-TAML/LDH. When the reaction is extended to 1800 s, similar debromination

344

ratios are also achieved for Fe(III)-TAML (21.1%) and Fe(III)-TAML/LDH (22.1%)

345

at pH 8 (Fig. S8(a)). The low oxidative debromination for TBBPA might be attributed

346

to the generation of the brominated organic carboxylic acids (Fig. 6), these

347

compounds are reported to be resistant to various oxidants (e.g. hydroxyl radicals and

348

ozone)21. The strong pH dependence for debromination rate and efficiency also results

349

from varied concentration of the most reactive species [Fe(III)-TAML(OH)]2- under

350

different pH values.

debromination

values

are

21.85,

23.41,

and

30.19%

for

351

Due to the low aqueous solubility of TBBPA under acidic condition60, after addition

352

of perchloric acid to terminate the reaction, TBBPA would precipitate, which might 16


Page 17 of 39


353

interfere with TOC measurement. Therefore, the mineralization efficiency during the

354

reaction was determined by calculating the difference between the initial and final

355

TOC (0 and 120 s) in reaction solution. As illustrated in Figure S9(a), TOC removals

356

after 120 s reaction are 29.22, 38.81 and 43.90% for Fe(III)-TAML under pH 8, 9,

357

and

358

Fe(III)-TAML/LDH. Similar results were also observed when the reaction was fully

359

completed (1800 s) (Fig. S9(b)). The results from debromination and mineralization

360

experiments show that in Fe(III)-TAML/LDH system, even there is no significant

361

increase for total debromination, higher TOC removal was observed compared to

362

Fe(III)-TAML. The discrepancy might be explained that during the experiment, the

363

reaction intermediates can be adsorbed on LDH, which could promote the reaction to

364

proceed, resulting in higher mineralization efficiency. On the other hand, these

365

brominated byproducts may accumulate on LDH external surfaces to retard further

366

debromination reaction. The adsorption of final reaction products on LDH could also

367

partially account for TOC removal. Therefore, the adsorption of TBBPA and the

368

reaction intermediates on LDH might affect the extent of degradation for TBBPA by

369

Fe(III)-TAML/LDH/H2O2 system, further study is needed to reveal the detailed

370

degradation mechanism by measuring the intermediates during the reaction and under

371

different reaction conditions.

10,

respectively,

and

increase

to

46.05,

54.33

and

62.30%

for

372

TBBPA degradation pathway. Based on the reaction products identified in our

373

experiment (Figs. S10 and S11) and previous studies17,21,61, the degradation pathway

374

of TBBPA is shown in Figure 6. The degradation is initiated by an electron 17



Page 18 of 39

375

deprivation reaction occurring on the hydroxyl group of TBBPA, generating the

376

phenoxy radical R1, which is then transformed to a more stable radical R2 by

377

resonance and leads to the relocation of the unpaired electron to the aromatic ring.

378

Subsequently, the carbocation intermediate radical R3 and a new radical R4 are

379

released by β-scission (cleavage between the isopropyl group and the benzene ring) of

380

R2. Similar results were also observed by previous researchers for TBBPA

381

degradation by manganese dioxide17, UV/Fenton reaction62 and ozonation21, they

382

proposed that the primary reaction includes the bond cleavage at the middle carbon

383

atom of TBBPA21. Substitution and deprotonation can transform the carbocation R3 to

384

4-(2-hydroxyisopropyl)-2,6-dibromophenol

385

4-isopropylene-2,6-dibromophenol (R5), respectively. Meanwhile, P1 can be

386

dehydrated

387

1-(3,5-dibromo-4-hydroxyphenyl)ethanone (P2, MW = 294). Although R5 was not

388

detected in our study, probably due to the rapid transformation upon contact with

389

Fe(III)-TAML, it has been proposed as the intermediate during the oxidation of

390

TBBPA by Lin et al.17 Our results further demonstrate that the reaction intermediate

391

P2 completely disappears after 600 s (Figs. S10 and S11), with the final products of

392

(Z)-4-acetyl-2-bromopent-2-enedioic acid (P3, MW = 251), CO2 or CO, and bromide

393

ion. Since product P3 has not been reported in the literature, the structure of P3 can be

394

only deduced from the information of mass fragments (Figs. S10 and S11). Similarly,

395

R4 may also be transformed to 2-bromosuccinic acid (P4, MW = 197), CO2 or CO,

396

and bromide ion by the oxidation of Fe(III)-TAML. In summary, the oxidative

to

form

R5,

which

(P1,

is

MW

further

18


=

310)

oxidized

and

to

Page 19 of 39


397

degradation of TBBPA is assumed to proceed via the cleavage of the middle carbon

398

atom of TBBPA, generating a carbocation intermediate R3 and a radical R4, which

399

are finally transformed to P3 and P4. The proposed degradation pathway is further

400

supported from the theoretical calculations. The calculated FED result for each atom

401

of TBBPA is listed in Figure S12(a), and the three-dimensional iso-surfaces of HOMO

402

(Fig. S12(b)) and LUMO (Fig. S12(c)) frontier orbitals are also exhibited for

403

visualization of the electron densities. As shown in Figure S12(a), 20O and 21O of

404

TBBPA have the highest 2FED2HOMO values (0.1558, 0.1558), suggesting that 20O or

405

21O should be the most possible reaction site, where one electron is extracted by

406

Fe-TAML, forming the TBBPA radical. It is consistent with the initial step in the

407

degradation pathway (Fig. 6). In addition to 20O and 21O, high 2FED2HOMO values

408

for 3C (0.1377) and 10C (0.1376) are also observed among all the C atoms of TBBPA,

409

indicating that these two sites are susceptible for attack by the reactive species of

410

Fe-TAML, which is confirmed by the reaction products, e.g., P1 and P2.

411

Reusability of Fe(III)-TAML/LDH. One of the advantages for immobilization of

412

bioactive materials is to increase their reutilization, hence in current study, we also

413

tested the reusability of the novel Fe(III)-TAML/LDH composite for degradation of

414

TBBPA. As illustrated in Figure S13(a), the activity of Fe(III)-TAML/LDH could be

415

retained for at least three cycles. The removal percentages of TBBPA are 100.0, 100.0,

416

44.6 and 0.4% for the first, second, third and fourth cycle, respectively, after 300 s

417

reaction for each cycle. The decrease of the reactivity for Fe(III)-TAML/LDH could

418

be attributed to the suicidal inactivation of Fe(III)-TAML58 and the release of 19



419

Fe(III)-TAML from LDH. Based on the results of stability tests for both

420

Fe(III)-TAML and Fe(III)-TAML/LDH (Fig. S14), the high activity of free

421

Fe(III)-TAML could be retained for four cycles, then decreases significantly, whereas,

422

the Fe(III)-TAML/LDH composite shows high reactivity until six cycles. The

423

decrease of activity can be explained by the suicidal inactivation of Fe(III)-TAML as

424

the substrate was quickly consumed during the reaction59,63. However, the presence of

425

LDH could increase the concentration of TBBPA in clay interlayer, which may

426

prolong the lifetime of Fe(III)-TAML63. More importantly, we found that

427

Fe(III)-TAML was released from LDH during the reaction by detecting the Fe in

428

solution, and the sum of Fe after each reaction cycle was 49.16 µg L-1 (0.88 µM) (Fig.

429

S13(b)), which is close to the total molar concentration of Fe(III)-TAML (1.0 µM)

430

used in the system. The release of Fe(III)-TAML could be due to the strong

431

competition for the adsorption sites in LDH interlayer with bromide ion (Br-), reaction

432

intermediates, e.g., carboxylic acids (P3 and P4), hydroxyl ion (OH-) and

433

bicarbonate/carbonate ion (HCO3-/CO32-). For example, at pH 10, the concentration of

434

OH- is 100 µM, which is 2 order of magnitude higher than Fe(III)-TAML.

435

Acute Toxicity. To demonstrate the “green catalyst” nature of Fe(III)-TAML, the

436

toxicity test was also conducted to assess the potential risk of reaction intermediates

437

produced from TBBPA degradation. Photobacterium phosphoreum (P. phosphoreum)

438

was used for the acute toxicity assay. As indicated in Figure S15, before the reaction,

439

the initial inhibitions of TBBPA mixed with Fe(III)-TAML or Fe(III)-TAML/LDH are

440

over 30%, after 600 s reaction period, the inhibition values decrease to less than 0.9%. 20


Page 20 of 39

Page 21 of 39


441

Significant decrease in the toxicity towards P. phosphoreum could be attributed to the

442

complete decomposition of TBBPA and the lower toxicity of the reaction products. In

443

addition, based on our observation, the toxicities of Fe(III)-TAML and

444

Fe(III)-TAML/LDH themselves towards P. phosphoreum are negligible (Fig. S15),

445

which is consistent with the results in the literatures.22,64 Gupta et al. reported that the

446

LONEC values (the highest observed concentrations of activator that show no

447

bacterial death) for Fe(III)-TAML activators were greater than 30 mg L-1 (~50 µM)

448

based on the luminescent bacteria test22. Truong et al.64 used the embryonic zebrafish

449

to assess the developmental toxicity of Fe(III)-TAML, observing no significant

450

mortalities and morphological malformations for embryonic zebrafish even exposed

451

to 250 µM Fe(III)-TAML.

452

Environmental Implications

453

In this study, a new Fe(III)-TAML/LDH composite was synthesized by

454

intercalating Fe(III)-TAML into the LDH interlayer via anion exchange. Both FTIR

455

and XRD spectra indicate that Fe(III)-TAML can form strong interaction with LDH,

456

and with the increase of adsorption, Fe(III)-TAML molecules tend to exist as the tilted

457

configuration. Compared to the degradation of TBBPA by Fe(III)-TAML,

458

Fe(III)-TAML/LDH showed significantly improved degradation efficiency, including

459

faster elimination of TBBPA and higher TOC removal. It could be attributed to the

460

strong adsorption of TBBPA onto LDH, which would increase the contact possibility

461

between TBBPA and Fe(III)-TAML.

462

Fe(III)-TAML has been considered as a green catalyst for rapid degradation of

463

many persistent organic contaminants with high inherent reactivity and low

464

environmental risks27,30,33. Our results indicate that the combination of Fe(III)-TAML 21



465

with LDH provides a new attempt to not only promote the degradation rate, but also

466

improve its reusability, which would extend the application of Fe(III)-TAML in a

467

wider range of pH for advanced treatment of recalcitrant pollutants with lower cost.

468 469

470

Acknowledgements

471

This work was financially supported by the National Key Basic Research Program of

472

China (2014CB441102), National Science Foundation of China (grants 21222704,

473

21237002 and 21477051) and the Collaborative Innovation Center for Regional

474

Environmental Quality. We thank the Analytical Center and High Performance

475

Computing Center of Nanjing University for the characterization of samples and

476

computational study.

477 478

Supporting Information

479

This material is available free of charge via the Internet at http://pubs.acs.org.

480

Texts S1-S8, Tables S1-S6, Scheme S1, and Figures S1-S15

481 482

References:

483

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Analytical and environmental aspects of the flame retardant tetrabromobisphenol A

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(2) Alaee, M.; Arias, P.; Sjödin, A.; Bergman, Å. An overview of commercially used

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716 717 718 719 720 721 722 723 724 725 726 727 728 729 730 731 732 733 734 735 736 737 738 739 740 741 742 743 744 745 746 747 748 749 750 751 752

Figure Legends Figure 1. The molecular structure (a) and optimized geometry (b) of the prototype Fe(III)-TAML activator. The optimized calculation was conducted at B3LYP/6-311G (d,p) level. All atoms except hydrogen are symbolled and labeled. Figure 2. X-ray diffraction patterns of LDH with Fe(III)-TAML loadings of (a) 0 µmol g-1, (b) 2.86 µmol g-1, (c) 12.32 µmol g-1, (d) 22.38 µmol g-1, (e) 87.15 µmol g-1 and (f) 176.82 µmol g-1, respectively. The basal spacings of LDH increase as 2 theta values decrease. Figure 3. Simulation of the existing configuration of Fe(III)-TAML anion in LDH interlayer with Fe(III)-TAML loadings of (a) 22.38 µmol g-1, (b) 87.15 µmol g-1 and (c) 176.82 µmol g-1. The Fe(III)-TAML molecule tends to arrange into tilted configuration as the increasing accommodation of more Fe(III)-TAML molecules. The distance between 22O and 24O is 7.7 Å. Figure 4. Calculated and experimental IR spectra: calculated IR spectrum of (a) Fe(III)-TAML; observed IR spectra of (b) Fe(III)-TAML in solid phase, (c) Fe(III)-TAML/LDH powder, and (d) dried synthesized LDH. The experimental IR spectra were all obtained using KBr pellet technique. Figure 5. Degradation kinetics of TBBPA catalyzed by Fe(III)-TAML and Fe(III)-TAML/LDH at (a) pH 8, (b) pH 9 and (c) pH 10. Lines represent the fitted pseudo-first-order kinetic curves for different pH levels. The determined coefficients of determination from the overall regression analysis R2 are greater than 0.96. Experimental conditions: the initial concentrations of TBBPA, Fe(III)-TAML, H2O2, and LDH were 18.4 µM, 1.0 µM, 2.0 mM and 0.045 g L-1, respectively. The Fe(III)-TAML/LDH with Fe(III)-TAML loading of 22.38 µmol g-1 was chosen to degrade TBBPA. To avoid any pH fluctuation, both the pHs of TBBPA/Fe(III)-TAML/LDH solution and H2O2 stock solution were pre-adjusted to desired pH by addition of 0.1 M NaOH or 0.1 M HClO4 solution. Error bars represent standard deviations (n = 3). Figure 6. Proposed catalytic pathway of TBBPA by Fe(III)-TAML/H2O2. Molecular ion clusters obtained from mass spectra are shown below their structures, respectively.

33



Figure 1.

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Figure 2.

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Figure 3.

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(d) LDH 1630 1370

Relative Intensity

(c) Fe(III)-TAML/LDH 1370

1459

1386

1641

1477 1562

(b) Fe(III)-TAML

1393 1626 1455 1575 (a) DFT: Fe(III)-TAML 1475 1513 1439 1379

1000

1200

1400

1636 1597

1600

1800 -1

Wavenumber (cm )

Figure 4.

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2000


1.2

Fe(III)-TAML Fe(III)-TAML/LDH Model Fit for Fe(III)-TAML Model Fit for Fe(III)-TAML/LDH

1.0 0.8 Ct/C0

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pH = 8 0.6 0.4 0.2 0.0 0

20

40

1.2

100

120


1.0 0.8 Ct/C0

60 80 Time (s) (a)

pH = 9 0.6 0.4 0.2 0.0 0

10

20

1.2

50

60


1.0 0.8 Ct/C0

30 40 Time (s) (b)

pH = 10 0.6 0.4 0.2 0.0 0

5

10

15 20 Time (s) (c)

25

Figure 5.

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Figure 6.

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Rapid Destruction of Tetrabromobisphenol A by Iron(III) - Doi.org

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