Efficient Sorption and Removal of Perfluoroalkyl Acids (PFAAs) from

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Environmental Science & Technology

236x139mm (96 x 96 DPI)

ACS Paragon Plus Environment


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Efficient Sorption and Removal of Perfluoroalkyl Acids (PFAAs) from Aqueous Solution by Metal Hydroxides Generated in situ by Electrocoagulation Hui Lina, b, Yujuan Wanga, Junfeng Niua*, Zhihan Yuea, Qingguo Huangb*

4 5 a

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Normal University, Beijing 100875, P.R. China

7

8 9

State Key Laboratory of Water Environment Simulation, School of Environment, Beijing

b

College of Agricultural and Environmental Sciences, Department of Crop and Soil Sciences, University of Georgia, Griffin, GA 30223, United States

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Hui Lin, E-mail: [email protected]

11

Yujuan Wang, E-mail: [email protected]

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Junfeng Niu, E-mail: [email protected]

13

Zhihan Yue, E-mail: [email protected]

14

Qingguo Huang, E-mail: [email protected]

1


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Abstract. Removal of environmentally persistent perfluoroalkyl acids (PFAAs), i.e., perfluorooctane

16

sulfonate (PFOS) and perfluorocarboxylic acids (PFCAs, C4~C10) were investigated through

17

sorption on four metal hydroxide flocs generated in situ by electrocoagulation in deionized water

18

with 10 mM NaCl as supporting electrolyte. The results indicated that the zinc hydroxide flocs

19

yielded the highest removal efficiency with a wide range concentration of PFOA/PFOS (1.5 μM ~

20

0.5 mM) at the zinc dosage < 150 mg L-1 with the energy consumption < 0.18 Wh L-1. The sorption

21

kinetics indicated that the zinc hydroxide flocs had an equilibrium adsorbed amount (qe) up to

22

5.74/7.69 mmol g-1 (Zn) for PFOA/PFOS at the initial concentration of 0.5 mM with an initial

23

sorption rate (v0) of 1.01×103/1.81×103 mmol g-1 h-1. The sorption of PFOA/PFOS reached

24

equilibrium within < 10 min. The sorption mechanisms of PFAAs on the zinc hydroxide flocs were

25

proposed based on the investigation of various driving forces. The results indicated that the

26

hydrophobic interaction was primarily responsible for the PFAAs sorption. The electrocoagulation

27

process with zinc anode may have a great potential for removing PFAAs from industrial wastewater

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as well as contaminated environmental waterbody.

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Keywords: Perfluoroalkyl acids (PFAAs); electrocoagulation; zinc hydroxide flocs; sorption

31

mechanisms

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Introduction

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Perfluoroalkyl acids (PFAAs) are a class of anthropogenic organofluorine compounds which

34

have a fully fluorinated alkyl chain of varying length with an acid headgroup such as sulfonic,

35

carboxylic, or phosphonic (1). Since 1950s, these compounds have been used extensively in a wide

36

range of industrial and medical applications, such as emulsification in fluoropolymer manufacturing,

37

foam forming in firefighting, and surface treatment in textile and semiconductor products (2, 3). The

38

widespread usage of these chemicals, in combination with their high environmental persistence, has

39

resulted in their frequent detections in various environmental and biological matrices, such as waters,

40

air, sediments, dusts, human blood, and wildlife (4-6). Moreover, perfluorooctanoic acid

41

(C7F15COOH, PFOA) and perfluorooctane sulfonate (C8F17SO3H, PFOS) have been considered to be

42

probable carcinogens (7).

43

PFAAs can enter the water environment during manufacturing processes, supply chains,

44

product use, and disposal of various industrial and consumer products (8, 9). Previous studies

45

demonstrated that the direct point source emission containing very high concentrations of PFAAs

46

was the main pathway of these chemicals releasing to the environment (2, 10). For example, the

47

typical concentration of PFOA in the untreated wastewater after emulsifying process in

48

fluoropolymer manufacturing plant was 0.34~3.35 mM (11). Historically, effluents from PFAA

49

production and usage were neither impounded nor pretreated prior to discharging to water treatment

50

systems or the environment, resulting in the contamination of groundwater and soil (12). For

51

example, PFAAs were used as surfactants in aqueous fire-fighting foams (AFFFs) that were

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extensively employed in U.S. military firefighting. Recent studies showed that the concentration of

53

PFAAs collected from the polluted groundwater ranged from several μg L-1 to a few mg L-1 (13, 14).

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Therefore, it is of great importance to develop effective techniques to eliminate PFAAs from

55

wastewater before they are discharged to the environment and remediate the environmental 3


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waterbody contaminated by PFAAs.

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Many treatment methods including photochemical oxidation (11), electrochemical oxidation

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(15, 16), and ultrasonic irradiation (17) were developed to degrade PFAAs in aqueous solution.

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However, these technologies are generally limited by their harsh treatment conditions, complicated

60

operation, chemical additives, low mineralization extents, low energy efficiency, or high operation

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costs (18). Therefore, the application of these technologies to treat large volumes of diluted PFAA

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wastewaters is not technically and economically feasible.

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It has been demonstrated in many cases that sorption may provide an effective option to remove

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PFAAs from aqueous streams at various concentrations. Many adsorbents including activated carbon

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(AC), carbon nanotubes (CNTs), resin, mineral material, biomaterials, and molecularly imprinted

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polymers have been evaluated (19, 20). Herein, granular activated carbon (GAC) is the most widely

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used sorbent in water purification, but it only exhibited limited sorption capacity of less than 0.4

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mmol g-1 for both PFOS and PFOA (21). Anion-exchange resins have relative high sorption

69

capacities for PFOS and PFOA, but their sorption rates were extremely slow with a

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pseudo-second-order kinetics constant of 10-5~10-4 g mg-1 h-1 (21). The sorbents having low sorption

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capacities or rates are easy to be penetrated, leading to failure in application. For instance, rapid

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penetration of perfluorinated surfactants through GAC filters in drinking water treatment plants was

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observed (22). Similar observations were also reported with trial ion-exchange resin column, the

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breakthrough of PFOA was reached at only about 45 bed volumes and the PFOA concentration rose

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up to influent levels (8 mg L-1) at less than 300 bed volumes, far below its sorption capacity (23).

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Moreover, the sorption capacity and sorption rate can significantly decrease in the presence of

77

effluent organic matter, further reducing the treatment efficiency in real-life scenarios (24). In

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addition, the used AC cannot be easily regenerated even by organic solvent wash and safe disposal

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of the spent sorbents is required (20). These limitations heightened a great interest in the 4



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development of novel cheap sorbents or techniques with fast sorption rate and high sorption capacity

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to remove PFAAs from aqueous solution.

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It is known that the metal hydroxide flocs formed during coagulation can strongly sorb certain

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pollutants and remove them from water (25). Electrocoagulation (EC) is very efficient in this regard

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because the metal hydroxide flocs are freshly formed in situ (25). Freshly formed amorphous metal

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hydroxide flocs such as aluminum and ferric are fractal and highly porous aggregates made up of

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many primary particles. These “sweep flocs” have large surface areas, which are beneficial for a

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rapid adsorption of soluble organic compounds and trapping of colloidal particles. To date, the

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literature is very limited and somewhat contradictory with respect to the sorption of PFAAs on Al/Fe

89

flocs during water treatments (26-29). Some recent investigations found that the coagulation unit in

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water treatment systems was inefficient in removing PFAAs (26, 27). Deng et al. (28) found that

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90% PFOA could be removed from aqueous solution during the coagulation process by using

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polyaluminum chloride (Al2O3, 29%) at a Al dosage of 10 mg L-1, primarily through the removal of

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PFOA-sorbed suspended solids, whereas, the formed aluminum hydroxide flocs alone were

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ineffective in removing soluble PFOA. However, Xiao et al. (29) found that 10~40% removal of

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PFOA/PFOS could be achieved with an enhanced coagulation at higher coagulant dosages (60~110

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mg L-1). The primary mechanism was sorption of PFOA/PFOS on the fine flocs rather than the

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co-removed turbidity particles, and the maximum sorption capacity was achieved at 2 min during the

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initial stage of coagulation. There was also a report on EC with aluminum anodes to purify

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fire-fighting foams containing fluorinated surfactants (30). To the best of our knowledge, there is no

100

report on systematically evaluating the removal efficiency and mechanisms of PFAAs in aqueous

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solution by the EC process.

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We reported here the rapid removal of PFAAs including PFOS, PFOA and other hydrophobic

103

perfluorocarboxylic acids (PFCAs, C7~C10) from aqueous solution by EC with high sorption 5


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capacities and rates. The primary objectives of this study were to investigate the removal efficiencies

105

of PFAAs by the EC process and explore the sorption mechanisms. This work focused on evaluating

106

the removal abilities of various sacrificial anode materials, including aluminum, iron, zinc, and

107

magnesium. Batch experiments were conducted to investigate the sorption kinetics and isotherms of

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PFAAs on metal hydroxides flocs generated in situ by EC. Furthermore, the possible mechanisms

109

were proposed by employing the possible interactions between PFAAs and metal hydroxide flocs.

110

Theoretical and Experimental

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Materials. All chemicals used in the experiments were reagent grade or higher and used as

112

received. Perfluorobutanoic acid (PFBA, 98%), perfluoheptanoic acid (PFHpA, 98%), PFOA (98%),

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PFOS (98%), perfluorononanoic acid (PFNA, 98%), and perfluorodecanoic acid (PFDA, 98%) were

114

from Sigma-Aldrich Chemical Co., Ltd. (St. Louis, MO, USA), and their properties are listed in

115

Table S1 of the Supporting Information (SI). The internal standard 13C4-PFOA and 13C8-PFOS were

116

obtained from Wellington Laboratories (Guelph, ON, Canada). Sodium chloride (NaCl) and

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ammonium acetate (CH3COONH4) were obtained from Sinopharm (Beijing, China). Milli-Q

118

(deionized, DI) water with conductance of 18.2 MΩ cm at 25 ± 1 ºC was prepared by a Millipore

119

water system and used in all experiments.

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Cell Construction and Experiments. Two different EC reactors were used to evaluate the

121

removal efficiency of PFAAs from deionized water (DI) water with 10 mM NaCl as electrolyte in

122

the batch experiments. The EC reactor (I) was a flat panel reactor with a 220 mL capacity (See

123

Figure S1 of the SI). Iron, magnesium, aluminum or zinc plate of 72 cm2 surface area acted as the

124

sacrificial anodes with a same dimension of 304 stainless steel plate as the cathode. The gap between

125

the two electrodes was fixed at 3.0 cm. In each run, an aqueous solution (180 mL) of 0.5 mM PFOA

126

was added into the cell and stirred continuously using a magnetic stirrer (IKA-RCT, Germany) with

127

the stirring rate of 800 r min-1. 6



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The EC reactor (II) was composed of a cylindrical EC cell (4 cm radius and 10 cm height) made

129

of organic glass with a 500 mL volume, and was equipped with an air stirring device (See Figure S1

130

of the SI). Once zinc hydroxide flocs has been identified as having the best anode for PFOA sorption,

131

the experiments were continued with the reactor (II) setup with zinc anode only. A zinc sheet of 200

132

cm2 surface area and 0.1 mm width was used as anode, while a 304 stainless steel rod of 3 mm

133

diameter was used as cathode. In each run, an aqueous solution (400 mL) of varying concentrations

134

of PFAAs was added into the cell, the solution initial pH was adjusted using 3 M HCl or NaOH.

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Subsequently, the EC system was operated in a batch mode at a constant current density with proper

136

agitation. In all cases, a direct current was supplied by a DC regulated power source (Beijing Dahua

137

Radio Instrument, China). Samples were taken at different time intervals and then filtered by 0.22

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μm syringe filters (acetate membrane). The recovery rates of PFOA/PFOS were more than 95% by

139

using this acetate membrane filter (See Figure S2 of the SI). All tests were triplicated and carried out

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at room temperature (25 ±1 oC).

141

Instrumental Analysis. High concentrations (mg L-1 level) of PFCAs (C4~C10) were analyzed

142

by high performance liquid chromatography (HPLC-UV, Dionex U3000, USA) equipped with a

143

WpH C18 column (4.6 mm × 250 mm, 5 μm). The following operating conditions were employed:

144

isocratic elution with acetonitrile / 20 mM NaH2PO4 (pH = 2) (50/50, v/v) at a flow rate of 1 mL

145

min-1, a sample injection volume of 25 μL, and a UV wavelength of 210 nm for the detector. More

146

details about the analytical method could be found in our previous study (31).

147

The concentration of PFOS including linear and branched isomers and μg L-1 level

148

concentrations of PFOA were measured using a LC system coupled with a triple-stage quadrupole

149

mass spectrometer (LC-MS/MS, API3200; Applied Biosystems, USA). The analysis was carried out

150

in multiple reaction monitoring (MRM) mode. Electrospray ionization (ESI) was operated in a

151

negative mode with the parameters set as capillary potential at -4.5 kV, source temperature at 120 °C 7


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and desolvation temperature at 450 °C. The pressure of sheath gas (N2) was 0.4 MPa. Each sample

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analyzed by LC-MS/MS was spiked with 5 mM

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The analytical method of PFAAs has been described in detail previously (31, 32). The concentration

155

of Zn ions in aqueous solution was measured by an inductively coupled plasma atomic emission

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spectrometer (ICP-AES, SPS 8000; Sea Light, Co., China) with a detection limit of 0.18 μg L-1.

13

C4-PFOA or 13C8-PFOS as the internal standard.

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Flocs Characterization. The zeta potentials of zinc hydroxide flocs at different

158

electrocoagulation time were measured with a zeta potential analyzer (Zetasizer Nano ZS90,

159

Malvern Instruments, UK). At the end of the run, the metal hydroxide flocs suspensions were filtered

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through a glass fiber filter of 0.22 μm pore size (Whatman, UK), and the flocs were then freeze-dried.

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The Brunauer-Emmett-Teller (BET) surface areas of the dried flocs were determined by nitrogen

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physisorption using an Autosorb-iQ system (Quantachrome, USA). Before each measurement, the

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sample was degassed at 60 °C for 12 h. The fourier transform infrared spectrum (FTIR) of the flocs

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was obtained in the frequency range of 400~4000 cm-1 using the Perkin-Elmer spectrum GX FTIR

165

spectrometer. The crystal structure of the flocs were analyzed using X' Pert Pro MPD (Panalytical

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Co., Holland) X-ray diffraction (XRD) with Cu Kα radiation at 40 KV/40 mA, and each sample was

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scanned from 5°to 80°. X-ray photoelectron spectroscopy (XPS) spectra of the flocs were measured

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by an ESCALAB 250Xi XPS system (Thermo Scientific Ltd, USA), using a monochromatic Al Kα

169

source.

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Theoretical Metal Dissolved Dosage and Energy Consumption Calculation. The electrical

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energy consumption (EEC) was calculated in terms of Wh L-1 of treated effluent using the equation

172

given below: EEC =

173

UI ×t V EC

(1)

174

where U is the average cell voltage (V), I is the input current (A), tEC is the EC treatment time (h),

175

and V is the volume (L) of effluents. 8



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The theoretical metal dissolved dosage (ω, mg L-1) in solution was calculated as a function of electrocoagulation time using the following equation: ω=1000

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It EC M ×η nFV

(2)

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where M is the relative molar mass of the metal concerned, n is the number of electrons in

180

oxidation/reduction reaction. F is the Faraday’s constant, 96 500 C mol-1. The current efficiency (η)

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of EC process was calculated using the following equation: η=

182

ΔM e x p ×1 0 0% ΔM t h e o

(3)

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where ΔMexp is the experimental measured amount of zinc dissolution during the EC process, ΔMtheo

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is the theoretical amount of zinc dissolution with the Faraday’s law with 100% current efficiency.

185 186

The adsorbed amount (qt, mmol g-1) of PFAAs on metal hydroxide flocs was calculated using the following equation: qt=1000

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C 0-Ct ω

(4)

where C0 and Ct are the PFAAs concentration in solution at initial and reaction time, respectively.

189 190

Results and Discussion

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Effect of Anode Materials. During EC process, soluble contaminants may be removed from

192

water mainly by the sorption on metal hydroxides generated in-situ from the sacrificial anodes. Since

193

the sorption efficiency is strongly dependent on the physical-chemical properties of the sorbents, this

194

study investigated four different sacrificial anode materials, including iron, aluminum, zinc, and

195

magnesium, which in-situ generated metal hydroxides of distinctive properties. As shown in Figure

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1a, the zinc anode was far more effective in removing PFOA than the other three anode materials.

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The removal efficiencies were 96.7%, 3.6%, 11.3%, and 10.6% for the zinc, magnesium, aluminum,

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and iron anodes, respectively, after 10 min of electrocoagulation. 9


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The dissolved amounts of different metal are different at the same electrocoagulation time. To

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illustrate it more clearly, the PFOA adsorbed amounts (qt) vs. time of different metal hydroxide flocs

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were calculated using Eq. 4. As shown in Figure 1b, the highest sorption amount of PFOA on

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aluminum hydroxide flocs is 1.76 mmol g-1 (Al), about 3 times that of iron and magnesium

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hydroxide flocs. An extremely high sorption capacity of PFOA (i.e., 4.71 mmol g-1 Zn), was

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achieved on zinc hydroxide flocs assuming a 100% Faraday current efficiency of these anodes. Thus,

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this study mainly focused on the sorption performances and removal mechanisms of PFAAs by the

206

zinc hydroxide in aqueous solution.

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The aluminum hydroxide flocs formed by coagulation (polyaluminum chloride) was ineffective

208

in removing soluble PFOA/PFOS, suggesting that the removal was caused by the sorption of

209

PFOA/PFOS on suspended solids (28). As shown in Figure 2, The FTIR spectra of adsorbed zinc

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hydroxide flocs showed significant adsorb vibration peaks of PFOA characteristic functional groups,

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i.e., CF2, CF3, and COO-, in the wave number region between 600~1800 cm-1. XRD spectra of zinc

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hydroxide flocs by 0.5 mM PFOS sorption also reflected distinct diffraction peaks of PFOS (see

213

Figure S3 of the SI). The PFOA solution after electrocoagulation was filtered through a vacuum

214

suction filter with a 0.22 μm glass fiber membrane (GF, Whatman, UK). The trapped zinc hydroxide

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flocs were collected and then dissolved by 0.1 M HCl solution. The results showed that the recovery

216

rate of the sorbed PFOA was 98.7 ± 3.2 % in triplicated experiments (for details, see Text S1 of the

217

SI). The PFOA/PFOS solution used in this study was prepared in DI water without suspended solids.

218

Therefore, the soluble PFOA/PFOS was certainly removed by sorption on the flocs.

219

Effect of Initial PFAA Concentration. The removal of PFOA/PFOS by EC was examined in

220

reactor (II) with a zinc anode. The initial concentrations ranged from 1.5 μM to 0.5 mM and the

221

initial pH was about 5. As illustrated in Figure 3a (Note: the theoretical Zn dosage was calculated by

222

Eq. 2, and the η value of zinc measured in this study was 127.5% ± 11.9), it is worth to note: (1) 10



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PFOA/PFOS can be quickly removed within 20 min over a wide concentration range (1.5 μM ~ 0.5

224

mM); (2) energy consumption were less than 0.18 Wh L-1; and (3) Zn dosage was relatively low, less

225

than 150 mg L-1 for complete removal of PFOA or PFOS. The results indicated that the EC method

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with zinc anode has a great potential for the removal of PFOA/PFOS from contaminated waters

227

across a wide concentration range. We also investigated the difference in the removal efficiencies

228

between linear structure PFOS (L-PFOS) and branched structure PFOS (M-PFOS). It is known that

229

the technical PFOS commonly used in industries is generally a mixture of both linear and branched

230

isomers (32). As shown in Figure 3b, the L-PFOS was removed more readily than M-PFOS.

231

Furthermore, an ICP-AES was employed to measure the zinc ion concentration. The results

232

showed that the residual zinc ion concentration was 0.88 mg L-1 after electrocoagulation. Zinc is an

233

essential semi-trace element and the drinking water ordinance limit of U.S. EPA is 5.0 mg L-1 for

234

zinc ion. It is thus safe to use zinc as the anode in EC process for water treatment.

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Sorption Kinetics and Isotherms. The sorption kinetics of PFOA/PFOS on zinc hydroxide

236

flocs are shown in Figure 3. It could be found from Figure 3 that the sorption proceeded rapidly and

237

the equilibrium was reached in less than 10 min. Similar result was also reported by Xiao et al. (29),

238

who found that the sorption of PFOA/PFOS by fine Al hydroxide flocs achieved equilibrium in 2

239

min during the initial stage of coagulation. In the EC process, the sorbent (zinc hydroxide flocs) was

240

produced gradually over time and the PFOA/PFOS amounts in solution were limited. The sorbent

241

produced later could not reach saturation adsorption. Thus, the sorbed amount of PFOA/PFOS

242

increased firstly and then decreased. It is unfortunate that the numbers of data points during the

243

initial sorption period of the experiments with the initial concentration of PFOA/PFOS at 1.5 μM

244

(see Figure 3a) were not sufficient to enable sorption kinetics fitting (See Figure S4), but those for

245

the experiments with the initial concentration of PFOA/PFOS at 0.5 mM were. Four sorption models:

246

pseudo-first-order kinetics, pseudo-second-order kinetics, Elovich and Intra-particle diffusion 11


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models (See Text S2 of the SI) were used to describe the experimental data with the initial

248

concentration of PFOA/PFOS at 0.5 mM (see Figure 3a). The fitting parameters are given in Table 1.

249

As shown in Figure 4 and Table 1, the R2 values of intra-particle diffusion model were relatively

250

lower than those obtained for the other models, indicating that the PFOA/PFOS sorption amounts

251

were high on the surface of the zinc hydroxide flocs rather than in the intra-particle pores. In

252

comparison to the pseudo-first-order kinetics and Elovich model, pseudo-second-order kinetics

253

models described the sorption process better. The initial sorption rates (v0) of PFOA and PFOS were

254

1.01×103 and 1.81×103 mmol g-1 h-1, respectively. The equilibrium sorbed amounts (qe) of PFOA

255

and PFOS were 5.74 and 7.69 mmol g-1 (Zn), respectively.

256

To further evaluate the sorption capacities of PFOA/PFOS and understand the sorbate-sorbent

257

interactions, the sorption isotherms of PFOA/PFOS on zinc hydroxide flocs in-situ generated by EC

258

process were studied. Since PFOA/PFOS could be quickly sorbed by zinc hydroxide flocs, the

259

sorption isotherm experiments were conducted as described following: the experiments were

260

conducted in the reactor (II) at the initial PFOA/PFOS concentrations ranging from 0.05 to 0.8 mM,

261

the electrochemical experiments were stopped after 3 min of electrocoagulation whereas the solution

262

was continuously stirred for another 5 min to ensure that the sorption equilibrium was reached. The

263

solution was then sampled for analysis. Figure S5 illustrates the Langmuir and Freundlich isotherms

264

(see Text S3 of the SI) of PFOA/PFOS. The corresponding parameters of the isotherms are shown in

265

Table 2. Based on the R2 values, Freundlich isotherm was observed to better fit the sorption

266

behaviors of both PFOA and PFOS, while the Langmuir isotherm could not describe the sorption

267

behavior well.

268

The sorption of PFOA/PFOS onto different sorbents reported in the literature is presented in

269

Table S2, which includes sorbent characteristics, experimental conditions, sorption capacity,

270

sorption equilibrium time, and initial sorption rates. The sorption capacities (qm) obtained by data 12



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fitting to Langmuir model of PFOA/PFOS on the most widely used sorbents, i.e., GAC and powder

272

active carbon (PAC), were 0.39/0.37 and 0.67/1.04 mmol g-1, respectively. The qe values of

273

PFOA/PFOS on the zinc hydroxide flocs in this study were 5.74/7.69 mmol g-1, which were

274

18.7/20.8 and 8.6/7.4 times higher than those of GAC and PAC, respectively. More importantly, the

275

sorption rates of PFOA/PFOS on zinc hydroxide flocs were several times of those reported with the

276

fastest PFAA sorption. The high sorption capacity and fast sorption rate made the EC process with

277

zinc anode to be a technology with great potential for rapid purification or remediation of PFAA

278

polluted waters, industry wastewater and environmental waterbody.

279

Sorption Mechanisms. It is important to investigate the driving force and rate-limiting step of

280

adsorption in order to understand the sorption mechanism. Because the sorption rates of PFAAs

281

were very fast, we focused on the driving force of adsorption. Some interactions including van der

282

Waals force, π-π bond, electrostatic interaction, hydrogen bond, ion and/or ligand exchange, and

283

hydrophobic effect possibly involved in the sorption process. Among these forces, the π-π bond is

284

impossible due to the absence of π electrons in both PFAA molecules and the zinc hydroxide flocs.

285

Similarly, van der Waals force is also unimportant because of the low polarizabilities of PFAAs

286

(20).

287

PFAA molecules, because of their charged head groups, -COO- or -SO3-, may be sorbed via

288

anion/ligand exchange with groups like -Cl, -CO3- on certain sorbents, such as anion exchange resin

289

and hydrotalcite. It was also postulated in previous studies (33, 34) that PFOA/PFOS may replace

290

the hydroxyl groups on metal oxides such as AlOOH and α-Fe2O3 by ligand exchange, as expressed

291

in the following reactions:

292

Al-OH + L- → Al-L + OH-

(4)

293

≡Fe-OH2+ + L- → ≡Fe-L + H2O

(5)

294

Theoretically, the density of hydroxyl groups on the surface of aluminum and iron hydroxide 13


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flocs are 1.5 times of that on zinc hydroxide flocs. However, their sorption capacities of

296

PFOA/PFOS were much lower than those of the zinc hydroxide flocs (see Figure 1). XPS analysis

297

was used to determine the density of hydroxyl groups on the surface of zinc hydroxide flocs before

298

and after PFAA sorption. The results showed that the density of hydroxyl groups on the surface of

299

the zinc hydroxide flocs after PFOA/PFOS sorption did not have obvious, if any, change (See

300

Figure S6 and Text S4). Thus, hydroxyl-based ligand exchange only has weak or no contribution to

301

the PFOA/PFOS sorption on the metal hydroxides generated in-situ by EC.

302

The zeta potential of the zinc hydroxide flocs as a function of electrolysis time during EC with

303

zinc as anode is depicted in Figure S7a (initial pH at 5). During the initial stage of the EC process,

304

the zeta potential on the zinc hydroxide flocs was slightly negative, and then decreased sharply to

305

-48.6 mV at 3 min and kept decreasing during the following 4 min, and then increased and reached

306

zero point at about 10.6 min. Therefore, the freshly formed zinc hydroxide flocs were negatively

307

charged at the initial 10.6 min of electrolysis. The electrostatic interaction between the

308

PFOA/PFOS anions and the fresh zinc hydroxide flocs was thus repulsive. In fact, most of PFOA

309

and PFOS were removed from water during the first 10 min (see Figure 3a). In addition, a test with

310

the initial solution pH varying from 5~10 revealed that pH had little effect on the PFOA removal

311

(see Figure S7b). The results indicated that PFOA/PFOS sorption on the zinc hydroxide flocs was

312

not attributed to the electrostatic attraction. Although O or S atoms in the hydrophilic functional

313

groups head of PFAAs, –COOH and -SO3-, were able to serve as the acceptors, hydrogen bonding

314

did not play a significant role in PFAA sorption on zinc hydroxide flocs, because the removal of

315

PFAAs with short C-F chain length (e.g., PFBA) was very limited (see Figure S8 and Table S3).

316

Hydrophobic interaction may also affect the sorption of PFOA/PFOS on hydrophobic sorbents

317

(20, 35). Previous studies (36) indicated that zinc oxide and many kinds of zinc hydroxides

318

including hydroxyl, chloride, carbonate and sulphate, have hydrophobic surfaces. Thus, 14



Page 16 of 29

319

hydrophobic interaction may contribute to the high sorption capacity of PFOA/PFOS on zinc

320

hydroxide flocs. In order to explore this possibility, sorption of six PFAAs with different C-F chain

321

length were compared. The results are shown in Figure S8. It can be seen that C-F chain could

322

significantly affect the sorption of PFAAs. PFAAs with long carbon chains including PFDA, PFNA,

323

and PFHpA could be quickly removed, while the shorter-chain PFAA, PFBA was removed slowly,

324

only 6.2% removal ratio after 20 min EC treatment. The sorption process could be well described

325

by pseudo-second-order model. The corresponding parameters are summarized in Table S3. It is

326

interesting that the sorption capacities on the zinc hydroxide flocs increased with the increasing

327

chain length, and the qe values of PFAAs followed the order of PFDA > PFNA > PFOS > PFOA >

328

PFHpA >> PFBA (see Table 1 and Table S3). We furthermore conducted the competitive sorption

329

experiments with PFOA, PFNA, and PFDA. The results showed that PFAAs with longer C-F chain

330

length were preferentially sorbed (see Figure S9). Since PFAAs with longer C-F chain length are

331

more hydrophobic, these findings clearly demonstrated that hydrophobic interaction plays a key

332

role on the high sorption capacities of hydrophobic PFAAs on the zinc hydroxide flocs. The

333

PFAAs sorption capacities on the metals hydroxide flocs followed the following order: zinc

334

hydroxide flocs >> aluminum hydroxide flocs >> magnesium and ferric hydroxide flocs. This

335

difference may dependent on their physical-chemical properties to some extent. Magnesium

336

hydroxide flocs are strongly hydrophilic (37) and ferric hydroxide species are also hydrophilic (38).

337

Aluminum hydroxide flocs are mainly composed by hydrophilic colloidal aluminum hydroxide

338

with small amount of hydrophobic tridimensional tactoids hydrated aluminum ions (39, 40), while

339

zinc hydroxide flocs exhibits a certain hydrophobicity (36). In addition, the sorption capacity of

340

PFAAs on zinc hydroxide flocs increased with the increasing chain length.

341

In addition, electric field in the EC process would enhance the concentrations of the PFAAs

342

and dissolved Zn2+ around the surface of anode. The locally higher PFAAs and Zn2+ concentrations 15


Page 17 of 29


343

could improve PFAAs enmeshment and sorption by the zinc hydroxide flocs. This electric field

344

assisted concentration phenomenon has been demonstrated in F ions removal by EC with aluminum

345

anode as well as virus removal by iron anode (41, 42).

346

The long-chain PFAAs such as PFOA and PFOS are not only hydrophobic but also

347

oelophobic. Recently, Meng et al. (35) reported that the air bubbles on the surface of the

348

hydrophobic carbonaceous sorbents played an important role in the hydrophobic sorption of PFOS.

349

The accumulation of PFOS at the interface of the air bubbles was primarily responsible for its

350

sorption. Numerous micron hydrogen bubbles were generated during the EC process, and many of

351

them were sorbed on the surface of the freshly formed metal hydroxide flocs. Thus, it is a key

352

question to answer if the hydrogen bubbles play an important role in the hydrophobic sorption of

353

PFAAs on the zinc hydroxide flocs generated in-situ during EC process. Low-frequently ultrasonic

354

(20 KHz, 60W) degassing treatment was conducted immediately after electrocoagulation. The

355

results showed that the sorbed PFOA/PFOS could not be released after 5 min treatment or even

356

more, indicating that the hydrogen bubbles sorbed on the zinc hydroxide flocs did not contribute to

357

the PFOA/PFOS sorption (for details, see Figure S10 and Text S5 of the SI).

358

PFAAs molecules would be flat adsorbed on the zinc hydroxide flocs surface to minimize

359

water-fluorine interactions. Based on the molecule size, the theoretical maximum number of PFOS

360

molecules per unit surface area for a monolayer of coverage was estimated about 4 molecules nm-2

361

assuming the long axis (C-F chain) of the molecule is parallel to the surface and no space exits

362

between molecules (43). The BET surface area of the zinc hydroxide flocs was measured as 48.7 m2

363

g-1 (see Figure S11), by unit conversion, the number of PFOS molecules sorbed per unit surface area

364

was 62.4 PFOS molecules nm-2 of the zinc hydroxide flocs surface according to the qe obtained by

365

pseudo-second-order kinetics (See Table 1).The results suggested that PFOS on the zinc hydroxide

366

flocs surface was multilayer sorption. However, it should be pointed out that the measured surface 16



Page 18 of 29

367

area of the freeze-dried zinc hydroxide flocs by BET method may be far less than the real surface

368

area of the fresh zinc hydroxide flocs in aqueous solution. For example, the BET surface area of

369

hydrous ferric oxide (HFO) was about 600 m2 g-1, while the surface area of fresh HFO was

370

determined to be 5500 ±170 m2 g-1 by dye adsorption method (44).

371

Environmental Implications. The findings of this study suggested that EC with zinc anode

372

may be a feasible approach for purification or remediation of PFAA contaminated waters. PFAAs

373

can be quickly sorbed on the surface of

374

mainly via hydrophobic interaction. Compared with previous reports on sorption or other

375

physical-chemical removal methods for PFAAs, the zinc hydroxide flocs in-situ generated in EC

376

process have much higher sorption capacity or faster sorption rate. The superior high sorption

377

capacity and extremely fast sorption rate allow this technique to be employed for removing PFAAs

378

from water at the concentrations varying from several hundred μg L-1 or even lower to hundreds of

379

mg L-1 within a short hydraulic retention time.

the zinc hydroxide flocs in-situ generated in EC process,

380

EC has been widely used in wastewater treatment for decades. It can be set up with great

381

flexibility, therefore is very applicable to be coupled with other treatment techniques, such as

382

membrane separation, electrochemical oxidation, and thermos-oxidation to enhance the removal

383

efficiency and reduce cost. It is estimated that the energy consumption for destructing PFAAs is

384

dependent on its concentration, with the energy required to degrade a mole of PFAAs decreasing by

385

at least one order of magnitude if the concentration increased by 2 orders of magnitude (18). The

386

EC process with zinc anode may be used as an approach to concentrate PFAAs in water and then

387

feed to the other treatment technologies to achieve acceptable cost-effectiveness. Unlike the

388

traditional sorbents, zinc hydroxide flocs can be easily dissolved in acid or base solution, so that the

389

sorbed PFAAs are easily released to solution again and thus concentrated, which can be treated by 17


Page 19 of 29


390

various oxidation methods such as electrochemical oxidation and UV/K2S2O8. Alternatively,

391

PFAAs can be incinerated at high temperature in halogen resistant incinerators, but water

392

incineration is not economically acceptable (30). The EC process developed in this study can be

393

used to extract PFAAs from water, and the zinc hydroxide flocs enriched with PFAAs can be then

394

incinerated cost effectively.

395 396

ASSOCIATED CONTENT

397

Supporting Information Available

398

Description of kinetics, isotherms, and physicochemical properties of PFAAs; the experimental

399

procedure in detail; the experimental apparatus; the zinc hydroxide flocs characterization and PFAA

400

sorption results under various conditions. This information is available free of charge via the Internet

401

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

402 403

AUTHOR INFORMATION

404

Corresponding Authors

405

Phone: +86-10-5880-7612; fax: 86-10-5880-7612; e-mail: [email protected].

406

Phone: 770-229-3302; fax: 770-412-4734; e-mail: [email protected]

407

Notes

408

The authors declare no competing financial interest.

409

ACKNOWLEDGMENTS

410

This study was financially supported by the Fund for Innovative Research Group of the

411

National Natural Science Foundation of China (No. 51421065), the National Natural Science

412

Foundation of China (No. 51378065) and the Fundamental Research Funds for the Central

413

Universities of China (No. 2012LZD03). 18



414

Literature Cited

415 416 417 418 419 420 421 422 423 424 425 426 427 428 429 430 431 432 433 434 435 436 437 438 439 440 441 442 443 444 445 446 447 448 449 450 451 452 453 454 455 456 457 458 459

(1)

(2) (3) (4) (5)

(6) (7)

(8)

(9)

(10)

(11)

(12)

(13)

(14) (15)

(16)

Page 20 of 29

Fujii, S.; Tanaka, S.; Lien, N. P. H.;Qiu, Y.; Polprasert, C. New POPs in the water environment: distribution, bioaccumulation and treatment of perfluorinated compounds—a review paper. J. Water Supply Res. Technol.—AQUA 2007, 56, 313-326. Prevedouros, K.; Cousins, I. T.; Buck, R. C.; Korzeniowski, S. H. Sources, fate and transport of perfluorocarboxylates. Environ. Sci. Technol. 2006, 40, 32-44. de Witte, J.; Piessens, G.; Dams, R. Fluorochemical intermediates, surfactants and their use in coatings. Surf. Coat. Int. 1995, 78, 58-64. Fiedler, S.; Pfishter, G.; Schramm, K. Poly- and perfluorinated compounds in household consumer products. Toxico. Environ. Chem. 2010, 92, 1801-1811. Borg, D.; Lund B.; Lindquist, N.; Håkansson, H. Cumulative health risk assessment of 17 perfluoroalkylated and polyfluoroalkylated substances (PFASs) in the Swedish population. Environ. Int. 2013, 59, 112-123. Ahrens, L. Polyfluoroalkyl compounds in the aquatic environment: a review of their occurrence and fate. J. Environ. Monit. 2011, 13, 20-31. Olsen, G. W.; Burris, J. M.; Ehresman, D. J.; Froehlich, J. W.; Seacat, A. M.; Butenhoff, J. L.; Zobel, L. R. Half-life of serum elimination of perfluorooctanesulfonate, perfluorohexanesulfonate, and perfluorooctanoate in retired fluorochemical production workers. Environ. Health Perspect. 2007, 115, 1298-1305. Ahrens, L.; Taniyasu, S.; Yeung, L. W. Y.; Yamashita, N.; Lam, P. K. S.; Ebinghaus, R. Distribution of polyfluoroalkyl compounds in water, suspended particulate matter and sediment from Tokyo Bay, Japan. Chemosphere 2010, 79, 266-272. Paul, A. G.; Jones, K. G.; Sweetman, A. J. A first global production, emission, and environmental inventory for perfluorooctane sulfonate. Environ. Sci. Technol. 2009, 43, 386-392. Clara, M.; Scheffknecht, C.; Scharf, S.; Weiss, S.; Gans, O. Emissions of perfluorinated alkylated substance (PFSA) from point soure – identification of relevant branches. Water Sci. Technol. 2008, 58, 59-66. Hori, H.; Hayakawa, E.; Einaga, H.; Kutsuna, S.; Koike, K.; Ibusuki, T.; Kiatagawa, H.; Arakawa, R. Decomposition of environmentally persistent perfluorooctanoic acid in water by photochemical approaches. Environ. Sci. Technol. 2004, 38, 6118-2614. Huang Q.G. Remediation of perfluoroalkyl contaminated aquifers using an in situ two-layer barrier laboratory batch and column study. http://serdp-estcp.org/Program-Areas/Environmental-Restoration/Contaminated-Groundwate r/Emerging-Issues/ER-2127/ER-2127/(language)/eng-US. Moody, C. A.; Hebert, G. N.; Strauss, S. H.; Field, J. A. Occurrence and persistence of perfluorooctanesulfonate and other perfluorinated surfactants in groundwater at a fire-training area at Wurtsmith Air Force Base, Michigan, USA. J. Environ. Monit. 2003, 5, 341-345. Schultz, M. M.; Barofshy, D. F.; Field, J. A. Quantitative determination of fluorotelomer sulfonates in groundwater by LC/MS/MS. Environ. Sci. Technol. 2004, 38, 1828-1835. Niu, J. F.; Lin, H.; Xu, J. L.; Wu, H.; Li, Y. Y. Electrochemical Mineralization of Perfluorocarboxylic Acids (PFCAs) by Ce-Doped Modified Porous Nanocrystalline PbO2 Film Electrode. Environ. Sci. Technol. 2012, 46, 10191-10198. Lin, H.; Niu, J. F.; Ding, S. Y.; Zhang, L. L. Electrochemical Degradation of Perfluorooctanoic Acid (PFOA) by Ti/SnO2-Sb, Ti/SnO2-Sb/PbO2 and Ti/SnO2-Sb/MnO2 19


Page 21 of 29

460 461 462 463 464 465 466 467 468 469 470 471 472 473 474 475 476 477 478 479 480 481 482 483 484 485 486 487 488 489 490 491 492 493 494 495 496 497 498 499 500 501 502 503 504 505 506

(17)

(18)

(19) (20)

(21)

(22) (23)

(24)

(25)

(26)

(27)

(28) (29)

(30)

(31)

(32)


anodes. Water Res. 2012, 46, 2281-2289. Moriwaki, H.; Takagi, Y.; Tanaka, M.; Tanaka, M.; Tsuruho, K.; Okitsu, K; Maeda, Y. Sonochemical decomposition of perfluorooctane sulfonate and perfluorooctanoic acid. Environ. Sci. Technol. 2005, 39, 3388-3392. Vecitis, C. D.; Park, H.; Cheng, J.; Mader, B. T.; Hoffman, M .R. Treatment technologies for aqueous perfluorooctanesulfaonate (PFOS) and perfluorooctanoate (PFOA). Front. Environ. Sci. Engin. China 2009, 3, 129-151. Lutze, H.; Panglisch, S.; Bergmann, A.; Schmidt, T. C. Treatment options for the removal and degradation of polyfluorinated chemicals. Handbook Environ. Chem. 2012, 17, 103-125. Du, Z. W.; Deng, S. B.; Bei, Y.; Huang, Q.; Wang, B.; Yu, G. Adsorption behavior and mechanism of perfluorinated compounds on various adsorbents—a review. J. Hazard. Mater. 2014, 274, 443-454. Yu, Q.; Zhang, R.Q.; Deng, S.B.; Huang, J.; Yu, G. Sorption of perfluorooctane sulfonate and perfluorooctanoate on activated carbons and resin: Kinetic and isotherm study. Water Res. 2009, 43, 1150-1158. Schaefer A. Perfluorinated surfactants contaminate German waters. Environ. Sci. Technol. 2006, 40, 7108~7109. Lampert, D. J.; Frisch M. A.; Speitel Jr., G. E. Removal of perfluorooctanoic acid and perfluorooctane sulfonate from wastewater by ion exchange. Pract. Period. Hazard. Toxic Radioact. Waste Manage. 2007, 11, 60-68. Yu, J.; Lv, L.; Lan, P.; Zhang, S. J.; Pan, B. C.; Zhang, W. M. Effect of effluent organic matter on the adsorption of perfluorinated compounds onto activated carbon. J. Hazard. Mater. 2012, 225-226, 99-106. Mollah, M. Y. A.; Morkovsky, P.;Gomes, J. A. G.; Kesmez, M.; Parga, J.; Cocke, D. L. Fundamentals, present and future perspectives of electrocoagulation. J. Hazard. Mater. 2004, B114, 199-210. Appleman, T. D.; Higgins, C. P.; Quiñones, O.; Vanderford, B. J.; Kolstad, C.; Zeigler-Holady, J. C.; Dickenson, E. R. V. Treatment of poly- and perfluoroalkyl substances in U.S. full-scale water treatment systems. Water Res. 2014, 51, 246-255. Rahman, M. F.; Peldszus, S.; Anderson, W. B. Behaviour and fate of perfluoroalkyl and polyfluoroalkyl substances (PFASs) in drinking water treatment: A review. Water Res. 2014, 50, 318-340. Deng, S. B.; Zhou, Q.; Yu, G.; Huang, J.; Fan, Q. Removal of perfluorooctanoate from surface water by polyaluminum chloride coagulation. Water Res. 2011, 45, 1774-1780. Xiao, F.; Simcik, M. F.; Gulliver, J. S. Mechanisms for removal of perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) form drinking water by conventional and enhanced coagulation. Water Res. 2013, 47, 49-56. Baudequin, C.; Couallier, E.; Rakib, M.; Deguerry, I.; Severac, R.; Pabon, M. Purification of firefighting water containing a fluorinated surfactant by reverse osmosis coupled to electrocoagulation-filtration. Sep. Purif. Technol. 2011, 76, 275-282. Lin, H.; Niu, J. F.; Xu, J. L.; Huang, H. O.; Li, D.; Yue, Z. H.; Feng, C. H. Highly efficient and mild electrochemical mineralization of long-chain perfluorocarboxylic acid (C9-C10) by Ti/SnO2-Sb-Ce, Ti/SnO2-Sb/Ce-PbO2, and Ti/BDD electrodes. Environ. Sci. Technol. 2013, 47, 13039-13046. Riddell, N.; Arsenault, G.; Benskin, J. P.; Chittim, B.; Martin, J. W.; Mcalees, A.; Mccrindle, R. Branched perfluorooctane sulfonate isomer quantification and characterization in blood serum samples by HPLC/ESI-MS(/MS). Environ. Sci. Technol. 2009, 43, 7902-7908. 20



507 508 509 510 511 512 513 514 515 516 517 518 519 520 521 522 523 524 525 526 527 528 529 530 531 532 533 534

(33) (34)

(35)

(36) (37) (38) (39) (40) (41)

(42) (43)

(44)

Page 22 of 29

Wang, F.; Liu, C.; Shih, K. Adsorption behavior of perfluorooctanesulfonate (PFOS) and perfluorooctanoate (PFOA) on Boehmite. Chemosphere, 2012, 89, 1009-1014. Gao, X. D.; Chorover, J. Adsorption of perfluorooctanoic acid and perfluorooctanesulfonic acid to iron oxide surface as studied by flow-through ATR-FTIR spectroscopy. Environ. Chem. 2012, 9, 148-157. Meng, P. P.; Deng, S. B.; Lu, X. Y.; Wang, B.; Huang, J. Wang, Y. J.; Yu, G.; Xing, B. S. Role of air bubbles overlooked in the adsorption of perfluorooctanesulfonate on hydrophobic carbonaceous adsorbents. Environ. Sci. Technol. 2014, 48, 13785-13792. Muster, T. H.; Neufeld, A. K.; Cole, I. S. The protective nature of passivation films on zinc: wetting and surface energy. Corros. Sci. 2004, 46, 2337-2354. Christopher, J. Reduced lime feeds: Effects on operational costs and water quality. Des Moines Water Works, Des Moines. Iowa, 2005. Ahlberg, E.; Forssberg, K. S. E.; Wang, X. The surface oxidation of pyrite in alkaline solution. J. Appl. Electrochem. 1990, 20, 1033-1039. Yariv, S.; Cross, H. Geochemistry of colloid systems: for earth scientists. Springer, Heidelberg: New York, 1979. Wefers, K.; Misra, C. Oxides and hydroxides of aluminum. Alcoa Technical Paper No. 19, Alcoa Laboratories: Pittsburg, PA, USA, 1987. Zhu B.T., Clifford D.A., Chellam S. Comparison of electrocoagulation and chemical coagulation pretreatment for enhanced virus removal using microfiltration membranes. Water Res. 2005, 39, 3098~3108. Zhu, J.; Zhao, H. Z.; Ni, J. R. Fluoride distribution in electrocoagulation defluoridation process. Sep. Purif. Technol. 2007, 56, 184-191. Johnson, R. L.; Anschutz, A. J.; Smolen, J. M.; Simcik, M. F.; Penn, R. L. The adsorption of perfluorooctane sulfonate onto sand, clay, and iron oxide surfaces. J. Chem. Eng. Data. 2007, 52, 1165-1170. Imre Takács. Experiments in activated sludge modelling. Ph.D. Dissertation, Ghent University, Belgium, 2008.

535

21


Page 23 of 29


536

Table 1. Sorption parameters of pseudo-first-order kinetics, pseudo-second-order kinetics, Elovich,

537

and intra-diffusion models for PFOA and PFOS. Pesudo-first-order kinetics qe (mmol g-1) PFOA PFOS

5.52±0.14 6.99±0.15

PFOA

α (mmol g-1 h-1) 1.25×105±7.1 5×103 2.11×105±1.4 6×104

PFOS

k1 (h-1) 7.68 12.33 Elovich β (mmol g-1)

Pesudo-second-order kinetics qe (mmol k2 (g mmol-1 v0 (mmol g-1) h-1) g-1 h-1) 5.74±0.08 30.69±3.09 1.01×103 7.69±0.04 30.62±1.01 1.81×103 Intra-particle diffussion

R2 0.950 0.965

R2 0.987 0.991

R2

Kid (mmol g-1 h-0.5)

I (mmol g-1)

R2

1.36±0.14

0.972

13.16±3.35

1.59±0.76

0.707

0.98±0.13

0.975

27.26±7.66

1.45±1.21

0.745

538 539

22



540

Page 24 of 29

Table 2. Parameters fitted by the sorption equilibrium data with Langmuir and Freundlich isotherms. Langmuir constants

PFOA PFOS

qm (mmol g-1)

b (L mmol-1)

R2

6.05±0.44 7.17±0.82

44.29±8.43 74.69±21.49

0.9598 0.9130

Freundlich constants K n R2 (1-1/n) 1/n -1 (mmol L g ) 10.34±0.54 0.40 0.9857 14.29±0.52 0.39 0.9944

541 542

23


Page 25 of 29


543

Figure Captions

544

Figure 1. (a) The effect of sacrificial anodes on the removal of PFOA (C0 = 0.5 mM, I = 0.1 A, pH =

545

5, 10 mM NaCl) by electrocoagulation; (b) the adsorbed amount of PFOA as a function of the metal

546

dissolved dosage.

547

Figure 2. Fourier transform infrared spectrum (FTIR) spectra of solid PFOA and the zinc hydroxide

548

flocs before and after PFOA sorption.

549

Figure 3. (a) Removal of PFOA/PFOS as a funciton of electrolysis/energy consumption (C0 = 1.5

550

μM / 0.5 mM, i = 0.5 mA cm-2, pH = 5, 10 mM NaCl) by zinc anode; (b) concentrations change of

551

linear and branched PFOS isomers during electrocoagulation process.

552

Figure 4. Soption kinetics of PFOA and PFOS on the zinc hyroxide flocs.

553 554 555

24


Environmental Science & Technology 100

6

Zinc Magnesium Aluminum Iron

(b)

5 4

(a)

-1

60

qt (mmol g )

PFOA Removal (%)

80

40

20

Zinc Magnesium Aluminum Iron

3 2 1 0

0 0

556

Page 26 of 29

2

4

6

8

10

0

20

40

60

80

100

-1

120

140

Dissovled Dosage (mg L )

Electrocoagulation Time (min)

557 558 559

Figure 1. (a) The effect of sacrificial anodes on the removal of PFOA (C0 = 0.5 mM, I = 0.1 A, pH =

560

5, 10 mM NaCl) by electrocoagulation; (b) the adsorbed amount of PFOA as a function of the metal

561

dissolved dosage.

562 563 564

25


-

v(COO )

-

v(COO )

PFOA solid (Purity>98%) Zinc hydroxide flocs adsorbed PFOA Zinc hydroxide flocs

500

1000

1500

vas(OH)

v(Zn(OH)2)

v(OH)

Intensity (cps)

vas(CF2+CF3)


v(C-C) vas(CF2)

Page 27 of 29

2000

2500

3000

3500

4000

-1

Wavenumber (cm )

565 566

Figure 2. Fourier transform infrared spectrum (FTIR) spectra of solid PFOA and the zinc hydroxide

567

flocs before and after PFOA sorption.

568

26



Page 28 of 29

569 1.5

0.5 mM PFOA 0.5 mM PFOS 1.5 μM PFOA 1.5 μM PFOS

40

0.6

L-PFOS

0.3 0

4

8 PFOS L-PFOS M-PFOS L-PFOS

0.4

20

-1

-1

Energy consumption (Wh L ) Electrocoagulation Time (min) 0

0.00 0

4 0.03 30

8 0.06

12 0.09

60

16 0.12

90

0.15 -1

120

0.6 0.4 0.2

0

3

0.5 0.0

0

570

M-PFOS

0.8

20 0.18

0.3 0.2

16

15

20

0.6 0.4 0.2

0.1

6 9 12 Time (min)

120.8

CL-PFOS /CPFOS

60

0.9

PFOS L-PFOS

CL-PFOS /CPFOS

umol L

-1

(a)

80

(b)

1.2

mmol L

PFOA/S Removal (%)

100

0

4

8 12 16 Time (min)

20

0.0

150

0

4

Theoretical Zn Dosage (mg L )

8

12

16

20

Electrocoagulation Time (min)

571 572

Figure 3. (a) Removal of PFOA/PFOS as a funciton of electrolysis/energy consumption (C0 = 1.5

573

μM / 0.5 mM, i = 0.5 mA cm-2, pH = 5, 10 mM NaCl) by zinc anode; (b) concentrations change of

574

linear and branched PFOS isomers during electrocoagulation process.

575 576

27


Page 29 of 29

Environmental Science & Technology 8 7

-1

qt (mmol g )

6 5 4 PFOA PFOS Pseudo-first order model Pseudo-second order model Elovich model Intra-particle diffussion model

3 2 1 0 0.00

577

0.05

0.10

0.15

0.20

0.25

0.30

Electrocoagulation time (h)

578 579

Figure 4. Sorption kinetics of PFOA and PFOS on the zinc hydroxide flocs.

580 581

28


0.35

Efficient Sorption and Removal of Perfluoroalkyl Acids (PFAAs) from

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