Sorption of perfluorooctane sulfonate and perfluorooctanoate on activated carbons and resin: Kinetic and isotherm study

water research 43 (2009) 1150–1158

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Sorption of perfluorooctane sulfonate and perfluorooctanoate on activated carbons and resin: Kinetic and isotherm study Qiang Yua,b,c, Ruiqi Zhanga, Shubo Denga,b,c,*, Jun Huanga,b,c, Gang Yua,b,c a

Department of Environmental Science and Engineering, Tsinghua University, Beijing 100084, PR China POPs Research Center, Tsinghua University, Beijing 100084, PR China c State Key Joint Laboratory of Environment Simulation and Pollution Control, Tsinghua University, Beijing 100084, PR China b

article info

abstract

Article history:

Perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) have increasingly attrac-

Received 5 August 2008

ted global concerns in recent years due to their global distribution, persistence, strong

Received in revised form

bioaccumulation and potential toxicity. The feasibility of using powder activated carbon

30 November 2008

(PAC), granular activated carbon (GAC) and anion-exchange resin (AI400) to remove PFOS

Accepted 1 December 2008

and PFOA from water was investigated with regard to their sorption kinetics and

Published online 13 December 2008

isotherms. Sorption kinetic results show that the adsorbent size influenced greatly the sorption velocity, and both the GAC and AI400 required over 168 h to achieve the equilib-

Keywords:

rium, much longer than 4 h for the PAC. Two kinetic models were adopted to describe the

PFOS

experimental data, and the pseudo-second-order model well described the sorption of

PFOA

PFOS and PFOA on the three adsorbents. The sorption isotherms show that the GAC had

Activated carbon

the lowest sorption capacity both for PFOS and PFOA among the three adsorbents, while

Anion-exchange resin

the PAC and AI400 possessed the highest sorption capacity of 1.04 mmol g1 for PFOS and

Sorption kinetics

2.92 mmol g1 for PFOA according to the Langmuir fitting. Based on the sorption behaviors

Sorption isotherm

and the characteristics of the adsorbents and adsorbates, ion exchange and electrostatic interaction as well as hydrophobic interaction were deduced to be involved in the sorption, and some hemi-micelles and micelles possibly formed in the intraparticle pores. ª 2008 Elsevier Ltd. All rights reserved.

1.

Introduction

Perfluorinated compounds (PFCs) have been widely used in industrial and commercial applications for about 50 years as surfactants, emulsifiers, fire retardants, polymer additives and etc. (Key et al., 1997; Fujii et al., 2007). The PFCs most commonly used and found in the environment are perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA), which may be directly discharged from pollution sources or generated by the microbial degradation of other perfluorinated compounds. Since they are found to be globally distributed, environmentally persistent, bioaccumulative and

potentially toxic, PFOS and PFOA have increasingly attracted global concerns in recent years and also have been proposed as the candidates of persistent organic pollutants (POPs) (Giesy and Kannan, 2002; Loos et al., 2008). Different from other typical POPs, PFOS and PFOA have high water solubility, and thus can exist and easily transport in water environments. So far, they have been detected in wastewater, surface water, groundwater and even tap water throughout the world (Fujii et al., 2007). Industrial wastewater has been implicated as a point source for PFOS and PFOA as well as their precursors entering into natural waters (Hansen et al., 2002; Prevedouros et al., 2006). Many researchers found

* Corresponding author. Department of Environmental Science and Engineering, Tsinghua University, Beijing 100084, PR China. Tel.: þ86 10 6279 2165; fax: þ86 10 6279 4006. E-mail address: [email protected] (S. Deng). 0043-1354/$ – see front matter ª 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2008.12.001

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the higher concentrations of PFOA and PFOS in the rivers near the PFCs-related factories in Germany (Loos et al., 2008) and America (Hansen et al., 2002; Sinclair et al., 2006). Besides, the PFOS concentrations at mg L1 level were detected in the river near a Canadian airport due to the accidental release of firefighting foam (Moody et al., 2002). Tang et al. (2006) reported that the concentration of PFOS in the original wastewater generated from photolithographic processes of semiconductor manufacture was up to 1650 mg L1, which would cause serious pollution if being discharged into the environment. Therefore, the development of effective techniques to remove PFOS and PFOA from industrial wastewater becomes crucial. Some conventional techniques including biological degradation, oxidation and reduction are difficult to destruct PFOS and PFOA in ambient environments due to their stable properties (Key et al., 1998; Schroder and Meesters, 2005). The recent studies show that some special techniques such as ultrasonic irradiation under argon atmosphere (Moriwaki et al., 2005), zerovalent iron in subcritical water (Hori et al., 2006), ultraviolet irradiation (Yamamoto et al., 2007) and vitamin B12/Ti-citrate reduction in anoxic environment (Ochoa-Herrera et al., 2008) may decompose PFOS or PFOA in solution, but the specific conditions and high energy consumption are required. Additionally, the commercial reverse osmosis and nanofiltration membranes were also used to separate PFOS from wastewater efficiently (Tang et al., 2006, 2007), and our previous study found that the chitosanbased molecularly imprinted adsorbents were effective for the selective removal of PFOS from water (Yu et al., 2008). It has been demonstrated in many cases that sorption is an effective and economical method to remove many pollutants from wastewater, but only few papers about PFOS or PFOA removal using some commercial adsorbents were published (Lampert et al., 2007; Ochoa-Herrera and Sierra-Alvarez, 2008). Some researchers reported that the perfluorinated surfactants could easily penetrate the granular activated carbon beds in German waterworks and thus it was doubtful whether activated carbon was effective for PFOS and PFOA removal (Schaefer, 2006). As no detailed investigation was conducted, it is unclear whether the conventional adsorbents are effective for PFOS and PFOA removal from water. Therefore, it is necessary to know which adsorbent should be used to effectively remove them once water is polluted by PFOS or PFOA. The objectives of this study are to investigate the sorption behaviors of PFOS and PFOA on the commercial adsorbents including activated carbons and resin, and evaluate their feasibility for PFOS and PFOA removal from water. The

sorption kinetics at different solution pH and sorption isotherms for PFOS and PFOA using the granular activated carbon, powder activated carbon and anionic resin were studied in detail. The possible interactions between the adsorbents and adsorbates were also discussed.

2.

Materials and methods

2.1.

Materials

Perfluorooctane sulfonate (PFOS, potassium salt) and perfluorooctanoate (PFOA, sodium salt) were purchased from Tokyo Kasei Kogyo (Japan), and their properties are summarized in Table 1. The Amberlite IRA 400 resin (AI400) and coalbased activated carbon were obtained from Sinopharm Chemical Regent Co., Ltd. (China) and Pengcheng Activated Carbon Co., Ltd. (China), respectively. HPLC-grade methanol was purchased from Fisher Chemical (USA). Other chemicals were of reagent grade.

2.2.

Adsorbent pretreatment

Prior to the use in the sorption experiment, the resin was first washed in deionized water to remove dirt and then dried at 50  C until constant weight. Similarly, the coal-based activated carbons were first rinsed with deionized water for several times and then washed in 80  C deionized water for 2 h to remove the impurities. After being dried in an oven at 105  C for 48 h, they were crushed by a mortar and screened. The activated carbons in the size range of 0.9–1.0 mm were used as the granular activated carbon (GAC), while the powder activated carbon (PAC) in this paper represented the particles below 0.1 mm.

2.3.

Characterization of activated carbons

The specific surface areas of activated carbons were determined by nitrogen adsorption at 77 K using a surface area analyzer (ASAP 2010, Micromeritics, USA). The determination of the point of zero charge (pHpzc) for the activated carbons was carried out as follows (Faria et al., 2004): 50 mL of 0.01 M NaCl solution was placed into each conical flask, and the solution pH was adjusted from 2 to 12 with 0.01 M HCl or NaOH solution. Thereafter, 0.15 g of activated carbon was added into each flask, and the flasks were sealed and shaken at 25  C for 48 h. Finally, the equilibrium solution pH was measured. The

Table 1 – Physicochemical properties of PFOS and PFOA. Compound PFOS PFOA

Mol. formula

Mol. weight

Mol. volume (cm3 mol1)

Water solubility (mg L1)

pKa

C8F17SO3K C7F15 COONa

538 436

257a 226a

570b 3400c

3.27b 2.5c

a Calculated by SPARC, http://ibmlc2.chem.uga.edu/sparc. b Brooke et al. (2004). c USEPA (2002).

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water research 43 (2009) 1150–1158

Table 2 – Characteristics of the three adsorbents used in this study. Adsorbent PAC GAC

Adsorbent AI400 a b c d

d

SBETa(m2 g1)

S1b (m2 g1)

S2c (m2 g1)

pHpzc

Size (mm)

812 712

466 313

346 399

7.5 7.5

<0.1 0.9–1.0

Exchange capacity 

3.0–3.5 eq kg 1

Functional group þ



–N R3 (Cl )

Matrix Polystyrene divinylbenzene

Size (mm) 0.3–1.2

BET surface area. Micropore area. Mesopore and macropore area. Properties given by the manufacturer.

pH point where pHinitial ¼ pHfinal was taken as the pHpzc of the activated carbon. The obtained results are shown in Table 2.

2.4.

Sorption experiments

Batch sorption experiments were carried out at 150 rpm in an orbital shaker with 0.01 g of adsorbents (0.005 g for the sorption kinetics of PFOA on the AI400) in the 250-mL flasks containing 100 mL PFOS or PFOA solution and 2 mM NaH2PO4 as pH buffer. The solution pH was adjusted using 0.1 M NaOH or HCl solution and the pH values were determined by an HQ40d Digital Multi-Parameter Meter from Hach (USA). In the sorption kinetic experiments, 50 mg L1 of PFOS or PFOA solution at initial pH 3 or 7 was used and the sorption experiments were conducted at 25  C. After the sorption experiments, the final solution pH was determined. The sorption isotherm experiments were conducted at the initial PFOS or PFOA concentrations ranging from 20 to 250 mg L1 at 25  C for 168 h (12 h for the PAC), and the solution pH was continuously adjusted to 5 and kept constant during the sorption. As a small volume of NaOH or HCl solution (less than 0.5 mL) was added, the effect of ionic strength on the sorption PFOS or PFOA was negligible. All experiments were conducted twice and the average value was adopted.

2.5.

PFOS and PFOA determination

After the sorption experiments, the mixture was filtrated by a filter with a 0.22 mm nylon membrane. The control experiments indicated that the adsorption of PFOS or PFOA on the membrane was negligible due to their high concentrations in solution. The concentrations of PFOS and PFOA were determined using an LC-10ADvp HPLC with a CDD-6A conductivity detector from Shimadzu (Japan). The TC-C18 column (4.6  250 mm) from Agilent Technologies (USA) was adopted and the mixture of methanol/0.02 M NaH2PO4 (70/30 for PFOS, 65/35 for PFOA, v/v) was used as the mobile phase at 1.5 mL min1 flow rate. The sample volume injected was 20 mL. In this study, the detection limits for PFOS and PFOA are about 1 and 0.7 mg L1, respectively. The sorption amount was calculated according to the difference of PFOS or PFOA concentrations before and after sorption.

3.

Results and discussion

3.1.

Sorption kinetics

Fig. 1 shows the sorption kinetics of PFOS and PFOA on the three adsorbents including the GAC, PAC and AI400. Although all the sorption processes are time dependent, their kinetic profiles are quite different. It can be seen that both PFOS and PFOA displayed the very slow sorption kinetics on the GAC and AI400, and the sorption equilibrium was achieved after at least 168 h, which may be the reason for the fast breakthrough of perfluorinated surfactants when the activated carbon filter was used in German waterworks (Schaefer, 2006). As the GAC and AI400 are granular porous adsorbents with large proportion of micropores and PFOS and PFOA molecules are about 1 nm in length (Erkoc and Erkoc, 2001; Johnson et al., 2007), it took long time for the adsorbates to diffuse into the intraparticle pores. In contrast, only about 4 h was required to reach the sorption equilibrium for the PAC, suggesting that the sizes of activated carbon influence the sorption velocity significantly. The smaller particles have larger external surface area and more functional groups are available for PFOS or PFOA sorption, causing the faster adsorption on the PAC than that on the GAC. To further understand the sorption kinetics, the pseudosecond-order model was selected to fit the kinetic data, which assumes that the sorption rate is controlled by chemical sorption and the sorption capacity is proportional to the number of active sites on the sorbent (Ho and McKay, 1999). This model has been successfully used in many adsorption processes over the whole time range (Wu et al., 2001; Chiou and Li, 2003), which can be expressed as follows (Ho and McKay, 1998). t 1 t 1 t ¼ þ ¼ þ qt k2 q2e qe y0 qe

(1)

where qe and qt are the amount of PFOS or PFOA adsorbed on the adsorbents at equilibrium and time t (mmol g1); k2 is the sorption rate constant (g mmol1 h1); v0 represents the initial sorption rate (mmol g1 h1). As shown in Fig. 1 and Table 3, the pseudo-second-order model fitted all the sorption data well according to the relatively high correlation coefficients (r2 > 0.93), indicating that the chemical interactions were possibly involved in the

water research 43 (2009) 1150–1158

a

0.5

GAC

qt (mmol g-1)

0.4

0.3

0.2 PFOS pH=3 PFOS pH=7 0.1

PFOA pH=3 PFOA pH=7

0.0

0

40

80

120

160

200

t (h)

b

1.0

PAC

qt (mmol g-1)

0.8

0.6

0.4 PFOS pH=3

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USEPA (2002) was adopted in this study. Higgins and Luthy (2006) also found that this interaction played an important role in the sorption of PFOA and PFOS on the sediments. As for the sorption of PFOA and PFOS on the AI400, ion exchange was expected to dominate the sorption, but the sorption amount of PFOA was much higher than that of PFOS on the resin, indicating that the sorption was complex and other interactions may also be involved in the sorption process. The reasons will be discussed in the final paragraph. In addition, it can be seen in Table 3 that the y0 value for the PAC is much higher than that for the GAC and AI400, indicating the fast sorption of PFOS and PFOA on the PAC. Generally, the adsorption process on a porous adsorbent can be divided into three stages. The first stage is called external diffusion, in which the adsorbates move from the bulk solution to the external surface of the adsorbent; the second stage is the intraparticle diffusion, and the adsorbates diffuse further within the adsorbent to the adsorption sites; in the last stage, the adsorbates are adsorbed at the active sites on the adsorbent, which is a fast step and usually can be negligible (Chingombe et al., 2006). As the pseudo-secondorder model cannot give a definite mechanism in the sorption process, the intraparticle diffusion model was adopted to fit the sorption kinetics, which can be expressed as (Boyd et al., 1947; Chiou and Li, 2003; Yang and Al-Duri, 2005)

PFOS pH=7 0.2

PFOA pH=3 PFOA pH=7

0.0

0

2

4

6

8

10

qt ¼

qt (mmol g-1)

AI400

3

2

PFOS pH=3 PFOS pH=7 1

0

PFOA pH=3 PFOA pH=7

0

40

80

120

160

(2)

12

t (h)

c

 1 6qe D 2 1 1 t2 ¼ kd t2 r p

200

t (h) Fig. 1 – Sorption kinetics of PFOS and PFOA on the (a) GAC, (b) PAC and (c) AI400 and modeling using the pseudosecond-order equation.

sorption processes. In consideration of the anion property of the sorbates due to their low pKa (3.27 for PFOS, 2.5 for PFOA, shown in Table 1) and positive surface charge of the activated carbons (pHpzc ¼ 7.5) in the pH range studied, the electrostatic interaction may take place between the sorbates and the sorbents. It should be pointed out that the pKa value for PFOA is hotly debated, and these reported values range from 0.5 to 3.8 (Burns et al., 2008). The value of 2.5 recognized by the

where kd (mmol g1 h0.5) is the intraparticle diffusion rate constant, related to the intraparticle diffusivity; r (mm) is the particle radius; qe (mmol g1) is the equilibrium sorption amount; D (mm2 h1) is the intraparticle diffusivity. The intraparticle diffusion model assumes that the external diffusion is negligible and intraparticle diffusion is the only rate-controlling step, which is usually true for the well-mixed solution (Yang and Al-Duri, 2005). The good linear relationship should be obtained in the plot of qt vs. t0.5 and the line should also pass through the origin if the intraparticle diffusion is the rate-controlling step. Fig. 2 shows the modeling result of PFOS and PFOA sorption on the three adsorbents using the intraparticle diffusion model. The good linear plots between the qt and t0.5 passing through the origin can be seen for the GAC and AI400 at the beginning, implying that the sorption kinetics of PFOS and PFOA on the granular porous adsorbents in the initial stage followed an intraparticle diffusion-controlled adsorption (except the sorption of PFOA on the GAC at pH 3). This stage lasted 72–144 h for PFOS or PFOA sorption at different solution pH. The kinetic plots for the GAC and AI400 in Fig. 2a and c exhibited the two-stage linearity, and the latter is final equilibrium stage where the intraparticle diffusion slowed down. However, this model failed to fit the sorption kinetics of PFOS and PFOA on the PAC since the plots not only displayed a bad linearity but also had significant positive intercepts (Fig. 2b). This result clearly demonstrates that the intraparticle diffusion was not the rate-controlling step in the sorption of PFOS or PFOA on the PAC. The adsorbates may easily diffuse into the inner pores due to the small PAC size, and the external and

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Table 3 – Kinetic parameters of the pseudo-second-order model for PFOS and PFOA sorption on the three adsorbents. Adsorbent

Adsorbate

Final pH

Pseudo-second-order parameter qe (mmol g )

y0 (mmol g1 h1)

k2 (g mmol1 h1)

r2

3.08 7.20 3.10 7.28

0.51 0.41 0.38 0.30

0.02 0.01 0.05 0.01

0.06 0.07 0.37 0.07

0.989 0.988 0.943 0.993

3.00 7.18 3.01 7.20

0.69 0.60 0.79 0.42

2.33 1.96 3.22 2.45

4.89 5.45 5.16 13.9

0.984 0.997 0.990 0.981

3.10 7.10 3.02 7.09

0.36 0.34 3.39 1.81

0.02 0.01 0.11 0.07

0.16 0.12 0.01 0.02

0.930 0.940 0.993 0.992

1

GAC

PFOS PFOA

PAC

PFOS PFOA

AI400

PFOS PFOA

intraparticle diffusion may be comparable or even the external diffusion becomes the rate-limited step. It can be seen in Fig. 1 that solution pH significantly affected the sorption of PFOS and PFOA on the three adsorbents. As mentioned before, the electrostatic interaction should be involved in the sorption of anionic PFOS and PFOA on the activated carbons, which were justified by the enhanced sorption of PFOS and PFOA on the GAC and PAC when solution pH decreased from 7 to 3. According to the obtained parameters in Table 3, the initial sorption rate increased greatly at lower pH as the parameter y0 for the GAC and PAC at pH 3 was much higher than that at pH 7. Solution pH not only influences the properties of the adsorbent surface, but also affects the adsorbate speciation in solution. In Fig. 1, it also can be found that solution pH had more obvious effect on the PFOA sorption onto the three adsorbents than PFOS, which is probably related to their speciation at different pH. As solution pH 3 is very close to the pKa of PFOA, some PFOA may exist in the form of neutral molecules, but all PFOS molecules are still in anionic form because of the negative pKa (3.27). Therefore, the enhanced sorption amount of PFOS at pH 3 can be attributed to the increased protonated groups on the GAC and PAC, while PFOS sorption on the AI400 changed little because no protonated groups were present on the resin. Under the same conditions, the enhanced sorption amount of PFOA on the adsorbents at pH 3 should be less than that of PFOS due to the decreased anionic PFOA, but the much higher enhanced sorption amount of PFOA on the three adsorbents at pH 3 was found in Fig. 1, indicating that other interactions must participate in the sorption at pH 3 and strengthen the sorption of PFOA. In consideration of the hydrophobic perfluorinated chain of the PFOA and the hydrophobicity of the activated carbons and polymer backbone of resin, the hydrophobic interaction should be also involved in the sorption process, especially this interaction becomes more obvious at pH 3 since more hydrophobic neutral PFOA exist in solution. Of course, the hydrophobic interaction is also expected to occur between PFOS and the adsorbents since PFOS has the similar perfluorinated chain as PFOA. At pH 3, the anionic PFOA species in solution account for about 76%, higher than the removal

percent of PFOA in all experiments, and thus the anionic PFOA species are enough to exchange with the anions on the resin in the sorption process. At the same time, the neutral PFOA species (about 24%) may adsorb on the adsorbent via the hydrophobic interaction. Therefore, it’s reasonable that the PFOA removal by the AI400 at pH 3 is much higher than that at pH 7 (shown in Fig. 1c), while the removal of PFOS by the resin hardly changed when solution pH decreased from 7 to 3 due to the constant anionic PFOS. It was reported that the hydrophobic interaction also played an important role in the sorption of PFOS or PFOA on the sediments, sand and clay (Higgins and Luthy, 2006; Johnson et al., 2007). The critical micelle concentration (CMC) for fluorinated surfactants in aqueous solution is mainly dependent on the fluorocarbon chain length and the counterion, while other factors such as temperature, pressure and electrolytes have little effect on the CMC values. For PFOS or PFOA, the CMC values are very different due to the different counterions. As sodium ion is the dominant counterion in solution in our study, the CMC values for PFOS and PFOA should be about 4573 and 15696 mg L1, respectively (Kissa, 2001). Because the PFOS or PFOA concentrations used in this study are far below their CMC values (less than the CMC by a factor of 100 in most cases), almost no micelles can form in solution. However, it is possible to form some hemi-micelles on the adsorbent surface when the PFOS or PFOA concentrations are in the range of 0.01–0.001 of the CMC (Johnson et al., 2007). Moreover, the hemi-micelles and even micelles may also form in the inner pores of the adsorbents after a large number of PFOS or PFOA molecules adsorb on the porous adsorbents, where the PFOS or PFOA concentrations are likely much higher than that in solution. Based on the above discussion, the possible adsorption models and sorbate–sorbent interactions are proposed in Fig. 3.

3.2.

Sorption isotherms

Sorption isotherm is critical to evaluate the sorption capacity of adsorbents as well as understand the sorbate–sorbent interactions. Fig. 4 shows the sorption amount of PFOS and PFOA on the GAC, PAC and AI400 at different equilibrium

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water research 43 (2009) 1150–1158

a

0.5

where qe is the equilibrium sorption amount (mmol g1), Ce represents the equilibrium concentration (mmol L1) of PFOS or PFOA in solution, qm is the maximum sorption capacity (mmol g1), b is the sorption equilibrium constant (L mmol1), K is a constant representing the sorption capacity (mmol11/n L1/n g1), and n is a constant depicting the sorption intensity. As shown in Fig. 4 and Table 4, the sorption isotherms of PFOS on the GAC can be fitted better by the Langmuir model than the Freundlich model, while the isotherms on the PAC were described better by the Freundlich model. Both the Langmuir and Freundlich models can well describe the sorption isotherms on the AI400. Obviously, it cannot be concluded that which model described all sorption isotherms better according to the obtained correlation coefficients. As the Langmuir equation is derived from the assumption of monolayer coverage, the good fitting results of the sorption of PFOS and PFOA on the adsorbents hint that the possible monolayer sorption occurred. It is reasonable that the monolayer

GAC

qt (mmol g-1)

0.4

0.3

0.2

PFOS pH=3 PFOS pH=7 PFOA pH=3

0.1

PFOA pH=7 Linear fitting 0.0

0

4

8

t

b

12

1/2

PAC

0.8

qt (mmol g-1)

0.6

a

B

0.4 PFOS pH=3 PFOS pH=7 PFOA pH=3

0.2

hemi-micelle

PFOA pH=7 Linear fitting 0.0

0

1

2

3

4

t1/2

c

3

B

AI400

B

micelle PFOS pH=3

A

PFOS pH=7

qt (mmol g-1)

PFOA pH=3 2

Activated carbon

PFOA pH=7 Linear fitting

b

-

Cl

N+R3

1

C

0

0

4

8

N+R3

12

N+R3

t1/2

hemi-micelle micelle

Fig. 2 – Intraparticle diffusion model for the sorption of PFOS and PFOA on the (a) GAC, (b) PAC and (c) AI400. concentrations. Two commonly used models, the Langmuir and Freundlich equations (Genc-Fuhrman et al., 2004) were adopted to describe the experimental data, which can be expressed respectively as

N+R3

Cl-N+R3

B

Anion-exchange resin bqm Ce qe ¼ 1 þ bCe 1

qe ¼ KCne

(3)

(4)

Fig. 3 – Schematic diagram of the sorption of PFOS and PFOA on the activated carbon (a) and resin (b) via some possible sorbate–sorbent interactions: A. electrostatic interaction; B. hydrophobic interaction; C. Ion exchange.

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a

water research 43 (2009) 1150–1158

0.4

GAC

qe (mmol g-1)

0.3

0.2 PFOS PFOA

0.1

Langmuir Freundlich 0.0 0.0

0.1

0.2

0.3

0.4

0.5

Ce (mmol L-1)

b

PAC

qe (mmol g-1)

1.0

0.5

PFOS PFOA Langmuir Freundlich

0.0 0.0

0.1

0.2

0.3

0.4

Ce (mmol L-1)

c

AI400

qe (mmol g-1)

3

PFOS

2

PFOA Langmuir Freundlich 1

0

0.0

0.1

0.2

0.3

Ce (mmol L-1) Fig. 4 – Sorption isotherms of PFOS and PFOA on the GAC, PAC and AI400 at 25 8C and modeling using the Langmuir and Freundlich equations. The error bars denote the standard deviation at each equilibrium concentration point. sorption of PFOS or PFOA on the AI400 occurred since ion exchange dominated the sorption, and the sorption amount reached the maximum values when the available sites were saturated with the PFOS or PFOA molecules. Due to the hydrophobic perfluorinated chain of PFOA and PFOS, the multilayer sorption also probably occurred at higher equilibrium concentration. The Freundlich model is an empirical

isotherm model usually used in heterogeneous surface energy systems. According to the results of Freundlich fitting, all the isotherms are nonlinear as the values of n1, an indicator of nonlinearity, are in the range from 0.05 to 0.28 (n1 ¼ 1 for a linear isotherm). Nonlinearity may result from many causes such as the sorption site heterogeneity and sorbate–sorbate interactions (Cheung et al., 2001). In this study, the sorbate– sorbate interactions such as electrostatic repulsion may be mainly responsible for the nonlinearity since the high concentrations of PFOS and PFOA were used and the electrostatic repulsion became significant in the process of intraparticle diffusion. It can be found in Fig. 4 that the sorption amount of PFOS and PFOA on the three adsorbents at the same equilibrium concentration followed the order of GAC < AI400 < PAC and GAC < PAC < AI400, respectively. This result is also in agreement with the order of qm and K, which are also the indicators of the sorption capacity of the adsorbents. Although the GAC and PAC were obtained from the same commercial activated carbons, their sorption capacities were quite different from each other. The differences can be partially attributed to their different physical properties such as surface area and pore distribution, shown in Table 2. The specific surface area of the PAC is a little higher than that of the GAC, which may explain the higher sorption capacity on the PAC to some extent, but the slightly higher surface area cannot lead to about double sorption capacity of the PAC for PFOS. Due to the big size of PFOS and PFOA molecules, some of micropores (less than 2 nm) available for N2 adsorption may be unaccessible for PFOS or PFOA molecules. Therefore, the available micropores for the sorption of PFOS and PFOA on the adsorbents are more important than the determined ones. Because of the smaller size of the PAC than the GAC, more micropores and surface functional groups on the PAC would be easily exposed to the sorbates and some active sites originally unapproachable for the sorbates in the GAC probably became available ones on the PAC, leading to the higher sorption capacity. Especially, as shown in Fig. 3a, the micropores on the surface may get blocked by the adsorbed PFOS or PFOA and even the formed micelles or hemi-micelles, making the intraparticle pores impassable for the coming adsorbates. This effect will become more obvious on the GAC than that on the PAC (Ahn et al., 2007). It is surprise to find that the sorption capacity of PFOA on the AI400 is over 5 times as large as that of PFOS, while the activated carbons prefer the sorption of PFOS to PFOA. If ion exchange is the dominant mechanism in the sorption process and all the exchange sites are available for the adsorbates, the sorption amount of PFOS and PFOA on the AI400 at pH 5 should be comparable. The significant difference in sorption capacity of PFOS and PFOA on the AI400 may be attributed to their different CMC and molecular size. As the PFOS molecular volume is bigger than PFOA (shown in Table 1) and the AI400 is a gel type micropore resin, the PFOS molecules are more difficult to diffuse into the AI400. Moreover, since the PFOS CMC is much lower than that of PFOA, it is easier for PFOS to form the hemi-micelles or micelles on the surface or inside the pores of the resin. Due to the bigger size of PFOS, the micelles or hemi-micelles formed by PFOS should have bigger sizes than that by PFOA, and easily block the pores and

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Table 4 – Calculated equilibrium constants using the Langmuir and Freundlich equations for PFOS and PFOA sorption on the three adsorbents. Adsorbent

Adsorbate

Langmuir constants 1

1

Freundlich constants 2

qm (mmol g )

b (L mmol )

r

(11/n)

K (mmol

L1/n g1)

n1

r2

GAC

PFOS PFOA

0.37 0.39

39 18

0.964 0.965

0.43 0.47

0.18 0.28

0.869 0.974

PAC

PFOS PFOA

1.04 0.67

55 59

0.835 0.911

1.27 0.83

0.18 0.20

0.960 0.950

AI400

PFOS PFOA

0.42 2.92

69 69

0.944 0.967

0.52 3.35

0.17 0.13

0.930 0.987

prevent the intraparticle diffusion of other PFOS molecules, leading to the much lower sorption amount of PFOS on the AI400. As perfluorinated chain is hydrophobic and oleophobic, the adsorbed PFOA molecules may adsorb other PFOA molecules and even form hemi-micelles or micelles at high concentrations, shown in Fig. 3b. The higher sorption capacity of PFOS on the GAC and PAC may be related to the more hydrophobic PFOS. As PFOS has longer perfluorinated chain than PFOA, the adsorbed PFOS molecules prefer to adsorb other PFOS molecules to form the hemi-micelles or micelles. In fact, the sorption mechanisms of PFOS and PFOA on the adsorbents are complex, and these possible mechanisms need to be verified in the near future.

Acknowledgments We thank the National Nature Science Foundation of China (project no. 50608045 and 50778095), special fund of State Key Joint Laboratory of Environment Simulation and Pollution (project no. 08Z04ESPCT), and National Outstanding Youth Foundation of China (50625823) for financial support, and the Program for New Century Excellent Talents in University is also appreciated. Additionally, the analytical work was supported by the Laboratory Found of Tsinghua University.

references

4.

Conclusions

The sorption kinetics and isotherms of PFOS and PFOA on the activated carbons and anion-exchange resin were investigated and the PAC was found to be the best adsorbent for PFOS in terms of sorption kinetics and sorption capacity, while the AI400 had the highest sorption capacity for PFOA. The sorption kinetic results reveal that the sorption of PFOS and PFOA on the granular porous adsorbents including GAC and AI400 was very slow, and the sorption equilibrium was achieved after at least 168 h, while the sorption equilibrium time was only about 4 h using the powder activated carbon. Obviously, the adsorbent size significantly affected the sorption velocity of PFOS and PFOA. The pseudo-second-order model can well describe the sorption kinetics of PFOS and PFOA on the three adsorbents, and the intraparticle diffusion model can well fit their sorption on the GAC and AI400 in the initial stage. The sorption isotherms show that the PAC and AI400 were the best adsorbents for PFOS and PFOA, respectively, and their maximum sorption capacities were 1.04 and 2.92 mmol g1 according to the Langmuir model. Additionally, their sorption capacity for PFOS and PFOA increased with decreasing solution pH. Besides the electrostatic interaction and ion exchange, the hydrophobic interaction was also possibly involved in the sorption. PFOS and PFOA may form the hemimicelles or micelles in the adsorbent pores, which significantly affected the sorption kinetic and sorption capacity. The appropriate PAC is the promising adsorbent for PFOS and PFOA removal from water.

Ahn, C.K., Kim, Y.M., Woo, S.H., Park, J.M., 2007. Selective adsorption of phenanthrene dissolved in surfactant solution using activated carbon. Chemosphere 69, 1681–1688. Boyd, G.E., Adamson, A.W., Myers Jr., L.S., 1947. The exchange adsorption of ions from aqueous solutions by organic zeolites. II. Kinetics. J. Am. Chem. Soc. 69, 2836–2848. Brooke, D., Footitt, A., Nwaogu, T.A., 2004. Environmental Risk Evaluation Report: Perfluorooctane Sulfonate (PFOS). UK Environment Agency. Burns, D.C., Ellis, D.A., Li, H.X., McMurdo, C.J., Webster, E., 2008. Experimental pKa determination for perfluorooctanoic acid (PFOA) and the potential impact of pKa concentration dependence on laboratory-measured partitioning phenomena and environmental modeling. Environ. Sci. Technol. 42, 9283–9288. Cheung, C.W., Porter, J.F., McKay, G., 2001. Sorption kinetic analysis for the removal of cadmium ions from effluents using bone char. Water Res. 35, 605–612. Chingombe, P., Saha, B., Wakeman, R.J., 2006. Sorption of atrazine on conventional and surface modified activated carbons. J. Colloid Interface Sci. 302, 408–416. Chiou, M.S., Li, H.Y., 2003. Adsorption behavior of reactive dye in aqueous solution on chemical cross-linked chitosan beads. Chemosphere 50, 1095–1105. Erkoc, S., Erkoc, F., 2001. Structural and electronic properties of PFOS and LiPFOS. J. Mol. Struct. (Theochem) 549, 289–293. Faria, P.C.C., Orfao, J.J.M., Pereira, M.F.R., 2004. Adsorption of anionic and cationic dyes on activated carbons with different surface chemistries. Water Res. 38, 2043–2052. Fujii, S., Polprasert, C., Tanaka, S., Lien, N.P.H., Qiu, Y., 2007. New POPs in the water environment: distribution, bioaccumulation and treatment of perfluorinated compounds – a review paper. J. Water Supply Res. Technol. – Aqua 56, 313–326.

1158

water research 43 (2009) 1150–1158

Genc-Fuhrman, H., Tjell, J.C., McConchie, D., 2004. Adsorption of arsenic from water using activated neutralized red mud. Environ. Sci. Technol. 38, 2428–2434. Giesy, J.P., Kannan, K., 2002. Perfluorochemical surfactants in the environment. Environ. Sci. Technol. 36, 146A–152A. Hansen, K.J., Johnson, H.O., Eldridge, J.S., Butenhoff, J.L., Dick, L.A., 2002. Quantitative characterization of trace levels of PFOS and PFOA in the Tennessee River. Environ. Sci. Technol. 36, 1681– 1685. Higgins, C.P., Luthy, R.G., 2006. Sorption of perfluorinated surfactants on sediments. Environ. Sci. Technol. 40, 7251–7256. Ho, Y.S., McKay, G., 1998. A comparison of chemisorption kinetic models applied to pollutant removal on various sorbents. Process Saf. Environ. Protect. 76, 332–340. Ho, Y.S., McKay, G., 1999. Pseudo-second order model for sorption processes. Process Biochem. 34, 451–465. Hori, H., Nagaoka, Y., Yamamoto, A., Sano, T., Yamashita, N., Taniyasu, S., Kutsuna, S., Osaka, I., Arakawa, R., 2006. Efficient decomposition of environmentally persistent perfluorooctane sulfonate and related fluorochemicals using zerovalent iron in subcritical water. Environ. Sci. Technol. 40, 1049–1054. Johnson, R.L., Anschutz, A.J., Smolen, J.M., Simcik, M.F., Penn, R.L., 2007. The adsorption of perfluorooctane sulfonate onto sand, clay, and iron oxide surfaces. J. Chem. Eng. Data 52, 1165–1170. Key, B.D., Howell, R.D., Criddle, C.S., 1997. Fluorinated organics in the biosphere. Environ. Sci. Technol. 31, 2445–2454. Key, B.D., Howell, R.D., Criddle, C.S., 1998. Defluorination of organofluorine sulfur compounds by Pseudomonas sp. strain D2. Environ. Sci. Technol. 32, 2283–2287. Kissa, E., 2001. Fluorinated Surfactants and Repellents, second ed. Marcel Dekker, New York. Lampert, D.J., Frisch, M.A., Speitel Jr., G.E., 2007. Removal of perfluorooctanoic acid and perfluorooctane sulfonate from wastewater by ion exchange. Pract. Periodical Haz., Toxic, and Radioactive Waste Mgmt. 11, 60–68. Loos, R., Locoro, G., Huber, T., Wollgast, J., Christoph, E.H., De Jager, A., Gawlik, B.M., Hanke, G., Umlauf, G., Zaldivar, J.M., 2008. Analysis of perfluorooctanoate (PFOA) and other perfluorinated compounds (PFCs) in the River Po watershed in N-Italy. Chemosphere 71, 306–313. Moody, C.A., Martin, J.W., Kwan, W.C., Muir, D.C.G., Mabury, S.C., 2002. Monitoring perfluorinated surfactants in biota and surface water samples following an accidental release of firefighting foam into Etohicoke Creek. Environ. Sci. Technol. 36, 545–551. Moriwaki, H., Takagi, Y., Tanaka, M., Tsuruho, K., Okitsu, K., Maeda, Y., 2005. Sonochemical decomposition of

perfluorooctane sulfonate and perfluorooctanoic acid. Environ. Sci. Technol. 39, 3388–3392. Ochoa-Herrera, V., Sierra-Alvarez, R., 2008. Removal of perfluorinated surfactants by sorption onto granular activated carbon, zeolite and sludge. Chemosphere 72, 1588–1593. Ochoa-Herrera, V., Sierra-Alvarez, R., Somogyi, A., Jacobsen, N.E., Wysocki, V.H., Field, J.A., 2008. Reductive defluorination of perfluorooctane sulfonate. Environ. Sci. Technol. 42, 3260–3264. Prevedouros, K., Cousins, I., Buck, R.C., Korzeniowski, S.H., 2006. Sources, fate and transport of perfluorocarboxyates. Environ. Sci. Technol. 40, 32–44. Schaefer, A., 2006. Perfluorinated surfactants contaminate German waters. Environ. Sci. Technol. 40, 7108–7109. Schroder, H.F., Meesters, R.J.W., 2005. Stability of fluorinated surfactants in advanced oxidation processes – a follow up of degradation products using flow injection-mass spectrometry, liquid chromatography–mass spectrometry and liquid chromatography–multiple stage mass spectrometry. J. Chromatogr. A 1082, 110–119. Sinclair, E., Mayack, D.T., Roblee, K., Yamashita, N., Kannan, K., 2006. Occurrence of perfluoroalkyl surfactants in water, fish, and birds from New York State. Arch. Environ. Contam. Toxicol. 50 (3), 398–410. Tang, C.Y.Y., Fu, Q.S., Robertson, A.P., Criddle, C.S., Leckie, J.O., 2006. Use of reverse osmosis membranes to remove perfluorooctane sulfonate (PFOS) from semiconductor wastewater. Environ. Sci. Technol. 40, 7343–7349. Tang, C.Y.Y., Fu, Q.S., Criddle, C.S., Leckie, J.O., 2007. Effect of flux (transmembrane pressure) and membrane properties on fouling and rejection of reverse osmosis and nanofiltration membranes treating perfluorooctane sulfonate containing wastewater. Environ. Sci. Technol. 41, 2008–2014. USEPA, 2002. Revised Draft Hazard Assessment of Perfluorooctanoic Acid and its Salts Available from: http:// www.ewg.org/files/EPA_PFOA_110402.pdf. Wu, F.C., Tseng, R.L., Juang, R.S., 2001. Kinetic modeling of liquidphase adsorption of reactive dyes and metal ions on chitosan. Water Res. 35, 613–618. Yamamoto, T., Noma, Y., Sakai, S.I., Shibata, Y., 2007. Photodegradation of perfluorooctane sulfonate by UV irradiation in water and alkaline 2-propanol. Environ. Sci. Technol. 41, 5660–5665. Yang, X., Al-Duri, B., 2005. Kinetic modeling of liquid-phase adsorption of reactive dyes on activated carbon. J. Colloid Interface Sci. 287, 25–34. Yu, Q., Deng, S.B., Yu, G., 2008. Selective removal of perfluorooctane sulfonate from aqueous solution using chitosan-based molecularly imprinted polymer adsorbents. Water Res. 42, 3089–3097.