Desalination 281 (2011) 68–74

Contents lists available at ScienceDirect

Desalination

j ourna l homepage: www.e lsev ie r.com/ locate /desa l

Adsorption of bisphenol A and 17α-ethinyl estradiol on single walled carbonnanotubes from seawater and brackish water

Lesley Joseph a, Jiyong Heo a, Yong-Gyun Park b, Joseph R.V. Flora a, Yeomin Yoon a,⁎a Department of Civil and Environmental Engineering, University of South Carolina, Columbia, SC 29208, USAb Environmental & Energy Research Team, GS E&C Research Institute, 417-1 Deokseong-ri Idong-myeon Cheoin-gu Yongin-si, Gyeonggi-do, 449-831, South Korea

⁎ Corresponding author. Tel.: +1 803 777 8952; fax:E-mail address: yoony@cec.sc.edu (Y. Yoon).

0011-9164/$ – see front matter © 2011 Elsevier B.V. Adoi:10.1016/j.desal.2011.07.044

a b s t r a c t

a r t i c l e i n f o

Article history:Received 16 April 2011Received in revised form 18 July 2011Accepted 20 July 2011Available online 11 August 2011

Keywords:Single walled carbon nanotubesSeawaterBrackish waterBisphenol A17α-ethinyl estradiolAdsorption

Recent studies have shown the presence of endocrine disrupting compounds (EDCs) in seawater and brackishwater, which could potentially complicate various seawater desalination treatment processes. In this study,the adsorption of bisphenol A (BPA) and 17α-ethinyl estradiol (EE2) by single walled carbon nanotubes(SWCNTs) was investigated. Solutions of artificial seawater, brackish water, and a combination of these twowaters were prepared, in accordance with previously published composition data. Overall, the removalefficiency for EE2 (95–98%) was higher than BPA (75–80%), possibly because of its higher log KOW value. Theadsorptive capacity of the SWCNTs remained relatively constant for the artificial source waters used in thisstudy, suggesting that the changes in the composition of the water did not affect the overall adsorption of theEDCs. Adjusting the pH of the solutions from 3.5 to 11 showed a 22–26% decrease in the adsorption of BPA,whereas no notable changes were found in the adsorption of EE2. Changes in the ionic strength of thesolutions by increasing the concentrations of Na+ and Ca2+ did not significantly affect the adsorption of BPAor EE2. The concentration of dissolved organic carbon (DOC), represented in this study by humic acid, hadnoticeable effects on the adsorption of BPA and EE2. As the concentration of DOC increased, the removal ofBPA and EE2 decreased by 5–15%, which could possibly be explained by competitive adsorption between theEDCs and humic acid. With increasing concentrations of SWCNTs, adsorption of DOC occurred with removalefficiencies of up to 95%. Hydrophobic interactions and π–π electron donor–acceptor (EDA) interactionsamong the EDCs, the DOC, and the SWCNTs have been hypothesized as the potential adsorption mechanismsfor BPA and EE2.

+1 803 777 0670.

ll rights reserved.

© 2011 Elsevier B.V. All rights reserved.

1. Introduction

Recently, the occurrence of endocrine disrupting compounds(EDCs) in the aquatic environment has caused concern for researchersin the scientific and public health communities. EDCs are chemicalsthat mimic or block the activity of natural hormones in humans andaquatic wildlife, thus disrupting their reproductive systems [1]. Themajority of EDCs that have entered the environment are pharmaceu-ticals excreted from humans at low concentrations, which aresubsequently discharged into water and wastewater streams [2].Various EDCs have been detected in surface water [3], wastewater [4],and landfill leachates [5,6]. Among the EDCs that are prevalent in theaquatic environment, bisphenol A (BPA), a common plasticizer, and17α-ethinyl estradiol (EE2), a synthetic estrogen, have been mostfrequently researched in water treatment [7]. With the reportedadverse health effects of exposure to EDCs, such as decreased spermcount, reduced fertility, and increased incidences of breast, ovarian,

and testicular cancers [8,9], effective removal of these contaminants isvital to maintaining public human health.

Currently, the majority of the EDCs that have been reported arefound in municipal wastewater [10]. However, researchers have alsodetected various EDCs in seawater sources; their presence couldpossibly be attributed to the transport of contaminated wastewatereffluent by rivers into larger water bodies such as oceans and seas[11]. Several EDCs, including nonylphenol mixtures, BPA, and EE2,have been found in concentrations ranging from 31 to 1777 ng/L, 11 to777 ng/L, and 10 to 269 ng/L, respectively [12]. The use of reverseosmosis (RO) as the primary seawater desalination method hasincreased greatly, due to its high rejection rate of monovalent ions(i.e., Na+ and Cl−), resulting in salt rejections of over 99% [13].However, a major issue with the use of RO membranes in seawaterdesalination is membrane fouling that can occur by the presence ofparticulate matter [14], dissolved organic matter [14], and organiccompounds [15]. To reduce membrane fouling, pretreatment is oftenconducted, using methods such as coagulation and flocculation [16],packed bed filtration [17,18], and precipitation processes [19].

Another popular form of pretreatment is the use of adsorbents,particularly powdered activated carbon (PAC) and granular

69L. Joseph et al. / Desalination 281 (2011) 68–74

activated carbon (GAC). Activated carbon has been shown toremove a wide variety of dissolved organic and inorganiccontaminants, which is ideal for membrane pretreatment[20,21]. Although activated carbon is a widely used adsorbent inwastewater treatment and seawater desalination, studies havebeen conducted to compare its removal efficiencies to those ofcarbon nanotubes (CNTs). CNTs have shown high adsorptioncapacities for phenols [22,23], heavy metals [24,25], and naturalorganic matter [26]. Studies have shown that CNTs can be moreeffective than activated carbon, due to their larger surface area[27] and shorter equilibrium times [28]. Recent studies have alsoexamined the adsorptive capacities and mechanisms of variousEDCs onto CNTs [29]. The objective of this study is to investigatethe adsorption of BPA and EE2 in seawater and brackish waterby single walled carbon nanotubes (SWCNTs). With the possibilityof membrane fouling by the presence of EDCs and natural organicmatter (NOM) in seawater and brackish water, analyzing theremoval efficiency of SWCNTs in highly saline source water (i.e.,waters with concentration of total dissolved solids greater that30,000 mg/L) is important with respect to the overall pretreat-ment process for membrane filtration. The effects of differentwater quality conditions (e.g., pH, ionic strength, and DOC) on theadsorption of the EDCs by SWCNTs must be investigated due tothe variability of seawater conditions around the world. SinceNOM is also considered to be a major foulant for membranes, itspotential removal and effects on overall adsorption must also beinvestigated.

2. Materials and methods

2.1. Materials

BPA (purity N99%), EE2 (purity N98%), and humic acid werepurchased from Sigma-Aldrich (St. Louis, MO, USA). Table 1 liststhe physicochemical properties of BPA and EE2. BPA and EE2 weredissolved separately in methanol as stock solutions of 1 mM each.Humic acid stock solution was prepared by adding 1 g of dryhumic acid powder to 1 L of deionized (DI) water and stirredovernight. The solution was then filtered through 0.7 μm glassmicrofiber filters (Whatman, Buckinghamshire, UK) to removeparticulate matter greater than 0.7 μm. Na2SO4 (purity N99%), KCl(purity N99%), NaHCO3 (purity N99.5%), MgCl2 (purity N96%), andCaCl2 (purity N93%) were purchased from Sigma-Aldrich. NaCl(purity N99%) was purchased from Fisher Scientific. SWCNTs(purity N90%) were purchased from Cheap Tubes, Inc. (Brattle-boro, Vermont, USA) and used without further purification.SWCNTs have a length of 5–30 μm and an outer diameter of 1–2 nm, as provided by the manufacturer [30].

Table 1Characteristics of bisphenol A and 17α-ethinylestradiol.

Compound Use Molecular weight (g/mol)

Bisphenol A Plasticizer 228.1

17α-ethinylestradiol

Ovulation inhibitor 296.2

2.2. Preparation of artificial seawater and brackish water solutions

To investigate the adsorption of BPA and EE2 from seawater andbrackish water, solutions of artificial seawater, brackish water, and acombination of both seawater and brackish water (SW/BW) wereprepared. Volumes of the BPA and EE2 stock solutions were placedtogether in separate beakers at concentrations equaling 1 and 10 μM,and methanol was evaporated from the stock solution at roomtemperature under a fume hood to remove its DOC. Dilutions of humicacid were taken from the stock solution and added to achieve thedesired concentration of DOC. From preliminary DOC analysis, thehumic acid used in this study was estimated to consist of 40% DOC.This percentage was used to estimate the DOC concentrations of theartificial seawater and brackish water solutions, which were thenconfirmed using the analytical methods described in this study (SeeSection 2.4 “Analytical methods”). Then, the other constituents in theartificial source waters were added using the concentrations listed inTable 2 and stirred overnight. The compositions of the artificialseawater and brackish water were determined on the basis of aliterature review of several studies related to the treatment andcharacterization of seawater [31,32] and brackish water [33,34]. Thecomposition of the combination of seawater and brackish water wasdetermined by using approximately median concentrations of thecomponents represented in the standard seawater and brackishwater. The pH was adjusted to the desired values using 1 M NaOH orHCl.

2.3. Adsorption experiments

A stock solution of 1000 mg/L of SWCNTs was prepared inultrapure DI water before conducting the experiments. Toreplicate its potential use in a commercial water treatment system,the SWCNTs used in this study were not subjected to anytreatment (e.g., sonication and heating). In present-day watertreatment applications, adsorbents (e.g., activated carbon) arepurchased from the manufacturer and placed directly into theexisting flocculation tanks to facilitate the removal of anycontaminants. This method was adopted for the batch experi-ments conducted throughout this study.

Adsorption isotherms were investigated using a batch tech-nique under ambient conditions, using the artificial seawater andbrackish water solutions, as described in Table 2. Applied doses ofSWCNTs ranged from 0 to 50 mg/L. SWCNTs were diluted usingvolumetric extractions from the continuously mixed stock solutionof SWCNTs. The samples containing BPA and EE2 were placed in40 mL screw-cap glass vials, spiked with the SWCNTs, and sealedwith Teflon screw caps. The vials were placed in a shaker for 4 h at13.9 rpm. Preliminary kinetic experiments showed that this

LogKOW pKa Structure

3.3 9.6–10.2

3.7 ~10.5

Table 2Characteristics of standard and artificial seawaters used in the study.

Chemical ion (mg/L)

Standard source water Artificial source water

Seawatera Brackishwaterb

Seawater SW/BW

Brackishwater

Chloride, Cl− 19,345 72–1867 19,346 5589 1302Sodium, Na+ 10,752 170–905 10,763 2940 828Sulfate, SO4

2− 2701 384–670 2711 1100 550Magnesium, Mg2+ 1295 58–132 1295 500 100

Calcium, Ca2+ 416 175–179 414 200 100Potassium, K+ 390 – 355 150 50Bicarbonate, HCO3− 145 146–260 142 142 142

Bromide, Br− 66 – – – –

Boromide, BO33− 27 – – – –

Strontium, Sr2+ 13 0–26.4 – – –

Fluoride, F− 1 – – – –

DOC – 0–1.4 2 2 2Totaldissolved solids

35,000 1320–3664 35,028 10,624 3075

a Range of values taken from Kester et al.[32] and Al-Rasheed and Cardin [31].b Range of values taken from Gaid and Treal [33] and Greenlee et al. [34].

70 L. Joseph et al. / Desalination 281 (2011) 68–74

amount of time was sufficient to attain an appropriate pseudo-equilibrium. Each sample was filtered with a 0.45 μm membranefilter and transferred to 2 mL amber vials for high-performanceliquid chromatography (HPLC) analysis. Control experimentsconfirmed that the membrane filters used in this study did notimpact the concentration of BPA and EE2 in the samples. Duplicateexperiments were conducted to account for experimental errorand to ensure the reproducibility of the results. The error barsshown on the figures represent the standard deviation of thesamples taken during the experiments. The amount of BPA and

3

3.5

4

4.5

5

0 1 2 3

log

q e (µ

g/g)

log Ce (µg/L)

a.1

0

0.005

0.01

0.015

0.02

0 0.1 0.2

Ce/

q e (m

g/g)

Ce(mg/L)

b.1

Fig. 1. Effect of various source waters on the adsorption of BPA and EE2 by SWCNTs (C0=1 μM(b.2). The composition of each source water is described in detail in Table 2. Lines denote li(Δ); SW/BW(○); brackish water (□).

EE2 adsorbed onto the CNTs was calculated using the followingequation:

q = C0−Ctð Þ Vm

ð1Þ

where q is the amount of BPA or EE2 adsorbed onto CNTs (mg/g); C0and Ct are the concentrations at the beginning and after a certainperiod of time (mg/L), respectively; V is the volume of the initialsolution (L); and m is the mass of the CNTs (g).

2.4. Analytical methods

Concentrations of BPA and EE2 were quantified using HPLC1200 Series (Agilent Technologies, Santa Clara, CA, USA). Detectionwas carried out using a fluorescence detector at an excitationwavelength of 280 nm and an emission wavelength of 310 for bothBPA and EE2. A Waters 5-μm LiChrosorb® RP18 analytical column(4.6 mm×100 mm) was used for reverse-phase separations with a100-μL sample loop. The mobile-phase solvent profile was 45% DIwater acidified with 10 mM H3PO4 and 55% MeOH for 30 min at aconstant flow rate of 1 mL/min. The detection limits were 0.88 nM(201 ng/L) for BPA and 0.96 nM (283 ng/L) for EE2. BPA and EE2eluted from the columns after 9.4 and 20.3 min, respectively. Thedescription of the HPLC method employed in this study has beendescribed elsewhere [35]. DOC was measured using ultraviolet(UV)–visible spectroscopy (Agilent Technologies, Santa Clara, CA,USA). Preliminary experiments showed that the correlationbetween DOC concentrations and absorbance at 254 nm provedto be nearly linear in both DI water and seawater water. For DOCanalysis with seawater samples, UV absorbance (UVA) is a goodalternative because it provides high precision at low DOC levels. Asimilar method was also performed in a previous study [31].

3

3.5

4

4.5

5

log

q e (µ

g/g)

log Ce (µg/L)0 1 2 3

a.2

0

0.005

0.01

0.015

0.02

Ce/

q e (m

g/g)

Ce(mg/L)0 0.1 0.2

b.2

; contact time=4 h). Freundlich: BPA (a.1) and EE2 (a.2); Langmuir: BPA (b.1) and EE2nearized model fitting for Freundlich and Langmuir isotherms. Source water: seawater

0

0.2

0.4

0.6

0.8

1

0 5 10 15

C/C

0C

/C0

C/C

0

pH

0 5 10 15

pH

0 5 10 15

pH

a

0

0.2

0.4

0.6

0.8

1

b

0

0.2

0.4

0.6

0.8

1c

Fig. 2. Effect of pH on the adsorption of BPA and EE2 by SWCNTs: (a) seawater; (b) SW/BWwater; and (c) brackish water (C0=10 μM; contact time=4 h; CNT dose=30mg/L). BPA(▲); EE2 (□).

71L. Joseph et al. / Desalination 281 (2011) 68–74

2.5. Isotherm modeling

The adsorption data from the experiments were fitted to twodifferent isotherm models:

FreundlichModel : qe = Kf C1=ne ð2Þ

LangmuirModel : qe =abCe

1 + bCeð3Þ

where qe is the solid-phase concentration (mg/g), Kf is the Freundlichadsorption coefficient ((mg/g)/(mg/L)1/n), Ce is the equilibriumsolution phase concentration (mg/L), n is the dimensionless numberrelated to surface heterogeneity, a is the maximum adsorptioncapacity, and b is the Langmuir fitting parameter (L/mg).

3. Results and discussion

3.1. Adsorption of BPA and EE2 in artificial source water

Linearized Freundlich and Langmuir isotherm data fitting wereused to show the adsorption of BPA and EE2 in the various types ofsource water by SWCNTs (Fig. 1). Langmuir and Freundlich isothermshave been frequently used to describe the adsorption of organicchemicals by CNTs [36–39]. Langmuir isotherms describe single-layeradsorption, where the chemical only interacts with the CNT surfacethroughout adsorption. Freundlich isotherms, however, describemulti-layer adsorption, where the chemical initially interacts withthe SWCNT surfaces, and then, the chemicals react with each other.Table 3 lists the fitting parameters of both isotherm models for thedifferent source waters used in this study. Based on the highcorrelation values (R2), the adsorption by SWCNTs could be explainedusing either isotherm model. The log Kf values for the Freundlichmodel and the a values for the Langmuir model did not varysignificantly among the different source waters, indicating that theadsorptive capacity of the SWCNTs was generally unaffected by theoverall composition of the water used in this study. However, theadsorptive capacity of the SWCNTs differed between BPA and EE2. Asshown in Table 3, the range of log Kf values for EE2 (3.39–3.71) wasslightly higher than that for BPA (3.20–3.47). Also, the range of avalues for EE2 (34.60–35.71) was much higher than that for BPA(13.39–16.05). The higher adsorptive capacity of EE2 than BPA canpossibly be explained by its larger log KOW values, which suggests thathydrophobic interactions between the EDCs and the SWCNTs may bethe dominant removal mechanism. This mechanism has been cited inseveral other adsorption studies using SWCNTs [40,41]. CNT surfaceshave been described as “evenly distributed hydrophobic sites” fororganic chemical adsorption [42], which reinforces the possibility ofhydrophobic interactions dominating the adsorption of BPA and EE2.However, another possible mechanism for adsorption of EDCs,particularly BPA and EE2, may be π–π electron donor–acceptor(EDA) interactions, as shown in previous studies [7].

Table 3Freundlich and Langmuir fitting parameters for adsorption of BPA and EE2 (C0=1 μM) ont

Chemical Source water Freundlich

log Kf n−

((mg/g)/(mg/L)1/n)

BPA Seawater 3.20 0.4SW/BW 3.47 0.2Brackish water 3.37 0.3

EE2 Seawater 3.39 0.4SW/BW 3.41 0.4Brackish water 3.71 0.3

3.2. Effect of pH and ionic strength

The effects of pH on the adsorption of BPA and EE2 from seawater,SW/BW, and brackish water are shown in Fig. 2a, b, and c,respectively. The pH of these source waters without adjustment was8.2±0.1, which is consistent with other published literature.Adjusting the pH of each type of source water from 3.5 to 8.5 didnot affect the adsorption of BPA. However, slightly decreasedadsorption was observed when the pH was increased to 11. Theadsorption of EE2 was not notably affected by changes in the pH.These trends were consistent with each type of source water. Also, noprecipitates (e.g., Mg(OH)2) were observed with the increase in pH;

o CNTs in various source waters.

Langmuir

1 R2 a b R2

(mg/g) (L/mg)

04 0.969 13.39 41.49 0.96381 0.988 13.83 40.17 0.98645 0.974 16.05 29.67 0.95763 0.903 35.46 14.84 0.91283 0.989 34.60 19.26 0.97647 0.972 35.71 31.11 0.931

0

0.2

0.4

0.6

0.8

1

0 0.2 0.4 0.6

C/C

0

0

0.2

0.4

0.6

0.8

1

C/C

0

Na+ conc. (M)

0 0.2 0.4 0.6

Na+ conc. (M)

a

b

Fig. 3. Effect of Na+ concentration on the adsorption of BPA and EE2 by SWCNTs inseawater with initial EDC concentrations of (a) 1 μM with CNT dose=15 mg/L and (b)10 μM with CNT dose=30 mg/L. Contact time=4 h. BPA (▲); EE2 (□).

0

0.2

0.4

0.6

0.8

1

0 0.01 0.02

C/C

0

0

0.2

0.4

0.6

0.8

1

C/C

0

0

0.2

0.4

0.6

0.8

1

C/C

0

Ca2+ conc (M)

Ca2+ conc (M)

Ca2+ conc (M)

a

0 0.005 0.01

b

0 0.005 0.01

c

Fig. 4. Effect of Ca2+on the adsorption of BPA andEE2by SWCNTs: (a) seawater; (b) SW/BWwater; and (c) brackish water (C0=10 μM; contact time=4 h; CNT dose=30mg/L). BPA(▲); EE2 (□).

72 L. Joseph et al. / Desalination 281 (2011) 68–74

therefore, the composition of the source waters remained constantthroughout the adsorption process. The contrast between thereduction in adsorption of BPA and the consistent adsorption of EE2with varying pH may be attributed to the different log KOW values ofeach compound. The higher log KOW value of EE2 would suggest thatits hydrophobic interactions aremore prominent and remain constantthroughout adsorption. The decrease in adsorption of BPA may be aresult of the increased ionization caused by the increase in pH, whichcan lead to reduced hydrophobic interactions with the SWCNTs [7]. Ifπ–π EDA interactions contribute to the adsorption of BPA, decreasedadsorption can be expected when the pH is higher than the pKa,because the chemical and the SWCNTs are both negatively chargedand experience increased electrostatic repulsion [42]. The surfacecharge (expressed as zeta potential, mV) of the SWCNTs used in thisstudy is negative (data available on request). Thus, the adsorption ofBPA remains fairly constant until the pH exceeds pKa for BPA, at whichpoint the adsorption is reduced.

The effects of ionic strength on the adsorption of BPA and EE2were investigated by adjusting the concentrations of Na+ inseawater with initial EDC concentrations of 1 and 10 μM, as shownin Fig. 3a and b, respectively. Due to the varying magnitude of EDCconcentrations, SWCNT doses of 15 and 30 mg/L were used forsolutions containing EDC concentrations of 1 and 10 μM, respec-tively. Increases in Na+ concentration from 0 to 0.5 M resulted insmall variations in the adsorption of BPA in seawater, with a slightincrease in adsorption occurring when the initial concentrationwas 1 μM (Fig. 3a) and a small decrease in adsorption occurringwhen the initial concentration was 10 μM (Fig. 3b). However,these increases in the Na+ concentration did not significantly alterthe adsorption of EE2. The effects of ionic strength were alsoinvestigated by changing the concentration of Ca2+ for seawater,SW/BW, and brackish water, as shown in Fig. 4a, b, and c,respectively. The overall adsorption of BPA and EE2 by SWCNTswas not influenced by increases in the concentration of Ca2+ in theseawater or SW/BW source water. However, an increase inadsorption was observed for both BPA and EE2 in brackish waterwith increasing Ca2+ concentration (Fig. 4c). In several analyses of

the effect of ionic strength on adsorption, it has been shown thatincreases in ionic strength potentially enhance the adsorption ofcompounds onto carbon materials due to the screening effect ofthe surface charge that is produced by adding salt [43]. On theother hand, many researchers have found that the effect of ionicstrength on the adsorption of organic chemicals is negligible[41,44]. The “salting-out” effect has also been discussed as amechanism that can potentially reduce the solubility of organiccompounds in salt solutions [45]. This effect has also been found tobe stronger with Ca2+ than with Na+[46]. This phenomenon,which is particularly prominent in seawater, could possibly makethe EDCs more accessible to the surface of the CNTs, and thusinfluence their overall adsorption. However, the data in this studysuggests that the “salting-out” effect is not significantly contrib-uting to the removal of the EDCs.

3.3. Effect of DOC

Fig. 5 shows the effects of increased concentrations of DOC on theadsorption of BPA and EE2 by SWCNTs. Adsorption of both BPA andEE2 was found to decrease linearly with an increase in theconcentration of DOC in each of the source waters. Similar reductionsin adsorption by SWCNTs and other carbon surfaces in the presence ofhumic acid were also found in previous studies [40,47,48]. Two

0

0.2

0.4

0.6

0.8

1

0 5 10

C/C

0

0

0.2

0.4

0.6

0.8

1

C/C

0

0

0.2

0.4

0.6

0.8

1

C/C

0

DOC conc. (mg/L)

0 5 10DOC conc. (mg/L)

0 5 10DOC conc. (mg/L)

a

b

c

Fig. 5. Effect of DOC on the adsorption of BPA and EE2 by SWCNTs: (a) seawater; (b) SW/BWwater; and (c) brackish water (C0=10 μM; contact time=4 h; CNT dose=30mg/L). BPA(▲); EE2 (□).

0

0.2

0.4

0.6

0.8

1

0 20 40 60

C/C

0C

/C0

CNT conc. (mg/L)

0 20 40 60

CNT conc. (mg/L)

a

0

0.2

0.4

0.6

0.8

1b

Fig. 6. Removal of NOM from artificial source waters by SWCNTs with initial EDCconcentrations of (a) 1 μM with DOC=2.42 mg/L (2 mg/L from humic acid and0.42 mg/L from BPA and EE2) and (b) 10 μM with DOC=6.2 mg/L (2 mg/L from humicacid and 4.2 mg/L from BPA and EE2). Seawater (•); SW/BW (□); brackish water (Δ).

73L. Joseph et al. / Desalination 281 (2011) 68–74

possible explanations have been proposed for this reduction inadsorption by SWCNTs in the presence of humic acid: (i) competitionbetween the EDCs and the humic acid for adsorption sites of theSWCNTs and (ii) pore blockage, which reduces the amount of surfacearea available for adsorption [44,49]. Based on the availableadsorption sites (i.e., interstitial channels of the SWCNTs) and thesize of the humic acid used in this study, it can be said that bothcompetitive adsorption and pore blockage can potentially inhibit theadsorption of BPA and EE2. Due to the low concentration of DOC in thesource waters, the overall reduction in the adsorption of BPA and EE2is small. However, the prohibitive effect of DOC on adsorption maybecomemore prominent when attempting to removemicropollutantsfrom waterways that contain higher quantities of natural organicmatter (e.g., landfill leachates and municipal wastewater).

Fig. 6 shows the removal of NOM, represented by theconcentration of hydrophobic DOC, at varying doses of SWCNTsfrom each solution with initial EDC concentrations of 1 and 10 μM,respectively. Significant removal was achieved with increasingamounts of SWCNTs, with the maximum removal of DOC reaching95% with the addition of 50 mg/L SWCNTs. The removal of DOCwas prominent, and it followed similar trends as those for theartificial source waters, with the minimum removal occurring inbrackish water and the maximum removal occurring in the SW/BW water. Several studies have demonstrated the ability of CNTsand other adsorbents (e.g., PAC and GAC) to remove humic acid

and other forms of natural organic matter [26,50,51]. Similarly,with the adsorption of EDCs, the interactions between humic acidand SWCNTs have often been explained by mechanisms such ashydrophobic interactions and π–π EDA interactions [26,48,52].With the hydrophobic nature of humic acid, competition can occurbetween the EDCs and the DOC for the adsorption sites of theSWCNTs, which can reduce the overall effectiveness of theSWCNTs to remove the DOC. High concentrations of humic acidcan cause pore blockage of the SWCNTs, which could significantlyreduce the adsorptive capacity of the SWCNTs.

4. Conclusions

SWCNTs showed high removal efficiencies of BPA and EE2 fromseawater and brackish water. Throughout the study, the adsorptivecapacity of the SWCNTs was higher for EE2 than for BPA, possiblybecause of its higher log KOW value, which may lead to an increase inits hydrophobic interactions with the SWCNTs. Changes in the waterchemistry conditions of the source waters did not significantly impactthe overall adsorption of the EDCs. While changing the pH of thesource water from 3.5 to 11 showed a reduction in the adsorption ofBPA, it did not affect the removal of EE2. Adjusting the ionic strengthusing different concentrations of Na+ and Ca2+ also had a negligibleeffect on the adsorption of BPA and EE2.

DOC was shown to be influential in the adsorption of BPA and EE2by SWCNTs. Decreased adsorption occurred as the concentration ofDOC increased. However, as the concentration of NOM increases inthese water sources, overall adsorption of EDCs is expected to beincreasingly affected, possibly due to direct competition for adsorp-tion sites on the SWCNTs or pore blockage by the NOM. Significantremoval of DOC also occurred with increasing doses of SWCNTs,which may be due to similar mechanisms that govern the adsorptionof EDCs (e.g., hydrophobic and π–π EDA interactions).

The use of SWCNTs as an adsorbent can be an effective approach forpretreatment of seawater and brackish water, prior to desalination. The

74 L. Joseph et al. / Desalination 281 (2011) 68–74

results of this study have shown that SWCNTs can remove organicchemicals such as EDCs and DOC associated with the presence of naturalorganic matter, both of which are major foulants in membrane filtration.SWCNTs can be useful in reducing the fouling potential of source watersand allowing the overall seawater desalination process to be moreefficient and effective. Future studies on seawater and brackish waterpretreatment using SWCNTs should investigate the ability of SWCNTs toremove low concentrations of other emerging contaminants and EDCs,along with combining SWCNTs with other pretreatment methods (e.g.,nanofiltration and ultrafiltration) to maximize the overall desalinationprocess.

Acknowledgments

This research was funded by GS E&C Research Institute and theUnited StatesDepartmentof Agriculture (USDAAward58-3148-0-167).

References

[1] S.A. Snyder, P. Westerhoff, Y. Yoon, D.L. Sedlak, Pharmaceuticals, personal careproducts, and endocrine disruptors in water: implications for the water industry,Environ. Eng. Sci. 20 (2003) 449–469.

[2] D.W. Kolpin, E.T. Furlong, M.T. Meyer, E.M. Thurman, S.D. Zaugg, L.B. Barber, H.T.Buxton, Pharmaceuticals, hormones and other organic wastewater contaminantsin U.S. streams, 1999–2000: a national reconnaissance, Environ. Sci. Technol. 36(2002) 1202–1211.

[3] G.R. Boyd, J.M. Palmeri, S.Y. Zhang, D.A. Grimm, Pharmaceuticals and personal careproducts (PPCPs) and endocrine disrupting chemicals (EDCs) in stormwatercanals and Bayou St. John in New Orleans, Louisiana, USA, Sci. Total. Environ. 333(2004) 137–148.

[4] J.Q. Jiang, Q. Yin, J.L. Zhou, P. Pearce, Occurrence and treatment trials of endocrinedisrupting chemicals (EDCs) in wastewaters, Chemosphere 61 (2005) 544–550.

[5] H. Asakura, T. Matsuto, N. Tanaka, Behavior of endocrine-disrupting chemicals inleachate from MSW landfill sites in Japan, Waste Manage. 24 (2004) 613–622.

[6] T. Yamamoto, A. Yasuhara, H. Shiraishi, O. Nakasugi, Bisphenol A in hazardouswaste landfill leachates, Chemosphere 42 (2001) 415–418.

[7] B. Pan, H. Mashayekhi, B. Xing, Adsorption and hysteresis of bisphenol A and 17α-ethinyl estradiol on carbon nanomaterials, Environ. Sci. Technol. 42 (2008)5480–5485.

[8] T. Colborn, F.S. Vom Saal, A.M. Soto, Developmental effects of endocrine-disrupting chemicals in wildlife and humans, Environ. Health Perspect. 101(1993) 378–384.

[9] P.T.C. Harrison, P. Holmes, C.D.N. Humfrey, Reproductive health in humans andwildlife: are adverse trends associated with environmental chemical exposure,Sci. Total. Environ. 205 (1997) 97–106.

[10] J. DeRudder, T. Van de Wiele, W. Dhooge, F. Comhaire, W. Verstraete, Advancedwater treatment with manganese oxide for the removal of 17α-ethinyl estradiol(EE2), Water Res. 38 (2004) 184–192.

[11] G. Ying, R.S. Kookana, Degradation of five selected endocrine-disruptingchemicals in seawater and marine sediment, Environ. Sci. Technol. 37 (2003)1256–1260.

[12] Y. Liu, Y.T. Guan, N.F.Y. Tam, T. Mizuno, H. Tsuno, W.P. Zhu, Influence of rainfalland basic water quality parameters on the distribution of endocrine-disruptingchemicals in coastal area, Water Air Soil Pollut. 209 (2010) 333–343.

[13] A. Brehant, V. Bonnelye, M. Perez, Comparison of MF/UF pretreatment withconventional filtration prior to RO membranes for surface seawater desalination,Desalination 144 (2002) 353–360.

[14] F. Reverberi, A. Gorenflo, Three year operational experience of a spiral-woundSWRO system with a high fouling potential feed water, Desalination 203 (2007)100–106.

[15] T. Tran, B. Bolto, S. Gray, M. Hoang, E. Ostarcevic, An autopsy study of a fouledreverse osmosis membrane element used in a brackish water treatment plant,Water Res. 41 (2007) 3915–3923.

[16] K.Y.J. Choi, B.A. Dempsey, In-line coagulation with low-pressure membranefiltration, Water Res. 38 (2004) 4271–4281.

[17] A.M. Eiguera, S.O.N. Baez, Development of the most adequate pre-treatment forhigh capacity seawater desalination plants with open intake, Desalination 184(2005) 173–183.

[18] H. Huang, K. Schwab, J.G. Jacangelo, Pretreatment for low pressure membranes inwater treatment: a review, Environ. Sci. Technol. 43 (2009) 3011–3019.

[19] A. Rahardianto, J.B. Gao, C.J. Gabelich, M.D. Williams, Y. Cohen, High recoverymembrane desalting of low-salinity brackish water: integration of acceleratedprecipitation softening with membrane RO, J. Memb. Sci. 289 (2007) 123–137.

[20] W. Tsujimoto, H. Kimura, T. Izu, T. Irie, Membrane filtration and pre-treatment byGAC, Desalination 119 (1998) 323–326.

[21] A. Yuasa, Drinkingwater production by coagulation–microfiltration and adsorption–ultrafiltration, Water Sci. Technol. 37 (1998) 135–146.

[22] D.H. Lin, B.S. Xing, Adsorption of phenolic compounds by carbon nanotubes: roleof aromaticity and substitution of hydroxyl groups, Environ. Sci. Technol. 42(2008) 7254–7259.

[23] K. Yang, W.H. Wu, Q.F. Jing, L.Z. Zhu, Aqueous adsorption of aniline, phenol, andtheir substitutes by multi-walled carbon nanotubes, Environ. Sci. Technol. 42(2008) 7931–7936.

[24] Y. Li, Z. Di, J. Ding, D. Wu, Z. Luan, Y. Zhu, Adsorption thermodynamic, kinetic anddesorption studies of Pb2+ on carbon nanotubes, Water Res. 39 (2005) 605–609.

[25] X.K. Wang, G. Chen, W. Hu, A. Ding, D. Xu, X. Zhou, Sorption of243Am(III) to

multiwall carbon nanotubes, Environ. Sci. Technol. 39 (2005) 2856–2860.[26] H. Hyung, J. Kim, Natural organic matter (NOM) adsorption to multi-walled

carbon nanotubes: effect of NOM characteristics and water quality parameters,Environ. Sci. Technol. 42 (2008) 4416–4421.

[27] M.S. Mauter, M. Elimelech, Environmental applications of carbon-based nanoma-terials, Environ. Sci. Technol. 42 (2008) 5843–5859.

[28] X.J. Peng, Y.H. Li, Z.K. Luan, Z.C. Di, H.Y. Wang, B.H. Tian, Z.P. Jia, Adsorption of 1,2-dichlorobenzene from water to carbon nanotubes, Chem. Phys. Lett. 376 (2003)154–158.

[29] B. Pan, B. Xing, Adsorption kinetics of 17α-ethinyl estradiol and bisphenol A oncarbon nanomaterials. I. Several concerns regarding pseudo-first order andpseudo-second order models, J. Soils Sediments 10 (2010) 838–844.

[30] Cheap Tubes, Inc, Single walled nanotubes—SWNTs, The Source for SingleWalled Nanotubes-SWNTs [Fact Sheet], 2009, Retrieved August 23, 2010, from,http://www.cheaptubesinc.com/swnts.htm#single_walled_nanotubes_swnts_90wt%_specifications.

[31] R. Al-Rasheed, D.J. Cardin, Photocatalytic degradation of humic acid in salinewaters. Part 1. Artificial seawater: influence of TiO2, temperature, pH, and air-flow, Chemosphere 51 (2003) 925–933.

[32] D.R. Kester, I.W. Duedall, D.N. Connors, R.M. Pytkowic, Preparation of artificialseawater, Limnol. Oceanogr. 12 (1967) 176–181.

[33] K. Gaid, Y. Treal, Reverse osmosis desalination: the experience of Veolia Water,Desalination 203 (2007) 1–14.

[34] L.F. Greenlee, D.F. Lawler, B.D. Freeman, B. Marrot, P. Moulin, Reverse osmosisdesalination: water sources, technology, and today's challenges, Water Res. 43(2009) 2317–2348.

[35] Y.M. Yoon, P. Westerhoff, S.A. Snyder, M. Esparza, HPLC-fluorescence detectionand adsorption of bisphenol A, 17β-estradiol, and 17α-ethynyl estradiol onpowdered activated carbon, Water Res. 37 (2003) 3530–3537.

[36] S. Agnihotri, M.J. Rood, M. Rostam-Abadi, Adsorption equilibrium of organicvapors on single-walled carbon nanotubes, Carbon 43 (2005) 2379–2388.

[37] G.H. Liu, J.L. Wang, Y.F. Zhu, X.R. Zhang, Application of multiwalled carbonnanotubes as a solid-phase extraction sorbent for chlorobenzenes, Anal. Lett. 37(2004) 3085–3104.

[38] C. Lu, Y.L. Chung, K.F. Chang, Adsorption of trihalomethanes from water withcarbon nanotubes, Water Res. 39 (2005) 1183–1189.

[39] F.S. Su, C.S. Lu, Adsorption kinetics, thermodynamics and desorption of naturaldissolved organic matter by multiwalled carbon nanotubes, J. Environ. Sci. Heal. A42 (2007) 1543–1552.

[40] J. Chen, W. Chen, D.Q. Zhu, Adsorption of nonionic aromatic compounds to single-walled carbon nanotubes: effects of aqueous solution chemistry, Environ. Sci.Technol. 42 (2008) 7225–7230.

[41] S. Gotovac, C.M. Yang, Y. Hattori, K. Takahashi, H. Kanoh, K. Kaneko, Adsorption ofpolyaromatic hydrocarbons on single wall carbon nanotubes of differentfunctionalities and diameters, J. Colloid Interf. Sci. 314 (2007) 18–24.

[42] B. Pan, B. Xing, Adsorption mechanisms of organic chemicals on carbonnanotubes, Environ. Sci. Technol. 42 (2008) 9005–9013.

[43] M.A. Fontecha-Camara, M.V. Lopez-Ramon, M.A. Alvarez-Merino, C. Moreno-Castilla, Effect of surface chemistry, solution pH, and ionic strength on the removalof herbicides diuron and amitrole from water by an activated carbon fiber,Langmuir 23 (2007) 1242–1247.

[44] S. Zhang, T. Shao, S.S.K. Bekaroglu, T. Karanfil, Adsorption of synthetic organicchemicals by carbon nanotubes: effects of background solution chemistry, WaterRes. 44 (2010) 2067–2074.

[45] W.H. Xie, W.Y. Shiu, D. Mackay, A review of the effect of salts on the solubility oforganic compounds in seawater, Mar. Environ. Res. 44 (1997) 429–444.

[46] M.A. Schlautman, S. Yim, E.R. Carraway, J.H. Lee, B.E. Herbert, Testing a surfacetension-based model to predict the salting out of polycyclic aromatic hydrocar-bons in model environmental solutions, Water Res. 38 (2004) 3331–3339.

[47] J.E. Kilduff, A. Wigton, Sorption of TCE by humic-preloaded activated carbon:incorporating size-exclusion and pore-blockage phenomena in a competitiveadsorption model, Environ. Sci. Technol. 33 (1999) 250–256.

[48] J.J. Pignatello, S. Kwon, Y.F. Lu, Effect of natural organic substances on the surfaceand adsorptive properties of environmental black carbon (char): attenuation ofsurface activity by humic and fulvic acids, Environ. Sci. Technol. 40 (2006)7757–7763.

[49] G. Newcombe, M. Drikas, R. Hayes, Influence of characterised natural organicmaterial on activated carbon adsorption. 2. Effect on pore volume distribution andadsorption of 2-methylisoborneol, Water Res. 31 (1997) 1065–1073.

[50] X. Wang, J. Lu, B. Xing, Sorption of organic contaminants by carbon nanotubes:influence of absorbed organic matter, Environ. Sci. Technol. 42 (2008) 3207–3212.

[51] K. Yang, B. Xing, Adsorption of fulvic acid by carbon nanotubes from water,Environ. Pollut. 157 (2009) 1095–1100.

[52] L. Joseph, Q. Zaib, I.A. Khan, N.D. Berge, Y.G. Park, N.B. Saleh, Y. Yoon, Removal ofbisphenol A and 17a-ethinyl estradiol from landfill leachate using single-walledcarbon nanotubes, Water Res. 45 (2011) 4056–4068.