w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4

Avai lab le at www.sc iencedi rect .com

journa l homepage : www.e lsev i er . com/ loca te /wat res

Partitioning of hydrophobic pesticides within asoil–water–anionic surfactant system

Peng Wang, Arturo A. Keller*

Bren School of Environmental Science and Management, University of California, Santa Barbara, CA 93106, USA

a r t i c l e i n f o

Article history:

Received 6 August 2008

Received in revised form

23 October 2008

Accepted 30 October 2008

Published online 12 November 2008

Keywords:

Hydrophobic organic compounds

Anionic surfactant

Surfactant-aided soil washing

Soil contamination

* Corresponding author. Tel.: þ1 805 893 754E-mail address: keller@bren.ucsb.edu (A.A

0043-1354/$ – see front matter ª 2008 Elsevidoi:10.1016/j.watres.2008.10.052

a b s t r a c t

Surfactants can be added to pesticide-contaminated soils to enhance the treatment effi-

ciency of soil washing. Our results showed that pesticide (atrazine and diuron) partitioning

and desorbability within a soil–water–anionic surfactant system is soil particle-size

dependent and is significantly influenced by the presence of anionic surfactant. Anionic

surfactant (linear alkylbenzene sulphonate, LAS) sorption was influenced by its complex-

ation with both the soluble and exchangeable divalent cations in soils (e.g. Ca2þ, Mg2þ). In

this study, we propose a new concept: soil system hardness which defines the total amount

of soluble and exchangeable divalent cations associated with a soil. Our results showed

that anionic surfactant works better with soils having lower soil system hardness. It was

also found that the hydrophobic organic compounds (HOCs) sorbed onto the LAS-divalent

cation precipitate, resulting in a significant decrease in the aqueous concentration of HOC.

Our results showed that the effect of exchangeable cations and sorption of HOC onto the

surfactant precipitates needs to be considered to accurately predict HOC behavior within

soil–water–anionic surfactant systems.

ª 2008 Elsevier Ltd. All rights reserved.

1. Introduction et al., 2003; Zhu et al., 2003a; Rodriguez-Cruz et al., 2006; Yang

Pesticide spills and accidents involving pesticide handling

take place each year on farms and pesticide formulating and

manufacturing plants (Mata-Sandoval et al., 2002). Many of

these pesticides are highly hydrophobic. The ability of

surfactants to enhance the water solubility of hydrophobic

organic compounds (HOCs) provides a potential means of

improving the treatment efficiency of ex situ soil washing

systems for remediating pesticide contaminated soils (San-

chez-Camazano et al., 2003; Chu, 2003; Yang et al., 2006; Paria,

2007).

Cationic surfactants, due to their significant sorption onto

soils, are not desirable for surfactant-aided soil washing

systems, while, the use of nonionic and anionic surfactants in

surfactant-aided soil washing systems has been investigated

to some extent (Mata-Sandoval et al., 2002; Sanchez-Camazano

8; fax: þ1 805 893 7612.. Keller).

er Ltd. All rights reserved

et al., 2006; Wang and Keller, 2008a,b,c). However, compared

with the number of studies reported on the use of nonionic

surfactants, the number of the experiments using anionic

surfactants in surfactant-aided soil washing systems is

surprisingly low (Rodriguez-Cruz et al., 2006) and thus there

are still some important questions to be answered. This study

presents the results of equilibrium partitioning of an anionic

surfactant and sorption and desorption of pesticide within

soil–water–anionic surfactant systems.

Anionic surfactant sorption mechanisms onto soil and

sediment have been studied to some extent. Some

researchers reported that, under low aqueous anionic

surfactant concentrations (<10�4 M), anionic surfactant

sorption onto soil and sediment was largely controlled by its

partitioning into soil organic matter (SOM) by hydrophobic

interactions (Di Toro et al., 1990; Westall et al., 1999; Garcia

.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4 707

et al., 2002; Higgins and Luthy, 2006, 2007). Electrostatic

interactions were also found to play a role for anionic

surfactant sorption, with the amount of surfactant sorbed

increasing with increasing aqueous Ca2þ concentration

(Higgins and Luthy, 2006, 2007; Westall et al., 1999; Garcia

et al., 2002). Others reported that, under high concentrations,

anionic surfactants formed precipitates with hardness

cations (Ca2þ, Mg2þ) in water and in soil solution (Stellner and

Scamehorn, 1989a,b; Jafvert and Heath, 1991; Jafvert, 1991;

Verge et al., 2001). Jafvert (1991) determined the distribution

of various polycyclic aromatic hydrocarbons (PAHs) between

several sediments or soils and aqueous phase containing

sodium dodecylsulfate (SDS) micelles. Unfortunately, only

the soluble divalent cations of the bulk sediment or soils

were explicitly considered in the study and the effect of

exchangeable divalent cations originally held on the cation

exchangeable sites on PAHs distribution was not clear (Jaf-

vert, 1991). Exchangeable divalent cations usually account for

a significant fraction of soil total cation content. For example,

a typical agricultural loam soil contains about 2–3 kg/m2 of

exchangeable cations in its root zone (0.5 m depth) (Bohn

et al., 2001). Since a high anionic surfactant concentration

(usually> 10 g/l) is usually applied for surfactant-aided soil

washing applications, the exchangeable divalent cations

might be exchanged out from these original sites and transfer

into the aqueous solution, where they may play a significant

role in surfactant precipitation. Also, although precipitation

can account for a significant anionic surfactant loss within

Table 1 – Selected properties of the bulk soils and their size fracapacities, maximal amount of pesticide sorbed in the presenc

Soils OC (%) SA (m2/g) CEC(cmol/kg)

ExchangeabCa2þþMg2

(cmol/kg)

Bulk

Ag#1 1.51 6.2 6.2 5.8

Ag#2 1.5 15.2 15.2 14.1

Ag#3 1.52 15.4 15.4 15.1

Clayey 1.37 15.7 15.7 15.2

Sediment 1.12 5.4 5.4 5.2

Clay

Ag#1 4.95 40.2 40.2 39.0

Ag#2 4.36 59 59.0 57.5

Ag#3 4.5 54.4 54.4 53.6

Clayey 1.8 50.4 50.4 48.5

Sediment 6.02 42.2 42.2 40.7

Silt

Ag#1 1.82 13 13.2 12.7

Ag#2 1.29 16 16.3 14.8

Ag#3 1.13 19 19.0 18.7

Clayey 0.66 18 18.0 17.5

Sediment 1.28 13 13.2 12.6

Sand

Ag#1 0.50 1.3 3.0 2.8

Ag#2 0.15 1.4 6.2 5.2

Ag#3 0.11 2.1 3.0 3.0

Clayey 0.50 1.0 8.2 7.4

Sediment 0.27 0.9 3.0 3.0

Note: OC: organic carbon; SA: surface area ; CEC: cation exchange capaci

soil washing systems (Stellner and Scamehorn, 1989a,b; Jaf-

vert and Heath, 1991; Jafvert, 1991; Verge et al., 2001), the

effect of the anionic surfactant precipitates on HOC parti-

tioning within these systems has not been explicitly inves-

tigated so far.

Our hypotheses are (1) both soluble and exchangeable

divalent cations in soils contribute to surfactant precipitation

within an anionic surfactant-aided soil washing system; (2)

the surfactant-divalent cation precipitate accounts for

a certain portion of HOC partitioning. To test these hypoth-

eses, the bulk soils were separated into primary soil size

fractions (clay, silt, and sand size fractions), which had no

soluble divalent cations associated with them and the effect of

exchangeable divalent cations on the LAS partitioning could

then be clearly investigated. To single out the effect of the

surfactant-divalent cation precipitate on HOC partitioning,

HOC, anionic surfactant, and divalent cations were allowed to

interact in a system without soils.

More specifically, in this study, linear alkylbenzene

sulphonate (LAS), the common ingredient of most commercial

detergents, was selected to study anionic surfactant parti-

tioning and its effect on the sorption and desorption of two

commonly used hydrophobic pesticides (atrazine and diuron)

from the bulk soils and their primary size fractions. The

objectives of this study were thus set to test the above-

mentioned hypotheses and to examine the soil particle size-

dependent sorption/desorption behavior of LAS and its effect

on HOC sorption and desorption.

ctions, linear alkylbenzene sulfonate (LAS) sorptione of LAS under the experimental conditions.

leþ

SolubleCa2þ þMg2þ

(cmol/kg)

LAS pseudosorption

capacities(mg/kg)

Maximal pesticidesorbed (mg/kg)

Diuron Atrazine

2.3 21,000 62.8 43.5

15.3 105,700 74.8 54.0

26.5 154,000 77.3 60.0

4.5 24,000 64.3 46.0

5.2 20,000 60.8 40.0

73,000 311.7 233.3

130,000 418.3 270.0

130,000 431.7 293.3

120,000 391.7 256.7

85,000 331.7 233.3

28,000 156.4 91.4

22,000 165.0 80.0

40,000 190.7 112.4

24,000 140.7 87.1

30,000 145.0 95.7

6100 49.5 29.0

9200 65.5 39.0

6900 70.5 34.0

9900 66.5 39.0

7000 57.5 27.0

ty.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4708

2. Materials and methods

2.1. Chemicals

Atrazine (2-chloro-4-ethylamino-6-isopropylamino-1,3,5-tri-

azine) was purchased from Supelco Inc. (Bellefonte, PA, USA)

with a reported purity>97%, and diuron (3-(3,4-dichlorofenyl)-

1,1-dimethylurea) was purchased from ChemService Inc.

(West Chestnut, PA, USA) with a reported purity >99%. These

pesticides were selected due to their high volume use world-

wide, and as representative of hydrophobic organic

compounds. Linear alkylbenzene sulphonate was purchased

from Sigma-Aldrich (St. Louis, MO, USA). LAS is commonly

used in pesticide formulations and has been proposed for

surfactant-aided soil washing. The chemicals were used as

received. Selected physicochemical properties of these

compounds can be found in Supporting Information (Tables S1

and S2).

2.2. Soils and soil size fractions

Four agricultural soils and one sediment (denoted as Ag#1,

Ag#2, Ag#3, Clayey, and Sediment) were collected from Santa

Barbara, California, USA. The water dispersible clay (< 2 mm),

silt (2–50 mm), and sand (>50 mm) fractions were separated

using a low energy method, which involved using only water

as dispersant, gentle mixing, repeated wet sedimentation,

dialysis desalination, and freeze-drying. The details of the size

separation can be found in Wang and Keller (2008a). The soil

organic carbon (OC), cation exchange capacity (CEC), BET

surface area (SA), and pH of the soils and the size fractions

were measured using the methods described by Carter (1993).

2.3. Pesticide solubility enhancement by LAS

The solubility enhancement experiments were conducted in

duplicate. A set of LAS solutions with varying concentrations

(0.25–8.0 g/l or 0.69–22.1 mM) was prepared and placed in

15 ml glass centrifuge tubes. This concentration range was

selected to evaluate the system behavior at aqueous concen-

trations lower and higher than the CMC of LAS, which is

0.45 g/l or 1.24 mM. Pesticide (atrazine or diuron) was subse-

quently added as a solid to each tube in an amount slightly

greater than required to saturate the solution. The tubes were

shaken end-over-end for 24 h at 22� 2 �C and then centri-

fuged at 5000� g for 1 h to separate the undissolved pesticide

solid. From each tube, 1.0 ml of the supernatant was then

taken and analyzed via high performance liquid chromatog-

raphy (HPLC) for pesticide concentrations.

2.4. Equilibrium surfactant and pesticide partitioning

The surfactant sorption was studied in duplicate by the batch

equilibration method. Initial LAS concentrations spanned

a large range (0–20.0 g/l) below and above the critical micelle

concentration (CMC¼ 0.45 g/l) of LAS. De-ionized (DI) water

was used for preparing all solutions. The pesticide concen-

trations used were 15.00 mg/l for atrazine and 15.95 mg/l for

diuron in all cases. A 0.01 M KCl background electrolyte was

used to minimize ionic strength change and 0.02% NaN3 was

used as microbial growth inhibitor in all cases. Aliquots of

2.00 g of a bulk soil, 0.30 g clay fraction, or 0.70 g silt fraction,

or 1.00 g sand fraction were treated with 10.0 ml of solution

containing the pesticide and LAS at varying concentrations in

15 ml glass centrifuge tubes. The amount of the size fractions

used was based on their weight percent in the bulk soils.

The tubes were shaken at 60 rpm for 24 h in an end-over-

end shaker at 22� 2 �C to reach LAS and pesticide sorption

equilibrium, and then centrifuged at 5000� g for 30 min at the

same temperature. A total of 1.5 ml of the supernatants was

then transferred into 1.5 ml microcentrifuge tubes which were

then centrifuged at 15,000� g for 20 min. Preliminary results

showed that 24 h were adequate for the sorption equilibrium

of the pesticide and surfactant to be reached (c.f. Supporting

information, Figure S-1). The sorption of pesticide and

surfactant on the centrifuge tubes was determined to be

negligible and the amount of the pesticide and surfactant

blank (with no soils) did not show any significant change

before and after mixing. 1.0 ml of the centrifuged supernatant

was then analyzed for LAS and pesticide concentrations via

HPLC. The amount of pesticide and surfactant sorbed was

calculated as the difference between the initial and final mass

in the aqueous phase.

2.5. Desorption of pesticide in the presence of LAS

Pesticide desorption experiments were conducted using the

same experimental procedure as for the sorption experiments

except that the desorption experiments consisted of one

sorption step followed by 5 consecutive desorption steps.

During the sorption step, aliquots of 2.00 g of a bulk soil, 0.30 g

clay fraction, or 0.70 g silt fraction were treated with 10.0 ml of

a solution containing only pesticide in 15 ml glass centrifuge

tubes. It should be noted that the sand fractions were not used

for the desorption experiments due to their low pesticide and

LAS sorption capacities. Initial pesticide concentrations used

were 15.00 mg/l for atrazine and 15.95 mg/l for diuron during

the sorption step in all cases. During each desorption step,

5.0 ml of the supernatant was replaced with 5.0 ml of surfac-

tant solution. The concentration of these 5.0 ml of surfactant

solution was so that the 10 ml of supernatant would be at the

desired surfactant concentration for each case. Two LAS

concentrations were used: 2.00 and 5.00 g/l. The blank exper-

iments were conducted with DI water only to which the cor-

responding background electrolytes were added. No

significant change in pH was observed before and after the

sorption/desorption experiments. In this paper, the amount of

the pesticide sorbed at end of the sorption cycle is denomi-

nated ‘presorbed pesticide’.

2.6. Pesticide sorption onto LAS precipitate

A set of batch experiments were conducted in which a pesti-

cide solution containing 0.01 M CaCl2 and atrazine at 14.0 mg/l

or diuron at 13.2 mg/l was treated at different LAS concen-

trations from 0.10 to 8.0 g/l for 24 h to reach the reaction

equilibrium using the same procedure as the sorption exper-

iments. The precipitate solid was separated out of the

0

2

4

6

8

0 2 4 6 8

Total LAS conc. (g/L)

S w* /S

w

AtrazineDiuron

S*w/Sw=1

CMC of LAS

Fig. 1 – Pesticide solubility enhancement as a function of

linear alkylbenzene sulfonate (LAS) concentration.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4 709

aqueous phase by centrifugation and the pesticide concen-

tration in the aqueous phase was measured.

2.7. HPLC analysis

A Shimadzu HPLC system (Shimadzu, Nakagyo-ku, Kyoto,

Japan) equipped with two LC-10AT VP pumps, a Sil-10AF

autosampler, a DGU-14A degasser, and a SPD-M10AVP diode

array detector was used. A Premier� C18 5 mm reverse phase

column (Shimadzu, Nakagyo-ku, Kyoto, Japan) was used with

a length of 250 mm and an inner diameter of 4.6 mm. The

analyses were performed at a constant flow rate of 1.0 ml/min.

The UV detector monitored the absorbance at 222 nm for

atrazine, 247 nm for diuron, and 220 nm for LAS. Some

samples were diluted as needed. The calibration was con-

ducted daily and R2 was greater than 0.98 in all cases.

3. Results and discussion

3.1. Characterization of the bulk soils and size fractions

The measured properties of the bulk soils and their size

fractions were presented in Table 1. Generally speaking, the

soil properties showed highly particle-size dependent

behavior. As the particle size decreases, the OC, SA, and CEC

increase. The clay fractions have consistently higher OC, SA,

and CEC than the bulk soils and the silt and sand fractions.

Divalent cations (Ca2þ and Mg2þ) were found to dominate the

CEC sites in all case (>92%). There was no significant presence

of trivalent cations within the soils. Also, due to the separation

process employed, the soluble cations were available only

Table 2 – Pesticide solubility enhancement by linear alkylbenz

Relationship R2

Atrazine S�wSw¼ 1þ 433� Xmc 0.99

Diuron S�wSw¼ 1þ 776� Xmc 0.98

a Lower 95% confidence interval–upper 95% confidence interval.

within the bulk soils and there were no soluble divalent

cations present with the different size fractions. A previous

study conducted by Jafvert and Heath (1991) showed that Ca2þ

and Mg2þ behave similarly within soil–water–anionic surfac-

tant systems. Thus, in this study, no effort was made to

differentiate between Ca2þ and Mg2þ, which are denominated

the divalent cations in the rest of the manuscript.

3.2. Pesticide solubility enhancement by LAS

The apparent water solubility enhancement of an HOC in

a surfactant solution can be expressed as (Zhu et al., 2003b;

Kile and Chiou, 1989; Wang and Keller, 2008a,b):

S�wSw¼ 1þ XmnKmn þ XmcKmc (1)

Where Xmn and Xmc are the surfactant monomer and micellar

concentrations in water respectively (g/l); and Kmn and Kmn are

the HOC partitioning coefficients with the surfactant mono-

mer and micellar phases respectively (l/g); S�w is the apparent

HOC solubility at a total surfactant concentration of

X¼XmnþXmc (mg/l); Sw is the intrinsic HOC solubility in

water in the absence of surfactant (mg/l). S�w and Sw were

measured directly on the HPLC. The results of the HOC solu-

bility enhancement experiments are presented in Fig. 1, which

showed that at concentrations below the CMC of LAS, there

was no interaction between LAS monomers and either of the

pesticides and thus no solubility enhancement was observed,

while, once the LAS concentrations exceeded the CMC,

a solubility enhancement was observed for either pesticide

but to a different extent. Thus, the measured HOC solubility

enhancement data was fitted with Equation (1) by conducting

liner regression and the fitted results are presented in Table 2.

Since the interaction between LAS monomers and both HOCs

has been determined insignificant, Kmn is zero in both cases

and is not presented.

The Kmc of diuron (776 l/kg) was much greater than that of

atrazine (443 l/kg) for LAS (Table 2), which is attributable to

diuron’s higher hydrophobicity, indicated by its higher octa-

nol-water partitioning coefficient, Kow, (Table S1), and thus its

higher affinity to the hydrophobic cores of the LAS micelles.

3.3. LAS sorption onto soils

LAS sorption onto the bulk soils and the size fractions showed

similar behavior, increasing before maximal sorption was

reached and then followed by a gradual decrease. To

demonstrate, the correlation between the equilibrium

aqueous LAS concentrations and the amount of LAS sorbed is

presented in Fig. 2 for the clay and silt fractions. The average

ene sulfonate (LAS).

Solubility(mg/l)

CMC(g/l)

Kmc

(l/kg)

33 0.45 433 (418–448)a

42 0.42 776 (753–798)a

a clay fractions

0

30000

60000

90000

120000

150000

0.0 0.5 1.0 1.5 2.0 2.5 3.0

LAS aqueous concentration (g/l)

LA

S so

rbed

con

c. (

mg/

kg)

0.0 0.5 1.0 1.5 2.0 2.5 3.0

b Silt fractions

0

10000

20000

30000

40000

50000

LAS aqueous concentration (g/l)

LA

S so

rbed

con

c. (

mg/

kg)

Ag#1 Ag#2 Ag#3

Clayey Sediment

Ag#1 Ag#2 Ag#3

Clayey Sediment

Fig. 2 – Linear alkylbenzene sulfonate (LAS) sorption

isotherms on (a) clay and (b) silt size fractions.

AqueousHOC

SOM

LASmonomers

Soil minerals

LAS precipitate

LASmicelles

Soil phase

Aqueous phase

Precipitate phase

Fig. 3 – Schematic diagram of linear alkylbenzene sulfonate

(LAS) and hydrophobic organic compound (HOC)

partitioning within soil–water–surfactant systems.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4710

of the duplicate measurements was used in preparing these

graphs. The standard errors were all smaller than 15% of the

averages; error bars are presented only for Ag#1 and Ag#3 for

clarity. If partitioning into SOM dominates LAS sorption onto

soils, the clay fractions would be expected to have much

higher sorption capacities than the bulk soils, which is

apparently not the case. Two major mechanisms have been

proposed to explain the interactions between LAS and the

divalent cations within soil–water–anionic surfactant systems

under high surfactant concentration: (1) LAS, represented as

RSO3Na could form a positively charged complex with Ca2þ (or

another divalent cation):

Ca2þ þ RSO�3 ¼ RSO3Caþ

that would then be preferentially adsorbed on the negatively

charged soil particles (Westall et al., 1999; Higgins and Luthy,

2006, 2007); (2) precipitation of Ca(RSO3)2 (Savitsky et al., 1981;

Matheson et al., 1985; Stellner and Scamehorn, 1989a,b; Jafvert

and Heath, 1991; Jafvert, 1991; Westall et al., 1999; Verge et al.,

2001):

Ca2þ þ 2RSO�3 ¼ CaðRSO3Þ2Y

The loss of LAS via interaction (1) could be viewed as part of

LAS soil sorption because the LAS-Caþ would be associated

with the soil particles, while that via interaction (2) could be

considered a LAS non-sorptive loss because Ca(RSO3)2 would

precipitate out as separate solid. The precipitates of anionic

surfactants can be redissolved as the aqueous surfactant

concentrations increases (Savitsky et al., 1981; Matheson

et al., 1985; Stellner and Scamehorn, 1989a,b; Jafvert and

Heath, 1991; Jafvert, 1991; Westall et al., 1999; Verge et al.,

2001) due to increasing concentration of Naþ.

As a result, the isotherms presented in Fig. 2 are not ‘true’

LAS sorption isotherms onto the soils; they just present the

correlation between LAS aqueous concentrations and the

amount of LAS associated with solid phases (soil particles and

surfactant precipitates). Some authors used term ‘abstraction’

to describe the combined surfactant loss due to sorption and

precipitation (Hanna and Somasundaran, 1979). The true

sorption capacity of LAS onto the soils and the size fractions

alone cannot be determined in the presence of divalent

cations, but in this study LAS pseudo sorption capacities were

still determined as the maximal points in Fig. 2, treating the

precipitates as part of the LAS soil sorption and normalizing

them for the amount of soil solids. The measured LAS pseudo

sorption capacities are presented in Table 1 along with the

measured soil properties.

The overall partitioning of LAS and HOC within a soil–

water–anionic surfactant system is depicted in Fig. 3. It should

be noted that the sorption of LAS onto the soil mineral phase

in Fig. 3 is in the form of RSO3Caþ. Interactions (1) and (2) can

take place simultaneously with LAS sorption onto SOM. It is

generally accepted that surfactant micelles do not sorb onto

solid surfaces themselves. The sorption of surfactant usually

occurs via monomers. The sorbed monomers can then form

admicelles on the solid surfaces onto which they sorb (Zhu

et al., 2003b; Jafvert and Heath, 1991; Jafvert, 1991). It is for this

reason that micelles-SOM and micelle-mineral interactions

were excluded from Fig. 3.

For the different size fractions, although no dissolved

divalent cations were present in the aqueous phase initially,

the presence of 0.01 M KCl as background electrolyte and Naþ

ions from LAS leads to the exchange of a certain portion of the

exchangeable divalent cations (Bohn et al., 2001; Yang et al.,

2007), resulting in the presence of Ca2þ and Mg2þ in the

aqueous phase. As interactions (1) and (2) proceed, more

divalent cations are exchanged out from their original sites

and into the aqueous phase to maintain equilibrium. This is

supported by the high correlation coefficient (0.98) between

a Atrazine with silt fractions

0

3

6

9

12

15

0.0 0.5 1.0 1.5

LAS aqueous concentration (g/l)

Atr

azin

e aq

ueou

s co

nc. (

mg/

l)

1.5

b Diuron with silt fractions

0

3

6

9

12

15

0.0 0.5 1.0

LAS aqueous concentration (g/l)

Diu

ron

aque

ous

conc

. (m

g/.l)

Ag#1 Ag#2 Ag#3

Clayey Sediment

Ag#1 Ag#2 Ag#3

Clayey Sediment

Fig. 5 – Pesticide aqueous concentration as a function of

linear alkylbenzene sulfonate (LAS) with the silt fraction

for (a) atrazine; (b) diuron.

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4 711

the exchangeable divalent cations concentrations of the size

fractions and their pseudo LAS sorption capacities.

Similar to aquatic chemistry, we proposed a new concept:

soil system hardness which is defined as the total amount of

divalent cations i.e., Ca2þ and Mg2þ, including both soluble

and exchangeable divalent cations associated with a soil. The

correlation coefficient is 0.93 between the LAS pseudo sorp-

tion capacities and the soil system hardness of the bulk soils

and their size fractions, suggesting that the interaction

between LAS and the divalent cations largely explains LAS

partitioning behavior for all bulk soils and their various size

fractions. The results also showed that the LAS sorption is

highly soil-particle size dependent, which is a direct result of

the particle-size dependence of soil exchangeable divalent

cation concentrations (Table 1).

3.4. Pesticide sorption onto LAS precipitates

The formation of LAS precipitates in the presence of Ca2þ can

be seen in the Supporting Information (Figure S-2). Fig. 4

presents the correlation between the amount of LAS precipi-

tated and the amount of pesticide sorbed with LAS precipitate,

i.e., Ca(RSO3)2. As can be seen, the aqueous atrazine and

diuron concentration decreased significantly in the presence

of Ca(RSO3)2, indicating that the sorption of either pesticide

with Ca(RSO3)2 is significant, although the overall pesticide

sorption efficiency decreases with increasing amount of

Ca(RSO3)2 precipitated. Diuron showed higher sorption onto

Ca(RSO3)2 precipitate than atrazine.

The results presented in Fig. 4 assume a uniform precipi-

tate size. Since the size of the precipitate is expected to affect

the HOC sorption efficiency of LAS precipitate, one should be

cautious in using Fig. 4 directly for an anionic surfactant soil-

washing system, in which particle size is likely to be more

heterogeneous. Smaller precipitates are likely to have higher

HOC sorption efficiency due to their higher surface area

available.

The difference between atrazine and diuron suggests that

the hydrophobicity of the pesticides can be used to indicate

the magnitude of the sorption.

0

2

4

6

8

10

0 10 20 30 40

LAS precipitated (mg)

Pes

tcid

e so

rbed

(m

g/g)

.

Atrazine

Diuron

Fig. 4 – Correlation between the amount of linear

alkylbenzene sulfonate (LAS) precipitated and the amount

of the pesticide sorbed onto the LAS precipitate.

3.5. Pesticide partitioning withinsoil–water–LAS systems

The dashed line in Fig. 3 indicates that without sorbed

RSO3Caþ onto the soil mineral phase, the sorption of HOCs

onto the soil minerals is insignificant. Also, although not

explicitly depicted in Fig. 3, the cation exchange interaction is

critical to better predict LAS and thus HOC partitioning within

such systems and should be included in the model.

The pesticide sorbed, defined as the pesticide associated

with the solid phases (soil organic matter, soil sorbed LAS, and

LAS precipitates), showed similar behavior in the presence of

the LAS for the bulk soils and their size fractions. For example,

Fig. 5 presents the aqueous pesticide concentrations in rela-

tion to the aqueous LAS concentrations for the silt fractions,

describing general pesticide sorption behavior. As can be seen,

the aqueous pesticide concentrations first decreased sharply

with increasing aqueous LAS concentrations before the CMC

was reached due to the soil sorption and co-precipitation of

LAS with the divalent cations, followed by partitioning of

pesticide into the sorbed and precipitated LAS.

After the CMC of LAS was reached, the increase in the

aqueous pesticide concentration was due to the increasing

a LAS (2.00 g/l)

0

100

200

300

0 4 8 12

Diuron aqueous concentration (mg/l)

Diu

ron

sorb

ed c

onc.

(m

g/kg

)

Ag#1 Ag#2

Ag#3 Clayey

Sediment

12

b LAS (5.00 g/l)

0

100

200

300

0 4 8

Diuron aqueous concentration (mg/l)

Diu

ron

sorb

ed c

onc.

(m

g/kg

)

Ag#1 Ag#2 Ag#3

Clayey Sediment

Fig. 6 – Diuron desorption isotherm from clay fractions

with (a) linear alkylbenzene sulfonate (LAS) at 2.00 g/l;

(b) LAS at 5.00 g/l.

Table 3 – Percentage of pesticide remaining sorbed (D) and deswith LAS at 2.0 and 5.0 g/l.

LAS conc. (g/l) Diuron

Presorbed (mg/kg) 0 LAS¼ 2.0 g/l LAS¼ 5

Dw Ds E Ds

Bulk

Ag#1 27.4 35% 30% 1.07 28%

Ag#2 28.5 32% 53% 0.69 31%

Ag#3 33.7 42% 53% 0.81 128%

Clayey 24.6 28% 35% 0.9 22%

Sediment 18.3 29% 33% 0.94 27%

Clay

Ag#1 145.6 43% 22% 1.36 30%

Ag#2 144 41% 26% 1.25 28%

Ag#3 164.1 47% 35% 1.22 34%

Clayey 117.9 39% 41% 0.96 45%

Sediment 194.3 35% 53% 0.73 32%

Silt

Ag#1 92.2 46% 25% 1.39 23%

Ag#2 55.3 53% 29% 1.51 32%

Ag#3 56.5 57% 59% 0.95 69%

Clayey 47.5 53% 39% 1.3 49%

Sediment 68.6 56% 30% 1.58 39%

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4712

micelle concentration and the release of the pesticides from

the precipitated LAS. Atrazine showed similar behavior to

diuron although both descending and rising limbs were less

steep due to its lower affinity both to the sorbed and precipi-

tated LAS and LAS micelles because of its lower hydropho-

bicity. The maximal amount of the pesticide sorbed was

defined as the amount of the pesticide sorbed with both the

sorbed and precipitated LAS, where the aqueous pesticide

concentration is the lowest in Fig. 5; it is presented in Table 1

along with the LAS sorption capacities. As indicated by the

maximal amount of the pesticide sorbed in the presence of

LAS, pesticide sorption in the presence of LAS showed highly

size-dependent behavior. On a unit weight basis, the smaller

the size fractions, the more effective they were in sorbing

pesticide in the presence of LAS. The correlation coefficient

between the LAS sorption capacities and the maximal amount

of pesticide sorbed was 0.66 for diuron and 0.67 for atrazine.

3.6. Pesticide desorption in the presence of LAS

The pesticide desorption isotherms showed similar trends for

bulk soils and their various size fractions. Fig. 6 presents

diuron desorption isotherms from the clay fractions in the

absence and presence of the surfactants, as an example.

Desorption steps were from higher to lower pesticide

concentrations, or from right to left in Fig. 6. In the presence of

LAS, the amount of the pesticide sorbed increased during the

first desorption step, followed by a sharp decrease and then

slower decreases thereafter. The initial increase was caused

by sorption and precipitation of LAS and the resulting parti-

tioning of the pesticide into the sorbed and precipitated LAS.

Once the maximal LAS sorption had been reached, the

amount of the pesticide sorbed decreased. These are consis-

tent with the results of the equilibrium pesticide sorption in

the presence of LAS.

orption efficiency coefficients (E ) after five desorption steps

Atrazine

.0 g/l Presorbed (mg/kg) 0 LAS¼ 2.0 g/l LAS¼ 5.0 g/l

E Dw Ds E Ds E

1.11 12.8 41% 155% �0.92 218% �1.99

1.01 14 33% 157% �0.86 191% �1.36

�0.49 15 35% 186% �1.32 258% �2.43

1.08 13.8 32% 147% �0.69 195% �1.39

1.02 10.9 38% 192% �1.48 210% �1.78

1.22 55.3 45% 125% �0.45 112% �0.22

1.22 52.9 37% 137% �0.58 138% �0.6

1.24 49.3 38% 166% �1.07 198% �1.59

0.9 38.6 35% 107% �0.11 125% �0.38

1.04 52.9 42% 112% �0.2 98% 0.03

1.42 25.2 28% 146% �0.64 144% �0.62

1.44 15.8 36% 143% �0.67 98% �0.01

0.72 15.1 49% 169% �1.36 195% �1.88

1.08 16.2 52% 121% �0.43 95% 0.1

1.37 17.4 45% 124% �0.44 108% �0.15

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4 713

A desorption efficiency coefficient, E, can be defined as

follows (Wang and Keller, 2008a):

E ¼ ð1� DsÞð1� DwÞ

(2)

where Ds and Dw are the fractions of pesticide remaining sor-

bed after a given number of desorption steps in the presence

(Ds) and absence (Dw) of surfactant relative to the initial amount

of pesticide presorbed. In this study we considered five

desorption steps. An E> 1 indicates enhanced pesticide

desorption, while E< 1 represents an inhibited pesticide

desorption. If at the end of five desorption steps the amount of

pesticide remaining sorbed is greater than the initial amount of

pesticide presorbed, a negative E value is generated, in which

case, instead of desorption, an overall enhanced sorption

occurs. The measured Ds, Dw, and E are presented in Table 3.

The E values for atrazine were consistently negative,

indicating that enhanced sorption occurred, while, in most of

cases (67%), enhanced desorption occurred (E> 1) for diuron.

This reflects the lower hydrophobicity of atrazine relative to

diuron as discussed earlier. Thus, enhanced atrazine desorp-

tion requires even higher LAS concentration.

Since dissolved Ca2þ and Mg2þ from the bulk soils were

absent for the size fractions, meaningful comparison between

the bulk soils and the size fractions cannot be made in terms of

pesticide desorption behavior in the presence of LAS. For the

bulk soils, the correlation coefficients between LAS pseudo

sorption capacities and Ds were 0.95 and 0.83 for LAS concen-

tration of 2.0 and 5.0 g/l, indicating the LAS sorption and

precipitation largely explains diuron desorption behavior.

However, those of atrazine were only 0.30 and 0.57, which

reflects that at the LAS concentrations used, instead of enhan-

ced desorption, enhanced atrazine sorption was obtained.

Statistically higher E values were observed for the clay fractions

than for the silt fractions with diuron, indicting pesticide

desorption in the presence of LAS is soil size dependent.

Interestingly, for diuron, with LAS of 5.0 g/l, statistically

lower E values were obtained than with LAS of 2.0 g/l, sug-

gesting that higher surfactant concentrations are not neces-

sarily more efficient than lower concentrations under

sequential washing conditions. Higher initial LAS concentra-

tion can lead to greater amount of LAS sorbed and precipitated

in the sorption step and thus greater sorption of the pesticide

onto the sorbed and precipitated LAS. Although even higher

LAS concentrations can solubilize more pesticide out of their

sorbed phase, but the efficiency is not necessarily optimal.

4. Conclusions

Soil decontamination within a soil–water–anionic surfactant

system is complicated by the precipitation of the anionic

surfactant as it interacts with divalent cations. In addition to

other interactions, the target HOC contaminants partition to

the precipitated surfactant. This results initially in increased

HOC sorption onto the solid phases, rather than the intended

decontamination. HOC sorption onto the precipitated anionic

surfactant is strongly related to pesticide hydrophobicity. The

CMC of the anionic surfactant needs to be exceeded, which

has to take into account the loss (abstraction) of surfactant

due to its precipitation. The pseudo sorption capacity of the

anionic surfactant is highly correlated with the soil system

hardness, which is the soil property that controls the treat-

ment efficiency of anionic–surfactant aided soil washing

systems. For soils with high soil system hardness, the loss of

anionic surfactant might be so significant that the use of the

anionic surfactant might not be as efficient as a nonionic

surfactant. However, for soils with low soil system hardness,

an anionic surfactant is still a good candidate.

Supplementary data

Supplementary data associated with this article can be found

in the online version, at doi:10.1016/j.watres.2008.10.052.

r e f e r e n c e s

Bohn, H.L., McNeal, B.L., O’Connor, G.A., 2001. Soil Chemistry,third ed. John Wiley & Sons, Inc., New York, USA.

Carter, M.R., 1993. Soil Sampling and Methods of Analysis. LewisPublishers, Boca Raton, FL.

Chu, W., 2003. Remediation of contaminated soils by surfactant-aided soil washing. Pract. Periodical of Haz. Toxic. andRadioactive Waste Mgmt. 7, 19–24.

Di Toro, D.M., Dodge, L.J., Hand, V.C., 1990. A model for anionicsurfactant sorption. Environ. Sci. Technol. 24, 1013–1020.

Garcia, M.T., Campos, E., Dalmau, M., Ribosa, I., Sanchez-Leal, J.,2002. Structure-activity relationships for association of linearalkylbenzene sulfonates with activated sludge. Chemosphere49, 279–286.

Hanna, H.S., Somasundaran, P., 1979. Adsorption of sulfonates onreservoir rocks. J. Colloid Interface Sci. 19, 221–232.

Higgins, C.P., Luthy, R., 2006. Sorption of perfluorinated surfactantson sediments. Environ. Sci. Technol. 40, 7251–7256.

Higgins, C.P., Luthy, R., 2007. Modeling sorption of anionicsurfactants onto sediment materials: an a priori approach forperfluoroalkyl surfactants and liner alkylbenzene sulfonates.Environ. Sci. Technol. 41, 3254–3261.

Jafvert, C.T., 1991. Sediment- and saturated-soil-associatedreactions involving an anionic surfactant (dodecylsulfate) 2.partition of PAH compounds among phases. Environ. Sci.Technol. 25, 1039–1045.

Jafvert, C.T., Heath, J.K., 1991. Sediment- and saturated-soil-associated reactions involving an anionic surfactant(dodecylsulfate) 1. precipitation and micelle formation.Environ. Sci. Technol. 25, 1031–1039.

Kile, D.E., Chiou, C.T., 1989. Water solubility enhancements ofDDT and trichlorobenzene by some surfactants below andabove the critical micelle concentration. Environ. Sci. Technol.23, 832–838.

Mata-Sandoval, J.C., Karns, J., Torrents, A., 2002. Influence ofRhamnolipids and Triton X-100 on the desorption of pesticidefrom soil. Environ. Sci. Technol. 36, 4669–4675.

Matheson, K.L., Cox, M.F., Smith, D.L., 1985. Interactions betweenlinear alkylbenzene sulfonates and water hardness ions: I effectof calcium ion on surfactant solubility and implication fordetergency performance. J. Am. Oil Chem. Soc. 62, 1391–1396.

Paria, S., 2007. Surfactant-enhanced remediation of organiccontaminated soil and water. Adv. Colloid Interface Sci., doi:10.1016/j.cis.2007.11.001.

Rodriguez-Cruz, M.S., Sanchez-Martin, M.J., Sanchez-Camazano, M., 2006. Surfactant-enhanced desorption of

w a t e r r e s e a r c h 4 3 ( 2 0 0 9 ) 7 0 6 – 7 1 4714

Atrazine and Linuron residues as affected by aging ofherbicides in soil. Arch. Environ. Contam. Toxicol. 50, 128–137.

Sanchez-Camazano, M., Rodriguez-Cruz, S., Sanchez-Martin, M.,2003. Evaluation of component characteristics of soil-surfactant-herbicide system that affect enhanced desorptionof linuron and atrazine preadsorbed by soils. Environ. Sci.Technol. 37, 2759–2766.

Savitsky, A.C., Wiers, B.H., Wendt, R.H., 1981. Adsorption of organiccompounds from dilute aqueous solutions onto the externalsurface of type A zeolite. Environ. Sci. Technol. 15, 1991–1996.

Stellner, K.L., Scamehorn, J.F., 1989a. Hardness tolerance ofanionic surfactant solutions 1. anionic surfactant with addedmonovalent electrolyte. Langmuir 5, 70–77.

Stellner, K.L., Scamehorn, J.F., 1989b. Hardness tolerance ofanionic surfactant solutions 2. effect of added nonionicsurfactant. Langmuir 5, 77–84.

Verge, C., Moreno, A., Bravo, J., Berna, J.L., 2001. Influence of waterhardness on the bioavailability and toxicity of linearalkylbenzene sulphonate (LAS). Chemosphere 44, 1749–1757.

Wang, P., Keller, A.A., 2008a. Particle-size dependent sorption anddesorption of pesticides within water-soil-nonionic surfactantsystems. Environ. Sci. Technol. 42, 3381–3387.

Wang, P., Keller, A.A., 2008b. Partitioning of hydrophobic organiccompounds within soil-water-surfactant systems. Water Res.42, 2093–2101.

Wang, P., Keller, A.A., 2008c. Soil particle-size dependentpartitioning behavior of pesticides within water–soil–cationicsurfactant systems. Water Res. 42 (14), 3781–3788.

Westall, J.C., Chen, H., Zheng, W.J., Brownawell, B.J., 1999.Sorption of linear alkylbenzenesulfonates on sedimentmaterials. Environ. Sci. Technol. 33, 3110–3118.

Yang, K., Zhu, L., Xing, B.S., 2007. Sorption of sodiumdodecylbenzene sulfonate by montmorillonite. Environ.Pollut. 145, 571–576.

Yang, K., Zhu, L.Z., Xing, B.S., 2006. Enhanced soil washing ofphenanthrene by mixed solutions of TX100 and SDBS.Environ. Sci. Technol. 40, 4274–4280.

Zhu, L.Z., Yang, K., Lou, B.F., Yuan, B.H., 2003a. A multi-component statistic analysis for the influence of sediment/soil composition on the sorption of a nonionic surfactant(Triton X-100) onto natural sediments/soils. Water Res. 37,4792–4800.

Zhu, L.Z., Chen, B.L., Tao, S., 2003b. Interactions of organiccontaminants with mineral-adsorbed surfactants. Environ.Sci. Technol. 37, 4001–4006.