Chemosphere 62 (2006) 17–25

www.elsevier.com/locate/chemosphere

Evaluation of soil flushing potential for clean-up ofdesert soil contaminated by industrial wastewater

Shai Arnon a,b,*, Zeev Ronen a, Alexander Yakirevich a, Eilon Adar a,b

a Department of Environmental Hydrology and Microbiology, Zuckerberg Institute for Water Research,

J. Blaustein Institutes for Desert Research, Ben-Gurion University of the Negev, Sede-Boqer Campus 84990, Israelb Department of Geological and Environmental Sciences, Ben-Gurion University of the Negev, Israel

Received 13 January 2005; received in revised form 28 March 2005; accepted 6 April 2005

Available online 9 June 2005

Abstract

The flushing potential of a desert loess soil contaminated by the flame retardant Tetrabromobisphenol A (TBBPA),

chloride (Cl�) and bromide (Br�) was studied in undisturbed laboratory column experiments (20 cm diameter, 45 cm

long) and a small field plot (2 · 2 m). While the soluble inorganic ions (Cl� and Br�) were efficiently flushed from

the soil profile after less than three pore volumes (PV) of water, about 50% of the initial amount of TBBPA in the soil

was also flushed, despite its hydrophobic nature. TBBPA leaching was made possible due to a significant increase in the

pH of the soil solution from 7.5 to 9, which increased TBBPA aqueous solubility. The remaining TBBPA mass in

the soil was not mobilized from its initial location in the topsoil due to the decrease in pH at this horizon. In situ soil

flushing demonstrated that this method is a feasible treatment for reducing soil contamination at this site.

� 2005 Elsevier Ltd. All rights reserved.

Keywords: Tetrabromobisphenol A; Loess; Leaching; Ionizable organic compounds

1. Introduction

In recent years, inappropriate waste disposal has

become a major environmental problem that requires

cost-effective remediation solutions. The ability of contami-

nants to reach groundwater and put water resources at

risk lies in their inherent physico-chemical properties,

as well as their resistance to removal by complex physi-

0045-6535/$ - see front matter � 2005 Elsevier Ltd. All rights reserv

doi:10.1016/j.chemosphere.2005.04.050

* Corresponding author. Address: Department of Civil and

Environmental Engineering, Northwestern University, 2145

Sheridan Road, Evanston, IL 60208-3109, USA. Tel.: +1 847

467 4980; fax: +1 847 491 4011.

E-mail addresses: s-arnon@northwestern.edu (S. Arnon),

zeevrone@bgumail.bgu.ac.il (Z. Ronen), alexy@bgumail.bgu.

ac.il (A. Yakirevich), eilon@bgumail.bgu.ac.il (E. Adar).

cal, chemical and biological reactions in unsaturated and

saturated geological media (Mackay et al., 1985; Saint-

Fort, 1991; Schwarzenbach et al., 1993). Applying

water, with or without additives, is frequently used for

soil flushing in order to clean up the vadose zone (Thom-

sen et al., 1989; MacKay et al., 1996). In situ soil flush-

ing can be limited due to several reasons, primarily: (1)

soils that have been contaminated for a long period of

time exhibit a bi-phasic pattern of desorption with an

initial fast stage (min–h) and a subsequent longer slow

phase (days–years) (Pavlostathis and Jaglal, 1991); (2)

the heterogeneous nature of soil implies that regions

with low hydraulic conductivity may exist, from which

contaminants must diffuse over a long period of time

to the more effectively flushed layers (MacKay et al.,

1996); and (3) the slow solubility rates of hydrophobic

ed.

18 S. Arnon et al. / Chemosphere 62 (2006) 17–25

compounds further increase the flushing duration. Iden-

tifying which process limits the clean-up of a particular

site may assist in suggesting approaches for enhancing

the clean-up. For example, MacKay et al. (1996) con-

cluded that the flushing pattern is highly dependent on

the desorption and solubility rates of the different tested

compounds (benzene, toluene, ethylbenzene and xylene).

They also found that the presence of a non-aqueous

phase liquid would substantially increase the site�s esti-mated flushing duration. Thomsen et al. (1989) demon-

strated efficient cleaning of the unsaturated zone from

volatile organic compounds by using in situ soil flushing

(with water), groundwater pumping, ex situ treatment

and recharge. In many other cases the use of specific

solutions has been suggested, such as water with surfac-

tants or organic solvents, rather than clean water, to en-

hance contaminant flushing (e.g., Bettahar et al., 1999;

Di Palma, 2003). Nevertheless, the use of water for

flushing, without additives, reduces operation activities

and costs.

In this study, we assess the flushing potential of

undisturbed contaminated desert soil at the laboratory

and small field scales, as an optional method for soil

remediation. The site, 25 acres of desert loess soil, 0.1–

2 m thick, overlying fractured chalk bedrock, was con-

taminated in the late 1980s as a result of industrial

wastewater disposal in a forced evaporation facility

(Nativ et al., 1999). The inorganic contamination was

dominated by Cl� and Br� ions and the major organic

contaminant was tetrabromobisphenol A (TBBPA).

Although soil flushing might increase the risk of ground-

water contamination, when combined with additional

groundwater treatment it can provide an inexpensive

method for clean-up of contaminated soils (Thomsen

et al., 1989; MacKay et al., 1996). An underground

drain, a pumping station and a treatment facility were

already in place at the downstream edge of this site

(Nativ et al., 2003); therefore soil flushing was proposed

as a solution to clean up the contaminated soil.

TBBPA (Fig. 1) is used as a flame retardant in elec-

tronic circuit boards and in the plastics industry (de

Wit, 2002). It is of concern since it was found in stream

sediments and municipal wastewater, which also makes

it a potential groundwater contaminant (Sellstrom and

Jansson, 1995; Ronen and Abeliovich, 2000; Oberg

Br

Br

Br

Br

HO OH

CH3

CH3

C

Fig. 1. Chemical structure of TBBPA.

et al., 2002). In some cases, increased levels of TBBPA

were linked to industrial sources (Oberg et al., 2002).

There is evidence that prolonged exposure of rats to

TBBPA disturbs the liver heme metabolism (Szymanska

et al., 2000), as well as the neural system (Eriksson et al.,

1998). In vitro assays demonstrated that TBBPA was up

to 25 times more potent in binding to human transthy-

retin (thyroid transport protein) than thyroxin (the na-

tive hormone) (Brouwer, 1998).

TBBPA aqueous solubility, sorption and bioavail-

ability are pH dependent since TBBPA can be partially

ionized (Lee et al., 1991). TBBPA is not easily

biodegraded and only recently was its mineralization

demonstrated through dehalogenation under anaerobic

conditions and further biodegradation of bisphenol A

under aerobic conditions (Ronen and Abeliovich,

2000; Voordeckers et al., 2002; Arbeli and Ronen,

2003). Ionized compounds such as TBBPA exhibit com-

plex transport behavior, as some of their physico-chemi-

cal properties are pH dependent. For example,

Schellenberg et al. (1984) demonstrated that the equilib-

rium partition coefficient (Kd) of chlorinated phenols

between the sorbent and the solution could be estimated

based upon their lipophilicity, as expressed by the octa-

nol/water partition coefficient (Kow), and on the organic

carbon contents (foc) of the sorbent. Mathematical rela-

tionships between Kd, Kow and foc have been derived for

various sets of compounds and natural sorbents:

Kd ¼ focbðkowÞa ð1Þ

where a and b are coefficients peculiar to the organic

compounds under consideration. For example, Schellen-

berg et al. (1984) found from their experiment with chlo-

rinated phenols that the values of a and b are 0.82 and

1.05, respectively. They also stated that if the pH of

the solution is not more than one unit above the acid

dissociation constant (pKa) of the compound, the contri-

bution of the phenolate ion to the overall adsorption of

the phenolic compound could be neglected. For pHs

higher than one unit above the pKa, the aforementioned

adsorption model (1) failed to describe phenolate

adsorption, as was shown by Fiore et al. (2003) and by

Shimizu et al. (1992). The latter have demonstrated that

the mineral constituents (mainly clay content) of the

soils controlled the adsorption of the ionic form of

pentachlorophenol.

2. Materials and methods

2.1. Sample collection and soil properties

The soil monoliths were collected from a contami-

nated site located in the northern Negev desert, Israel.

The sampling technique was designed to collect soil

cores, 10 cm in diameter and �45 cm long, in an attempt

S. Arnon et al. / Chemosphere 62 (2006) 17–25 19

to represent the soil structure at the field site. The col-

umns consisted of a PVC cylinder with a removable

sharpened steel edge attached to its base. As the sharp-

ened column edge was pressed down by weight, shovels

were used to remove the soil around the exterior of the

steel edge to lower the soil resistance and compaction

near the monolith edge. At all times, 5 cm of soil re-

mained constantly around the column edge to prevent

separation of the soil and the steel edge. When the col-

umn was filled, the steel edge was removed, and a

2 cm layer of glass fibers was positioned at both core

ends to prevent solid particles from being washed out

of the column during the flushing experiment. The col-

umn was then sealed with caps and transported to the

laboratory. Soil samples adjacent to the sharpened edge

were taken during the collection of the monoliths for

assessment of the initial contaminant concentrations

and soil properties.

Selected soil properties appear in Table 1. The major

soil constituent was quartz (>50%). Other minerals that

were found in the soil, in order of abundance, are: cal-

cite, feldspar, gypsum, kaolinite and illite, as determined

by X-ray diffraction analysis (Philips XRD diffractome-

ter). Halite was found only in the upper section of the

soil. Soil particle distribution was measured using a laser

diffraction system (Malvern mastersizer). Specific sur-

face area was also measured using the Malvern master-

sizer. Natural organic matter was measured using the

dichromate oxidation method (Lowell, 1993). Bulk den-

sity was found by measuring the dry mass of a known

structured sample with a defined volume, and porosity

was estimated using the bulk density and particle density

(taken as 2.65 gcm�3) (Lowell, 1993).

2.2. Batch desorption kinetics

Five grams (dry weight) of contaminated sieved soil

(<2 mm) were mixed with 15 ml of double distilled water

inside 20 ml glass vials with Teflon screw caps (24 sam-

ples). The vials were shaken at 200 rpm at 25 �C. At each

time interval, two vials were removed and centrifuged

Table 1

Soil properties

Parameter Value

Sand % (50–2000 lm) 13 ± 3

Silt % (2–50 lm) 63 ± 4

Clay % (<2 lm) 25 ± 6

Specific surface (m2 g�1) 3.9 ± 1.3

Natural organic

carbon content

(%) 0.14 ± 0.03

Porosity 0.4 ± 0.02

Bulk density (g cm�3) 1.54

The averaged results were calculated from soil samples collected

at depths of 0–50 cm.

for 15 min at 3000 rpm to separate the supernatant

and the soil. TBBPA concentrations were quantified at

14 time intervals, not equally distributed, over 100 h

(0.016, 0.083, 0.25, 0.5, 0.75, 1, 1.5, 2, 3, 5, 13, 24, 50

and 100 h). Biodegradation of TBBPA was not pre-

vented in this experiment, either by soil sterilization or

by biocide addition, as it was previously shown that

TBBPA is not degraded in this soil under aerobic condi-

tions (Ronen and Abeliovich, 2000).

2.3. Soil extracts

Soil extractions were performed for measurement of

the major ions, TOC (total organic carbon) and pH by

mixing air-dried sieved soil (<2 mm) and double distilled

water at a 1:1 weight ratio (40 g) inside 125 ml flasks

(triplicates). The soil–water mixtures were shaken at

200 rpm at 25 �C for 24 h and filtered through GF/C fil-

ters (Whatman) prior to the analysis.

TBBPA was extracted from contaminated sieved soil

(<2 mm) according to method no. 3550 (EPA, 1997).

Ten grams of contaminated soil and 50 ml of ethyl-ace-

tate were added to 250 ml flasks equipped with Teflon

screw caps (triplicates). The mixtures were sonicated

for 15 min and shaken at 200 rpm at 25 �C for 12 h.

The solution was separated from the soil by filtration

through GF/C filters (Whatman) and concentrated via

evaporation to 1 ml. O-hydroxybiphenyl (97%, Aldrich)

was added to the soil before extraction, as an internal

standard for the assessment of the TBBPA recovery.

2.4. Batch solubility test

TBBPA solubility was examined under a pH range of

7–9. Five different buffer solutions were prepared from

Tris-HCl 0.1 M with pH values of 7, 7.5, 8, 8.5 and 9.

TBBPA in acetone was added to 20 ml glass vials

equipped with Teflon screw caps in quantities which,

after acetone evaporation and addition of 10 ml of buf-

fer solutions and upon complete dissolution, yielded

maximum concentrations of 100, 200, 300, 400 and

500 mg l�1 (i.e. each concentration was examined at five

pH values). Duplicate samples were shaken at 200 rpm

at 25 �C for 10 h. The 10 ml solutions were filtered

(25 mm; 0.45 lm; Gelman) and the last 1 ml was taken

for analysis by high-performance liquid chromatogra-

phy (HPLC), after preliminary experiments verified that

losses of TBBPA by sorption onto the filter in this pro-

cedure are negligible.

2.5. Column experiments

Soil flushing with tap water was investigated during

column experiments using three undisturbed soil cores.

The composition of the tap water used for flushing

was as follows (mg l�1): 68.1 (Ca2+), 40.5 (Mg2+), 120

0

20

40

60

80

100

7 7.5 8 8.5 9

100 mg/l200 mg/l300 mg/l400 mg/l500 mg/l

Solu

ble

fract

ion

(%)

pH

Initial concentration

Fig. 2. The effect of pH on TBBPA solubility. The relative

amount of soluble TBBPA is displayed (soluble fraction) at five

different initial concentrations. The results are averages ± the

SDs of duplicates.

20 S. Arnon et al. / Chemosphere 62 (2006) 17–25

(Na+), 5.35 (K+), 220 (Cl�), 41.4 (SO2�4 ) and pH of 7.5.

The columns were placed upside down and flow was in

an upward direction (from the surface soil into the soil

profile), to diminish air trapping. The soil was saturated

by introducing water in the direction of flow under rela-

tively small head differences (2–4 cm). Water level was

increased based on the predicted movement of the water

inside the column (based on the hydraulic conductivity

of this soil), until water exited the column. During the

flushing experiments, a constant head controlled the

flow, and Darcy�s law was used to calculate the hydrau-

lic conductivity based on timed collections of fixed

volumes of water under three different hydraulic gradi-

ents. The effluent was analyzed for Cl�, Br� and TBBPA

concentrations until TBBPA concentration fell below

the detection limit. The pH, EC (electrical conductivity)

and TOC were also measured.

2.6. In situ soil flushing

A ‘‘double ring-like’’ infiltration pond was con-

structed at the contaminated site. The dimension of

the outer pond was 2 · 2 m, and the inner pond was

1.2 · 1.2 m. Both ponds were filled with tap water grad-

ually over 4 h, maintaining the same levels at both sec-

tions, until the water level reached 45 cm above the

surface. This level was kept constant for two more hours

and then the experiment was terminated (i.e. water was

drained). EC and TOC profiles in the soil were measured

by analyzing soil solution extracts using the same method

discussed earlier.

2.7. Chemical analyses

The following measurements were carried out in the

water samples: EC, pH and TOC with a Dohrmann

DC-190 (Rosemount). Cl� and Br� were measured with

an ion chromatograph (Dionex, 4500i). TBBPA concen-

trations in the soil (extracted by ethyl-acetate) were ana-

lyzed with a gas chromatograph (GC). The GC in use

was an HP 5490 equipped with a 15 m capillary column

(SPB5), 0.53 mm inner diameter and 1.5 lm film thick-

ness (Supelco). The temperature program began at

100 �C for 4 min, followed by a ramp at 10 �C min�1

until 280 �C, where the temperature was held for 8 min.

Helium was used as a carrier gas at 12 ml min�1. The

compounds were detected with a flame ionization detec-

tor at a temperature of 250 �C. TBBPA concentrations

were corrected according to the internal standard recov-

eries. Error in analysis was estimated at ±5% and a con-

centration of 1 mg l�1 was considered the low detection

limit based on the signal to noise ratio.

TBBPA concentrations from the desorption and sol-

ubility experiments were analyzed with HPLC (Kon-

tron), equipped with a 4.6 mm by 25 cm C18 column

(Supelcosil), and with UV detection (Diode array 440)

at a wavelength of 290 nm. The mobile phase consisted

of two solutions: (A) methanol and acetic acid (1%)

and (B) ammonium acetate 0.018 M-H2O with acetic

acid (1%). The initial A:B ratio was 60%:40%. The

A:B ratio was changed at a constant rate during

13 min until 100% A was reached. The final stage was

maintained for 1 min. Error in analysis was estimated

at ±2% and a concentration of 0.2 mg l�1 was consid-

ered the low detection limit based on the signal to noise

ratio. All values in the chemical analysis are given as

mean ± standard deviation (SD).

3. Results and Discussion

3.1. TBBPA solubility

The effect of the solution pH on TBBPA solubility

was investigated at a variety of pHs and initial TBBPA

masses (Fig. 2). Below pH 7.5 TBBPA was not detected

(concentrations <0.2 mg l�1, i.e. below the detection

limit), whilst above pH 8 there was a sharp increase in

its solubility, reaching complete solubility of the com-

pound at pH 9 (up to 500 mg l�1). In addition, the solu-

ble fraction of TBBPA increased with higher initial

TBBPA input under the entire examined pH range (7–9).

When the pH is increased by one unit above the pKa,

�90% of the hydrophobic ionizable organic compounds,

such as TBBPA, is in its ionized form and its solubility

in water is expected to rise by several orders of magni-

tude compared to its solubility at pH values below the

pKa (Schwarzenbach et al., 1993). Recently, Eriksson

et al. (2004) reported that although TBBPA has two hy-

droxy-groups, the differences in the pKa�s (7.5 and 8.5,

WHO, 1995) is relatively small, indicating that the two

rings may function, at least in part, independently of

each other. In another report that was recently submit-

ted to the EPA by the Brominated Flame Retardant

S. Arnon et al. / Chemosphere 62 (2006) 17–25 21

Industry Panel (BFRIP) (BFRIP, 2004), the pKa of

TBBPA was reported to be 9.4. While the pKa�s of

TBBPA in the aforementioned studies (BFRIP, 2004;

Eriksson et al., 2004) differ significantly from each other,

our experimental results suggest that the values reported

by Eriksson et al. (2004) better describe the behavior of

TBBPA, since TBBPA concentrations increase signifi-

cantly above pH 8 (up to 500 mg l�1). The BFRIP report

also states that the solubility of TBBPA at pHs 5, 7, and

9 was 0.148, 1.26 and 2.34 mg l�1, respectively. These re-

sults also differ from our results shown in Fig. 2. One

possible reason for the difference between our results

and those reported by the BFRIP might lie in the specific

details of the experimental procedure (e.g., what is the

maximum TBBPA concentration that might be expected

from the amount of TBBPA that was added to the

water).

3.2. TBBPA desorption

The kinetics of TBBPA emergence in the solution,

integrating both the release from the sorbent (desorp-

tion) and the dissolution (if TBBPA is in solid form)

as a single process, are illustrated in Fig. 3. During the

first five min of the experiment the concentrations

reached �70–80% of the maximum released concentra-

tion, observed 90 min after the start of the experiment.

Over the next 4 d, the concentrations did not differ from

the value observed after 90 min (data not shown). The

solution pH increased after 5 min from 7.5 to 8.5 and re-

mained constant until the end of the experiment. When

desorption controls the appearance of a compound in a

soil–water solution, we usually observe bi-phasic behav-

ior involving an initial fast rise and a second slower rise

(e.g., Pavlostathis and Jaglal, 1991; Opdyke and Loehr,

1999). This behavior can be seen clearly in Fig. 3, where

the initial fast increase was in the order of min and a fur-

ther increase, up to the maximum concentration after

0

0.2

0.4

0.6

0.8

1

0 20 40 60 80

TBBPA

Rel

ativ

e co

ncen

tratio

n

Time (min)

Fig. 3. TBBPA desorption kinetics from the contaminated soil.

The concentrations are normalized to the maximum observed

concentration (15 mg l�1, after 90 min). The results are averages

± the SDs of duplicates.

90 min, was much slower. The fast increase in TBBPA

also implies that if dissolution is taking place, its kinetics

do not have a major impact on the release of TBBPA

into the solution.

The second sample (after 2 min) was exceptional

since it did not follow the general smooth increase as ex-

pected from a desorption experiment and as seen for the

rest of the data. Given that we used 24 replicates, in

which duplicates were sacrificed at each time measure-

ment, it might be that some heterogeneity was apparent

despite the initial soil mixing, causing the non-smooth

increase in the first 20 min.

3.3. Soil flushing

The breakthrough curves (BTCs) of Cl�, Br� and

TBBPA during the soil flushing are presented in Fig.

4. Since similar observations were documented for

all tested cores (Arnon, 1996), the results presented in

Figs. 4–7 were obtained only from column CL2, to avoid

redundancy. High Cl� and Br� concentrations (accoun-

ting for more than 95% of the water EC) were observed

soon after the water started to leach out from the col-

umn, while maximum concentrations appeared simulta-

neously after 0.42 pore volumes (PV). The saturation

procedure of the initially dried soil caused high solute

concentrations soon after water started to flow out

(Fig. 4). Since the water for saturation was applied in

the direction of the flow, the first PV that was introduced

in order to saturate the column leached considerable

amounts of the contaminants toward the outlet. Never-

theless, the volume of water that was used for saturation

is not shown in Fig. 4, since the PV was calculated based

on the accumulated effluent volume (and soil porosity

data).

The transport of TBBPA was slow compared with

that of Cl� and Br�, with calculation of a retardation

factor based on the relative timings of Cl�, Br� and

TBBPA maximum concentrations. The averaged retar-

dation factor was 3.2 (±0.6), as calculated using data

from all three tested columns. After approximately

0

50

100

150

200

250

300

0

4000

8000

12000

16000

0 1 2 3 4 5

TBBPACl-Br-

TBBP

A (m

g/l)

Cl - and Br - (m

g/l)

Pore volume

Fig. 4. BTCs of Cl�, Br� and TBBPA during the continuous

flushing of initially dry, undisturbed contaminated soil columns

using tap water.

0 50 100 150 200 250 300 350 400

Before flushingAfter flushing

TBBPA (mg/kgdry soil)

b

0 1000 2000 3000 4000 5000

0

10

20

30

40

Cl- - before flushingCl- - after flushingBr- - before flushingBr- - after flushing

Concentration (mg/kgdry soil)

Dep

th (c

m)

a

Fig. 5. Soil concentration profiles of Cl�, Br� (a) and TBBPA (b) before and after the continuous flushing of an undisturbed

contaminated soil column (Fig. 3). Each profile was constructed by five measurements evenly distributed along the profile. The results

are averages ± the SDs of triplicates.

0

20

40

60

80

100

7.5

8

8.5

9

0 0.5 1 1.5 2 2.5 3 3.5

TBBPA pH

TBBP

A re

lativ

e m

ass

rem

oval

(%)

pH

Pore volume

Fig. 6. The change in effluent pH along with the TBBPA mass

that was flushed out of the column. TBBPA mass is calculated

relative to its initial amount in the soil. The results are averages

± the SDs of triplicates.

7 7.5 8 8.5 9 9.5 10

0

10

20

30

40

Before flushingAfter flushing

pH

Dep

th (c

m)

Fig. 7. Soil extract pH before and after the continuous flushing

of an undisturbed contaminated soil column (Fig. 3). Each

profile was constructed from five measurements evenly distrib-

uted along the profile. The results are averages ± the SDs of

triplicates.

Table 2

Hydraulic conditions during the soil flushing

Pore volume Hydraulic

gradient

Flux

(ml h�1)

Hydraulic conductivity

(cm s�1)

0–0.92 0.22 23.1–100 1.15 · 10�3– 4.3 · 10�3

0.92–1.69 0.44 19.1– 43.8 4.78 · 10�4– 1.09 · 10�3

1.69–6 0.66 31.8– 46.9 5.29 · 10�4–7.81 · 10�4

22 S. Arnon et al. / Chemosphere 62 (2006) 17–25

5 PV, the TBBPA concentration declined to below its

detection limit (0.2 mg l�1) and the experiment was

terminated.

The fluxes, hydraulic gradients and hydraulic con-

ductivity during the flushing experiment are summarized

in Table 2. The flow rate through the columns was not

constant as the hydraulic conductivity decreased by a

factor of 4, from �4 · 10�3 to �1 · 10�3 cm s�1 during

the first PV (Table 2), probably due to swelling of clay

minerals.

3.4. Soil flushing efficiency

The concentrations of Cl�, Br� and TBBPA in the

soil before and after the flushing are illustrated in

Fig. 5. Most of the initial contaminant mass was concen-

trated in the upper part of the soil. This phenomenon is

typical for arid climates where evaporation is dominant,

rather than infiltration and deep percolation (Allison

and Hughes, 1983; Scanlon, 1991). Fig. 5a shows that

Cl� and Br� were completely flushed from the soil pro-

file. The average Cl� concentration in the soil profile

after the flushing was 56 mg kg�1, since the injected

water contained 220 mg l�1 of Cl�. Conversely, TBBPA

was only partly flushed from the soil (Fig. 5b), with the

remaining mass being retained in the topsoil. The mass

recoveries of TBBPA, Cl� and Br� for the three tested

columns are summarized in Table 3. Mass balance calcu-

lations were made according to:

Mass balance ð%Þ ¼ F l � InBf � Af

� �� 100 ð2Þ

Table 3

Mass balance for Cl�, Br� and TBBPA during the column experiments

Column number CL2 CL3 CR3

Contaminant Cl� Br� TBBPA Cl� Br� TBBPA Cl� Br� TBBPA

Mass balance (%) 103 84 110 105 88 116 109 81 114

Retained TBBPA in the soil

(% of the initial mass)

12 17 42

S. Arnon et al. / Chemosphere 62 (2006) 17–25 23

where Fl is the mass that was leached out of the column

(M), In is the mass of the solute that was introduced by

the water (relevant only for chloride) (M), Bf is the mass

in the soil before flushing (M) and Af is the mass in the

soil after flushing (M).

None of the above mass balance calculations exceeds

±20% from the ideal mass recovery. Error in mass bal-

ance was probably due to the non-dense sampling in

the upper part of the soil where most of the mass was

concentrated. Imperfect soil extraction may also have

contributed to the error. Nevertheless, the results pre-

sented in Fig. 5 and Table 3 clearly illustrates the effi-

cient salt flushing and the incomplete TBBPA flushing.

3.5. The effect of pH on the flushing of TBBPA

The effluent pH rose above 8 after approximately

1.2 PV; at the same time a sharp increase in TBBPA

mass removal was observed (Fig. 6). Despite the fact

that the effluent pH reached values between 8.5 and 9,

a significant fraction of TBBPA was retained in the top-

soil (Fig. 5). The profound effect of pH on TBBPA sol-

ubility (Fig. 2) and desorption (Fig. 3) was discussed

earlier in the text. This is also the reason for the incom-

plete flushing of TBBPA during the column experiments,

which is further illustrated in Fig. 7. Fig. 7 shows the pH

of the soil extract at the beginning and end of the col-

umn experiments. The pH was slightly reduced in the

upper section of the soil and significantly increased in

the regions closer to the outlet at the end of the experi-

0 5 10

0

20

40

60

80

100

120

Before flushingAfter flushing

EC (mS/cm)

Dep

th (c

m)

a

Fig. 8. Soil EC (a) and TOC (b) profiles before and after the continuo

soil. Each profile is generated from two separate soil cores taken at th

The results are averages ± the SDs of duplicates.

ments. The reduced potential for pH increase in the top-

soil caused incomplete dissolution of TBBPA and was

the reason why significant mass of TBBPA was not

mobilized.

Another major factor that may affect TBBPA trans-

port is adsorption to the soil and natural organic car-

bon. The Kd of TBBPA on the loess soil (at pH 8.5)

was found to be 0.09 ml g�1 (Arnon, 1996). This value

alone could not account for the delayed BTC of TBBPA

as compared with that of Cl� and Br� (Fig. 4). The Kd

of TBBPA is affected by its ionic and non-ionic species

as discussed previously with respect to other ionized or-

ganic compounds (Lee et al., 1991; Shimizu et al., 1992;

Fiore et al., 2003). As the pH increases we expect the Kd

to be less dependent on the natural organic content of

the soil (Schellenberg et al., 1984; Lee et al., 1991) and

to be more influenced by the mineralogy of the soil,

mainly by the clay content (Shimizu et al., 1992). Thus,

we suggest that pH changes in the soil columns with time

affected the Kd and sorption of TBBPA and thus played

a major role in controlling its transport. However, the

lack of information on the spatial and temporal pH val-

ues prevents a quantitative assessment of the relative

role of Kd on TBBPA transport.

3.6. Soil flushing at the field site

The flushing potential was also examined at a con-

taminated field site from which the columns for the

laboratory experiments were taken. In this experiment,

15 10 20 30 40 50 60 70

Before flushing

After flushing

Concentration (mg/Kgdry soil)

b

us flushing of a pilot scale (2 · 2 m), undisturbed contaminated

e outer ring (before flushing) and the inner ring (after flushing).

24 S. Arnon et al. / Chemosphere 62 (2006) 17–25

TOC was monitored as a marker of organic contami-

nant (mainly TBBPA) concentrations, while the EC

represented the soluble salt (Cl� and Br�) concentra-

tions. Both tracers, TOC and EC, exhibited linear cor-

relations to TBBPA and to Cl� + Br� (R2 = 0.8 and

0.94, respectively), as measured during the laboratory

experiments. Fig. 8 shows the EC and TOC profiles

in the soil before and after flushing. No comprehensive

effort was made to deal with the spatial distribution of

the soil contaminants. Therefore, the results in Fig. 8

should be treated as qualitative, expressing mainly the

major patterns of the contaminants� distribution in

the soil.

The amount of water that percolated into the soil was

equal to 1 PV of a soil section up to a depth of 0.9 m.

This was calculated based on the volume of water that

percolated through the inner pond (1.2 · 1.2 m) and a

porosity of 0.4 (Table 1). The salts front (EC) leached

into the soil profile up to 90 cm, leaving the topsoil al-

most completely leached (Fig. 8a). The calculated PV

and the depth of the front verify that the salts were effi-

ciently flushed from the soil, as observed in the labora-

tory studies. The TOC profile suggests that a major

fraction of the organic contaminants still remained in

the topsoil (Fig. 8b), similar to the laboratory experi-

ments (Fig. 5). Nevertheless, a significant amount of

TOC was still abundant in the entire soil profile as a re-

sult of the relatively small volume of water that was used

(1 PV).

4. Summary and conclusions

The results of this study suggest that soil flushing can

be efficiently applied to remediate contaminated desert

soils. The laboratory experiments show that complete re-

moval of Cl� and Br� from the soil required about 3 PV,

while at the same time more than 50% of TBBPA initial

mass was also removed. Nevertheless, after 5 PV,

TBBPA leaching became inefficient due to inability to

increase the pH to the value needed for dissociation of

TBBPA in the topsoil. Thus, the remaining TBBPA

mass was sustained in the uppermost part of the soil pro-

file rather than being leached to the lower parts of the

profile. An artificial increase in the flushing solution

pH has the potential to increase the efficiency of this

treatment.

A pilot scale field experiment demonstrated that in

situ soil flushing is a feasible treatment for reducing con-

tamination levels in desert loess soils, when accompa-

nied by groundwater extraction and treatment (‘‘pump

& treat’’). Since saline soils are ubiquitous in arid zones

it is expected that high pH will be developed upon wet-

ting. (Hillel, 1998), which is favorable for the flushing of

ionogenic compounds. Nevertheless, further study is

needed, mainly on the two-dimensional flow patterns

that are expected to develop when irrigating with water

on a highly saline soil such as loess.

Acknowledgements

We thank Zoe Grabinar for editorial assistance and

Dr. Helen Graber for constructive comments.

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