Electrokinetic Enhancement of Phenanthrene Bio Degradation in Creosote-polluted Clay Soil

8/14/2019 Electrokinetic Enhancement of Phenanthrene Bio Degradation in Creosote-polluted Clay Soil

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Electrokinetic enhancement of phenanthrene biodegradationin creosote-polluted clay soil

Jose-Luis Niqui-Arroyo, Marisa Bueno-Montes, Rosa Posada-Baquero,Jose-Julio Ortega-Calvo*

Instituto de Recursos Naturales y Agrobiologa, C.S.I.C., Apartado 1052, E-41080-Seville, Spain

Received 5 May 2005; received in revised form 29 September 2005; accepted 2 October 2005

Electrokinetic bioremediation is a potentially effective technology to treat PAH-polluted, clay-rich soils.

Abstract

Given the difficulties caused by low-permeable soils in bioremediation, a new electrokinetic technology is proposed, based on laboratory

results with phenanthrene, to afford bioremediation of polycyclic aromatic hydrocarbons (PAH) in clay soils. Microbial activity in a clay

soil historically polluted with creosote was promoted using a specially designed electrokinetic cell with a permanent anode-to-cathode flow

and controlled pH. The rates of phenanthrene losses during treatment were tenfold higher in soil treated with an electric field than in the control

cells without current or microbial activity. Results from experiments with Tenax-assisted desorption and mineralization of 14C-labeled phenan-

threne indicated that phenanthrene biodegradation was limited by mass-transfer of the chemical. We suggest that the enhancement effect of the

applied electric field on phenanthrene biodegradation resulted from mobilization of the PAH and nutrients dissolved in the soil fluids.

2005 Elsevier Ltd. All rights reserved.

Keywords: Polycyclic aromatic hydrocarbons; Bioremediation; Electrokinetics; Electro-osmosis; Desorption

1. Introduction

Creosote is a complex mixture of organic chemicals, mainly

polycyclic aromatic hydrocarbons (PAH). Used worldwide as

a wood preservative, its accidental spillage and improper use

and handling at processing sites has led to the contamination

of underlying soils and groundwater. Bioremediation of creo-

sote-polluted sites is considered a realistic alternative to other

remediation methods, as it has the advantages of relatively lowcost and reasonable execution periods (Mueller et al., 1989).

A major factor limiting the success of bioremediation of

PAH is the presence in soil of a high proportion of clay-sized

particles. From the experience already gained in bioremedia-

tion technology, a high clay content in a contaminated soil pla-

ces serious doubts on the final success of bioremediation alone

as a treatment strategy. Clay-rich soils may present a limited

bioavailability of PAH, because of their high surface area

available for sorption (Lahlou and Ortega-Calvo, 1999), and

a difficulty for bacterial transport through the soil, what limits

the access to the source of hydrophobic substrates (Lahlou

et al., 2000). The oxygen and nutrient supply to the degrading

populations may also be limiting, due to slow diffusion and

low hydraulic conductivity. Associated operations such as han-

dling, excavation, and nutrient amendments may be physicallyhampered because of the high consistency of clay-rich soils,

which are also prone to bacterial clogging (Kaufman, 1994).

Successful attempts at bioremediation in creosote-polluted

sites have been documented only with sandy soils (Breedveld

and Karlsen, 2000; Breedveld and Sparrevik, 2000; Carriere

and Mesania, 1995; Eriksson et al., 2000).

Electroremediation, consisting of the controlled application

of low-power DC electric fields to polluted soils, is especially

indicated for clay soils. This technology, which relies on

three processesdelectromigration (movement of charges),* Corresponding author. Tel.: 34 95 462 4711; fax: 34 95 462 4002.

E-mail address: [email protected] (J.-J. Ortega-Calvo).

0269-7491/$ - see front matter 2005 Elsevier Ltd. All rights reserved.

doi:10.1016/j.envpol.2005.10.007

Environmental Pollution 142 (2006) 326e332www.elsevier.com/locate/envpol
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electro-osmosis (water), and electrophoresis (charged parti-

cles)dhas already been used to remove heavy metals and

organic pollutants from soil (Virkutyte et al., 2002).

Its mobilizing potential can also be coupled to biodegradation

processes, as has been shown in laboratory studies where

(i) bacterial strains have been transported electrokinetically

through diesel-contaminated soil (Lee and Lee, 2001) andmodel aquifer material (DeFlaun and Condee, 1997; Wick

et al., 2004), (ii) the organic pollutant itself was transported

towards soil zones harboring microbial populations able to

degrade the pollutant, such as 2,4-dichlorophenoxyacetic

acid (Jackman et al., 2001) or p-nitrophenol (Ho et al.,

1995), and (iii) a co-metabolic substrate was injected into

soil to promote TCE biodegradation (Rabbi et al., 2000). To

our knowledge, there are no studies about the effect of this

technology on the biodegradation of PAH, possibly due to the

limited transport that these chemicals often exhibit in clay

soil during electrokinetic treatment (Saichek and Reddy, 2003).

This research focuses on the effect of an electric field on the

biodegradation of PAH present in a clay-rich, creosote-pollutedsoil. We employed a historically polluted soil containing

a high load of native PAH, of which phenanthrene was fol-

lowed as a target compound. Compound disappearance was

measured in electrokinetic cells designed to promote microbial

activity, and the kinetics of phenanthrene desorption and bio-

degradation were determined in solid phase and soil slurries.

Our main objectives were (i) to determine if low-voltage DC

currents promote phenanthrene biodegradation in clay-rich

soil, and, if so, (ii) to determine the possible mechanism(s)

involved.

2. Materials and methods

2.1. Soil

The soil used in this study was a clay soil, classified as a calcaric fluvisol,

provided by EMGRISA (Madrid, Spain) from a wood-treating facility in

Andujar (Jaen, southern Spain), with a record of pollution by creosote exceed-

ing 100 years. The site geology consists of a superficial granular fill of up to

1 m in thickness, a horizon of silts and clays (2e3 m), a water-bearing horizon

of sandy gravel, and compact marls. Groundwater is located at a depth of ap-

proximately 4 m. After sampling different locations, which served as prelim-

inary tests, a homogenous sample (50 L) of the silty clay layer from a heavily

contaminated point was prepared by air drying for 2 weeks, thorough mixing,

and sieving (2 mm mesh). The resulting soil sample was characterized accord-

ing to standard methods of soil analysis (Klute, 1986; Page et al., 1982), and

had the following characteristics: 6.6% moisture; pH 7.92; 23.4% CaCO3;

3.26% organic matter; 0.106% organic nitrogen (Kjeldahl); 0.9 mg kg1 avail-

able phosphorus; particle size distribution: 1.1% coarse-grained sand, 2.4%

fine-grained sand, 37.0% silt, and 60.0% clay; 2777 mg kg1 total petroleum

hydrocarbons; 4501 mg kg1 total PAH (sum of 16 EPA PAH). Phenanthrene

content was 1319.5 78.5 mg kg1 dry soil. The number of indigenous mi-croorganisms able to grow with phenanthrene, estimated as colony-forming

units on solid medium containing the chemical as the sole source of carbon

and energy (Vila et al., 2001), was 2 104 cells g dry soil1.

2.2. Electrokinetic treatment

The cell designed, shown in Fig. 1, consisted of a polyethylene body

protected with an inner layer of glass, to prevent phenanthrene loss due to

sorption. Two cylindrical electrodes, made of stainless steel, were inserted

in the soil inside cylindrical glass filtercandles (Robuglasfilter-Gerate

GmbH, Hattert, Germany), which served as electrode reservoirs. The amount

of dry soil packed in the cell was approximately 700 g. The filtercandles had

porous walls (160e250 mm pore size), allowing the exchange of electrode

solution into and out of the reservoirs. The separation distance between the

electrodes was 16 cm. To promote maximum microbial activity in the soil

by a permanent anode-to-cathode flow and controlled pH, these reservoirs

were kept filled with a buffer solution, which was recirculated independently.

The reservoirs were connected to a peristaltic pump, and 1 M phosphate buffer

(K2HPO4/KH2PO4) adjusted to pH 8.00 (anode) or 5.80 (cathode) was recircu-

lated at a constant flow rate of 12 mL min1. This buffer concentration was

chosen not only because it allowed the efficient control of pH in the electrode

reservoirs, but also to minimize possible changes induced in the pore fluid

properties due to the P consumption associated to microbial assimilation,

which is typical of bioremediation of hydrocarbons (Huesemann, 1994). In ad-

dition, the soil was packed in the cell in layers after saturating with inorganic

salts solution, which contained KH2PO4 (0.9 g L1), K2HPO4 (0.1 g L

1),

NH4NO3 (0.1 g L1), MgSO4 $ 7H2O (0.1 g L

1), CaCl2 (0.080 g L1),

FeCl3 $ 6H2O (0.01 g L1), and 1 mL L1 of a microelements stock to obtain

the final concentrations of 0.0014 g L1 for Na2MoO4 $ 2H2O and 0.002 g L1

for each of the following: Na2B4O7 $ 10H2O, ZnSO4 $ H2O, MnSO4 $ H2O

and CuSO4 $ 5H2O. This solution presented a pH value of 5.8. The porosity

of the soil packed in this way ranged from 0.158 to 0.188 (determined gravi-

metrically). During packing, the soil was inoculated, along the line between

the electrodes, with the bacterium Novosphingobium sp. LH128, which was ca-

pable of using phenanthrene as a sole carbon and energy source for growth.

The bacterium was cultured and prepared for the experiments as previously de-

scribed (Garcia-Junco et al., 2003), and was added at a cell density of 2.6 107 cells g1. The test cell was treated for a total of 337 h (14 days), during

which the voltage applied was 0.5e0.6 V cm1 (77 h) and 0.2e0.3 V cm1

(260 h), in alternate periods. The DC power supply used was a Freak

HY3005D-3 model unit. A control cell was maintained under exactly the

same conditions (including saturation with water, inoculation, and recircula-

tion of electrode fluids), but without an electric field. An abiotic control was

also run with soil autoclaved three times, which received no inoculum, and

was treated for 135 h at 0.7e0.8 V cm1. After treatment, soil cores (25 g

dry soil) were taken from the cell along the anode-cathode axis and analyzed

in duplicate for residual phenanthrene content. Statistical comparisons wereperformed with analysis of variance and Scheffe post hoc test at p 0.10and p 0.05.

2.3. Electro-osmotic flow

Percolated columns were used in a similar way to that one previously de-

scribed for the study of bacterial transport through clay-rich porous media

(Lahlou et al., 2000; Ortega-Calvo et al., 1999). The soil was wet-packed in

glass columns with an inside diameter of 0.9 cm (cross-sectional

area 0.125 cm2), and a length of 10 cm. A portion of soil (1 cm) next tothe cathode was spiked during packing with 27,000 dpm (266.3 ng) of

[14C]phenanthrene (8.3 mCi mmol1, radiochemical purity >98%, SigmaChemical Co., St. Louis, MO) completely dissolved in 0.5 ml of inorganic

salts solution. Hollow, stainless steel cylindrical electrodes were connected

to each side of the columns. A reservoir with inorganic salts solution was con-

nected to the anode side to keep the soil saturation conditions. The formation

of an hydraulic gradient was prevented by height adjustment of the reservoir.

A constant electric field of 1 V cm1 was applied for 5 h, and the column

effluent passing through the cathode was collected and weighted. Then, it

was mixed with 5 mL of liquid scintillation cocktail (Ready Safe, Beckman

Instruments, Fullerton, CA, USA), and radioactivity was measured with a

liquid scintillation counter (Beckman Instruments Inc., Fullerton, Calif.;

model LS5000TD).

2.4. Desorption

Phenanthrene desorption kinetics were determined in duplicate with the

Tenax solid-phase extraction method (Cornelissen et al., 1998). Briefly, 1 g

dry soil, 70 mL milli-Q water, 0.35 mL formaldehyde (40%), and 1.5 g Tenax

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TA beads were placed in a separation funnel. The funnel was continuously

shaken at 23 2 C on an orbital shaker operating at 170 rpm. After certaintime intervals, the Tenax was separated from the soil suspension, and replaced

by fresh Tenax. The sorbent was extracted by shaking with 50 mL of

hexane for 48 h. The extract was evaporated to near dryness, redissolved in

acetonitrile, and filtered. Analysis of phenanthrene was performed by HPLC

as described later. The total mass of phenanthrene desorbed (extractedby Tenax) plus the amount still present in the soil at the end of the

experiments was 97.3 1.7% of the initial mass determined by whole-soilextractions.

2.5. Biodegradation

Biodegradation was measured in solid phase and in slurries. Solid-phase

experiments were performed in closed biometer flasks (Bellco glass, NJ)

with 25 g of dry soil adjusted to 80% of field capacity with sterile, distilled

water. A portion of the water used to adjust humidity was an aqueous solution

of approximately 100,000 dpm of [9-14C]phenanthrene (8.3 mCi mmol1,

radiochemical purity >98%, Sigma Chemical Co., St. Louis, MO). This guar-anteed the homogenization of the labeled compound with native phenanthrene.

The flasks were closed with Teflon-lined stoppers, and incubated at 23 2 C.

Production of14CO2 was measured as radioactivity appearing in the alkali trap

of the biometer flasks. The trap contained 1 mL of 0.5 M NaOH. Periodically,

the solution was removed from the trap and replaced with fresh alkali. The

NaOH solution was mixed with 5 mL of liquid scintillation cocktail, and the

mixture kept in darkness for about 8 h for dissipation of chemiluminescence.

Radioactivity was measured with a liquid scintillation counter. Residual con-

tents of native phenanthrene were determined in separate flasks that were in-

cubated under the same conditions but contained no 14C-labeled compound.After certain time intervals, duplicate flasks were sacrificed and kept frozen

at 80 C until analysis for phenanthrene content by HPLC.For experiments with soil slurries, 15 g of soil was placed in 250-mL

Erlenmeyer flasks, 1 mL of distilled water containing 100,000 dpm of radiola-

beled phenanthrene was added to the soil, and the mixture was homogenized.

A sterile, inorganic salts solution (pH 5.8), described above, was added to

complete a final volume of 100 mL. The slurries were then inoculated with

Novosphingobium sp. LH128, which was cultured and prepared for minerali-

zation experiments as previously described (Garcia-Junco et al., 2003). Each

flask received an inoculum of approximately 107 cells g1. The flasks were

then closed with Teflon-lined stoppers, from which an 8-mL vial containing

1 mL of 0.5 M NaOH was suspended to trap 14CO2. The flasks were incubated

at 23 2 C on an orbital shaker operating at 100 rpm. Measurements ofmineralization of the radiolabeled phenanthrene and residual concentration

of the native compound were carried out as above.

B

Filtercandle

Peristaltic pumpA

Phosphate

buffer pH 8.00

Phosphate

buffer pH 5.80

Contaminated clay soil

Glass body cell

Water flow

Stainless steel

cathode

Stainless steel

anode

Magnetic stirrer

Tygon tubes

+ -

Fig. 1. Design of electroremediation cell used to treat creosote-polluted clay soil. (A) Schematic diagram. (B) Detail.

328 J.-L. Niqui-Arroyo et al. / Environmental Pollution 142 (2006) 326e332


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2.6. Data analysis

Desorption and mineralization/biodegradation data were analyzed in two

ways. On the one hand, mineralization and biodegradation rates were calculated

and compared as the slope of the regression lines drawn with the points

belonging to the phase of maximum mineralization or disappearance in experi-

ments with radiolabeled and native phenanthrene, respectively (Ortega-Calvo

et al., 1995). On the other hand, data of desorption and biodegradation ofnative phenanthrene could be empirically described by the following two-

compartment kinetic model:

St=So Frap expKrapt

Fslow expKslowt 1

where St and So (g) are the soil-sorbed amounts of phenanthrene at time t(h)and

at the start of the experiment, respectively (Cornelissen et al., 1998), Frap and

Fslow are the rapidly and slowly desorbing (or biodegrading) fractions, Krapand Kslow (h

1) are the rate constants of rapid and slow desorption/biodegrada-

tion. Frap, Fslow, Krap and Kslow were obtained by minimizing the cumulative

squared residuals between experimental and calculated values of ln(St/So).

The software usedfor the minimization was MicrosoftExcel 97 (Solver option).

It should be noted that the use of this empirical, two-compartment model with

desorption and biodegradation data does not necessarily imply a direct depen-

dence or influence of one of these two processes on the other.

2.7. Phenanthrene analysis

Soil samples were defrosted, dried at 40 C, ground with a mortar and pes-

tle, and extracted in a Soxhlet with 100 mL dichloromethane for 8 h. Then, the

organic solvent was evaporated to near dryness, and the residue was dissolved

in 15 mL acetonitrile and filtered through Whatman grade 1 filter paper. Phen-

anthrene analysis was carried out using a Waters HPLC system (2690 separa-

tions module and 474 scanning fluorescence detector. Column: Nova-Pak C18Waters, 3.9 150 mm; flow: 1 mL min1; mobile phase: 60% acetonitrile40% water). Analysis of phenanthrene content of a reference material (diesel

particulate matter 2975, National Institute of Standards and Technology, Gai-

thersburg, USA) was in good agreement with the certified value (the measured

value being 10.8 2.8 mg kg1, and the certified value 17.0 2.8 mg kg1).A control was also run for possible losses during drying. The soil sample

was wetted with distilled water up to 80% field capacity, and processed as

described above. The concentration of phenanthrene was 1111.5 89.1 mg kg1. In experiments with soil slurries, the solids were collected by

centrifugation at 7000 rpm for 10 min, and treated as described above.

3. Results

3.1. Electrokinetic bioremediation

An electroremediation cell was designed to test the effect of

electric fields on phenanthrene biodegradation in soil. Prelim-

inary tests showed that for a successful electrokinetic bioreme-

diation it was crucial to avoid the generation of extreme pH

values in the electrode reservoirs (acidic pH at the anode

and basic pH at the cathode) due to electrolytic reactions: Mi-

gration of protons and/or hydroxyl ions into soil would have

prevented biodegradation, due to unfavorable ecological con-

ditions for phenanthrene-degrading bacteria and cessation of

electro-osmotic flow. This was achieved by continuous renewal

of electrode reservoir solutions, maintaining their pH at a con-

stant value. Other relevant operating aspects were (1) the

progressive increase in current intensity observed during the

treatment, which accompanied the application of a constant

potential (0.8 V cm1), and was possibly caused by a higher

conductivity of the soil, and (2) an increased temperature of

the treated soil (35 C, consistently 5e

9 C above the

temperature in the control cell without current). Voltage mea-

surements along the anode-cathode axis at points located in

the same sections of the cells showed a linear relationship to

the distance from the electrodes, and confirmed the monodi-

mensional distribution of voltage within the cell.

Fig. 2 shows the results obtained in a representative exper-

iment. Phenanthrene losses after treatment were significantlyhigher in the electroremediated soil than in the control cells

without current or microbial activity. The enhancement effect

of the electric current on biodegradation was observed in the

middle of the cell and close to the cathode, where a significant

reduction (p 0.10) of phenanthrene concentration was ob-served after treatment, whereas little or no compound was

lost without current. Assuming that biodegradation occurred

linearly during that period, the rate of compound disappear-

ance in this region was 1.320 0.450 mg kg1 h1 withcurrent, tenfold higher than the rate without current (0.140 0.002 mg kg1 h1). No significant compound dissipation

was detected in the abiotic control. There was a slight com-

pound depletion observed close to the anode, but it was notstatistically significant (p 0.10 and p 0.05). This showsthat the electric current did not cause any chemical degrada-

tion of phenanthrene.

We considered two hypotheses that could explain the

observed electrokinetic stimulation: (i) Direct effect of an

increased soil temperature observed during electrokinetic

treatment; (ii) a positive influence of the electro-osmotic

flow through soil caused by the current.

3.2. Effect of temperature

The possible contribution to the observed electrokinetic stim-ulation by heating of the soil was examined in a experiment in

Distance to anode (cm)

2 10 12 14

[phen](mg/Kg)

600

800

1000

1200

1400

1600

+ FIELD

- FIELD

ABIOTIC

4 6 8

Fig. 2. Effect of electric field on phenanthrene biodegradation in creosote-

polluted clay soil. Profiles of phenanthrene concentrations in soil treated in

electrokinetic cells under current application (FIELD), as compared witha control without current (FIELD), and sterilized soil (ABIOTIC). Initialphenanthrene content is indicated by the dotted line. Error bars correspond

to one standard deviation of duplicate measurements.



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which [14C]phenanthrene mineralization was compared in du-

plicate samples of soil maintained at 25 and 35 C. The soil

(25 g) was inoculated with 8 107 cells of Novosphingobiumsp. LH128. Mineralization of the labeled compound was deter-

mined as described in Section 2 for solid-phase experiments.

The results showed that the increase in temperature observed

during electrokinetic treatment of the soil had little influenceon biodegradation (Fig. 3). At 25 C, the maximum rate of

mineralization was 1.03 0.03 mg kg1 h1, while at 35 Cthe rate increased only to 1.11 0.11 mg kg1 h1.

3.3. Electro-osmotic mobilization of phenanthrene

The electro-osmotic flow of water through soil is a well-

documented process (see Section 1). Indeed, during electro-

kinetic treatments we could observe macroscopically the

movement of water in the anode-to-cathode direction. However,

the design of the cell did not allow measurements of the

mobilized volume of water. Therefore, to test whether the ap-

plied electric field induced the electro-osmotic flow of water

through this soil, and the mobilization of the phenanthrene dis-

solved in the soil fluid, a different experimental setup was used.

The system was a column-type cell in which the electric field

was applied between two hollow, cylindrical electrodes. The

portion of the soil close to the cathode was spiked with radiola-

beled phenanthrene. In this way, we could estimate an electro-

osmotic flow toward the cathode of approximately 0.03 ml/h.

The amount of [14C]phenanthrene eluted after 5 h was 1.1 ng.

3.4. Phenanthrene desorption and biodegradation

A series of desorption and biodegradation experiments wasperformed in the absence of an electric field to understand pos-

sible effects on bioavailability of phenanthrene by the observed

electro-osmotic flow. The kinetics of desorption, determined by

Tenax extraction for 41 days in sterilized soil suspensions,

showed the existence of two first-order kinetic fractions

(Fig. 4). The size of the rapidly desorbing fraction calculated

by the model accounted for 96.1 0.1% of the initial amountof phenanthrene present in thesoil.The rate constant forthe rapid

fraction (Krap) was 0.38 0.00 h1, and several orders of

magnitude larger than the corresponding Kslow (2.70 0.03 103 h1). The dissipation of the native compound ob-

served in biodegradation experiments, performed in solid- andslurry-phase conditions, also followed first-order kinetics

(Fig. 5). The final extents of compound disappearance were

very close to the size of the rapidly desorbing fraction deter-

mined in Tenax experiments. Furthermore, Fig. 5 shows good

agreement between phenanthrene concentrations determined

experimentally in soil during biodegradation and those pre-

dicted by the first-order model of Eq. (1). The calculated values

forKrap were 4.05 0.21 103 and 45.95 5.80 103 h1

in solid- and slurry-phase biodegradation experiments, respec-

tively, while the corresponding Kslow were 1.74 2.46 105

and 1.35 0.21 103 h1. The calculated values for Frap insolid- and slurry-phase experiments were, respectively,

96.10 2.82% and 95.40 0.59%.

4. Discussion

The enhancement effect of an applied electric field on phen-

anthrene biodegradation wasvery likely caused by the mobiliza-

tion of the soil fluids associated to electro-osmosis. This is

supported by (i) the results from mineralization experiments,

showing that the enhancement of biodegradation by the electric

field was not due to an increased temperature, (ii) the observed

mobilization of water and associated [14C]phenanthrene due to

electro-osmotic flow, and (iii) the occurrence of a significant

amount (more than 95%) of phenanthrene as a fast-desorbingfraction, as revealed by Tenax-assisted desorption experiments.

The electrokinetic mobilization of phenanthrene in soils

has been observed in several studies that used solubility-

enhancing agents (surfactants, co-solvents, and cyclodextrins)

in the flushing solutions. These studies have shown that, in

spite of the limited solubility of the compound (1.1 mg L1

(Schwarzenbach et al., 2003)), which restricts its transport

Time (days)

0 2 8 10 12 14 16 18

%1

4Cmineraliz

ed

0

2

4

6

8

10

12

14

16

4 6

Fig. 3. Mineralization of phenanthrene in inoculated, creosote-polluted clay

soil maintained at 25 C (circles) and 35 C (triangles).

TIME (h)

0 200 400 600 800 1000

LnSt/S0

-6

-5

-4

-3

-2

-1

0

Fig. 4. Kinetics of phenanthrene desorption from creosote-polluted clay soil,

as measured by Tenax extraction. The dashed line represents the curve fitting

the two-phase desorption model in Eq. (1).

330 J.-L. Niqui-Arroyo et al. / Environmental Pollution 142 (2006) 326e332
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through the soil, phenanthrene can be mobilized even in the

absence of these agents. For example, Saichek and Reddy

(2003) observed the nearly complete removal of 500 mg kg1

phenanthrene in the region adjacent to the anode after the elec-

trokinetic treatment of kaolin soil for 60 days with only deion-

ized water. In the rest of the soil profile, however, the

phenanthrene concentration remained unaltered. A significant

removal of compound (up to 35%) was also reported in another

study where kaolinite samples were initially amended with

1.9 mg kg1 phenanthrene and treated electrokinetically for

6 days with a NaCl purging solution (Ko et al., 2000). In these

studies, performed with column-type reactors, the uneven dis-

tribution of phenanthrene through the soil after electrokinetic

treatment evidences that the electro-osmotic flow may

generally not be uniform. Some sections present variable

flow rates and pore pressures, probably as a result of the com-

plex nature of the factors that control the electro-osmotic flow,

including zeta potential and electrical gradient. This may ap-

ply to our study, and explain the differences in phenanthrene

biodegradation along the anode-cathode axis.

Although we observed mobilization of [14C]phenanthrene

induced by electro-osmotic flow, the electrokinetic treatment

in the absence of biological activity (Fig. 2) did not cause signif-

icant changes in the phenanthrene concentration profile.

However, the first-order kinetics of biodegradation observed

in separate biodegradation experiments are indicative of con-

centration-dependent biodegradation. It is therefore possible

that the electrokinetic treatment caused a mobilization of

0 25 50 75 100 1 25 1 50 1 75 2 00 2 25 250

0

10

20

30

40

50

B

TIME(h)

TIME(h)

0 500 1000 1500 2000 2500 3000 3500 4000 4500

%1

4Cmineraliz

ed

%1

4Cmineralized

0

10

20

30

40

50

60

mg/Kgsoil

0

200

400

600

800

1000

1200

1400

1600

1800

mg/Kgsoil

0

200

400

600

800

1000

1200

1400

1600

1800

A

Fig. 5. Mineralization of [14C]phenanthrene (circles) and evolution of concentration of native phenanthrene (triangles) by indigenous microbial population in

creosote-polluted clay soil (A) and inoculated soil slurries (B). The dashed line represents the curve fitting the two-phase desorption/biodegradation model in

Eq. (1). Error bars correspond to one standard deviation of measurements in duplicate flasks.



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Electrokinetic Enhancement of Phenanthrene Bio Degradation in Creosote-polluted Clay Soil