remediation of saturated soil contaminated with petroleum … · remediation of saturated soil...

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Journal of Environmental Management 83 (2007) 339–350

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Remediation of saturated soil contaminated with petroleumproducts using air sparging with thermal enhancement

A.M.I. Mohamed�, Nabil El-menshawy, Amany M. Saif

Mechanical Power Department, Faculty of Engineering, Suez Canal University, Egypt

Received 18 April 2005; received in revised form 3 April 2006; accepted 4 April 2006

Available online 17 July 2006

Abstract

Pollutants in the form of non-aqueous phase liquids (NAPLs), such as petroleum products, pose a serious threat to the soil and

groundwater. A mathematical model was derived to study the unsteady pollutant concentrations through water saturated contaminated

soil under air sparging conditions for different NAPLs and soil properties. The comparison between the numerical model results and the

published experimental results showed acceptable agreement. Furthermore, an experimental study was conducted to remove NAPLs

from the contaminated soil using the sparging air technique, considering the sparging air velocity, air temperature, soil grain size and

different contaminant properties. This study showed that sparging air at ambient temperature through the contaminated soil can remove

NAPLs, however, employing hot air sparging can provide higher contaminant removal efficiency, by about 9%. An empirical correlation

for the volatilization mass transfer coefficient was developed from the experimental results. The dimensionless numbers used were

Sherwood number (Sh), Peclet number (Pe), Schmidt number (Sc) and several physical-chemical properties of VOCs and porous media.

Finally, the estimated volatilization mass transfer coefficient was used for calculation of the influence of heated sparging air on the

spreading of the NAPL plume through the contaminated soil.

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Keywords: NAPLs; Air sparging; Volatilization mass transfer coefficient; Soil

1. Introduction

Non-aqueous phase liquids (NAPLs) pose a significantthreat to ground water resources. When NAPLs infiltrateto the subsurface they descend as an immiscible phase. Incases where of the spilled quantity exceeds the retentioncapacity of the ground water saturated zone; the NAPLswill reach the capillary fringe. A NAPL less dense thanwater (LNAPL) will form pools at the water table, while amore dense one (DNAPL) will move through the saturatedzone, spreading along the less permeable layers and leavingbehind a portion of its volume as immobilized pockets ofliquid called residual saturation (Okeson et al., 1997).

In the case of both DNAPL and LNAPL, pumping toremove free product within a highly NAPL saturated lenscannot completely recover the NAPL. With the rising ofthe water table or DNAPL lens migration, NAPL will

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ess: [email protected] (A.M.I. Mohamed).

become trapped during pumping as a discontinuousresidual. Entrapped NAPLs act as long-term sources ofgroundwater contamination, (Fisher et al., 1999; Spragueand Delahaye, 1996; Held and Celia, 2001).The present research work aims at the enhancement of

the air sparging remedial technology. Air sparging is a cost-effective, time-efficient system for the remediation ofvolatile and/or biodegradable contaminants. This techni-que involves introducing forced air into the saturated zoneof an aquifer to encourage volatilization of contaminantsinto the unsaturated zone where the contaminants can thenbe removed with another complementary technology suchas soil vapor extraction (SVE), bioventing, horizontalwells, or heating (Sprague and Delahaye, 1996; Reddyet al., 1999).The airflow behavior induced by air sparging is typically

characterized by a conical air plume, often known as theradius of influence (Johnson, 1998; Mohtar et al., 1996). Jiet al. (1993) visualized the steady state air distributionpatterns using a thin Plexiglas tank with uniform lighting

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Nomenclature

A surface area, m2

a specific interfacial area, m2/m3

C concentration, mg/cm3

Cs aqueous saturation concentration, mg/cm3

D molecular diffusion coefficient, cm2/sdp particle diameter of the soil, mmd0 normalized mean particle sized1 pipe diameter, ‘mmH dimensionless Henry’s constanth Henry’s constant, atm.m3/molK permeability, mm2

kL liquid phase volatilization mass transfer coeffi-cient, cm/min

kG gas phase volatilization mass transfer coeffi-cient, cm/min

ms initial mass of VOC injected into the employedsoil

_m rate of mass transferred, mg/minPe Peclet numberQ aeration rate, L/minSc Schmidt numberSh Sherwood numberT temperature, Kt time, sUC uniformity coefficient of the porous mediau air velocity, cm/sVa air phase volume, cm3

x distance in x-dir., cmz depth in z-dir., cmX soil reactor width, cmZ soil reactor depth, cm

Greek letters

Zrem The contaminant removal efficiencyl weight factor for the finite difference techniquet tortuosity factor of the porous mediae porosity of the porous median kinematic viscosity, cm2/sym normalized mean temperatureo uncertainty of measured or estimated values

Subscripts

a air phasediff diffusiondiss dissolutioni interfacialin inletG gas phaseL liquid phaseNAPL non-aqueous phase liquidout outletref referencew water phase

A.M.I. Mohamed et al. / Journal of Environmental Management 83 (2007) 339–350340

behind the tank and glass beads as porous media. Theyobserved two patterns of airflow in soils. The first patternwas observed for medium to coarse grained media wherethe airflow is characterized by a plume of discrete airbubbles. The second pattern describes the airflow regimefor coarse to fine grained media that resemble the texturesof sand, silts and clays of natural aquifer material. Thisflow regime was dominated by channel flow where airplumes are formed from discrete and continuous airchannels. As the injection rate increases air channelsincrease in number and grow into a condensed continuouscone-shape cavity.

Braida and Ong (2000) conducted experiments using asingle air channel setup to study the influence of porousmedia properties and air velocity on the fate of NAPLs underair sparging conditions. Their study showed that the presenceof advective airflow in air channels controlled the spreadingof the dissolved phase but the overall removal efficiency wasindependent of the air flowrate. In addition, they noted thatthe removal efficiencies and dissolution rates of the NAPLwere strongly affected by the mean particle size of the porousmedia during air sparging. This agrees with the investigationsmade by Reddy and Adams (1998). They performed a seriesof one dimensional column experiments to study the effects ofsoil type, air injection mode, and the synergistic effects of co-

contaminants on air sparging removal efficiency. They foundthat there is a threshold value for the effective particle size(dp10), which is equal to 0.2mm; above this threshold value,the rate of removal is linearly proportional to the (dp10) value;while below this value, there is a drastic increase in the timerequired for contaminant removal. Additionally, they foundthat the pulsed air injection mode has no advantage overcontinuous injection for coarse sand; however, pulsed airinjection led to substantial reductions in system operatingtime for fine sand. Also they observed a slight increase inremoval rate when benzene and toluene coexisted in the testsoil compared to when they existed alone.Chao et al. (1998) developed non-equilibrium water-to-

air mass transfer experiments for six volatile organiccompounds during air sparging in soil columns packedwith coarse, medium, or fine sand or glass beads. Theyperformed a numerical study and assumed that theconcentration in the bulk phase remained constant due toslow diffusion of VOCs in the aqueous phase to theair–water interface as compared to the rapid volatilizationof VOCs at the air–water interface. Therefore, theymodeled the interface mass transfer alone.At an in-situ air sparging remediation site contaminated

through the spillage of a LNAPL waste, the influence ofthe system design parameters in terms of contaminant

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Fig. 1. Schematic diagram of the approach model.

A.M.I. Mohamed et al. / Journal of Environmental Management 83 (2007) 339–350 341

removal time was reported by Benner et al. (2000). Theysuggested that only the type of sparging operation (i.e.pulsed or continuous) was significant in terms of totalcontaminant removal time, while both the spargingoperation and air injection rate were significant in termsof removal of critical xylene species.

However, the contaminants within the radius of influ-ence will not be removed with equal efficiencies due to thenatural inherited heterogeneity of the porous medium thatyields non-uniform and asymmetrical air plumes (Ahlfeldet al., 1994). Rahbeh and Mohtar (2001) studied theinfluence of porous media heterogeneity and air channeli-zation on contaminant removal by air sparging. Theirresults showed that the contaminant removal is propor-tional to channel spacing, and controlled by the process ofdiffusion between air channels. On the other hand, thecontaminant removal is inversely proportional to thespatial variability of the flow field.

Many investigators such as Hussein (1997, 1999),Chrysikopoulos and Kim (2000), Chrysikopoulos et al.(2002), Ghoshal and Luthy (1998), and Illangasekare et al.(2000) have demonstrated the complex flow and transportbehavior of non-aqueous phase liquids in porous media.

Kueper and Frind (1989) developed a two-dimensionalvertical section finite difference model to study thesimultaneous movement of a dense, non-aqueous phaseliquid and water under the water table in heterogeneousporous media. The model was validated against a parallelplate laboratory experiment involving the infiltration oftetrachloroethylene into a heterogeneous sand pack initi-ally saturated with water.

Imhoff et al. (1995) investigated experimentally the hotwater flooding effects on the remediation of porous mediacontaminated with tetrachloroethylene at residual satura-tion. Interfacial tension measurements indicated that therewas no change in the tetrachloroethylene -aqueous inter-facial tension over the temperature range examined in thisstudy (10–40 1C). On the other hand, flushing with hotwater increased the mass transfer rate coefficient byapproximately a factor of two as the aqueous phasetemperature was increased from 5 to 40 1C. Hot waterflushing was suggested to be used before using air spargingremediation technology.

The main objective of this work is to investigatenumerically as well as experimentally the influence of soilgradation, aeration rate and sparging air temperature onthe removal efficiency of NAPLs, such as petroleumproducts, from contaminated soil. Further, the effect ofsparging air temperature on the fate of NAPLs is alsoinvestigated numerically.

2. Mathematical formulation and numerical model

2.1. Mathematical formulation

As mentioned by Ji et al. (1993), the steady state airdistribution pattern for fine to coarse media is dominated

by channel flow where air plumes are formed from discreteand continuous air channels. The principle mechanism ofmass transfer behind in-situ air sparging technology is thevolatilization of dissolved VOCs and NAPLs into the airphase. The single-air channel approach has been success-fully used for providing information on contaminantremoval occurring at the air-channel level during air-sparging (Braida and Ong, 1998, 2000). Since a constantair–water interfacial area is available through this ap-proach, it is convenient for studying the influence of soilgradation, aeration rate and sparging air temperature onthe removal efficiency of NAPLs at the level of air-channelsformed during in-situ sparging of clean air into thecontaminated soil. The conceptual experimental set-upused in the current study is a two-dimensional dissolution-diffusion-volatilization model shown schematically inFig. 1. The model is designed for a thin air channel lo-cated above the water-saturated soil. The one componentNAPL (source point) is injected into the reactor at thepoint (x0, z0).In the case of the NAPL–water interface it reaches

equilibrium. Several mathematical models describing thedissolution of residual NAPLs in porous media employ theassumption that the dissolved concentration along theNAPL–water interface is equal to the solubility or aqueoussaturation concentration Cs. The results from manyprevious experimental studies associated with residualNAPL dissolution support the applicability of thisassumption (Chrysikopoulos and Kim, 2000). Then, thedissolved NAPL disperses through the saturated soil bymolecular diffusion. When the dissolved NAPL moleculesreach the plane of the air–water interface (at z ¼ 0), thevolatilization of the dissolved contaminant occurs.Batch tests determined that the adsorption of organic

contaminants onto the soil solids was negligible (Semer andReddy, 1998). Further, due to the low organic carboncontent, sorption of benzene is not considered in theformulation of the model. Mixing effects as a result oftransfer of momentum from the air phase to the aqueousphase are negligible. Assuming stagnant conditions for theaqueous phase. The dissolved concentration along theNAPL–water interface is equal to the solubility or aqueoussaturation concentration (Cs).

ARTICLE IN PRESSA.M.I. Mohamed et al. / Journal of Environmental Management 83 (2007) 339–350342

Considering the two-dimensional diffusion–dissolutionequation (Braida and Ong, 2000) for any interior node inthe reactor we obtain:

�qCw

qt¼ tD

q2Cw

qxxþ

q2Cw

qzx

� �� J kdiss

ANAPL

V elementðCw � CsÞ,

(1)

where J ¼ 1, in the elements where the NAPL is present,and 0 otherwise. Eq. (1) describes the two dimensionaldiffusion-dissolution of a contaminant through the watersaturated contaminated soil.

As shown in Fig. 1 the NAPL is injected at point (x0, z0),therefore, the injected NAPL will be assumed as a sphericalglob of radius, r.

The above equation is subject to the following boundaryconditions. Considering Eq. (1), the initial conditions are:

C0wðx0 � r; zÞ ¼ Cs

C0wðx0 þ r; zÞ ¼ Cs

9=;; ðz0 � rÞpzpðz0 þ rÞ

C0wðx; z0 � rÞ ¼ Cs

C0wðx; z0 þ rÞ ¼ Cs

9=;; ðx0 � rÞpxpðx0 þ rÞ

C0wðx; zÞ ¼ 0; elsewhere; ð2Þ

and the boundary conditions are:

qCtwð0;zÞqx¼ 0

qCtwðX ;zÞqx¼ 0

9=;t � 0; 0pzpZ̄

qCtwðx;ZÞqz¼ 0o

t � 0; 0pxpX̄

(3)

At the air–water interface, the volatilization of thedissolved NAPL (at z ¼ 0) will be as follows:

qCtwðx; 0Þ

qz¼ �

kL

tDCwðx; 0Þ �

Ca

H

� �t40; 0pxpX̄ . (4)

2.2. Numerical technique

For solving the basic governing equations of the systemintroduced above, a finite-difference procedure has beenused which is based on a method published by Zheng andBennett (1995). An explicit-implicit Crank-Niclson form ofthe finite difference technique is used (Croft and Lilley,1977). This method was found to be much more stable witha minimum round error compared with the other methods.

Table 1

Parameters used by Braida and Ong (2000)

Experiment number Soil

1 Medium sand (dp50 ¼ 0.305mm)

2 Fine sand (dp50 ¼ 0.168mm)

A regular grid and a Cartesian coordinate system wereapplied in the present study. The finite difference formula-tions were derived for each node of the finite differencemesh. This generates the matrix of the solution, which wassolved iteratively to obtain the contaminant concentrationdistribution through the contaminated soil reactor.A computer program was developed to simulate the

behavior of the time variant contaminant concentrationdistribution through the contaminated soil. The code waswritten in FORTRAN programming language. It consistsof a main program and different subprograms forcalculating the volatilization mass transfer coefficient anditerating the solution of the concentration distribution.

2.3. Model validation

The numerical model results were validated against theresults of two-dimensional sand pack experiments, whichwere conducted by Braida and Ong (2000) for the sameboundary and physical conditions. Benzene was used as aNAPL while medium silica sand of (dp50 ¼ 0.305mm,UC ¼ 1.41) and fine silica sand of (dp50 ¼ 0.168mm,UC ¼ 1.64) were used as the contaminated soil.The dissolution mass transfer coefficient, kdiss, and the

liquid phase volatilization mass transfer coefficient, kL, aretaken from the reference experimental data (Braida andOng, 2000). Table 1 contains the main operating condi-tions, which were used in the air sparging experiments forbenzene NAPL with an aeration rate of 27.5mL/min.A comparison between the predicted contaminant

concentrations and experimental data obtained fromBraida and Ong (2000) is demonstrated in Fig. 2(a–c) for24, 48, and 72 h from the beginning of benzene injection inthe medium silica sand composed of mean particle size of0.305mm under an aeration rate of 27.5mL/min at theair–water interface. The maximum discrepancy between thepresent theoretical results and the available experimentaldata from Braida and Ong (2000) was about 20%.The comparison between the present numerical model

results and the reference experimental results shows, ingeneral, acceptable agreement. The discrepancies betweenthe two comparative groups of data, the numerical modelresults and the data obtained from Braida and Ong (2000),for the two types of silica sand are expected due touncertainties in the measured quantities obtained byBraida and Ong (2000) and the simplifying assumptionsin the numerical solution in addition to the computationalerrors.

e t kL (cm/min) kdiss (cm/min)

0.37 0.51 1.16� 10�3 0.227

0.4 0.47 1.23� 10�3 0.0041

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Fig. 2. Comparison of the present model results of theoretical concentration distribution with experimental data obtained from Braida and Ong (2000)

through contaminated soil of (dp50 ¼ 0.305mm) under aeration rate of 27.5mL/min after 24, 48 and 72 h.


2.4. Numerical model results

Two-dimensional isoconcentration lines of benzene forthe numerical model runs were drawn. By assuming abenzene source with a concentration equal to the benzenesolubility, the software used a Linear Variogram KrigingProcedure and the estimated concentration from thenumerical model to draw the isoconcentration lines. Theisoconcentration lines for benzene through the contami-nated soil of mean particle size 0.305mm at an aerationrate of 27.5mL/min at 24, 48, and 72 h are presented inFig. 3(a–c).

A lateral symmetrical diffusion of benzene from theNAPL can be observed. It can be seen that theisoconcentration lines through the contaminated soil getwider with the increase in time. The 1mg/L isoconcentra-tion line of the plume extends further in both the lateraland vertical directions after 72 h than at 48 and 24 h.

3. Experimental setup

A schematic description of the experimental test rig isshown in Fig. 4. The airflow discharged from the blower iscontrolled by a by-pass system via control valves. The air isthen passed through an air filter, which is fixed inside the

main pipe. The filtered air is then passed to the electricheating system and finally to the contaminated soil which islocated in the reactor test section. A slurry of the soil andwater is packed into the soil reactor. The reactor consists ofa box made of Plexiglas, which has dimensions of 0.20mlong, 0.11m high, and 0.05m wide. The design of theexperimental set-up allows the air to pass over thesaturated soil through an air channel. A sampling pointfor the measurement of the sparged VOC vapor concen-tration in the exhaust air is included.The employed soils were washed thoroughly and dried in

an oven at 105 1C. The contaminated soils employed in theexperimental study were fine sand and medium sand withan average particle size (dp50) of 0.278 and 0.39mm,respectively. Three VOCs were used in the experimentalstudy of the volatilization mass transfer process under airsparging conditions. The employed VOCs were benzene,toluene and m-xylene.A digital unidirectional hot-wire anemometer (Testo

435) was used to measure the air velocity with an accuracyof 0.01m/s and the air temperature with an accuracy of70.5 1C. The effluent air VOC concentration was mea-sured by using a multi-gas monitor (Q-RAE PLUS—PGM-2000). The readings are displayed as a percentage ofthe LEL (lower explosive limit) or as a percentage by

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Fig. 3. Isoconcentration lines (mg/L) for benzene NAPL through contaminated soil of (dp50 ¼ 0.305mm) under aeration rate of 27.5mL/min after 24, 48

and 72 h.

Fig. 4. Schematic diagram of the test rig, 1—holder, 2—blower, 3—by-

pass valve, 4—control valve, 5—air filter, 6—electric heaters, 7—

insulation, 8—thermostats, 9—orifice meter, 10—soil reactor, 11—stand,

12—U-tube manometer.


volume when the combustible gases go beyond thelower explosive range. A volume of 1.5 cm3 of theemployed contaminant was injected into the soil reactorat approximately 20mm below the air–water interface,100mm from the air inlet, and 25mm from the front andback side walls.

A set of experiments were conducted with the single-airchannel setup for different soils (Medium and Fine),different types of VOCs (benzene, toluene, and m-xylene),different air velocities (from 1 to 6 cm/s) and different airinlet air temperatures (varied from 16 to 38 1C). During theexperiments, air samples were taken from the effluentsampling port every 5–15min and analyzed.

The contaminant removal efficiency and the volatiliza-tion mass transfer coefficient can be calculated from theexperimental data. The following equation represents afirst-order kinetic process, which models the non-equili-brium mass transfer between air and water phases (Chaoet al., 1998):

_maw ¼ kGAiðHCwi � CaÞ. (5)

The air phase contaminant concentration is described bythe following equation (Chao et al., 1998):

VadCa

dt¼ kGAiðHCwi � CaÞ � _QCa. (6)

The air temperature was normalized using a referencetemperature, Tref. Therefore, the obtained results arepresented in the form of the normalized mean temperature,ym, which is defined as

ym ¼ðTain þ TaoutÞ=2

T ref. (7)

The contaminant removal efficiency, Zrem is defined asfollows:

Zrem ¼_QR t

0 Ca dt

mc� 100ð%Þ. (8)

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4. Results and discusion

4.1. Experimental results

The contaminant mass removal is believed to becontrolled by two distinct processes of advection anddiffusion. Initially the mass of the removed contaminant iscontrolled by advection. Contaminant is removed byrelatively quick volatilization from the air–water interfaceand subsequent advection by air, until the contaminantconcentration in the air channel decreases below that in theaqueous phase. Then at that time, the diffusion process willbegin to dominate the contaminant removal. These tworegions of flow regimes are seen as initial peaks followed byasymptotic behavior in all of the obtained results for thecontaminant removal curves.

Fig. 5 shows the variation of benzene concentration inthe effluent air against the elapsed time for two types ofcontaminated soils at 6 cm/s air velocity while thetemperature of the air sparging over the contaminated soilwas maintained approximately constant at 16 1C (ym ¼ 1).The removed amount of benzene from the contaminatedsoil that had a mean particle size of 0.39mm was higherthan that of 0.278mm for approximately the first half hourof the experiment. This trend was then reversed for theremaining elapsed time of the experiment where thecontaminated soil of mean particle size of 0.278mm hadhigher benzene concentrations than that of mean particlesize 0.39mm. This indicates that contaminated soil that hasa smaller mean particle size will delay the spreading of thedissolved benzene plume to the air–water interface and willin turn result in initially less volatilization at the air–waterinterface. The total benzene mass removal efficiency formedium grained contaminated soil of mean particle size0.39mm, after the elapsed time of the experiment (5 h), wasabout 75.65%. However for the finer contaminated soil ofmean particle size 0.278mm, the mass removed was 68.6%.

Fig. 5. Effluent air benzene concentration against elapsed time for various

types of contaminated soils (6 cm/s and ym ¼ 1).

These results show that the long-term contaminant massremoval is dependent on the aqueous diffusion of thedissolved phase towards the air–water interface.Fig. 6 presents the variation of the contaminant

concentration in the effluent air against elapsed time forbenzene, toluene, and m-xylene in contaminated soil ofmean particle size 0.39mm at 6 cm/s air velocity while thetemperature of the air sparging over the contaminated soilwas maintained approximately constant at 16 1C (ym ¼ 1).Benzene has the lowest Henry’s constant but the highestsolubility and diffusion coefficients in air and water in thisgroup of VOCs. This explains the reason for the increasedbenzene concentration in the effluent air compared totoluene and m-xylene over the elapsed time of theexperiment and consequently the total contaminant re-moval efficiency. The total benzene removal efficiencyreached 75.65% while the total removal efficiencies fortoluene and m-xylene were only 67% and 52.9%,respectively. These results clearly indicate that the volati-lization mass transfer increases with the increase in the airand water diffusion coefficients of the VOCs. Furthermore,regarding the Henry’s law constant of the VOCs, aninversely proportional trend is observed.Fig. 7 demonstrates the variation of benzene concentra-

tion in the effluent air against elapsed time and the totalbenzene mass removal efficiency for 6, 2.5 and 1.5 cm/s airvelocities in contaminated soil of mean particle size0.39mm while the temperature of the air sparging wasmaintained approximately constant at 16 1C (ym ¼ 1). Itcan be seen that as the air velocity increased, the benzeneconcentrations in the effluent air decreased. However, theair flowrates for these three velocities were not the same.Therefore, in order to investigate the effect of increasing airvelocity and hence the air flowrates on the benzeneremoval, the removed mass of benzene was integrated toestimate the total benzene mass removal efficiencies. Thetotal benzene removal efficiency for an air velocity of

Fig. 6. Effluent air VOC concentration against elapsed time in con-

taminated soil of dp50 ¼ 0.39mm at 6 cm/s and ym ¼ 1.

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Fig. 7. Effluent air benzene concentration against elapsed time for various

air velocities (contaminated soil of (dp50 ¼ 0.39mm) at ym ¼ 1).Fig. 8. Effluent air benzene concentration against elapsed time for

different normalized mean temperatures in contaminated soil of

dp50 ¼ 0.39mm and 2.5 cm/s.

Fig. 9. Effluent air benzene concentration against elapsed time for

different normalized mean temperatures in contaminated soil of

dp50 ¼ 0.39mm and 6 cm/s.


1.5 cm/s (air flowrate ¼ 0.27L/min) was only 34.5% andfor an air velocity of 2.5 cm/s (air flowrate ¼ 0.46L/min) itwas 49.6% but for an air velocity of 6 cm/s (airflowrate ¼ 1L/min) it was 75.65%. These results indicatethat an increase in air flowrate produces an increase in theoverall long term mass removal and this might be related toan increase in the mass transfer zone affected by the airsparging channel.

Fig. 8 illustrates respectively the effluent air benzeneconcentration against elapsed time and the total benzeneremoval efficiency for various inlet air temperatures (16 1C(ym ¼ 1) and 24 1C (ym ¼ 1.019)). A sparging air velocity of2.5 cm/s was attained over the contaminated soil of meanparticle size 10.39mm. It can be seen that initially, as theair temperature increased the benzene air phase concentra-tion increased. This behavior is attributed to the change ofthe properties of the benzene at the air–water interfacewhere increased air temperature increases the benzene airdiffusion coefficient. However, after about 40min from thebeginning of benzene injection, the effluent air benzeneconcentration for (ym ¼ 1.019) was lower than that at(ym ¼ 1). However, the later change in the benzeneconcentration may have occurred because the rate ofvolatilization at the air–water interface is greater than therate of diffusion in the aqueous phase, and this would inturn decrease the average water concentration at theair–water interface. The total benzene removal efficiencyfor a normalized mean temperature of (ym ¼ 1) was about49.6% while for a normalized mean air temperature of(ym ¼ 1.019) it was 53.8%. It is clear that raising the inletair temperature results in an improvement in the con-taminant removal efficiency.

This observation is repeated in Fig. 9 which shows theresults of three air sparging runs at 6 cm/s air velocity incontaminated soil of mean particle size 0.39mm. Concen-trations of the sparged contaminant, benzene, in theeffluent air, as a function of time for inlet air temperatures

of approximately 16 1C (ym ¼ 1), 33 1C (ym ¼ 1.037), and38 1C (ym ¼ 1.051) are compared. It is clear that as theair temperature increased the benzene concentration inthe effluent air increased for about 50min from thebeginning of benzene injection and then decreased laterally.This may occur for the reason given for the previousgroup of curves in Fig. 8. Furthermore, the aeration ratefor this group of curves was higher than that of theprevious one. The total benzene removal efficiency for theinlet air temperature of 16 1C and normalized meantemperature of (ym ¼ 1) was about 75.65% while for theinlet temperature of 33 1C and normalized mean tempera-ture of (ym ¼ 1.037) it was 80.65% and for the inlet airtemperature of 38 1C and normalized mean temperature of(ym ¼ 1.051) it was 82.4%. Therefore, to improve thecontaminant removal efficiency, it is desirable to promote a

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Fig. 10. Comparison of experimentally determined Sherwood numbers

with predicted Sherwood numbers.


relatively high-normalized mean temperature for the airsparging over the contaminated soil.

4.2. Air–water mass transfer correlation

The gas phase mass transfer coefficient is calculated fromEq. (6). When the contaminant glob is injected close to theair–water interface, the dissolved phase concentration atthe air–water interface, Cwm, is assumed to be slightly lessthan the value of the contaminant solubility concentration,Cs. The value of the contaminant concentration at the airwater interface is assumed equal to the solubility concen-tration at a certain time, when the air phase concentrationreaches its peak value. The estimated mass transfercoefficients, KG, are correlated with several dimensionlessnumbers. The dimensionless numbers used were the Sher-wood number (Sh), Peclet number (Pe), Schmidt number(Sc) and several physical-chemical properties of VOCs andporous media. The definitions of the Sherwood number,Peclet number and Schmidt number are as follows:

Sh ¼KG dp50

DG, (9)

Pe ¼ua dp50

DG, (10)

Sc ¼n

DG. (11)

Both the Henry’s law constant and the normalized meanparticle size are considered for the correlation. Thenormalized mean particle size is defined as

d0 ¼ dp50=dm, (12)

where (dm ¼ 0.05 cm) is the mean grain size of mediumsand. The effect of sparging air temperature is introducedin the correlation by the normalized mean temperature, ym.The empirical dimensionless correlation may be presentedas

Sh ¼ b0Peb1Scb2db30 Hb4yb5m , (13)

where b0, b1, b2, b3, b4 and b5 ¼ constants. The parametersbi are estimated by stepwise multiple regression analysis toobtain the best fit parameters for the log-linearized form ofEq. (13). The best fit correlation (R2

¼ 0.98) is

Sh ¼ 6:2E� 4Pe0:96Sc�0:04d1:030 H�1:133y2:32m . (14)

Eq. (14) shows that mass transfer is affected by thephysical properties of the contaminated soil such as theparticle size, air flowrate, the diffusivity of the VOCs,sparging air temperature and the volatility of the VOCs asrepresented by the dimensionless Henry’s law constant.

Fig. 10 presents a comparison of experimentally deter-mined Sherwood numbers with Sherwood numbers predictedby the empirical formula. As illustrated in Fig. 10, thepredictions of the correlation show good agreement with theexperimentally determined Sherwood numbers.

4.3. Uncertainty analysis of experimental results

Eq. (8) was used to determine the contaminant removalefficiency, Zrem. The contaminant removal efficiency is afunction of the initial mass of injected contaminant, theaeration rate—which depends on the measured air velocityand the diameter of the pipe at the measuring point—andthe contaminant concentration in the effluent air which isbased on the LEL reading of the Q-RAE PLUS multi-gasmonitor:

Zrem ¼ f ðmc; u; d21; CaÞ. (15)

The uncertainty of the contaminant removal efficiency isequal to the square root of the sum of the squares of theuncertainties of the separated terms (Holman, 1986):

oZrem ¼

ffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiqZremqmc

omc

� �2þ

qZremqu

ou

� �2þ

qZremqd1

od1

� �2þ

qZremqCa

oCa

� �2s.

(16)

Table 2 summarizes the expected individual uncertaintiesof the measured quantities:From Eq. (16), the uncertainty of the contaminant

removal efficiency could be written as

oZremZrem

¼

ffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiomc

mc

h i2þ

ou

u

h i2þ 2

od1

d1

� �2þ

oCa

Ca

� �2s. (17)

Substituting the individual expected uncertainties of themeasured quantities in Eq. (17) yields that the systematicuncertainty in the contaminant removal efficiency will benot greater than about 71.953%.

ARTICLE IN PRESS

Table 2

Expected accuracy of the measured quantities

Measured quantity Expected accuracy

Initial mass of injected contaminant, mc 71.67%

Sparging air velocity, u 70.025%

Pipe diameter at the measuring point, d1 70.08%

Air phase contaminant concentration, Ca 71%

Fig. 11. The effect of normalized mean air temperature on the benzene

concentration for (z ¼ 0 cm) at 24 h (Silica sand of (dp50 ¼ 0.39mm) and

aeration rate of 27.5mL/min).


concentration for (z ¼ �0.25 cm) at 24 h (Silica sand of (dp50 ¼ 0.39mm),

and aeration rate of 27.5mL/min).


4.4. Theoretical results

The effect of the normalized mean air temperature on thetransient behavior of the dissolved NAPL concentrationthrough the water-saturated soil was studied theoretically.This was achieved by using the estimated volatilizationmass transfer coefficient, which was predicted from theexperimental work, with the theoretical model. Theoreticalstudy was conducted to investigate the benzene NAPLconcentration through the saturated silica sand of meanparticle size diameter 0.39mm. An aeration rate of27.51mL/min was employed for different normalized meanair temperatures. The dissolution mass transfer coefficientwas used from the literature and was equal to 0.227 cm/minfor the current studied conditions.

Fig. 11 presents the effect of normalized mean airtemperature on the benzene concentrations at the air–waterinterface (z ¼ 0 cm) through silica sand of dp50 ¼ 0.39mmunder an aeration rate of 27.5mL/min after 24 h fornormalized mean air temperatures of 1, 1.05, 1.1, 1.15 and1.2. It could be noticed that the changes in the benzeneconcentrations for different normalized mean air tempera-tures are concentrated at x ¼ 4.5 cm to x ¼ 12.5 cm. Inaddition, the benzene concentrations are considered to besymmetrical about the vertical plane of (x ¼ 8.5 cm) whichis the vertical plane of the injection of the NAPL. For aninlet temperature of the sparging air equal to the referencetemperature of 16 oC (ym ¼ 1), the maximum recordedconcentration was 54.7mg/L while for an inlet temperatureof 80 1C (ym ¼ 1.2), the maximum concentration ofbenzene was 37.8mg/L. For a horizontal plane of (z ¼�0.25 cm), Fig. 12 presents the effect of normalized meanair temperatures of 1, 1.05, 1.1, 1.15 and 1.2 on the benzeneconcentration through silica sand of (dp50 ¼ 0.39mm) withan aeration rate of 27.5mL/min after 24 h from thebeginning of the injection of the NAPL. Again, it couldbe noticed that the changes in benzene concentrationsfor different normalized mean air temperatures areconcentrated at x ¼ 4.5 cm to x ¼ 12.5 cm. Also, the dis-tribution of benzene concentrations is consideredto be symmetrical about the vertical plane of NAPLinjection (x ¼ t8.5 cm). For an inlet temperature of thesparging air equal to the reference temperature of 16 1C(ym ¼ 1), the maximum benzene concentration was 181mg/L while for an inlet temperature of 80 1C (ym ¼ 1.2) themaximum achieved concentration of dissolved benzene was168mg/L.

The benzene concentrations for the planes (x ¼ 8.5 and6 cm) through silica sand of (dp50 ¼ 0.39mm) with anaeration rate of 27.5mL/min after 24 h for normalizedmean air temperatures of 1, 1.05, 1.1, 1.15 and 1.2 arepresented in Figs. 13 and 14 respectively. Clearly, the effectof normalized mean air temperature on the distribution ofthe dissolved benzene concentrations for these planes getslower with the increase in depth into the saturated soil. InFig. 13, for (x ¼ 8.5 cm), the benzene concentrations areapproximately similar below the horizontal plane of(z ¼ �1 cm). But for (x ¼ 6 cm) as presented in Fig. 14,the benzene concentrations are approximately similarbelow the horizontal plane of (z ¼ �1.75 cm). This meansthat for the horizontal planes in the zone affected by theheated air, the increase of sparging air temperature hasmore effect for points far from the point of injection of theNAPL glob than the other points near it. This may happen

ARTICLE IN PRESS


concentration for (x ¼ 8.5 cm) at 24 h (Silica sand of (dp50 ¼ 0.39mm)

and aeration rate of 27.5mL/min).


concentration for (x ¼ 6 cm) at 24 h (Silica sand of (dp50 ¼ 0.39mm) and

aeration rate of 27.5mL/min).


because the benzene diffusion from the NAPL plumedominates over the volatilization mass transfer processresulting from the sparging air for the neighborhood closeto the NAPL glob.

The above results show that heating the sparging airaffects the spreading of the dissolved NAPL plume throughthe saturated contaminated soil. This behavior seems to bein agreement with the experimental study where promotinga relatively high normalized mean temperature for thesparging air over the contaminated soil improves thecontaminant removal efficiency.

5. Conculsions

Experimental and theoretical investigations were con-ducted for predicting the dynamic behavior of thecontaminant concentration through water saturated con-

taminated soil under different hot air sparging conditionsfor different NAPLs and soil properties. The conclusionsare as follows:

�
The long-term mass removal for a glop of NAPLlocated close to an air channel was dependant on theaqueous diffusion of the dissolved phase towards theair–water interface which in turns increases with theincrease in the mean particle size of the contaminatedsoil. � An increase in air flowrate produces an increase in the
overall mass removal and this might be related to anincrease of the mass transfer zone affected by the airsparging channel.
� The volatilization mass transfer increased with the
increase of the air and water diffusion coefficients ofthe VOCs. With regard to the Henry’s law constant ofthe VOCs, an inversely proportional trend was obtained.
� Using hot sparging air through the single air channel
setup provided higher contaminant removal efficiencies,as deduced from the experimental study, by about 9%.
� Promoting a relatively high-normalized mean tempera-
ture for the air sparging over the contaminated soilimproved the contaminant removal efficiency as a resultof enhancing the volatilization mass transfer process.This would, in turn, affect the spreading of the dissolvedNAPL plume through the mass transfer zone affected bythe sparging air.

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Further reading

Julio, S., Drucker, A.S., 2000. Air sparging remediation: A study on

heterogeneity and air mobility reduction. Proceedings of the 2000

Conference on Hazardous Waste Research.

http://www.ce.vt.edu/programarea-areas/environmental/teach/gwprimer/airsparg2/airsparg.html



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