FACULTEIT LANDBOUWKUNDIGE ENTOEGEPASTE BIOLOGISCHE WETENSCHAPPEN

Academiejaar 2003-2004

Modelling diffusion and reaction of gases inpetroleum hydrocarbon contaminated soils.

Case study of the biopile treatment technology

Modelleren van diffusie en reactie van gassen inmet minerale olie verontreinigde bodems.Gevalstudie van remediatie in biobedden

door

ir. Joke Van De Steene

Thesis submitted in fulfilment of the requirements for the degree of Doctor(Ph.D.) in Applied Biological Sciences, Option: Land Management and

Forestry

Proefschrift voorgedragen tot het bekomen van de graad van Doctor in deToegepaste Biologische Wetenschappen, Optie: Land- en bosbeheer

op gezag vanRector: Prof. Dr. A. De Leenheer

Decaan: Promotor:Prof. Dr. ir. H. Van Langenhove Prof. Dr. ir. H. Verplancke

De auteur en de promotor geven de toelating dit werk voor consultatie beschikbaar testellen en delen ervan te kopiëren voor persoonlijk gebruik. Elk ander gebruik valt onderde beperkingen van het auteursrecht, in het bijzonder met betrekking tot de verplichtinguitdrukkelijk de bron te vermelden bij het aanhalen van resultaten uit dit werk.

The author and the promotor give permission to consult and to copy parts of this work forpersonal use. Any other use is limited by the Laws of Copyright. Permission to reproduceany material contained in this work should be obtained from the author.

Gent, september 2004

De auteur De promotor

ir. Joke Van De Steene Prof. Dr. ir. H. Verplancke

DankwoordEen doctoraatsthesis schrijven kan je niet alleen. Daarom wil ik iedereen danken die deafgelopen jaren een bijdrage heeft geleverd. Een aantal mensen wil ik hier in hetbijzonder vermelden.

Eerst en vooral wil ik mijn promotor, Prof. Dr. ir. H. Verplancke bedanken voor de kansdie hij me bood om dit onderzoek uit te voeren. Bij hem kon ik terecht met al mijn vragenen bedenkingen en hij gaf me een grote vrijheid om in te werken, die ik enorm hebgewaardeerd. Een speciaal woord van dank gaat uit naar Dr. ir. Alex De Visscher, Prof.Dr. ir. Jo Dewulf en Dr. Johan Gemoets voor hun kritische bemerkingen die verwerkt zijnin dit proefschrift.

Mijn collega's wil ik bedanken voor de prettige samenwerking, in het bijzonder Prof. Dr.ir. G. Hofman en Dr. ir. Wim Cornelis voor het lezen van mijn proefschrift en Dr. ir.Stefaan De Neve en ir. Philippe Verschoore voor de vele tips die ze me gaven.

Mijn thesisstudenten, ir. Els de Brabanter, ing Hannelore Van Vooren, ir. RobertLandazuri en ing Dirk De Bruyne, wil ik bedanken voor hun medewerking en hun inzet,die zeer waardevol zijn gebleken. Verder een bijzonder woord van dank aan Eric,Thérèse, Jan voor de hulp bij de analyses, Luc, Matthieu voor het grondverzet en Anita,Elly en Patrick voor de administratieve hulp.

Voor de kans die mij werd geboden om praktijkproeven uit te voeren in het GRC, wil ikhet milieubedrijf DEC danken. Gunther, Stany, Dirk en Katelijne, de feedback enpraktijkervaring die jullie me gaven, was erg waardevol.

Tijdens mijn onderzoek ben ik bij heel wat mensen op de deur gaan kloppen. Devolgende mensen hebben allemaal op één of andere manier bijgedragen tot het welslagenvan dit werk: Prof. Dr. ir. H. Van Langenhove, Dr. ir. Jan de Myttenaere, ir. An Adamsen ir. Karlien De Roo en Dhr. Theo De Smet.

Tenslotte wil ik in het bijzonder mijn ouders bedanken. Jullie waren er altijd op deachtergrond, waarbij jullie mij steunden en stimuleerden in alles wat ik deed. Bedanktvoor al jullie vertrouwen. En last but not least, Gaetan, dank je voor je onvoorwaardelijkesteun en omdat je me samen met de kleine Noah zo gelukkig maakt.

Gent, 23 september 2004.

Content____________________________________List of symbolsChapter 1 Introduction 1

1.1 Soil pollution 31.2 Soils polluted with petroleum hydrocarbons 41.3 Remediation of petroleum hydrocarbon polluted soils 61.4 Objectives and outline 8

Chapter 2 Literature review 92.1 Petroleum 11

2.1.1 Introduction 112.1.2 Petroleum hydrocarbons 11

2.1.2.1 Composition and characterisation 112.1.2.2 Petroleum distillates 142.1.2.3 Diesel fuel 142.1.2.4 Toxicity of diesel fuel 16

2.1.3. Petroleum hydrocarbons in soil 172.1.3.1 Behaviour of petroleum hydrocarbons in soil 172.1.3.2 Sorption of HOCs to soil aggregates 18

2.1.4 The Flemish statute on soil clean up 222.1.5 Analysis of petroleum hydrocarbons 23

2.1.5.1 Gas chromatography 242.1.5.2 IR spectroscopy (Weisman, 1998) 262.1.5.3 Gravimatric method (Weisman, 1998) 262.1.5.4 Immunoassay (Weisman, 1998) 27

2.2 Biodegradation of petroleum hydrocarbons 282.2.1 Introduction 282.2.2 Biodegradability of petroleum 282.2.3 Factors influencing biodegradation of petroleum hydrocarbons 29

2.2.3.1 Soil water content 292.2.3.2 Oxygen 302.2.3.3 Temperature 302.2.3.4 pH 312.2.3.5 Nutrients 322.2.3.6 Bioavailability of petroleum hydrocarbons 33

2.2.4 Measurement of petroleum biodegradation 342.2.5 Petroleum hydrocarbon biodegradation rates in soil 352.2.6 Modelling biodegradation of petroleum hydrocarbons in soil 37

2.2.6.1 Microbial kinetics 372.2.3.2 Biodegradation models 38

2.3 Gas transport through soil 40

2.3.1 Introduction 402.3.2 Gas conservation equation 402.3.3 Gas convection in soil 412.3.4 Gas diffusion in soil 41

2.3.4.1 Free gas diffusion 412.3.4.2 Diffusion in porous media 432.3.4.3 Equations for relative effective diffusivity in soil 45

2.4 Bioremediation 502.4.1 Introduction 502.4.2 Bioremediation technologies 50

2.4.2.1 Ex situ processes 502.4.2.2 In situ processes 51

2.4.3 Biopiles (Batelle, 1996) 532.4.4 Biopile operation 54

2.4.4.1 General construction 542.4.4.2 Aeration 552.4.4.3 Moisture addition 572.4.4.4 Nutrient addition 582.4.4.5 Microbial amendment 592.4.4.6 Temperature 592.4.4.7 Metal content 59

2.4.5 Respiration testing 60

Chapter 3 Oxygen diffusion coefficient and oxygen diffusionrate in a loamy sand soil 63

3.1 Introduction 653.2 Effective diffusion coefficient in a loamy sand soil 65

3.2.1 Introduction 663.2.2 Diffusion measurement methodologies 66

3.2.2.1 Steady state methods 663.2.2.2 Non-steady-state methods 673.2.2.3 Radial diffusion methods 71

3.2.3 Material and methods 723.2.4 Results and discussion 76

3.2.4.1 Limitations and accuracy of the 763.2.4.2 Effective diffusivity ξ as a function of air-filled porosity 78

3.3 Oxygen diffusion rate (ODR) in a loamy sand soil 803.3.1 Introduction 803.3.2 ODR principal and practical limitations 803.3.3 Material and methods 853.3.4 Results and discussion 86

3.4 Oxygen diffusion coefficient versus ODR in a loamy sand soil 884.4.1 Introduction 88

4.4.2 Materials and Methods 884.4.3 Results and discussion 90

3.5 Conclusions 92

Chapter 4 Optimal soil water content for diesel degradationin a loamy sand soil 95

4.1 Introduction 974.2 Measurements of soil water content 974.3 Factors influencing the microbial activity to water relation 100

4.3.1 Undisturbed versus disturbed soils samples 1004.3.2 Diffusion limitation 1014.3.3 Fungal versus bacterial activity 1014.3.4 Substrate induced respiration (SIR) 102

4.4 Optimal soil water content 1024.5 Conceptual model of optimum soil water content for aerobic

microbial activity 1024.6 Material and methods 107

4.6.1 Soil physical measurements 1074.6.2 Incubation 108

4.7 Results and discussion 1104.7.1 Soil respiration 1104.7.2 Methodological aspects 1114.7.3 Diesel induced soil respiration 1154.7.4 The respiratory quotient r 1184.7.5 The Skop model 120

4.7.5.1 Solute diffusion 1204.7.5.2 Oxygen diffusion 1234.7.5.3 Optimal soil water content 123

4.8 Conclusions 128

Chapter 5 Applicability of Fick’s law to gas diffusion in soilcontaminated with petroleum hydrocarbons 129

5.1 Introduction 1315.2 Theory 132

5.2.1Continuity equation 1355.2.2 Fick’s law 1395.2.3 Stefan-Maxwell Equations 1395.2.4 Adjusted Fick's law 142

5.3 Material and Methods 1435.3.1 Calculation of Yi(z) profiles 1445.3.2 Calculating production rates SFi from the Fickian solution and

SAFi from the adjusted Fickian solution 1455.3.3 Selection of the model hydrocarbons 145

5.4 Results and discussion 1475.4.1 Steady state transport in a hypothetical biopile with toluene

biodegradation 1475.4.2 Steady state transport in a hypothetical biopile with

biodegradation of different petroleum hydrocarbons 1565.4.3 Deviations made when using Fick’s Law and the adjusted Fick's law

to predict respiration rates in the hypothetical biopile 1585.4.4 Deviations made when using the adjusted Fick's law and neglecting

the presence of a volatile hydrocarbon 1625.5 Conclusions 163

Chapter 6 Estimating hydrocarbon degradationfrom N2, O2 and CO2 profiles 165

6.1 Introduction 1676.1.1 Transient methods 1676.1.2 Steady-state methods 168

6.2 Mathemathical model 1696.3 Material and methods 170

6.3.1 Soil pretreatment 1706.3.2 Design column experiment 1706.3.3 Parameters column experiment 172

6.4 Results 1736.4.1 Soil water content θ and relative effective diffusivity ξ 1736.4.2 Oxygen and carbon dioxide profiles 1766.4.3 O2 consumption and CO2 production 1796.4.4 Diesel degradation 183

6.4.4.1 TPH analysis 1836.4.4.2 Calculating diesel degradation rates from respiration data186

6.5 Conclusions 189

Chapter 7 Petroleum hydrocarbon remediation in biopiles 191

7.1 Introduction 1937.2 Mathematical model 1947.3 Material and methods 194

7.3.1 The biopile cell 1957.3.2 Pretreatment of the soil 1967.3.3 Sampling 1967.3.4 Respiration testing 199

7.4 Results 2007.4.1 Hydrocarbon degradation 2007.4.2 Bulk density 2007.4.3 Soil water content 2017.4.4 Temperature 2037.4.5 Aeration system 2097.4.6 Soil respiration 2107.4.7 Hydrocarbon degradation rate 214

7.5 Conclusions 216

Chapter 8 Conclusions and perspectives 2178.1 Introduction 2198.2 Comparing the effective oxygen diffusion coefficient to theoxygen diffusion rate in a loamy sand soil 2208.3 Finding the optimal soil water content for diesel degradation in aloamy sand soil 2218.4 Estimating petroleum hydrocarbon degradation rates from O2,CO2 and N2 versus depth profiles in soils 222

Appendix A 227

Samenvatting/Summary 229

References 243

Curriculum Vitae 265

List of symbols

a empirical constant Eq. [4.3],-a1 empirical constant Eq. [2.13],-a2 empirical constant Eq. [2.17],-A cross sectional area of the sample, m2

Ael electrode surface, cm2

A(αn) function defined by Eq. [3.6]Al area per unit soil mass through which the flux of soluble substrate

occurs, m2 kg-1

AO2 area per unit soil mass trough which the flux of oxygen occurs, m2 kg-1

b Campbell (1974) water retention parameter,-b' slope of linear regression curveb1 empirical constant Eq. [2.13],-B biomass concentration, mg biomass L-1

B(αn) function defined by Eq. [3.7]B2 parameter Eq. [5.32]B3 parameter Eq. [5.38]B4 parameter Eq. [5.40]c = p/RT, molar density of the gas mixture, mol gas m-3 airci dimensionless concentration, = Ci,E/Ci,E0c1 empirical constant Eq. [2.13],-C substrate concentration, mg L-1

Cg soil gas concentration, mol gas m-3 soil airCi, concentration of component i in the gaseous phase, mol m-3 airCi,I concentration of gas component i in the injection chamber, mg L-1

Ci,E concentration of gas component i in the exit chamber, mg L-1

Ci,E0 concentration of gas component i in the exit chamber at t = 0, mg L-1

Ci, ∞ concentration at equilibrium or infinite time, mg L-1

Cl substrate concentration within microbially inaccesible regions in thesoil, mol m-3soil

Cmax,s maximum concentration of the hydrocarbon in the soil solid phase, molkg-1 soil

Cmax,w maximum concentration of the hydrocarbon in the soil water, mol m-3

soil waterCO2 oxygen concentration in the atmosphere, mol m-3 airCorg concentration of the organic component in the gas phase, mol m-3 airCorg,s concentration of the organic component in the solid phase, mol kg-1 soilCorg

sat saturated organic vapour concentration of component i, mol m-3 airCorg,w concentration of the organic component in the soil water, mol m-3 soil

waterCs equilibrium sorbent phase solute concentration, mg kg-1

CT totale solute concentration, mol m-3 soil

Cw equilibrium aqueous phase solute concentration, mg L-1

C1 integration constant,-d empirical constant Eq. [4.3],-d' boundary layer thickness Eq. [5.18], cmdi diffusion volume of the ith molecule,-Ddryg effective diffusion coefficient for the dry bed, m2 day-1

DeF,i effective Fickian diffusion coefficient of gas component i, m2 day-1

Deg effective diffusion coefficient in the gaseous phase, m2 day-1

De,ij effective binary diffusion coefficient of gas components i and j, m2 day-1

De,im effective binary diffusion coefficient of gas component i in a mixture,m2 day-1

Del effective diffusion coefficient in the aqueous phase, m2 day-1

De,O2 effective diffusion coefficient of oxygen in soil, m2 day-1

Dg100 effective diffusion coefficient at –100 cm H20, m2 day-1

Dij binary diffusion coefficient of gas component i and j, m2 day-1

Dim binary diffusion coefficient of gas component i in a mixture, m2 day-1

D0F Fickian gaseous diffusion coefficient in free air, m2 day-1

D0F,i Fickian diffusion coefficient of gas component i in free air, m2 day-1

D0g effective diffusion coefficient in free air, m2 day-1

D0,O2 diffusion coefficient of oxygen in free air, m2 day-1

D0l effective diffusion coefficient in water, m2 day-1

D0,org effective diffusion coefficient of the hydrocarbon in free air, m2 day-1

e parameter Eq. [2.33],-f empirical constant Eq. [4.3],-fCO2 diffusional flux density of carbon dioxide, mol m-2 day-1

fg soil gas flux, mol m-2 day-1

fi diffusional flux of gas component i, mol m-2 day-1

fN2 diffusional flux of nitrogen, mol m-2 day-1

fOC fraction of organic carbon in soil, g g-1

forg diffusional flux of the hydrocarbon, mol m-2 day-1

fO2 diffusional flux of oxygen, mol m-2 day-1

fvol volatilisation flux of the hydrocarbon, mol m-2 day-1

F Faraday constant, 96 500 C mol-1

g empirical constant Eq. [4.3],-H parameter Eq. [4.12], [4.13] and [4.14],-i indexI parameter Eq. [4.12], [4.13] and [4.14],-j indexJ oxygen reduction current, µAJl diffusional flux of soluble substrate, mol m-2 day-1

JO2 diffusional flux of oxygen, mol m-2 day-1

k indexkHC the hydrocarbon degradation rate, mg HC equivalent kg-1 soil day-1

kT thermal conductivity, W m-1 K-1

k0 zero-order biodegradation rate constant, mg L-1 day-1

k1 first-order biodegradation rate constant, L day-1

K parameter Eq. [3.1],-K1 parameter Eq. [5.25],-K2 parameter Eq. [5.31],-K3 parameter Eq. [5.37],-K4 parameter Eq. [5.39],-Kd equilibrium partitioning coefficient, L kg-1

KH Henry’s law constant for gas component i, mol m-3 gas (mol m-3 water)-1

Km Michaelis-Menten constant, mg L-1

KOC normalised partitioning coefficient, L kg-1

KOW octanol-water partitioning coefficient,-Kp concentration independent partitioning coefficient, L kg-1

Ks half saturation constant for growth, mg L-1

l linear distance across the porous medium, m soill' equivalent diffusion distance, mle tortuous path length across the pore space, m airL Length of the sample, mm parameter Eq. [2.33],-M Van Genuchten parameter, Eq. [4.8],-Mij parameter Eq. [2.15],-n parameter Eq. [2.1], Eq. [2.33],-ne number of electrons used per molecule of oxygen electrolysed, ne = 4N van Genuchten parameter Eq. [4.8],-ODF oxygen flux density, g cm-2 s-1

ODR oxygen diffusion rate, g cm-2 s-1

p total pressure, Papsat saturated vapour pressure, PaP potential microbial activity, mol kg-1 soil day-1

PO2 partial pressure of O2 in the N2 vessel, PaPO2,0 partial pressure of O2 in the N2 vessel at time t = 0, Paq heat flux, W m-2

r respiration quotient, mol mol-1

R universal gas constant, 8.3144 J mol-1 K-1

Rl soluble substrate transfer resistance, m-1

RO2 oxygen transfer resistance, m-1

s = SO2/SCnHm, hydrocarbon degradation quotient, mol mol-1

S = θ/φ, degree of saturation or relative soil water content, m3 water m-3

voidSA the water-filled volume fraction of the intra-aggregate pore space, θA/φA,

m3 water m-3 voidSAFi production rate of gas component i calculated with the adjusted Fick's

law, mol m-3 soil day-1

Saq aqueous solubility, mol m-3 waterSA surface area, m2 g-1

SAvol surface area on a volumetric basis, m2 m-3

SCnHm hydrocarbon consumption rate, mol m-3 soil day-1

SCO2 carbon dioxide production rate, mol m-3 soil day-1

SFi production of gas component i calculated with Fick's law, mol m-3 soilday-1

Sg gas reaction loss rate, mol m-3 soil day-1

Si production rate of gas component i, mol m-3 soil day-1

Sorg (= Svol - Sorg,bio), mol m-3 soil day-1

Sorg,bio hydrocarbon degradation rate, mol m-3 soil day-1

SO2 oxygen consumption rate, mol m-3 soil day-1

Sp the water-filled volume fraction of the inter-aggregate pore space,θp/φp,m3 water m-3 void

Svol production rate of hydrocarbon by volatilisation, mol m-3 soil day-1

t time, daysT absolute temperature, Ku experimental parameter Eq. [2.22], Eq. [2.34],-u1, u2 parameters Eq. [4.14],-U electrical potential polarising the indicator electrode, VUe electrical potential of electrolysis of water, V. For Pt, as referred to SCE,

equal to about -0.7 VUfix fixed level electrical potential, Vv experimental parameter Eq. [2.22], Eq. [2.34],-v' velocity, m day-1

V volume of the chamber, m3

Vmax maximum reaction velocity, mg L-1 day-1

Vs volume of the sample, m3

Wg gravimetric soil water content, g g-1

WHC water holding capacity, %Wp water potential, kPaWFP(S) water filled porosity or water filled pore space, %x distance, cmx1, x2, x3 parameters Eq. [2.23],-y = x/L, relative distance,-Y yield coefficient, mg L-1 biomass produced per mg L-1substrate degradedYi mole fraction of gas component i, mol mol-1

Yi,z=0 mole fraction of gas component i at the soil surface, mol mol-1

z height or depth, mzL height at the bottom of the soil profile, mz0 height at the top of the soil profile, mα van Genuchten parameter, Eq. [4.8],-αn nth root of Eq. [3.8], parameter of Eq. [3.8],-β = H/Lε, length ratio, (Eq. [3.3] and [3.5]),-γ = V/AL, volume ratio (Eq. [3.3] and [3.5]),-δ constrictivity,-ε air filled porosity, m3 soil air m-3 soilε100 air filled porosity at –100 cm H20, m3 soil air m-3 soilφ total soil porosity, m3 void m-3 soil

φA intra-aggregate porosity, m3 intra-aggregate void m-3 soilφp interaggregate porosity, m3 inter aggregate void m-3 soilλn nth root of Eq. [3.4], parameter of Eq. [3.3],-µmax maximum specific growth rate, day-1

ν number of species in gas mixture,-θ volumetric soil water content, m3 water m-3 soilθA volumetric soil water content of the intra-aggregate pore space, m3 m-3

θopt optimal volumetric soil water content for microbial activity, m3 m-3

θp volumetric soil water content of the inter-aggregate pore space, m3 m-3

θr residual soil water content, m3 water m-3 soilθs saturated soil water content, m3 water m-3 soilθS = 1 – ε, volume solid per total bed, m3 m-3

θth threshold water content at which solute diffusion is effectively zero, m3

m-3

ρb soil bulk density, Mg m-3

ρw density of water, Mg m-3

τ tortuosity,-τ' = Degt/ εL2, dimensionless time,-ωi molecular weight of component i, g mol-1

ωO2 molecular weight of oxygen, 32 g mol-1

Ω ratio = aAlRlClD0l/dAO2RO2CO2D0,O2,-ξ relative effective diffusivity (section 2.3.4.1)ξ' effective diffusivity of the bed Eq. [2.21],-ξT total effective diffusivity Eq. [2.27],-ψ soil matric potential, kPa

Chapter 1

Introduction

3

Chapter 1

Introduction

1.1 Soil pollution

Years and years, soil and groundwater pollution problems have had a silent existence,when only shocking cases couldn't escape the public eye. Incidents as the "Lommel zincdesert" in Flanders and the "Lekkerkerk incident" in the Netherlands during the 80's,accelerated the awareness of these environmental problems. However, only after issuingthe Soil Remediation Decree in 1995, the soil pollution problem was addressed inFlanders (Ide and Ectors, 1996). Nowadays, preliminary surveys of contaminated sitesare far advanced in many European countries, but further stages are proceeding slowly(EEA, 2003).

From national as well as international experience, lessons have been learned thatinsufficient attention in the past could lead to serious situations. Many times, peopledidn't even realise that a soil could be polluted (Ide and Ectors, 1996). Looking at thenumber of sites identified as being polluted, the extent of this problem becomes clearer.In the European Union (EU) efforts were made to assess the number of potentiallycontaminated and contaminated sites in 15 European countries (Table 1.1), resulting in atotal number of 1 417 742 contaminated by 2001 (EEA, 2002). More recently, thenumber of potentially contaminated sites was estimated to be 76 200 in Flanders only. Bythe end of 2002, the Flemish Public Waste Company (Openbare VlaamseAfvalmaatschappij, OVAM) had processed investigations on 16 688 of these sites, ofwhich 13 305 proved to be polluted (MIRA, 2003).

4

Table 1.1 - Available data on the number of contaminated sites and remediated sites in 16European countries.

Country number of contaminates sites(sites included in nationalregisters after a preliminarysurvey)

number of remediated sites(sites where remediationactivities have been completed)

referenceyear

Austria 80 000 29 2001Belgium (Fl) 53 000 60 2000Denmark 30 000 4 800 2000Finland 25 000 1 000 2000France 300 000-400 000 466 2000Germany 362 000 3 000 1998Iceland n.i. 1 2000Ireland 2 500 n.i. 2000Italy 100 000 500 2001Liechtenstein 100 9 1999Netherlands 175 000 7 100 2000Spain 18 142 23 2000Sweden 22 000 200 2000Switserland 50 000 n.i. 2000UK 100 000 n.i. 2000n.i. no information available, Fl data for Belgium include only Flanders region

1.2 Soils polluted with petroleum hydrocarbons

Petroleum, or crude oil, continues to be the world's most important primary energysource, accounting for 38.5 percent of the world primary energy production in 2001(USEIA, 2003). Furthermore, chemicals derived from petroleum or natural gas, so-called"petrochemicals", are being used in an incredible variety of areas, from household goods,to medicine and from leisure to highly specialised fields like archaeology or crimedetection.

Considering the scale of the petroleum use and transport, emissions to the environmentare substantial. This causes mineral oil to be one of the major pollutants of soil and water,being emitted to the environment by handling losses, defects or accidents (EEA, 2001).

Tanker oil spills continue, causing enormous amounts of petroleum hydrocarbons to bespilled in sea. One of the worst ecological disasters ever recorded in Galicia, in Spain, inEurope, and even world wide, started in November 2002, when the Prestige oil tanker,suffering from a turbulent sea and high winds started to spill fuel. The Prestige was

5

transporting more than 77 000 metric tons of fuel oil, twice as much oil as the infamousExxon Valdez, which sank in Alaskan waters in 1989 (ADEGA, 2003).

However, spills on land can also cover substantial surfaces. In February 1994, an oilexploration well blew out during drilling in Trecate, northern Italy. Approximately15 000 m3 of crude oil was expelled from the well and deposited over approximately5 km2 of agricultural lands (Reisinger et al., 1996). Other sources of petroleumcontamination on land are fuel storage tanks, handling of fuels and leaking pipelines.

In the USA more than 439 000 releases from leaking underground storage tanks (LUSTS)have been reported as of September 2003. Cleanups have been initiated at more than403 000 of these sites, and more than 303 000 sites have been cleaned up (USEPA,2004).

In Flanders mineral oil, defined as all non-aromatic hydrocarbons, is the second mostprevalent pollution type (Fig. 1.1).

heavy metals25%

mineral oil17%

PAH's16%

BTEX8%

chlorinated solvents

5%

other29%

Figure 1.1 - Total number of polluted sites in Flanders in 2002 divided by pollution type(MIRA, 2003)

6

Mineral oil pollution is found at big professional market places, like fuel distributors andpetrol stations, as well as companies using these chemicals as fuels or raw material or infuel storage tanks of private individuals (MIRA, 2003). The prevalence of potentiallypolluted sites of petrol stations in Belgium, and the high costs associated withremediating these sites, resulted in the founding of a non profit organisation. Thisorganisation named "Bodemsaneringsfonds voor Tankstations" (BOFAS) was recognisedby the Inter Regional Soil Remediation Committee (Interregionale Bodemsanerings-commissie, IBC) in Belgium in March 2004. The mission of BOFAS is to structurallyapproach the soil contamination problem of petrol stations in Belgium.

1.3 Remediation of petroleum hydrocarbon polluted soils

Today, remediation of contaminated soils is as diverse as the types of contaminants andthe site conditions (Kujat, 1999). Traditional clean-up technologies used to remediateLUSTs in the USA were pump-and-treat for groundwater and excavation and disposal forsoil. With so many sites requiring remediation at such an enormous cost, theEnvironmental Protection Agency (EPA) actively promoted faster, more effective andless costly alternatives to traditional cleanup methods. Alternative cleanup technologiesbeing applied in situ include soil vapor extraction (SVE), bioventing, air sparging,biosparging, in-situ groundwater bioremediation, dual-phase extraction and monitorednatural attenuation. Ex-situ alternatives are low-temperature thermal desorption,landfarming, biopiles and soil washing (USEPA, 2004).

By 2002 in Flanders the most popular remediation technique remained the off-siteremediation, comprising the excavation, transport of the soil and treatment by an ex-situtreatment technology (Table 1.2). Ex-situ technologies include physicochemicaltechniques as well as biological techniques.

Biopiles are used to reduce concentrations of petroleum constituents in excavated soilsthrough the use of biodegradation. This technology involves heaping contaminated soilsinto piles and stimulating aerobic microbial activity within the soils through the aerationand/or addition of minerals, nutrients and moisture.

7

Table 1.2 - Overview of remediation techniques of soil remediation projects in Flanders until2002 (MIRA, 2003)

Remediation type Remediation technique Numberex situ soil remediation off-site remediation 760

on-site remediation 10groundwater remediation disposal/ex-situ treatment 748

re-infiltration 12in-situ remediation soil vapour extraction 187

airsparging 67bioventing/biosparging 15reactive barier 4natural attenuation 67fytoremediation 2bioprecipitation 1immobilisatio,/neutralisation 5ORC/HRC 9chemical oxidation 6steam injection 1surfactants 1civil technical 54geohydrological 15

Biodegradation has been found to be an efficient method for the reduction ofhydrocarbons in soil. Although, it is by far the most elegant and pure cleanup solution,resulting in the transformation of toxic compounds to carbon dioxide and water, theoverall process can be unpredictable and unreliable due the variety of physical, chemicaland biological factors (Kujat, 1999).

Oxygen is the key parameter in the aerobic microbial degradation of the pollutants.Detailed knowledge of the spatial variability of oxygen penetration and biodegradationrates over time is required to determine the efficiency of the remediation treatment andenable decisions of the end-point of treatment. Few data are available on the long-termmonitoring of degradation rates within biopiles (Patterson et al., 1999). Although theimportance of soil water content for oxygen diffusion has been mentioned (Huesemannand Truex, 1996), the optimal range of soil water content for biopile treatment has beenvaguely described to be situated between 40 and 85 percent of field capacity (USEPA,2004).

8

1.4 Objectives and outline

Petroleum biodegradation in biopiles, is affected by oxygen diffusion and reaction. Theprimary objective of this thesis was to gain a better understanding of the diffusion andreaction of gases associated with petroleum hydrocarbon biodegradation.

Background information on biodegradation of petroleum hydrocarbons, gas transport andbioremediation is provided in the literature review (Chapter 2). This is followed by adiscussion and a comparison of two diffusion measurement methodologies, being theoxygen diffusion coefficient and the oxygen diffusion rate (ODR) (Chapter 3).

Soil water content has a marked influence on soil microbial activity. A second objectivewas to gain a better understanding of the relationship between this parameter andpetroleum hydrocarbon degradation. The influence of soil water content on dieseldegradation was studied in laboratory incubations (Chapter 4).

A third objective was to study the use of oxygen, carbon dioxide and nitrogenconcentration versus depth profiles to predict hydrocarbon degradation rates in situ. InChapter 5 a theoretical analysis is given of models describing diffusion of gasesassociated with petroleum hydrocarbon degradation in a hypothetical biopile. The modeldeveloped in Chapter 5, was evaluated with experimental results from laboratory columnstudies (Chapter 6) and a biopile treatment facility (Chapter 7).

Conclusions and perspectives are given in Chapter 8.

Chapter 2

Literature review

11

Chapter 2

Literature review

2.1 Petroleum

2.1.1 Introduction

To date, petroleum is still the most important world energy source. Because of the scaleof petroleum consumption and petroleum transport the emissions into the environment asa result of human errors, defects or spillages may be small as a percentage, but inabsolute terms they can be significant (Van Eyk, 1997).

2.1.2 Petroleum hydrocarbons

2.1.2.1. Composition and characterisation

The origin of petroleum has long been a topic of research interest and discussion. It isnow evident that petroleum formation is associated with the development of fine-grainedsedimentary rocks, deposited in a marine or near-marine environment and that petroleumis the product of plant and animal debris, incorporated in these sediments at the time ofdeposition. However, the details of this transformation and the mechanism by whichpetroleum is expelled from the source sediment and accumulates in the reservoir rock arestill uncertain (Speight, 1991).

On a molecular basis, petroleum is a complex mixture of hydrocarbons with 1 to 60carbon atoms and compounds containing nitrogen, sulphur and oxygen (NSO). Metallicconstituents may also be present, particularly vanadium, nickel, iron, and copper but onlyto a minor extent. The hydrocarbon content of petroleum may be as high as 97%, forexample in the lighter paraffinic petroleums, or as low as 50% or less as illustrated by theheavier asphaltic crude oils. From the fragmentary data that is available it appears that theproportions of the elements in petroleum vary only slightly over fairly narrow limits(Table 2.1) (Speight, 1991).

12

The components of crude oil can be split into a number of fundamental chemical groupson the basis of their structures, namely the aliphatic hydrocarbons, the cycloaliphatichydrocarbons, the aromatic hydrocarbons and the NSO compounds. The distribution ofthe various compound types throughout the boiling range of petroleum is given in Figure2.1.

Aliphatic hydrocarbons or paraffins include n-alkanes, branched alkanes (isoparaffins)and unsaturated hydrocarbons, alkenes or olefins and alkynes. The latter are rare in crudeoil but may be common in some refined products.

Table 2.1 - Proportion of the different elements in crude oil (Speight, 1991)Element Proportion (%)

Carbon 83.0-87.0Hydrogen 10.0-14.0Nitrogen 0.1-2.0Oxygen 0.05-1.5Sulphur 0.05-6.0

The proportion of cycloaliphatic compounds or naphthenes varies with the type of crudeoil. Cyclopentane, cyclohexane and decahydronaphtalene derivatives are largelypresented in oil fractions. Multi-ring compounds, such as terpenes are also common(Speight,1991).

Aromatics are a very important class of compounds that occur in petroleum. There is ageneral increase in proportion of aromatic hydrocarbons with increasing molecularweight, but aromatic hydrocarbons without the accompanying napthene rings or alkyl-substituted derivatives seem to be present in appreciable amounts only in the lowerpetroleum fractions (Speight, 1991). Polycyclic aromatic hydrocarbons (PAHs) constitutea class of organic chemicals consisting of two or more fused benzene rings in linear,angular and cluster arrangements (Cerniglia, 1992).

Oxygen in organic compounds can occur in a variety of forms. Although alicyclic(napthenic) acids appear to be the more dominant, aliphatic acids are also present. In

13

addition to the carboxylic acids and phenolic compounds, the presence of ketones, esters,ethers and anhydrides accounted for a variety of petroleums (Speight, 1991).

Figure 2.1 – The distribution of the various compound types throughout the boiling range ofpetroleum (Speight, 1991).

The sulphur content of crude oil ranges from less than 2 to 60 g kg-1, depending on theorigin of the oil. The sulphur is present not only as sulfide but also as mercaptans,thiophenes, and more complex organic sulphur compounds (Environmental HealthCriteria 20, 1982). The level of organic nitrogen compounds in most crude oils is lessthan 1 g kg-1, but some may occasionally contain as much as 20 g kg-1. Nitrogen inpetroleum may be classified arbitrarily as basic or non-basic. The basic nitrogencompounds, which are composed mainly of pyridine homologs and occur throughout theboiling ranges, have a tendency to exist in the higher boiling fractions and residua. Thenon-basic nitrogen compounds, which are usually of the pyrrole, indole, and carbazoletypes, also occur in the higher boiling fractions and residua. In general the nitrogencontent of petroleum is low and falls within the range 0.1 to 2.0%. (Speight, 1991).

There also exists a fraction termed “resins” or “asphaltenes”, which is defined as asolubility class that is precipitated from petroleum, heavy oil, and bitumen by addition of

14

an excess of a liquid paraffinic hydrocarbon. They are recognised as being a complexmixture of species with varying molecular weight and polarity. (Speight, 1999).

2.1.2.2 Petroleum distillates

In the refinery, crude oil is distilled by heating it to a liquid temperature of approximately340 °C. When the temperature is further increased, the crude oil starts to disintegrate(known as "cracking") (Van Eyck, 1997). The main products, after the first majortreatment (distillation) of crude oil at the refinery, ranked according to increasing boilingranges are:

- LPG or liquefied gas,- naphthenes; used as a feedstock for the chemical industry,- gasoline; for internal combustion engines,- jet fuels; for jet engines,- paraffin-oil; for heating, illuminating and cooking purposes,- diesel oil; for diesel engines,- gasoil or domestic oil; for central heating,- fuel oils with different densities, heavy fuel oil; used for ship engines,- lubricating oil; for lubrication,- bitumen; for roads, roof cover, etc.

Table 2.2 lists some of the pertinent properties of the various fractions listed above. (VanEyck, 1997).

2.1.2.3 Diesel fuel

Since our research focuses on diesel fuel pollution, this fraction of crude oil will bediscussed in more detail.

Diesel fuel is a complex mixture of 60 to more than 90 percent normal, branched, andcyclic alkanes, having a hydrocarbon chain length usually between C9 and C30, about 20percent aromatic compounds, especially alkylbenzenes and 0 to 10 percent alkenes,obtained from the middle-distillate, gas-oil fraction during petroleum separation.

15

Table 2.2 - Some distillates from petroleum refining (after Van Eyck, 1997)Fraction Density

at 15°Ckg m-3

ChainLength*

Aromaticstotal %

AromaticsBTX%

Boilingrange°C

Vapourpressureat 15°CkP

LPG 510-580 C2-C5 0 0 -42-0 180-720Gasoline 720-760 C4-C12 40 15-40 25-200 16-40Naphthenes 650-760 C5-C10 10-40 0-5 40-200 7-40Paraffin oil 780-820 C7-C15 20 <1 140-270 0.1-0.25Jet fuel JP-4** 790-820 C7-C15 20 - 140-270 0.1-0.25Diesel fuel 820-860 C9-C30 20 - 180-380 <0.2Domestic fuel 800-870 C9-C25 20 - 180-370 <0.2Fuel oil 950-995 C20-C40 65 - 300+ <0.01Lubricating oil 860-990 C20-C40 0-60 - 300+ Var.

* This represents the main range. In gasoline constituents up to C15 and diesel fuel constituents upto at least C25 can be detected on a chromatogram. ** Jet fuel JP-4 is fuel used for civil airplanes.

Benzene, toluene, ethylbenzene, and xylenes and polycyclic aromatic hydrocarbons(PAHs), especially naphthalene and its methyl-substituted derivatives, may be present atlevels of parts per million in diesel fuel. The sulphur content of diesel fuels depends onthe source of crude oil and the refinery process. It is regulated by law in a number ofcountries and is usually between 0.05 and 0.5 weight percent.

Additives are used to influence the flow, storage, and combustion of diesel fuel, as wellas to differentiate products, and meet trademark specifications. At room temperature,diesel fuels are generally moderately volatile, slightly viscous, flammable, brown liquidswith a kerosene-like odour. The boiling ranges are usually between 180 and 380°C(Marchal et al., 2003), at 20°C the density is 0.87 to 1.0 g cm-3 and the water solubility is0.2 to 5 mg l-1. The quality and composition of diesel fuel influence the emissions ofpollutants from diesel engines considerably.

Important variables are ignition behaviour, expressed in terms of cetane number, density,viscosity, and sulphur content (Environmental Health Criteria 171, 1996). The cetanenumber is a measure of the tendency of a diesel fuel to knock in a diesel engine. Thespecifications of commercial diesel fuel vary considerably from country to country.

16

In Europe the specifications for diesel fuel for transportation purposes, are given inEuropean standard EN 590 (European Committee for Standardization, 1993). In the USthree qualities of diesel fuel are available commercially: diesel fuel No. 1, diesel fuel No.2, and diesel fuel No. 4 (Table 2.3). Diesel fuel No 1, with relatively high volatility, isrecommended for high-speed engines and for use in cold temperatures. Diesel fuel No. 2is used mainly as automobile fuel. The more viscous diesel fuel No 4 is suitable for low-to-medium-speed engines, such as those in ships (Environmental Health Criteria 171,1996). The physical and chemical properties of the three qualities of diesel fuel areshown in Table 2.3.

2.1.2.4 Toxicity of diesel fuel

Direct dermal exposure to diesel, diesel fuel exhaust exposure, and the inhalation ofdiesel fuel vapours, can have effects on human health. Although only very lowconcentrations of vapours are likely to be produced at room temperature because of itslow volatility. Non-occupational exposure to diesel fuel can occur during manual fillingof fuel tanks. The primary source of dermal exposure is accidental spills, which result inimmediate high levels of exposure but are of short duration. After accidental dermalcontact, anuria, renal failure, gastro-intestinal symptoms, and cutaneous hyperkeratosishave been reported. Toxic lung disease has been observed after accidental ingestion ofdiesel fuel and subsequent aspiration. Persistent productive cough has been reported afterinhalation (Environmental Health Criteria 171, 1996).

As diesel fuel contains trace quantities of benzenes, xylenes, toluene, or n-hexane, and asmall percentage of potentially neurotoxic C9-C12 fractions, there would appear to be onlyminimal human neurotoxicity, or neurobehavioral risk, from dermal or oral exposure tothe raw fuel, or from respiratory exposure to the fuel vapour (Ritchie et al., 2001).

Diesel fuel is more toxic than crude oil to aquatic organisms and plants. The ecotoxicityof diesel fuel is generally attributed to soluble aromatic compounds, but insolublealiphatic hydrocarbons may also be implicated. Of the aromatic compounds,monoaromatics are the least toxic, their acute toxicity increasing with molecular mass upto the four- to five-ring compounds, although these are poorly soluble in seawater. In

17

some animals, e.g. fish and birds, physical coating of the body surface by the fuel canproduce toxicity and mortality (Environmental Health Criteria 171, 1996).

Table 2.3 – Physicochemical properties of diesel fuel (after Environmental Health Criteria171, 1996)

Property Diesel fuelNo. 1

Diesel fuelNo. 2

Diesel fuelNo. 4

Melting point (°C) -34 -18 -29 -9Boiling range (°C) 145 - 300 282 - 338 101 - 588Density (g cm-3) 0.81 - 0.94

(15°C)0.87 - 0.95(20°C)

0.81 - 0.94(15°C)

Viscosity (mm2 s-1) 1.3 - 2.4(40°C)

1.9 - 4.1(40°C)

5.5 - 45.0(40°C)

Vapour pressure (kPa) 2.83 - 35.2(21°C)

2.83 - 35.2(21°C)

2.83 - 35.2(21°C)

Water solubility (mg l-1) ± 5 (20°C) ± 5 (20°C) ± 5 (20°C)Henry’s law constant (Pa m3 mol-1)(20 °C)

6.03-7.5 x 105 6.03-7.5 x 105 6.03-7.5 x 105

n-Octanol-water partitioncoefficient (log KOW)

3.3 - 7.06 3.3 - 7.06 3.3 - 7.06

Soil sorption coefficient (log KOC) 3.0 - 6.7 3.0 - 6.7 3.0 - 6.7

Reported studies on the response of microbial populations to the exposure of petroleumhydrocarbons, in general, show that the total number of micro-organisms as well asmicrobial respiration, increase in spite of an occasional initial drop in numbers.Ammonification and nitrification on the other hand, are found to decrease (Van Eyck,1997).

2.1.3 Petroleum hydrocarbons in soil

2.1.3.1. Behaviour of petroleum hydrocarbons in soil

Petroleum hydrocarbons can be introduced into the environment, by accidental release ona large scale, such as during a tanker disaster, pipeline leaks, blowouts of terrestrial oilwells, or on a smaller scale, from contamination of soil around factories, fuel stations orspillage during fueling and leakage of underground storage tanks (Environmental HealthCriteria 171, 1996).

The hydrophobic organic contaminants (HOCs) can be present as a non-aqueous-phase-liquid (NAPL), either as a film layer on soil particles or as dispersed droplets. Once a

18

NAPL has spilled onto the soil surface, gravity causes the HOCs to migrate into thesubsurface. Transport of the NAPL through the soil matrix, is influenced by the viscosityof the hydrophobic pollutants. Depending on the retention capacity of the soil, the NAPLwill leave behind residual liquid, because migration will only proceed, if the retentioncapacity is exceeded (Guigard et al., 1996). This residual liquid is known as the residualoil saturation or the amount of oil that is permanently retained by the soil. The fraction ofthe pollutant that is sorbed on the soil matrix is not available for transport. If theinfiltration of the HOC stops and the oil saturation in the percolation zone drops belowthe minimum oil saturation value, the existing flow channels will break up into separateoil droplets (Van Eyck, 1997). These NAPL droplets may serve as reservoirs forhydrophobic contaminants. The spread or dispersion of these contaminants may becontrolled by their release from the NAPL phase (Schaefer et al., 1998). Transport of theHOCs will be as dissolved compounds to the groundwater by mass flow or as vapour bydiffusion.

Fuel contamination in soil is often a long-term source of groundwater contamination, dueto leaching of dissolved compounds, transport and dissolution of NAPLs, transport ofdissolved organic matter (DOM), or gaseous transport of volatile organic compounds(VOCs) (Pasteris et al., 2002). NAPLs that are denser than water (DNAPLs) like tars willsink and, through dissolution of some constituents, generate contaminated groundwaterplumes (Luthy et al., 1994). NAPLs lighter than water (LNAPLs) such as petroleum,spread out as the groundwater table is reached and the more polar and low-molecular-mass components slowly dissolve and leach out from the formed slick (Atlas and Bartha,1992), again forming a contaminated groundwater plume. Volatile organic carbon (VOC)vapours may also volatilise into the atmosphere, thereby creating a potential health threatto individuals living in the vicinity of emission sources (Jin et al., 1994).

2.1.3.2 Sorption of HOCs to soil aggregates

Sorption to natural solids is an underlying process affecting the transport, degradationand biological activity of organic compounds in the environment. Although oftenregarded as instantaneous for modelling purposes, sorption may require weeks or manymonths to reach equilibrium (Pignatello and Xing, 1996).

19

Luthy et al. (1997) give a conceptual model of the chemical interactions of hydrophobicorganic contaminants (HOCs) with soils and sediments (geosorbents).

The conceptual model is presented in Figure 2.2. The types of reactivity are distributedacross three principal domains. The first of these is the mineral domain with surfacereactivity attributable to (i) exposed external mineral surfaces at the particle scale andsurfaces within macropores; (ii) interlayer surfaces of swelling clays at the nanometerscale; and (iii) the surfaces within mesopores and micropores of inorganic mineralmatrices. The soft and hard carbon soil organic matter (SOM) components constitute asecond principal domain at the nanometer scale of the composite geosorbent. Adherent orentrapped NAPLs constitute the third domain and may function in this regard as softcarbon organic matter except possibly for highly weathered material or interfacial films.Combustion residue, e.g., soot, is another type of organic matter, which might act thesame as hard carbon. Also shown is natural organic matter that may be accessible toHOCs because of encapsulation.

Sorption of HOCs to the mineral fraction in soil (domain D and E) is typically a linearand reversible process with equilibrium attained essentially instantaneously undercompletely mixed conditions (Luthy et al., 1997). Solution pH or sorbent surface chargedensity does not affect sorption, indicating that there is not a direct interaction betweenthe mineral surface and non-ionic HOCs. It is suggested that HOCs have only a weak andnon-specific interaction with mineral surfaces (Mader et al., 1997).

Sorption to the mineral fraction is usually of minor importance compared to organicmatter sorption (Luthy et al., 1997). However, Löser et al. (1999) showed thathydrocarbons may be strongly sorbed even on coarse grained and organic-free soils bymicroporosity.

The Freundlich isotherm is commonly used to relate geosorbent and aqueous-phase HOCconcentrations of a compound at equilibrium and widely used to quantitatively describesorption of HOCs to soil:

nwds CKC 1= [2.1]

20

where Cs is the equilibrium sorbent phase solute concentration (mg kg-1), Cw is theequilibrium aqueous phase solute concentration (mg L-1), Kd the equilibrium partitioningcoefficient (L kg-1) and 1/n a constant relating to the surface adsorption capacity of thesolid phase (Means et al., 1980).

Figure 2.2 – Conceptual model of geosorbent domains (Luthy et al., 1997)

Linear isotherm models (i.e., 1/n = 1) have been used to describe sorption over limitedconcentration ranges (Chiou et al., 1979; Karickhoff et al., 1979, Means et al., 1980;Schwarzenbach and Westall, 1981). Assuming linearity, the resulting equilibriumcoefficient Kd, can represent a concentration-independent partition coefficient Kp

(L kg-1), commonly used to quantify the distribution of inorganic pollutants between theaqueous phase:

wps CKC = [2.2]

21

When Kp is normalised to the organic carbon content fOC (g g-1) of the soil, the resultingpartitioning coefficient KOC (L kg-1) becomes nearly independent of the properties of thesoils investigated (Walter et al., 2000):

OC

pOC f

KK = [2.3]

Due to the linearity of the sorption isotherm, uptake of neutral organic chemicals by soilis considered a process of partitioning rather than physical adsorption (Chiou et al.,1979). Karickhoff et al. (1979) demonstrated that there is a significant relationshipbetween the octanol-water partition coefficient KOW and the KOC value for a compound.Predictions of the partition coefficients Kp or KOC, made from the solvent-waterdistribution coefficient KOW and/or the aqueous solubility Saq are given in Table 2.4.

Sorption tends, however, to be non-linear for HOC sorption to soil if a wideconcentration range is considered. Moreover Pignatello and Xing (1996) concluded thatinstantaneous equilibrium is inadequate to describe the sorption isotherm because of slowsorption kinetics, and they observed that the measured linear adsorption coefficient Kd, insome systems increased by up to ten-fold between short and long contact times. Thepotential causes of slow sorption are pore and matrix diffusion limitations, activationenergy of sorptive bonds and/or sterical hindrance to adsorption, the former being ofmajor importance (Pignatello and Xing, 1996). Systems which have been polluted for along time, may contain pollutants co-encapsulated with the soil organic matter (SOM) inwhich they were originally associated. The encapsulation of such SOMs and claysurfaces, when it occurs after HOC sorption might trap sorbed molecules within matricesfrom which they cannot readily escape and contribute significantly in some cases to the”aging” phenomenon commonly observed in field samples (Luthy et al., 1997).

Considering this, at present, soil organic matter is proposed as a dual-mode sorbent(Pignatello and Xing, 1996) with a rubbery or amorphous compartment (amorphousSOM, domain A) and a glassy or condensed compartment (dense SOM, domain B)(Gamst et al., 2001). Sorption to the rubbery or amorphous compartment is thenconsidered instantaneous or fast and fully reversible, while sorption to the glassy orcondensed compartment is rate limited by an intra-organic matter diffusion process that

22

controls slow sorption and desorption nonsingularity or hysteresis. (Gamst et al., 2001).Like a polymer, SOM may exhibit a transition from its condensed state to a rubbery stateas temperature increases (Luthy et al., 1997).

Table 2.4 - Predictions of partition coefficients Kp and KOC from octanol-water distributioncoefficient (KOW) and Solubility (Saq, mg L-1)

Equation R2 Reference

KOC = 0.63 KOW 0.96 Karickhoff et al., 1979Log KOC = 1.00 log KOW – 0.21 1.00 Karickhoff et al., 1979Log KOC = log KOW – 0.317 0.98 Means et al., 1980Log Kp = 0.72 log KOW + log fOC + 0.49* 0.95 Schwarzenbach and Westall, 1981Log KOC = 0.62 log KOW + 0.70 0.91 Walter et al., 2000Log KOC = -0.54 log Saq + 0.44 0.94 Karickhoff et al., 1979Log KOC = -0.82 log Saq + 4.070 1.00 Means et al., 1980Log KOC = -0.729 log Saq + 0.001 0.996 Chiou et al., 1983

* for fOC > 0.001,

Underprediction of solute retention indicates the influence of additional retentionmechanisms not traditionally taken into consideration. Adsorption of organic vapour towater surfaces (soil gas – bulk water interface) has been proposed as a mechanismaccounting for much of the additional retention observed in unsaturated systems. The realextent of the gas-water interface in a soil system plays a major role in dictating itsimportance as a source of contaminant retention (Costanza and Brusseau, 2000).

2.1.4 The Flemish statute on soil clean up

Since the Belgian constitutional reforms in 1980, Belgium is a federal state composed ofthree regions: the Flemish Region, the Walloon Region and the Brussels MetropolitanRegion. Regional authorities regulate almost all sectors of environmental policy, with thefederal government having jurisdiction over the transportation of waste, radioactivematerials and the protection of the North Sea. All regions have enacted legislationregulating the emission of air and water pollution and the disposal of hazardous wastes(Soler, 1997).

23

On February 22nd 1995, the Flemish Government issued a decree concerning soilremediation and laid down soil and groundwater quality criteria that are dependent on thedestination of the soil. The Decree grants authorities the power to force a polluter or landowner to remediate contaminated land that poses a serious threat to human health or theenvironment. It provides for the establishment of a Register of Contaminated Land for thepurpose of prohibiting transfers of land to unsuspecting buyers. As such, contaminatedland cannot be sold without a certificate from regional authorities notifying the buyer ofexisting environmental problems.

The Decree distinguished two types of pollution. Recently polluted soil, i.e. soil pollutedafter 29 October 1995, should be remediated if certain soil or groundwater criteria areexceeded. The operator of the company on the site, or the owner or the user of thepolluted site is legally bound to remediate. Historically polluted soil, i.e. polluted before29 October 1995, should be remediated if the pollution forms a serious threat to publichealth or if the polluted site occurs on a list published by the administrative body the"Openbare Vlaamse Afvalstoffenmaatschapij" (OVAM). The persons or companiesresponsible for the pollution are legally bound to remediate the site. If those causing thepollution or existing land owners are insolvent or refuse to conduct cleanup activities, orif no responsible party can be found, regional authorities may conduct cleanup activitiesand attempt to obtain reimbursement at a later date.

Soil remediation is aimed at realising background values for soil and groundwaterquality. If these values cannot be reached by the best available technique, soil andgroundwater quality should be at least better than the intervention values. When even thisis impossible, safety measures and restrictions for use are imposed.

2.1.5 Analysis of petroleum hydrocarbons

Total petroleum hydrocarbons (TPH) are sometimes referred to as mineral oil,hydrocarbon oil, extractable hydrocarbon and oil and grease. There are many analyticaltechniques available that measure TPH concentrations in the environment. No singlemethod measures the entire range of petroleum-derived hydrocarbons. The techniquesvary in the way hydrocarbons are extracted, cleaned up, and detected, and therefore each

24

measure slightly different subsets of the petroleum-derived hydrocarbons present in asample.

Figure 2.3 shows the relationship between boiling range and carbon number for somecommon petroleum products. This figure clearly shows the overlap between carbonranges of different products as well as the overlap in corresponding analytical methods.For example, Figure 2.3 shows that an analytical method designed for gasoline mayreport some of the hydrocarbons present in diesel fuel. The same is also true foranalytical tests for diesel range hydrocarbons that will identify some of the hydrocarbonspresent in gasoline contaminated soils. The definition of total petroleum hydrocarbonsTPH depends on the analytical method used because the TPH measurement is the totalconcentration of the hydrocarbons extracted and measured by a particular method. Thefour most commonly used testing methods include, gas chromatography (GC), infraredspectrometry (IR), gravimetric analysis and immunoassay (Weisman, 1998).

2.1.5.1 Gas chromatography

Gas chromatography techniques separate a complex mixture of organic materials intotheir individual components. The sample extract is injected into a heated chamber, inwhich the mixture is concentrated at the head of a separating column. The mixture is thencarried through the column by an inert gas (mobile phase). As the column is heated thecompounds pass through absorbent materials (stationary phase). From the chromatogramand the digital information contained in the quantification report, compounds containedin the sample can be accurately identified and quantified. The flame ionisation detector(FID) responds to ions produced by the burning of compounds (separated by gaschromatography) in a H2/air flame. FID response depends on the number of ionsproduced by a compound. Since this varies considerably between compound classes, theFID response factors demonstrate this variability (Korda et al., 1997).

GC can be accompanied by Mass spectrometry (MS). In the MS-chamber, compoundsseparated by GC are bombarded by electrons and broken down into characteristicfragments called ions. Compounds are identified on the basis of the electrical currentrequired to ionise the fragments. The relative intensity of this current over all thedifferent masses recorded for a particular compound generates its mass spectrum. The

25

pattern of fragmentation ions in a mass spectrum has a characteristic appearance used todistinguish one compound from another. In addition, the intensity of the current recordedfor one characteristic ion over time gives rise to its mass chromatogram, which is used toquantify the concentration of the compound as it elutes from the gas chromatograph. Thischaracteristic ion is called the quantification ion. The mass chromatograms for all ionsdetected in a sample can be superimposed onto a reconstructed ion chromatogram (RIC)also called a total ion chromatogram. The RIC is a graphic display of the total ionisationcurrent recorded from all mass fragments of all compounds detected throughout theanalysis (Korda et al., 1997).

Carbon numbers

C2 C4 C6 C8 C10 C12 C14 C16 C18 C20 C22 C24 C26 C28 C30

69 402343126 216 449

Gasoline

Jet fuel/Kerosene

Standard solventsNapthas

JP-4Diesel fuel/middle distillates

Lubricating oil, motor oil,

Fuel oils

Boiling point (°C)

Figure 2.3 – Approximate carbon and boiling ranges of product types derived frompetroleum (Weisman, 1998)

Some investigators (Huesemann, 1995) utilise a comprehensive petroleumcharacterisation procedure involving group-type-separation analyses, boiling-pointdistributions, and hydrocarbon typing by field desorption mass spectroscopy to determineinitial and final concentrations of specified hydrocarbon classes.

26

2.1.5.2 IR spectroscopy (Weisman, 1998)

Infrared spectroscopy (IR) measures the vibration (stretching and bending) that occurswhen a molecule absorbs energy (heat) in the infrared region of the electromagneticspectrum. Different functional groups and bond types have different IR absorptionfrequencies and intensities.

IR-based TPH methods measure the absorbance of the C-H bond. The IR absorbance is ameasurement of the sum of all the compounds contributing to the TPH result. IR-basedTPH methods cannot provide information on the type of hydrocarbon contamination. Theextraction solvent for measuring TPH in soil must not contribute any C-H stretching tothe measurement. Carbon tetrachloride has been used for IR-based methods but is inlimited use in the U.S.A. because it is a known carcinogen and reputed to affect the ozonelayer. Tetrachloroethene (also known as perchloroethylene, or PERC) is currently beingused by some American labs. Solvents such as methanol, methylene chloride, or hexaneare not suitable for an IR-based method because they contain C-H bonds.

For all IR-based TPH methods, the C-H absorbance is quantified by comparing it to theabsorbance of standards of known concentration. An assumption is made that thestandard has an aliphatic-to-aromatic ratio and IR response similar to that of the sample.Consequently, it is important to use a calibration standard, which mimics the type ofcontamination.

2.1.5.3 Gravimetric method (Weisman, 1998)

Gravimetric methods measure anything extractable by a solvent, that is not removedduring solvent evaporation, and capable of being weighed. Some gravimetric methodsinclude a cleanup step to remove biogenic material. Those that do have this step, areconsidered TPH methods. Those that do not, are considered oil and grease (O&G)methods. The advantage of gravimetric methods is that they are simple, quick, andinexpensive. Detection limits are approximately 5-10 mg l-1 in water and 50 mg kg-1 insoils. These methods are not especially suitable for measurement of light hydrocarbonsthat volatilise at temperatures below 70-85°C. They are recommended for TPHmeasurement only for very oily sludges, samples containing heavy molecular weighthydrocarbons, or aqueous samples when hexane is preferred as the solvent. Gravimetric

27

methods provide no information on the type of fuel present, nor information about thepresence or absence of toxic compounds, and no specific information concerningpotential risk associated with the contamination.

2.1.5.4 Immunoassay (Weisman, 1998)

Immunoassay methods correlate TPH with the response of antibodies to specificpetroleum components. A number of different testing kits based on immunoassaytechnology are available for rapid determination of TPH. The kits are self-contained,portable systems designed to conduct analytical work in the field. They includecomponents for sample preparation, instrumentation to read assay results, andimmunoassay reagents. Currently, most of these methods measure only aromatics.

Immunoassay is used as a screening technique due to its precision and accuracy beinglower than standard laboratory methods, such as GC/FID or IR. Immunoassaymeasurements may be reported as a range or a single value. Typical detection limits forTPH range from 10-500 mg kg-1 in soil and 200 to 500 µg l-1 in water.

Antibodies are made of proteins that recognise and bind with foreign substances(antigens) that invade host animals. Synthetic antibodies have been developed to combinewith petroleum constituents. The antibodies in the test kit are immobilised on the walls ofa special cell or filter membrane. Water samples are added directly to the cell embeddedwith antibodies, while soils must undergo an extraction process before analysis. A knownamount of labelled compound is added after the sample. The label is typically an enzymewith an affinity for the antibody. The sample compound competes with the enzyme-labelled compound for sites on the antibodies. After equilibrium is established, the cell iswashed to remove non-specific sample or labelled enzyme. Colour development reagentsthat react with the labelled enzyme are added. A solution that stops colour development isadded at a specified time, and the optical density (colour intensity) is measured. With thecolouring agent reacting with the labelled enzyme, samples with high optical densitycontain low concentrations of compounds. Concentration is inversely proportional tooptical density.

28

2.2 Biodegradation of Petroleum hydrocarbons

2.2.1 Introduction

Biodegradation of petroleum ultimately results in CO2, H2O and microbial biomass.Partially oxygenated biodegradation intermediates of hydrocarbons are fatty acids andphenolic substances. Depending on the extraction and measurement technique, some ofthese intermediates may or may not be assessed as residual petroleum material. Somepetroleum-carbon may become part of soil humus, directly or indirectly via the microbialbiomass (Atlas, 1984).

2.2.2 Biodegradability of petroleum

When an oil spill occurs, a broad range of hydrocarbon substrates become simultaneouslyavailable. Predictably, interactions between these substrates occur, ranging from thesparing of iso-alkanes in the presence of n-alkanes, to the cometabolic oxidation of non-substrate hydrocarbons in the presence of substrate hydrocarbons.

Compared to an aquatic environment, the soil environment will favour biodegradationbecause of two typical soil features. First, the infiltration of low-molecular-weighthydrocarbons into deeper soil layers minimises their evaporation and thereforebiodegradation rather than evaporation becomes the principal mechanism for theirremoval as long as oxygen remains available. Second, the great absorption capacity ofsoil for both polar and non-polar materials reduces the effective toxicity of thecontaminant hydrocarbons (Atlas, 1984).

Huesemann (1995) provided evidence that the extent of hydrocarbon biodegradation islargely affected by molecular composition of the soil contaminant. From petroleumcomponents, n-alkanes, n-alkylaromatic, and aromatic compounds of the C10-C22 rangeare the least toxic and most readily biodegradable

The n-alkanes, alkylaromatic and aromatic hydrocarbons in the C5-C9 range haverelatively high solvent-type membrane toxicity. They are biodegradable at lowconcentrations by some microorganisms, but in most environments they are removed byvolatilisation rather than by biodegradation. Gaseous n-alkanes (C1-C4) are biodegradable

29

but utilised only by a narrow range of specialised hydrocarbon degraders. The n-alkanes,alkylaromatics and aromatic compounds above C22, have low toxicity, but their physicalcharacteristics, such as, their extremely low water solubility and their solid state atphysiological temperatures are unfavourable for biodegradative transformations.

Branched alkanes and cycloalkanes of the C10-C22 range are less biodegradable than theirn-alkane and aromatic analogues. Branching creates tertiary and quaternary carbon atoms

that constitute a hindrance to β-oxidation. Geerdink et al. (1996) indicated that the lag

phase for degradation of branched alkanes in the presence of linear alkanes is caused by ahigher affinity of the linear alkanes for one or more of the enzyme systems involved.

The biodegradation of cycloalkanes requires synergistic co-operations of two or moremicrobial species, and cycloalkanes of C10 and below have high solvent-type membranetoxicity (Atlas, 1984).

Generally, the rate of degradation of PAHs is inversely proportional to the number ofrings in the PAH molecule. Thus, the lower weight PAHs, are biodegraded more rapidlythan the higher weight compounds. Four and five-ring PAHs are very recalcitrant due totheir low water solubilities and the resonance energies of their structures (Cerniglia,1992).

2.2.3 Factors influencing biodegradation of petroleum hydrocarbons

Petroleum biodegradation is influenced by factors affecting the rate of microbial growthand enzymatic activities (Atlas, 1984). The overall rates of hydrocarbon degradation arelimited by soil water content, oxygen, temperature, pH and inorganic nutrients. Othervariables that influence the rate of degradation are contaminant bioavailability or abiotictransformations (Gallego et al., 2001).

2.2.3.1 Soil water content

An extensive discussion on the effect of soil water on biodegradation of petroleumhydrocarbons will be given in Chapter 4.

30

2.2.3.2 Oxygen

The initial steps in the catabolism of aliphatic, cyclic, and aromatic hydrocarbons bybacteria and fungi involve the oxidation of the substrate by oxygenases, for whichmolecular oxygen is required. Moreover, oxygen is used by oxidases as the terminalelectron acceptor in the respiratory chain. The availability of oxygen in soils depends onrates of microbial oxygen consumption, the type of soil, soil water content, and thepresence of utilisable substrates, which can lead to oxygen depletion (Leahy and Colwell,1990).

Oxygen can be supplied by tillage (Rhykerd et al., 1999), aeration with air (known as"bioventing") (Lee and Swindoll, 1993), pure oxygen (Morgan and Watkinson, 1992) andH2O2 (Hinchee et al., 1991) or by a commercial Oxygen Release Compound (ORC)(Fischer et al., 2001). When H2O2 decomposes in the soil, it releases water and molecularoxygen, but if the rate of oxygen generation exceeds the rate of oxygen utilisation,oxygen may escape in gaseous form because of its limited solubility in water. Gaseousoxygen may form bubbles, which may be transported in groundwater resulting ininefficient oxygen delivery in the contaminated zone (Hinchee et al., 1991) and poreblockage (Morgan and Watkinson, 1992). Moreover, H2O2 is toxic to microorganisms.However, it is claimed that adaptation to increasing concentrations of hydrogen peroxidecan occur and that concentrations of several hundreds mg l-1 can be employed (Morganand Watkinson, 1992).

2.2.3.3 Temperature

Hydrocarbon biodegradation can occur over a wide range of temperatures, withpsychrotrophic, mesophilic and thermophylic hydrocarbon-utilising microorganismsisolated (Atlas, 1984). Temperatures can have a marked effect on the rates ofhydrocarbon degradation. There are seasonal shifts in the composition of the microbialcommunity, which can be reflected in the rates of hydrocarbon metabolism at a giventemperature.

Microbial respiration has been reported at temperatures as low as – 5°C in arctic andsubarctic soils (Clein and Schimel, 1995). In soil, particularly in northern industrialisedcountries in Europe and North America, the soil temperature during a large part of the

31

year is too low for efficient microbial degradation of soil contaminants. The same mayalso be true for deeper soil layers in other parts of the world. Rates of degradation aregenerally observed to increase with increasing temperature, this is believed to be a resultprimarily of increasing rates of enzymatic activity, or the "Q10" effect (Atlas and Bartha,1992). The Q10 factor is the factor by which the enzymatic activity increases, for a 10degrees increase in temperature.

The optimum temperature for biodegradation of petroleum products has been found to be

35°C (Yeung et al., 1997) or in the range of 20-30 °C (Atlas and Bartha, 1992) or 30-

40°C (Leahy and Colwell, 1990). However, Dibble and Bartha (1979) reported thehighest biodegradation rate for oil sludge at 20°C and Beaudin et al. (1999) at 23°C forweathered hydrocarbon contaminated soil. Thermophilic degradation has been observedat 50°C (Beaudin et al., 1999).

A number of studies have been conducted on the fate of oil in cold Arctic and sub-Arcticsoils (Ferguson et al., 2003). Coulon and Delille (2003) indicated that a temperatureincrease, can induce an order of magnitude increase in hydrocarbon degrading microbialnumbers in diesel and crude oil contaminated sub-arctic soils.

It is apparent that the influence of temperature on hydrocarbon degradation is morecomplex than simple considerations of Q10 values. The effects of temperature areinteractive with other factors, such as the quality of the hydrocarbon mixture and thecomposition of the microbial community (Atlas, 1984).

2.2.3.4 pH

Most bacteria prefer a neutral pH and unlike many fungi, most bacteria cannot endureacidic conditions for long periods of time. There are few reports on the effects of pH onthe biodegradation of oil in soil (Van Eyk, 1997). Soil pH is important in determining themicrobial population that will be active. Hydrocarbons are mineralised most rapidly at apH between 6.5 and 8.0 (Gallego et al., 2001).

If soil polluted with petroleum hydrocarbons, is low in calcite, the production of CO2 andpossible organic acids following the biodegradation of petroleum constituents may result

32

in an increase in acidity, which will eventually significantly reduce contribution of soilbacteria to the bioremediation process. Although fungi will in part gradually replace thebacteria, the overall rate will be lower than at neutral pH (Atlas, 1984). The positiveeffect of liming has been demonstrated by Dibble and Bartha (1979). The highest pHtested (pH 7.8) in their experiment, was suggested to be close to the optimum forbiodegradation.

2.2.3.5 Nutrients

The importance of nutrients to microbial processes has long been known. Nitrogen (N) isrequired in amino acids, and phosphorus (P) is involved in energy transport as adenosinetriphosphate. Compositional analysis of microbial biomass indicates that carbon (C), Nand P are present in the ratio of ca. 106:16:1, respectively. Contaminated soils that haveintrinsically low N and P values will require nutrient additions to allow sufficientincrease in biomass for environmentally significant hydrocarbon degradation to occur. Toestimate how much nutrient amendment is required, the so-called Redfield ratio(106:16:1) is commonly cited as an optimal C/N/P target. While this theoreticallydetermines the nutrients required for total hydrocarbon conversion to biomass, theapplication of the Redfield ratio to mineralisation experiments ignores that the majorityof the carbon mineralised, is generally converted to CO2 and is then lost from the system.Also inorganic species of nitrogen can be lost from the system through nitrification-denitrification processes. (Ferguson et al., 2003). In practise C/N ratios ranging from 14:1(Møller et al., 1996) up to 560:1 have been proposed as suitable or ‘optimal’ forbiodegradation.

The results of nutrient amendments to hydrocarbon degradation rates are varied.Seklemova et al. (2001) demonstrated inhibition or no significant effect. Gallego et al(2001) reported a higher proportion of specialised micro-organisms in the soil samplessupplemented with inorganic nutrients. Whereas other authors (Brag et al, 1994; Mohn etal., 2001; Ferguson et al., 2003) suggested higher degradation rates in soil after nutrientamendment.

Important differences, between two common mineral fertilisers, ammonia and nitratehave been demonstrated. In poorly buffered media, acid production associated with

33

ammonia metabolism can reduce the culture pH to a level that inhibits oil biodegradation.Conversely, when the culture pH is controlled, biodegradation starts more quickly in thepresence of ammonia than in the presence of nitrate (Wrenn et al., 1994).

Due to the intimate relationship between nutrient and water concentrations in the soilmatrix, maintenance of sufficient N and P, must be accompanied by an adequate soilwater content (Ferguson et al., 2003).

2.2.3.6 Bioavailability of petroleum hydrocarbons

The term bioavailability designates the fraction of a chemical that is available for uptakeand/or transformation by living organisms (Scow and Johnson, 1997). Two mechanismscontrol the bioavailabilty of oil. At higher concentrations, the oil is mainly present as anon-aqueous phase liquid (NAPL), the bacteria are present near the NAPL-waterinterfaces (Atlas, 1984; Li et al., 2000), and bioavailability is controlled by NAPL-watersolubilisation (De Jonge et al., 1997). At lower concentrations, the contaminant is nolonger present as a NAPL. In this stage sorption/desorption processes affectbioavailability of HOCs in the subsurface in two important ways. Firstly, sorption causeshigh contaminant concentrations in impermeable zones, to which bacterial access isobstructed. When hydrocarbons associate with clay surfaces and become immobilised inthe interlayer space, they are physically inaccessible as the dimension of bacterial cells ismore than two orders of magnitude larger than the interlayer separation (Theng et al.,2001). Secondly, because desorption and diffusion must occur before degradation canproceed, the overall rate of bioremediation can be limited or even controlled by thesemass transfer processes (Robinson et al., 1990). The aqueous solubility is expected tohave a large impact on the diffusion rates (De Jonge et al., 1997). Factors influencingsorption and desorption (see 2.1.3.2) like organic matter content and composition, pH,ionic strength, temperature and contact time also affect bioavailability.

After extensive weathering, petroleum hydrocarbons occur in the environment as tarballs. Limited bioavailability of tar balls, is due to their chemical structures, and anunfavourable surface area-to-volume ratio (Atlas, 1981).

34

Since availability of increased surface area should accelerate biodegradation (Atlas,1981), (bio)surfactants have been used to enhance the bioavailability of petroleumhydrocarbons. Surfactants are organic molecules that usually consist of a hydrophobicpart and a hydrophilic part. The hydrophilic part makes surfactants soluble in water,while the hydrophobic part causes them to concentrate at interfaces. In the presence of aNAPL, surfactants concentrate at the liquid-liquid interface, reducing the interfacialtension, causing dispersion of NAPL droplets. Furthermore, the presence of micelles,formed by surfactants, leads to an increase in the apparent solubility of HOCs, alsoreferred to as ‘solubilisation’. Disadvantages of using surfactants in bioremediation arethe potential degradation product and the toxic effects on soil bacteria. Therefore, the useof biologically produced surfactants, or biosurfactants, which occur naturally in soil, maybe more acceptable from a social point of view (Volkering et al., 1998). Biosurfactantsproduced by microorganisms actively metabolising petroleum hydrocarbons are morelikely to increase the bioavailability of oil than synthetic surfactants that simply increasethe surface area by affecting the droplet size (Uraizee et al., 1997).

2.2.4 Measurements of petroleum biodegradation

Numerous authors have observed oil biodegradation by essentially indirect methods, suchas, an increase of oil-degrading micro-organisms (Raymond et al., 1976) and/or acharacteristic change in oil composition (Li et al., 2000).

Mineralisation of hydrocarbons in soils may also be monitored by measuring the excessO2 consumption or CO2 production, as compared with the untreated soil sample of similarsize (Li et al., 2000). For long-term measurements the monitoring of CO2 production,either static or in flow-through devices, tends to be more reliable. Carbon dioxide may bemeasured titrimetrically, by infrared detectors or by gaschromatography. Gas exchangemeasurements do not destroy the sample and thus lend themselves to continuousmonitoring. Such monitoring may be supplemented by periodic solvent-extraction of areplicate sample for analysis of hydrocarbon biodegradation (Atlas, 1984).

Mineralisation measurements can be obscured by an anomalous burst of CO2 evolutionupon remoistening of air-dried soil. This burst of CO2 represents primarily mineralisationof soil biomass killed by the drying process. Physical changes caused by drying and

35

remoistening may contribute to some desorption and greater availability of non-biomasssoil organic matter (El-din Sharabi and Bartha, 1993).

A refinement of the mineralisation measurement is the use of radiolabelled hydrocarbons,usually 14C. The advantage is greater precision and specificity, since the 14CO2 from theradiolabeled hydrocarbon can be easily distinguished from the background CO2 evolutionof the soil. Complex crude petroleum may be spiked with representative radiolabelledhydrocarbons, and the resultant 14CO2 evolution can be interpreted to indicate the overallrate of petroleum mineralisation (Atlas, 1984). If appropriately 14C labelled test materialis available, the measurement and their interpretations are relatively straightforward. Thisis unfortunately balanced by the technical difficulty and expense involved in obtaininglabelled material (El-din Sharabi and Bartha, 1993).

As compared with laboratory studies, measurement of hydrocarbon biodegradation in thefield is more difficult and less reproducible. Some authors measured CO2 evolution in thefield, but most workers rely on solvent extraction and analysis of the residual oil by oneof the described techniques. The extracted material includes hydrocarbon biodegradationintermediates, therefore, some authors refer to it as extractable organic carbon rather thanoil or hydrocarbon. In field experiments, it is usually not possible to distinguish betweenoil removal by biodegradation and by abiotic mechanisms such as evaporation, runoffand leaching. Difficulties in even application and representative sampling are inherent tofield experiments (Atlas, 1984).

Madsen et al. (1991) found high protozoa numbers indicative of rapidly growingpopulations of bacteria in situ. To the extent that their prey are growing on contaminantcompounds, the elevated protozoan biomass reflects in situ biodegradation activity.

2.2.5 Petroleum hydrocarbon biodegradation rates in soil

Oil biodegradation rates in soil are of great interest because they determine the timecourse of the recovery of accidentally contaminated land. In ex-situ bioremediation ofsoil contaminated soils, the rate of biodegradation determines residence time of thecontaminated soil in the treatment plant and thus the capacity of the treatment plant foroil contaminated soil (Atlas, 1984).

36

Generally, the rates of biodegradation of individual hydrocarbons are believed to berelated, to their chemical structure, degree of aromaticity, concentration andphysicochemical properties. Biodegradation of a component in an industrial mixture alsodepends to a large extent, on the mixture, pre-exposure conditions and concentration(Cerniglia, 1992).

However, any attempt to integrate information on rates is hampered by tremendousdiversity in measurement techniques, soils, environmental conditions, as well as, thequality and quantity of the petroleum product. Since our study will focus on dieseldegradation, we report diesel mineralisation rates, found by different authors in the rangeof 2 to 26 mg hydrocarbon (HC) per kg soil per day in Table 2.5. The rates are assumedto be linear, zero-order kinetics.

The diesel biodegradation rates listed in the Table 2.5 represent essentially ultimatebiodegradation of hydrocarbons to CO2, H2O, microbial biomass and humus. They arebased on the decrease of solvent-extractable material or on evolution of CO2.

Biodegradation rates in bioventing demonstrations have ranged between 1 and 20 mg kg-1

day-1 (Lee and Swindoll, 1993).

Table 2.5 - Rates of diesel mineralisation in soilMineralisation ratemg HC kg-1 soil day-1

Fertiliser Inoculum Venting Reference

6 - 10 - - - Møller et al., 19960.2 - 6 - - + Møller et al., 19960.7 - 23 + - + Møller et al., 19961.1 - 26 + + + Møller et al., 19962.51 - BC - Marquez-Rocha et al. (2000)3.11 - P. ostreatus - Marquez-Rocha et al. (2000)4.28 - BC + P. ostreatus - Marquez-Rocha et al. (2000)2.4-6.5 + - - Davis et al. (2003)

BC = bacterial consortium

37

2.2.6 Modelling biodegradation of petroleum hydrocarbons in soil

Extensive studies have been reported on bacterial growth, microbial uptake andtransformation kinetics and many of these investigations were designed to serve asmodels to predict what will occur in natural ecosystems. Since these studies were mostlycarried out in batch growth or continuous culture growth, with only a small number ofdifferent microbial species, the results can not easily be translated to the prediction ofchemical fate in natural ecosystems. Many factors such as substrate composition, varietyand toxicity, pH, temperature, sorption, solubility, mass transport limitation, diffusion,predation of bacteria by protozoa and the interaction between all these potentiallyimportant factors, may severely reduce the practical value of attempts to predict themineralisation kinetics of a particular pollutant. As a result, few rate constants weremeasured using natural soils and indigenous bacterial populations (Van Eyck, 1997).

According to Scow and Johnson (1997), the biodegradation of a sorbed chemical isactually a coupled process. This process includes a biological component, such as, themetabolism of the chemical and a physical/chemical component, being, the distributionand movement of the chemical in the physical environment in relation to the microbialpopulation able to degrade it. The relative importance of these processes depends on howstrongly sorbed and how rapidly degraded the particular compound is in a given soil.

2.2.6.1 Microbial kinetics

A limited number of kinetic expressions have been developed to model metabolism orsubstrate disappearance. The most commonly used forms are the Monod (for growth-linked metabolism) and Michaelis-Menten (for non-growth- linked metabolism)equations, and the simple expressions of first- or zero-order kinetics that can be derivedfrom both equations (Table 2.6) (Scow and Johnson, 1997).

Monod kinetic parameters and the effect of substrate interactions to Monod kineticparameters have been determined for various microorganisms growing in liquid cultureson volatile organic carbons (VOCs) or n-alkanes. All this work was based onmicroorganisms that have been isolated from aquatic systems such as, sewage or aquifersediments, and were cultivated in the laboratory, usually in liquid media at relatively highcarbon substrate concentrations (Höhener et al., 2003).

38

If large amounts of petroleum are spilled on the soil, soil bacteria could, in principle,grow under conditions of excess substrate (Van Eyck, 1997). However, autochtonous soilmicroorganisms have to cope with carbon limitation, exhibit moderate to low specificactivity and may comprise a large number of inactive (dormant) cells. Moreover, in theunsaturated zone, microbial populations generally live attached to surfaces, which mayresult in drastically reduced substrate availability (Höhener et al., 2003).

Table 2.6 – Monod-derived Biodegradation Kinetics (Scow and Johnson, 1997)Model Equation

Monod

)(max

CKYBC

dtdC

s += µ Eq. [2.4]

Michaelis-Menten

)(max

CKCV

dtdC

m +=

Eq. [2.5]

First orderCk

dtdC

1=Eq. [2.6]

Zero order0k

dtdC =

Eq. [2.7]

Note. C, substrate concentration (mg L-1); t, time (days); µmax, maximum specific growth rate

(day-1), B, biomass concentration (mg biomass L-1), Y, yield coefficient (mg L-1 biomass producedper mg L-1substrate degraded); Ks half saturation constant for growth (mg L-1); Vmax, maximumreaction velocity (mg L-1 day-1), Km, Michaelis-Menten constant (mg l-1); k1, first-orderbiodegradation rate constant (L day-1); ko, zero-order biodegradation rate constant (mg L-1 day-1).Units of mg L-1 refer to substrate concentration unless otherwise specified.

2.2.6.2 Biodegradation models

Scow and Johnson (1997) give an extensive overview of coupled sorption/desorption andbiodegradation models.

Simkins and Alexander (1984) proposed six kinetic models derived from the Monodequation, incorporating only the variables of substrate concentration and cell density andused these models to fit mineralisation rates of [14C]benzoate by an induced population ofPseudomonas sp. at different initial substrate concentrations. The limitations of these six

39

models are the assumption of a constant cell yield, Y, over time and their inapplicabilityto accommodate significant induction periods (Van Eyck, 1997).

Brunner and Focht (1984) proposed a single deterministic model, which does not demanda prior assignment of a lag phase. This model contains only two interdependent constantsin contrast to four in the Monod equation and is suitable for substrate consumption with(pseudo-second order) or without (pseudo-first order) growth. Substrate utilisation isexpressed as carbon dioxide production, which can be easily measured. Implicit in thismodel is the assumption that the carbon flow into biomass, or into transientintermediates, is relatively small. An interesting conclusion from these modelling resultsis that all data, except that for irradiated soils, fits better by assuming that growth waslinear, rather than exponential. Linear growth would be caused by the fact that contrary togrowth in fermentors, growth in soil is limited by diffusion of both substrate and nutrientsas a direct result of their occurance within the soil matrix (Van Eyck, 1997).

Many authors have observed a short lag phase, followed by an exponential increase in themicrobial degradation of hydrocarbons (Geerdink et al., 1996; Harms, 1996; Löser et al.,1999). The end of the exponential growth phase, marks the moment when the demand ofthe microbes for a specific compound, is greater than its maximum mass transfer. Thisholds especially true for compounds that are consumed to a large extent, as well as, thosewhich have a low water solubility, such as, oxygen and hydrocarbons (Löser et al., 1999).

In mathematical modelling of the reactive transport of VOCs in the unsaturated zone,typically first-order kinetics have been used to represent biodegradation (Höhener et al.,2003). First–order reactions are popular because of simplicity. They assume constantbiomass, but do not reflect biological phenomena such as, dependence on substrateconcentration, inhibition or preferential substrate utilisation (Schirmer et al., 1999).

40

2.3 Gas transport through soil

2.3.1 Introduction

The transport of gases through the soil air space and the exchange of gases between thesoil and the atmosphere are important aspects in bioremediation of the unsaturated zone.Two important gases in bioremediation are oxygen, which is consumed by the soilmicroorganisms when petroleum hydrocarbons are degraded and carbon dioxide, whichis produced in soil by the same processes.

2.3.2 Gas conservation equation

The gas conservation equation in differential form is given by (Jury et al., 1991)

0=+∂∂

+∂

∂g

gT Sz

ft

C [2.8]

where CT (mol m-3) is the total solute concentration, fg the soil gas flux (mol m-2 day-1), Sg

is the gas reaction loss rate (mol m-3 soil day-1), z is the height (m) and t is time (days).For an insoluble gas, all of the solute mass resides in the soil air phase, and the totalsolute concentration CT reduces to

gT CC ε= [2.9]

where ε is the air-filled porosity (m3 soil air m-3 soil) and Cg is the soil gas concentration

(mol gas m-3 soil air). With Eq. [2.9] the gas conservation equation Eq. [2.8] reduces to

( ) 0=+∂∂

+∂∂

gg

g Sz

fC

tε [2.10]

Before Eq. [2.10] can be applied to solve gas transport problems, a gas flux law must bespecified, relating fg to the gas concentration Cg.

41

2.3.3 Gas convection in soil

In principle gases in soil can be displaced by bulk movement of the soil air phase inresponse to differences in the total air pressure. Within the air phase of the soil, thesepressure changes could be induced by soil temperature changes, barometric pressurefluctuations in the atmosphere above the surface, wind blowing over the soil surface, andinfiltration. Passive soil venting or barometric pumping has been is used as a remediationtechnology for VOCs (Zhang et al., 2004).

2.3.4 Gas diffusion in soil

Since convection is usually negligible except under the special circumstances asmentioned above, the major mechanism of gas transport in soil is diffusion of vapourwithin the soil air space (Jury et al., 1991).

2.3.4.1 Free gas diffusion

In free air, the gas diffusion flux fg is expressed by Fick’s law of diffusion:

zC

Df gFg ∂

∂−= 0 [2.11]

where D0F (m2 day-1) is the Fickian gaseous diffusion coefficient in free air. Its valuedepends on the chemical diffusing in the air and on the temperature and air pressure (Birdet al., 1960).

Fick’s First law is only an approximation in the case of gas-phase diffusion. Frommolecular statistics, a more accurate equation can be derived for diffusion of compound i

in a gas mixture of ν components (Jaynes and Rogowski, 1983)

∑≠=

−=−

υ

ijj ij

ijjii

DYfYf

dzdY

RTp

1

[2.12]

42

where P is total pressure (Pa), R, the universal gas constant (= 8.3144 J mol-1 K-1), T istemperature (K), Yi is the mole fraction of compound i, fi is the molar flux of compound i(mol m-2 day-1) and Dij is the binary diffusion coefficient of compound i and j (m2 day-1).

Marrero and Mason (1972) have compiled an extensive tabulation of binary diffusioncoefficients taken from literature, based on the Chapman-Enskog equations and extensiveexperimental relations for the first approximation of over 70 gas pairs. Jaynes andRogowski (1983) present equations for the calculation of Dij for pairs of commonatmospheric gases, as a function of temperature:

TcTbaijD 111 )ln()ln(ln −+=

[2.13]

Diffusion coefficients for gas pairs of interest in soil research are tabulated in Table 2.7.The temperature range over which these coefficients were found and the uncertaintylimits in Dij at the lower end of the temperature range are also given in Table 2.7.

Another equation for calculating Dij pairs of gases as a function of temperature andpressure was proposed by Fuller et al. (1966)

( )23/13/1

2/175.1310

ji

ijij

ddp

MTD

+= − [2.14]

with

ji

jiijM

ϖϖϖϖ +

= [2.15]

where T is temperature (Kelvin), p is total pressure (bars), and di is the diffusion volume

of the ith molecule obtained by summing atomic diffusion volumes and ϖi is the

molecular weight of component i (g mol-1). Table 2.8, taken from Reid et al (1987),provides data for obtaining di.

43

Table 2.7 – Coefficients for calculating Dij for gas pairs of interest in soil research (Jaynes andRogowski, 1983)

Gas couple a × 107

M2 kPa s-1 K-1b c

KT range Uncertainty ‡ limits

%Ar-CH4 0.792 1.785 0.0 307-10-4 3Ar-N2 0.913 1.752 0.0 244-104 2Ar-O2 0.987 1.736 0.0 243-104 3Ar-air 0.926 1.749 0.0 244-104 3Ar-CO2 1.76 1.646 89.1 276-1800 3CH4-N2 1.01 1.750 0.0 298-104 3CH4-O2 1.68 1.695 44.2 294-104 3CH3-air 1.04 1.747 0.0 298-104 3N2-O2 1.14 1.724 0.0 285-104 3N2-H2O 0.188 2.072 0.0 282-373 4N2-CO2 3.18 1.570 113.6 288-1800 2O2-H2O 0.191 2.072 0.0 282-450 7O2-CO2 1.58 1.661 61.3 287-1083 3Air-H2O 0.189 2.072 0.0 282-450 5Air-CO2 2.73 1.866 102.1 280-1800 3H2O-CO2 9.33 1.590 307.9 296-1640 10CO2-N2O 0.284 1.500 0.0 195-550 3

‡ Uncertainty in Dij term at lower end of listed temperature range.

2.3.4.2 Diffusion in porous media

Equation [2.11] and equation [2.12] will overestimate the flux of gas through porousmedia, like soils, for three reasons. Firstly, gas must diffuse through a longer path lengthto get from one point to the other. The square of the ratio between the length of thetortuous path between two points and a straight line connecting these points, is called thetortuosity.

Secondly, the cross-sectional area available for flow is reduced by solid and liquidbarriers. The ratio between the diffusivity and the diffusivity calculated by taking theaverage cross-section of the pores is called the constrictivity. And thirdly, diffusion only

occurs in a fraction of the soil, the air-filled pore space ε.

44

Table 2.8 – Diffusion volumes and structural elements (from Reid et al., 1987)Diffusion Volumes

di for simple moleculesHe 2.67Ne 5.98H2 6.12N2 18.5O2 16.3CO 18.0CO2 26.9NO 35.9NH3 20.7H2O 13.1

∆d Structural Increments (di = Σ∆d)

C 15.9H 2.31O 6.11N 4.54Aromatic ring -18.3Heterocyclic ring -18.3Cl 21S 22.9

Currie (1960) introduced the relative effective diffusivity οr diffusibility ζ as the ratio

between the diffusion coefficient in soil air Deg (m2 day-1), also called the effectivediffusivity and the diffusion coefficient in free air D0g (m2 day-1):

ετδδε =

=

2

0 eg

eg

ll

DD

[2.16]

where l is the linear distance across the porous medium (m soil), le, the tortuous path

length across the pore space (m air), τ, the tortuosity (-), δ, the constrictivity (-) and ε, the

air-filled porosity (m-3 soil air m-3 soil).

45

2.3.4.3 Equations for relative effective diffusivity in soil

Due to the importance of diffusion in the gaseous phase, several investigators haveattempted to find a relationship between the diffusivity and the volumetric soil air

content, ε (e.g. Penman, 1940; Millington and Quirk, 1961; Troeh et al., 1982; Currie,

1961). It is generally assumed that the ratio Dsg/D0g is independent of the nature of thediffusing gases (Curie, 1984). Shimumara (1992) investigated this for three differentnonreactive tracers (N2, CH4, and H2) and found good agreement. Theoretically, it will

therefore be sufficient to determine the Dsg/D0g-ε relationship for one gas and use that to

determine the behavior of other gases knowing only their respective diffusion coefficientsin free air.

The simplest assumption is that ξ is a linear function of the air-filled pore space ε :

ε20

aDD

g

eg = [2.17]

where a2 is a constant. Penman (1940) studied the diffusion of carbon disulfide through

packed soil cores with a range of 0.195 < ε < 0.676 and recommended 0.66 as an average

value for a2. An other and probably the most commonly used gas diffusion model in soilsis the Millington-Quirk model (Millington and Quirk, 1961)

2

310

φεξ = [2.18]

where φ is total soil porosity (m3 void m-3 soil). The two above models have been widely

used due to their simplicity and minimum parameter requirements. However, the Penman

model has been found to greatly overestimate ζ (Sallam et al., 1984; Jin and Jury, 1996)

and the Millington-Quirk model slightly underestimated the ratio Ds,g/Da,g (Sallam et al.,1984).

By fitting to his experimental data, Curie (1961) found the following relation for beds ofsolid particles

46

( )41 SDD drygeg −= [2.19]

where S is the degree of saturation or water filled volume fraction of the total pore space

(S = θ / φ, m3 water m-3 void) and Ddryg is the effective diffusion coefficient for the dry

bed (m2 day-1).

Substituting Ddryg by Eq. [2.16] gives

( )4

01 S

S−

==

ετδξ [2.20]

For beds of porous particles Curie (1961) reports that the relation

4

'

=

pφεξξ [2.21]

fitted his experimental data in the moisture content range where the larger (intercrumb)

pores were drained. The φp value is the inter-aggregate porosity (m3 inter aggregate void

m-3 soil) and ξ' is the effective diffusivity for the bed when the intercrumb pore space is

completely drained and the pores in the crumbs are water-filled. No such relationship wasfound, over the range in which the smaller pores (crumb pores) were drained.

Troeh et al. (1982), combining both the linear (Penman, 1940) and curvi-linear(Millington and Quirk, 1961) relations, proposed a new empirical model of the form:

v

uu

−−=

1εξ [2.22]

The parameter u, represents blocked pores and ineffective passages of pore space, while vcontrols the curvature of the line representing the equation in a graph. The following

limits are given for the parameters u en v, 0 ≤ u < 1, u ≤ ε ≤ 1 and 1 ≤ v ≤ 2.

47

Xu et al (1992) give an overview of several models and found that their data were bestfitted by exponential relationships of the form given by Currie (1961) and Troeh et al.(1982).

For aggregated media Millington and Shearer (1971) describe the effective diffusivitytrough the gas-filled pores in the bed by

( ) ( )( ) ( )[ ]

( ) ( ) ( ) ( )[ ]( ) ( ) 3

32

1

32

1

22

222

2

222

2

1

11

11x

PPPx

PPPPx

P

x

SA

AAA

xPPPP

xP

x

SA

AAA

S

S

Sθφ

θφθφφφφθφ

θφθφφφφθφ

ξ −−+

−−−+−

+−−

−−−−

+−−

= [2.23]

where the exponents x1, x2 and x3 are given by equations:

11112

=

+−

−+

+−

x

SA

AAx

SA

AA

φφθφ

φφθφ

[2.24]

( ) 11 222 =−+ xP

xP φφ [2.25]

( ) ( )[ ] 11 332 =−−+− xPP

xPP θφθφ [2.26]

where φS = 1 – φ, is volume solid per total bed volume, φP, the interaggregate porosity, i.e.

volume pores between the particles per total bed volume, φA, the intra-aggregate porosity,

i.e. volume pores within the particles per total bed volume, SP, the water-filled volumefraction of the interaggregate pore space, SA, the water-filled volume fraction of the intra-

aggregate pore space, θp = φPSP and θA = φA SA.

To use the estimation method for aggregated media from Millington and Shearer (1971)

(Eq. [2.23]) one must know, not only the different porosities, φP and φA, but also the

water distribution between the two pore classes, for example SP and SA.

At very high moisture contents (saturation S = θ / φ > 0.95) diffusion through the water-

filled pore space contributes significantly to the total diffusional transport. The relation

48

proposed by Millington and Shearer (1971) (Eq. [2.23]) was modified to include thediffusion through the water-filled pore space by Colin and Rasmuson (1988). Assumingthat the diffusion through the water-filled pore space only affects the total effectivediffusivity at very high moisture contents, i.e., almost saturated media. Collin andRasmuson (1988) give the following expression for the total effective diffusivity:

100

0

0 =

+=

Sl

el

g

lH

g

egT D

DDD

KSDD

ξ [2.27]

where the first term of the right hand side is the diffusion in air-filled spaces (Eq. [2.23]),the second term is the diffusion in the water-filled spaces, S is the degree of saturation(m3 water m-3 void), D0l is the diffusion coefficient in water and D0g is the diffusioncoefficient in free air (m2 day-1).

Comparing gas diffusivity models with measured data for a number of differentlytextured sieved and repacked soils, Jin and Jury (1996) concluded that the hithertooverlooked Millington and Quirk (1961) model:

32

2

φ

εξ = [2.28]

best described the measured data as compared to the classical models.

In dry (void of water), sieved and repacked porous media Moldrup et al. (2000b) foundthat gas diffusivity was best described by the Marshall (1959) model

5.1εξ = [2.29]

In completely dry soil, ε will equal the total porosity, φ. Adding a linear reduction term

(ε/φ, where φ is soil total porosity) to account for water-induced changes in air-filled pore

shape and configuration in a wet compared with a completely dry soil, the water-inducedlinear reduction (WLR) model

49

φεεξ 5.1= [2.30]

was found to accurately describe diffusivity in sieved and repacked soils at different soilwater contents and total porosities (Moldrup et al., 2000b).

In the case of undisturbed soil, Moldrup et al. (2000a) found that both soil type andcontent of large pores apparently influenced gas diffusivity. At a soil water contentcorresponding to -100 cm H2O of soil water matric head, the following expression wasfound to well describe gas diffusivity for soils with different texture, from different soilhorizons and representing different soil management

1003100

0

100 04.02 εε +=g

g

DD

[2.31]

where Dg100 is the gas diffusion coefficient at –100 cm H2O, and ε100 is the air-filled

porosity at –100 cm H2O (corresponding to the volumetric content of soil pores with an

equivalent pore diameter > 30 µm). By combining Eq. [2.25] with the Campbell (1974)-

Burdine (1953) permeability) water retention model, modified for gas diffusivity(Moldrup et al. 1996), a diffusivity model for undisturbed soils was derived (Moldrup etal., 2000a):

( ) b

g

eg

DD

32

100100

3100

004.02

+

+=

εεεε [2.32]

where b is the Campbell (1974) PSD index (corresponding to the slope of the soil waterretention curve in log-log coordinate system).

50

2.4 Bioremediation

2.4.1 Introduction

Bioremediation may be defined as the use of "enhanced" biodegradation processes(usually by means of microorganisms) to clean up soils and waters polluted by hazardoussubstances. Bioremediation technologies are today well established for the clean-up ofchemically contaminated land, and many technologies are applied commercially on largescale. The first bioremediation technologies that developed were ex-situ technologies, i.e.the treatment of excavated soil in contrast to in-situ technologies which aim at treatmentwithout excavation and often take care of both groundwater and soil pollution (Jørgensenet al., 2000).

Although a large range of microbial genera have been reported to degrade hydrocarbons(Atlas, 1984; Rosenberg et al., 1992), contaminated soils are often poor in organic matterand have a low microbial activity. Usually the bacterial community is adapted to thepresence of the contaminant, but other environmental conditions such as nutrientavailability and oxygen concentration may be unfavourable, and thus the microbialdegradation of the contaminant is slow in situ.

The important questions for full-scale bioremediation applications are; how low aconcentration of the contaminant can be obtained (bioavailability, microbial activity),what is the fate of the contaminant (mineralisation, biotransformation, evaporation, build-up of microbial biomass, incorporation to the bound residue, etc.), how much time isneeded to obtain the set goal (degradation rate), and (4) what the costs are (Batelle,1996).

2.4.2 Bioremediation technologies

2.4.2.1 Ex situ processes

These processes require excavation and remediation of the contaminated soil either on-site or off-site in a soil remediation plant (Agathos and Reineke, 2002). Ex situbioremediation technologies include both slurry- phase remediation, where a water phaseis added to enhance the physical mixing and solid-phase remediation, where the humidity

51

of the soil to be treated corresponds approximately to 50 to 70% of its maximum watercapacity.

In slurry-phase remediation, extremely fine textured or cohesive soil, or sludge, e.g.products from pre-treatment steps such as soil washing, are treated in impeller-typemixing vessels or fluidised-bed reactors.

Biopiles refer to the piling of the material to be biotreated by adding nutrients and air intopiles or windrows usually to a height of 2 to 4 m. Biopiles may be static with installedaeration piping or they may be turned or mixed by special devices for this purpose.Biopiles may be amended with a bulking agent, usually with straw, saw dust, bark orwood chips or some other organic material. Bulking agents are materials of low densitythat lower the soils bulk density, increase porosity, may increase oxygen diffusion andmay help form water stable aggregates when added to the soils (Rhykerd et al., 1999).

If organic material is added, the technology is termed composting. By adding an organicmatrix to contaminated soil the general microbial activity is enhanced and also theactivity of specific degraders, which may be found in the contaminated soil or introducedalong with the organic material. A more advanced type of composting is a drumcompostor, which is closed and has a continuous feed and output.

Landfarming is the term for the older practice of treating oil wastes by adding, forexample, oil sludge and nutrients to agricultural land and mixing by agriculturalpractices.

Basically, it is possible to add microbial inocula, i.e. bioaugmentation, to all these typesof technologies.

2.4.2.2 In situ processes

The decisive viewpoint for an in situ treatment, is that no soil masses need to be removedand that the saturated or unsaturated zone of a contaminated site itself is used as theintegral reactor for the microbial degradation of contaminants (Agathos and Reineke,2002).

52

Bioventing is a venting remediation system, designed and operated to maximizebiodegradation and minimize the volatilization of constituents. Bioventing could result insubstantial reductions in remediation costs by reducing or eliminating the need of off-gastreatment (Lee and Swindoll, 1993).

Airsparging is an in situ technology that reduces concentrations of volatile compoundsthat are dissolved in groundwater, by injection of air into the subsurface saturated zone.Field tests have demonstrated that air sparging can, under the proper conditions, enhancethe recovery of some contaminants through a combination of effects, includingcontaminant partitioning into the vapour phase and acceleration of bacterial degradationof contaminants by increasing dissolved oxygen concentrations. The primary advantagesoffered by airsparging are that surface water treatment equipment and water disposal areeliminated, and that the remediation of sorbed contaminants and contaminants in thecapillary fringe is accelerated (Lundegard and LaBrecque, 1995). Biosparging is thetechnology where the injection of air under the groundwater table is aimed to stimulatethe aerobic biodegradation of the contaminant.

Natural attenuation (NA) is defined as the loss of contaminant mass and concentrationover time and space in a contaminant plume. Monitored natural attenuation (MNA) refersto the reliance on natural processes to achieve site-specific remedial objectives. Naturalattenuation processes include a variety of physical, chemical or biological processes that,under favourable conditions, act without human intervention to reduce the mass, toxicity,mobility, volume or concentration of contaminant in soil or ground water. Theseprocesses include biodegradation, dispersion, dilution, sorption, volatilisation andchemical or biological stabilisation, transformation or destruction of contaminants(USEPA, 1999). NA continues to gain increasing attention from site managers andregulators as a viable alternative to more costly engineered site remediation approaches(Leeson et al., 2001).

Phytoremediation uses the microbial-enhancing processes within the rhizosphere ofplants and trees to stimulate biodegradation. Advocates of phytoremediation suggest thatplants enhance the oxygenation of contaminated soils through two mechanisms. First,specially adapted plants use aerenchyma, channels of reduced air resistance, to transport

53

oxygen to the root zone, enhancing aerobic biological degradation. And second, soildewatering and fracturing increases soil porosity, allowing increased diffusion ofatmospheric oxygen (Rentz et al., 2003).

2.4.3 Biopiles (Batelle, 1996)

Many organic contaminants have successfully been bioremediated in biopiles. Thistechnology has been demonstrated to function in field pilot or full scale especially forpetroleum hydrocarbons (Carrera et al., 2001; Davis et al., 2003), polyaromatichydrocarbons (PAHs) (Hansen et al., 2004), chlorophenols (Long et al., 2001), and fornitro-aromatics.

Biopile technology involves forming petroleum-contaminated soils into piles or cellsabove ground and stimulating aerobic microbial activity within the soils through aeration.Microbial activity can be enhanced by adding moisture and nutrients such as nitrogen andphosphorus. The aerobic microbial activity degrades the petroleum based constituentsadsorbed to soil particles, thus reducing the concentrations of these contaminants.Biopiles are typically constructed on an impermeable base to reduce the potentialmigration of leachate to the subsurface environment. A perforated piping networkinstalled above the base is connected to a blower that facilitates the aeration of the pile. Insome cases, a leachate collection system is constructed, especially if a moisture additionsystem is being considered for the pile. The piles are generally covered with animpermeable membrane to prevent the release of contaminants and/or contaminated soilto the environment and to protect the soil from wind and precipitation. Biopiles operateeffectively in temperate climates but can be operated in colder climates by introducingwarm air through the aeration process.

Biopiles have a lot of advantages compared to other ex situ technologies. Thecontaminants are destroyed, in a relatively easy to design and construct system, in arelatively short time (3 to 6 months) and future containment of the treated soil is notrequired. Biopiles offer a cost-competitive technology compared to e.g. thermaldesorption and can be engineered to be potentially effective for any combination of siteconditions and petroleum products.

54

There are however some limitations to the biopile technology. First biopiles may not beeffective for high contaminant concentrations (>50 000 mg kg-1 total petroleumhydrocarbons). However, such levels are not common in underground storage tanks(UST) sites. During excavation, the peak contaminant levels are reduced, because highlycontaminated soil becomes mixed with surrounding soil that is less contaminated.Second, the presence of significant heavy metal concentrations (>2 500 mg kg-1) mayinhibit microbial growth.

The TPH treatment target levels usually range from 300 to 1 000 mg kg-1 of soil inFlanders. The actual target levels depend on the destination and re-use of the soil.

2.4.4 Biopile operation

2.4.4.1 General construction

Most soils require the addition of water and nutrients before being treated. In some cases,the nutrient and water contents of the soil to be treated are adequate and the grain size iscoarse enough to provide good air permeability without adding anything to the soil. Insoils with a high clay content, soil shredding and/or blending with a bulking agent may beneeded to improve the soil structure and porosity. Typical agents are wood chips or sand(Brown and Cartwright, 1990).

Prior to any soil shredding, a screening step should occur to remove rocks and debris.The typical treatment train to prepare soils requiring the addition of bulking agentsincludes the use of a set of parallel metal bars for bulk separation followed by soilshredding and then addition of water and nutrients. Crushing and mixing of the soil maybe desirable to increase the contaminant homogeneity and improve the soil permeabilityand may eliminate the need to add a bulking agent (Batelle, 1996).

The soil, as received or as processed (shredding or addition of bulking agents), is mixedwith the appropriate amount of nutrients and water as the soil lifts are added to the biopilefoundation. The nutrients may be added as part of a water-spray solution or may be addeddry in measured quantities as each additional bucket of soil is unloaded on the biopile.The pile is then arranged to allow efficient aeration while minimising contaminant andodour release and controlling the internal temperature.

55

The construction details of biopiles are flexible. Effective biopiles have been built in awide variety of sizes and shapes. Historically, biopile dimensions have been constrainedmore by space availability and logistics than by size-based performance limitations.Although a wide range of sizes and shapes have been used, biopile constructioncomplexity increases significantly when the pile dimensions exceed the reach of a front-end loader. Thus, biopile dimensions usually do not exceed a height of 2.4 m. There areno general length or width restrictions, but the front-end loader must avoid driving overprevious lifts.

The task of evenly aerating the pile influences the size and shape of the biopile. Tall piles(>3 m) generally require more than one level of aeration pipes, thus complicating theconstruction process. When installing multilevel aeration systems, the aeration pipes arefrequently placed on top of each layer of soil and covered by the next layer so that thepipes are located at various heights within the pile. Experience, however, indicates that asingle set of aeration pipes located at the bottom of the pile is adequate for piles up toabout 2.4 to 3.0 m high.

After the pile is formed, it should be covered. The cover serves to protect thebiodegrading soil from the elements to retain moisture, heat, prevent excessive, suddenwater addition from rain, prevent wind from blowing dust from the pile and preventcementation of the upper soil layer from wetting and drying. Plastic sheeting materialheld down by old tires, sand bags, or weighted netting is often selected as a low-cost,effective approach for biopile protection (Brown and Cartwright, 1990). In a few cases,an existing building has been used, or an inexpensive structure such as a sheet metalbuilding has been built in lieu of a cover.

2.4.4.2 Aeration

The biopile must be adequately aerated to support efficient degradation of contaminantsby microorganisms. Of all the metabolic factors, oxygen is the most important, soefficient aeration is essential to biopile success. Both active and passive air supplysystems have been used successfully (Brown and Cartwright, 1990).

56

The simplest method of pile aeration is a passive system. Drain tiles, perforated tubing, orslotted pipes are placed at various heights throughout the pile. The tiles or pipes are longenough so their ends stick out of the pile, allowing air transfer and venting due to naturalcurrents. The passive method reduces capital and operating costs because no blower isrequired, and reduces the potential for drying the pile because the airflow rate is low.

Despite the lower cost for passive aeration systems, active aeration is preferred, becauseit gives more complete and more controllable airflow that speeds bioremediation in thepile. Two active aeration configurations have been used, being air injection and airextraction. Both configurations have perforated pipes located in the pile that areconnected to a blower to push air into the pile (injection) or to pull air through and out ofthe pile (extraction).

Typically, airflow rates through the pile are just great enough to keep the soil aboveoxygen-limiting conditions. Such flow rates prevent excess volatilisation ofhydrocarbons, thus reducing the vapour emissions from the pile. One method fordetermining the required air flow rate is presented in Leeson and Hinchee (1995).Although this method was developed for in situ treatment of hydrocarbons, it can be usedfor biopile facilities.

The contaminant volatilization rate is also dependent on the type of contaminant presentin the biopile. Operating in the extraction configuration is often preferred whenvolatilization of organic compounds is a concern. In the extraction configuration,emissions from the pile can be collected and controlled. Extracted air from the bloweroutlet is passed through a treatment system to destroy contaminant vapours. Granularactivated carbon (GAC) historically has been used at biopile sites for the treatment ofdischarge vapours. When the TPH contamination is a heavier fuel (diesel or heavier), off-gas treatment may not be necessary. The TPH concentration in the biopile exhaust airwill rapidly decline as the minor, lighter hydrocarbon fraction is depleted. Vapourtreatment can be stopped once TPH concentrations in the biopile exhaust have decreasedbelow a negotiated level.

Discontinuous aeration can prove to be more cost effective than continuous aeration.Cassidy and Irvine (1997) showed that an operating strategy consisting of alternating 15

57

minute periods with and without aeration provided a rate and extent of overall diesel fuelremoval comparable to continuous aeration. Volatile losses of diesel fuel wereapproximately three times higher for continuous aeration.

Aeration can also be achieved through tillage. Apart from an increase in aeration, thepositive effect of tillage on hydrocarbon disappearance is also attributed to an increase inbioavailability of oil by redistributing the oil in the soil which increased the area ofexposure of oil to microorganisms (Rhykerd et al., 1999).

2.4.4.3 Moisture addition

Water must be available in the biopile, but the amount must not be excessive.Microorganisms require water to transport nutrients, to carry out metabolic processes,and to maintain cell structure. However, excessive water is undesirable because, whenwater occupies a high fraction of the pore space in the soil, the air permeability declines,reducing aeration efficiency, and excess water will increase leaching of contaminants andnutrients from the pile.

The water content of soils to be treated in a biopile can readily be adjusted during theinitial preparation of the soil for the biopile. At some sites, little or no initial wateraddition will be needed. If the soil water content is too high, dry bulking agents can bemixed with the soil. The bulking agent can be selected to increase or decrease waterretention. The soil water content of the biopile may change as the remediation proceeds.Air normally will enter the biopile at less than 100% relative humidity. The air will tendto remove water as it moves through the biopile and become saturated with water, thusreducing the soil water content. However, at the same time, the biodegradation process isconverting hydrocarbons to CO2 and H2O, thus renewing the moisture content to somedegree.

Depending on the site conditions, it may be necessary to add water during biopileoperation. However, dry ambient air conditions and low initial hydrocarbon contenttogether or individually tend to increase the need for water addition. Unless the feed air isdry, the aeration rate is excessive, or the soil organic content is low, an initial adjustmentof water content usually is sufficient to eliminate the need for water addition during

58

operation. Under normal conditions, a covered biopile system should be expected to lose1 to 2% of soil water over a 3- to 4-month operating period. Generally, the soil watercontent is kept between 40 and 85% of field capacity throughout the remediation process.

Out of 123 sites surveyed in a bioventing field study, the soils at 114 sites containedbetween 5% to 25% water by weight. A slight increase in biodegradation with increasingwater content was detected, but the results did not show a strong correlation between thebiodegradation rate and moisture content (Leeson and Hinchee, 1995).

2.4.4.4 Nutrient addition

Biopiles work to degrade contaminants by means of the microorganisms in the pile thatuse the contaminants as a source of carbon and energy. The organisms need a supply ofcarbon to build biomass. The contaminants and natural organic compounds in the soiltypically provide an adequate amount of carbon, but the availability of other essentialnutrients such as nitrogen, phosphorus, or potassium may be insufficient compared to thequantity of carbon. Typically, the C:N:P ratio is brought to within the range of 100:10:1to 100:10:0.5 (Bossert et al., 1984).

Many authors found a positive effect of nutrient addition to the biodegradation of crudeoil (Chaineau et al., 2003) and no effect on the biodegradation of PAH and resins(Chaineau et al., 2003). The lack of effect of nutrients on PAH assimilation can beattributed to the non-stimulation of the aromatic-degrading microorganisms.

In general, the soil should be amended with nutrients prior to biopile construction (Brownand Cartwright, 1990). The nutrients may be either, dissolved in water and sprayed ontothe soil prior to construction of the pile or applied in granulated form and mixed with thesoil while the pile is being constructed. Nutrients may also be added during operation.The nutrient addition is combined with a moisture addition system. When the pile beginsto dry out, a dilute solution of nutrients in water is applied to the top of the pile usingsprays or drip irrigation systems. The nutrient solution then percolates down through thepile (Brown and Cartwright, 1990).

Although the air supplied to the biopile and the contaminants are consumed by biologicalaction, the inorganic nutrients are recycled by the ecosystem. As a result, the nutrients do

59

not have to be continually replenished. After the initial inorganic nutrient amendment ismade (if needed), no further nutrient additions will be required.

2.4.4.5 Microbial amendment

Some biopile designs have included the addition of microbes along with the nutrients.The microbial amendment is added to the nutrient solution and is sprayed onto the soil inpreparation for biopile construction. Microbial amendments increase the overall cost andhave not been clearly demonstrated to improve the degradation of petroleumhydrocarbons. Amendments such as white-rot fungi may be necessary to degraderecalcitrant compounds, but most biopile users reject the addition of exogenous microbes.

2.4.4.6 Temperature

Proper construction of the biopiles is important to avoid excessive internal temperatures.Biological degradation of contaminants releases heat in the same way combustion would,but at a lower rate. The heat release increases the biopile temperature during operation.Some temperature rise is desirable to enhance the microbial degradation rate, but anexcessive temperature increase is undesirable because bioactivity declines after theoptimum temperature is reached. The typical target temperature falls in the range of 20 to

40°C.

2.4.4.7 Metal content

High concentrations of metals will retard the bioremediation process. Traceconcentrations of some metals are essential to growth but high concentrations will have adetrimental effect. Other metals, such as arsenic and mercury, have no nutrient value andmay reduce biological activity when present at low concentrations. As a general rule, thetotal transition and heavy metal concentration in soil to be treated should be less than 2500 mg kg-1. Higher concentrations of cationic metals may be tolerated if the soil pH isgreater than 6.5 or the cation exchange capacity is high.

60

2.4.5 Respiration testing

The respiration test is performed to obtain data for calculating the TPH degradation ratesin the biopile soil. In the respiration test, O2 levels are measured in soil gas sampled fromthe monitoring points installed in various locations of the biopile. Readings are generallytaken until oxygen concentrations drop below 7% or until the O2 concentration no longerdecreases. If O2 decreases rapidly, more frequent readings will be necessary than if O2

decreases slowly. To determine the oxygen utilisation rate, oxygen percentage is plottedagainst time. The slope of this line is referred to as the oxygen utilisation rate and isreported as change of oxygen percent per day.

If low oxygen levels become a limiting factor for biodegradation, the slope of the linewill level off and will no longer be indicative of oxygen consumption relative to TPHdegradation. In this case, only the linear portion of the curve, generally limited to datapoints at or above 12% O2, will be used to calculate biodegradation rates.

The process of aerobic decomposition of hydrocarbons can be described by twotransformation reactions. Firstly, mineralisation is given by (Freijer, 1996)

OemHenCOOmneHeC mn 222 21

41 +→

++ [2.33]

Secondly, conversion of hydrocarbons to cell material and metabolites can be describedby

( ) ( ) ( ) ( ) OvHeOHCeOvueHCe uvmnmn 22 2111

41

2111 −+−→

+−+− − [2.34]

The factor e, the mineralisation quotient denotes the fraction of hydrocarbons that ismineralised. According to chemical kinetics, the production and consumption rates of allsubstances involved are related via molar conversion terms. This yields the followingequation for a closed aqueous system:

( ) 224

12

14

1 )1()(11

OCOHC Svuemne

Sen

Smn +−++

−==− [2.35]

61

where SCnHm is the consumption rate of the hydrocarbon (mol m-3 day-1), SO2 the oxygenconsumption rate (mol m-3 day-1) and SCO2 the carbon dioxide production rate (mol m-3

day-1).

From Eq. [2.35] it follows that the respiration quotient r is given by:

( )vuemneen

SS

rO

CO

41

21

41 )1()(

2

2

+−++−=−= [2.36]

If no biomass is produced (e = 1) and n and m are derived from the C/H ratio of thealkanes (0.50) the respiration quotient r equals 0.67.

From Eq. [2.35] a hydrocarbon degradation quotient s can be defined by:

( )vuemneSS

smnHC

O4

12

14

1 )1()(2 +−++== [2.37]

From Eq. [2.37], the biodegradation rate in milligram hydrocarbon per kilogram of soilper day can be estimated by:

sS

kb

OOHC ρ

ϖε22= [2.38]

where kHC is the hydrocarbon degradation rate (mg HC equivalent kg-1 soil day-1), SO2,

the oxygen consumption rate (mol O2 m-3 soil air day-1), ρb, the soil bulk density

(Mg m 3), ωO2, the molecular weight of O2 (g mol-1), ε, air-filled porosity (m3 soil air

m-3 soil) and s is the mg of O2 used to degrade 1 mg HC equivalent according to Eq.[2.33] (= 3.31).