1. ENHANCING IN SITU PAH
BIODEGRADATION
The Effects of Amendments on Bench-Scale Bioremediation
Systems
by
Avery Gottfried
A thesis submitted in partial fulfilment of the requirements for the degree of
Master of Engineering,
Department of Civil and Environmental Engineering.
The University of Auckland, 2009.

2. ii
ABSTRACT
Current research in the field of bioremediation is uncovering a growing number of
microorganisms with the metabolic potential to degrade PAHs in soil and water. In situ
bioremediation is based on encouraging the growth of microorganisms, either indigenous or
introduced, to improve the degradation of contaminants without excavating or transporting
the soil. The majority of PAHs sorb strongly to soil organic matter posing a complex barrier
to biodegradation. Biosurfactants can increase soil-sorbed PAHs desorption, solubilisation,
and dissolution into the aqueous phase, which increases the bioavailability of PAHs for
microbial metabolism. In this study, biosurfactants, carbon sources, metabolic pathway
inducers, and oxygen were tested as stimulators of microorganism degradation.
Phenanthrene served as a model PAH and Pseudomonas Putida ATCC 17484 was used as the
naphthalene and phenanthrene degrading microorganism for the liquid solutions, soil
slurries and column systems used in this investigation. Bench-scale trials demonstrated that
the addition of rhamnolipid biosurfactant increases the apparent aqueous solubility of
phenanthrene, and overall degradation by at least 20% when combined with salicylate and
glucose. In soil slurries containing salicylate, the effects of biosurfactant additions were
negligible as there was greater than 90% removal, regardless of the biosurfactant
concentration. An in situ enhancement strategy for phenanthrene degradation could focus
on providing additional carbon substrates to induce metabolic pathway catabolic enzyme
production, if degradation pathway intermediates are known. The results of experiments
performed in this study provide further evidence that future studies should focus on
enhancing the metabolic processes responsible for successful in situ bioremediation.

3. iii
ACKNOWLEDGEMENTS
I would like to thank the Commonwealth Scholarship and Fellowship Plan which awarded
me the funding to come to New Zealand and pursue my research interests in this unique
part of the World. While here I was lucky enough to be part of a diverse Environmental
Engineering research group headed by Dr. Naresh Singhal, who also provided supervision for
this work. Special thanks to Abel Francis, our laboratory technician, for providing training,
maintaining equipment, and helping with experimental setup; Dr. Simon Swift in the
Molecular Medicine and Pathology Department for guidance, laboratory training, and
laboratory facilities to develop the biological aspects of this research; and Roy Elliot for
sharing his microbiology expertise in designing experiments and assisting with many
laboratory techniques over the research period.

4. iv
TABLE OF CONTENTS
ABSTRACT ........................................................................................................................II
ACKNOWLEDGEMENTS....................................................................................................III
LIST OF FIGURES ..............................................................................................................VI
LIST OF TABLES..............................................................................................................VIII
ABBREVIATIONS...........................................................................................................IX
CHAPTER 1 INTRODUCTION AND THESIS OBJECTIVES........................................................1
1.1 INTRODUCTION ..................................................................................................1
1.2 THESIS OBJECTIVES .............................................................................................3
1.3 ORGANIZATION OF THE THESIS ...........................................................................4
CHAPTER 2 LITERATURE REVIEW.......................................................................................5
2.1 CONTAMINATED SITES IN NEW ZEALAND............................................................5
2.2 BIOREMEDIATION...............................................................................................6
2.3 POLYCYCLIC AROMATIC HYDROCARBONS ...........................................................8
2.4 IN SITU BIOREMEDIATION................................................................................. 11
2.5 BACTERIAL DEGRADATION................................................................................ 13
2.5.1 PHENANTHRENE METABOLISM .........................................................................13
2.6 LIMITING FACTORS AND STRATEGIES FOR BIOREMEDIATION ............................ 18
2.6.1 BIOAVAILABILITY OF PAH CONTAMINANTS.......................................................20
2.6.2 BIOSTIMULATION AND BIOAUGMENTATION....................................................21
2.6.3 SURFACTANTS AND BIOSURFACTANTS..............................................................22
2.6.4 SORPTION AND DESORPTION ............................................................................27
2.6.5 IONIC STRENGTH AND PH EFFECTS ON BIOSURFACTANTS ...............................33
2.6.6 BIOSURFACTANT MICROBUBBLE DISPERSIONS.................................................34
2.6.7 METABOLIC PATHWAY INDUCERS.....................................................................37
2.7 DETERMINING TRANSPORT PARAMATERS ........................................................ 41
CHAPTER 3 MATERIALS AND METHODS .......................................................................... 45
3.1 LABORATORY SUPPLIES AND MICROOGANIM STORAGE.................................... 45
3.1.1 CHEMICALS.........................................................................................................45
3.1.2 BIOSURFACTANT................................................................................................45
3.1.3 MICROORGANISMS............................................................................................45
3.1.4 MEDIA AND NUTRIENT SUPPLY .........................................................................46
3.2 CELL CULTURING............................................................................................... 48
3.2.1 AGAR PLATES .....................................................................................................48
3.2.2 INOCULANT PREPARATION AND HARVESTING..................................................48
3.2.3 PLATE COUNTS...................................................................................................49
3.2.4 OPTICAL DENSITY...............................................................................................50

5. v
3.3 SOIL METHODS ................................................................................................. 51
3.3.1 SOIL PROPERTIES................................................................................................51
3.3.2 SOIL CONTAMINATION ......................................................................................52
3.3.3 CONTAMINANT EXTRACTION ............................................................................53
3.4 EXPERIMENTAL PROCEDURES ........................................................................... 55
3.4.1 OBJECTIVE 1: LIQUID MEDIUM TESTS................................................................55
3.4.1.1 Inoculant culture growth ............................................................................55
3.4.1.2 Phenanthrene dissolution...........................................................................56
3.4.1.3 Degradation trials.......................................................................................56
3.4.2 OBJECTIVE 2: SOIL SLURRY TESTS ......................................................................58
3.4.2.1 Phenanthrene Desorption in the Presence of Surfactants..........................58
3.4.2.2 Degradation Trials ......................................................................................59
3.4.3 OBJECTIVE 3: COLUMN TESTS............................................................................61
3.4.3.1 Experimental Apparatus.............................................................................61
3.4.3.2 Pressure Measurement...............................................................................62
3.4.3.3 Micro-foam Generation and Stability.........................................................62
3.4.3.4 Column Packing and Unpacking.................................................................63
3.4.3.5 Experimental Operating Conditions............................................................64
3.5 ANALITICAL METHODS...................................................................................... 66
3.5.1 PAH DETECTION.................................................................................................66
3.5.2 BIOSURFACTANT DETECTION ............................................................................67
3.5.3 CHLORIDE ANION...............................................................................................67
CHAPTER 4 RESULTS AND DISCUSSION............................................................................ 68
4.1 OBJECTIVE 1: LIQUID CULTURES ........................................................................ 68
4.1.1 BACTERIA GROWTH ON VARIOUS SUBSTRATES................................................68
4.1.2 EFFECT OF INOCULANT ACCLIMATIZATION AND PRE-TREATMENT..................70
4.1.3 BIOSURFACTANT SOLUBILITY ENHANCEMENT..................................................73
4.1.4 PHENANTHRENE DEGRADATION.......................................................................75
4.2 OBJECTIVE 2: SOIL SLURRIES.............................................................................. 78
4.2.1 PHENANTHRENE DISTRIBUTION IN THE PRESENCE OF BIOSURFACTANTS .......78
4.2.2 SOIL DEGRADATION...........................................................................................83
4.3 OBJECTIVE 3: COLUMN TESTS............................................................................ 91
4.3.1 TRACER AND BIOSURFACTANT BREAKTHROUGH CURVE FITTING....................91
4.3.1.1 Tracer Breakthrough Curves.......................................................................91
4.3.1.2 Biosurfactant breakthrough .......................................................................92
4.3.2 PRESSURE DROP ASSOCIATED WITH MICROBUBBLE DISPERSION PUMPING...97
4.3.3 BIODEGRADING TRIALS IN CONTINUOUS FLOW SYSTEMS..............................100
CHAPTER 5 CONCLUSIONS AND RECOMENDATIONS...................................................... 106
5.1 RECOMMENDATION FOR FUTURE WORK ........................................................ 109
CHAPTER 6 WORKS CITED............................................................................................. 112

6. vi
LIST OF FIGURES
Figure 2.1 Factors that influence biodegradation systems in bioremediation. Adapted
from Singh and Ward (2004)............................................................................................7
Figure 2.2 The diagrammatic structures of 16 U.S EPA priory pollutant PAH compounds.
Modified from Rogers et al. (2002)..................................................................................9
Figure 2.3 Bay, K and L regions of PAHs involved in the formation of metabolically active
epoxides. Adapted from Chauhan et al. (2008) .............................................................14
Figure 2.4 Illustration of common steps in the upper pathway for aerobic metabolism of
phenanthrene (Samanta, Chakraborti et al. 1999; Chauhan, Fazlurrahman et al.
2008; Jun 2008) ..............................................................................................................16
Figure 2.5 Illustration of common steps in aerobic metabolism of naphthalene and one of
the lower pathways for aerobic metabolism of phenanthrene (Samanta,
Chakraborti et al. 1999; Chauhan, Fazlurrahman et al. 2008; Jun 2008).......................17
Figure 2.6 Schematic diagram of physical changes that occur due to surfactant addition
above the CMC. Adapted from (Mulligan, Yong et al. 2001) .........................................24
Figure 2.7 Examples of typical glycolipid biosurfactants produced by Pseudomonas
aeruginosa......................................................................................................................25
Figure 2.8 Schematic diagram of abiotic processes in a soil-aqueous-surfactant system
containing a non-ionic surfactant, phenanthrene and soil organic matter. Adapted
from Edwards et al. (1994).............................................................................................30
Figure 2.9 Plasmid-encoded naphthalene (upper pathway) and salicylate (lower pathway)
degradation genes of NAH7 catabolic plasmid for Pseudomonas sp. Genes nahA-D
encode the upper pathway operon which encodes enzymes for the degradation of
naphthalene to salicylate and genes nahG-M encode the lower pathway operon,
where salicylate is further degraded to pyruvate and acetylaldehyde.The product
from nahR (a trans-acting positive control regulator) is the positive regulator for
both operons and is induced by salicylate. The location of each respective operon
promoter is shown and locations of genes encoding the naphthalene dioxygenase
complex are indicated....................................................................................................40
Figure 3.1 Schematic diagram for serial dilutions to determine cfu/mL of solution ..............50
Figure 3.2 Particle size distribution..........................................................................................52
Figure 3.3 Soil column setup for uplflow pumping experiments ............................................61
Figure 4.1 Partial growth curves for P.Putida until early stationary phase in four growth
medias of glucose (2g/L); naphthalene (0.5g/L); salicylate (0.5g/L); and naphthalene
(0.5g/L) + biosurfactant (1g/L) .......................................................................................69
Figure 4.2 Naphthalene degradation and cell growth in liquid cultures containing
different bacteria inoculant seeds which were pre-grown in seven different
solutions (s1 BHB+naphthalene+glucose grown for 1 week; s2 BHB+glucose; s3
BHB+naphthalene+glucose; s4 BHB+salicylic acid+glucose; s5 LB; s6
LB+naphthalene; s7 LB+salicylic acid; s2 - s7 grown overnight approximately 20
hours growth).................................................................................................................71
Figure 4.3 Phenanthrene solubility enhancement as a function of biosurfactant
concentration. The equation refers to the fit of data above the CMC and ࢟ ൌ ࡿ࢝ ‫כ‬
ࡿ࢝...................................................................................................................................73

7. vii
Figure 4.4 Phenanthrene degradation in liquid cultures containing BHB and/or
biosurfactant (1000mg/L), salicylate (100mg/L), and glucose (100mg/L). Data
presented is the average of triplicate measurements taken at 22 and 46 hours after
inoculation......................................................................................................................75
Figure 4.5 Phenanthrene desorption from soil in the presence of biosurfactant.
Desorption partitioning coefficient Kd calculated from the linear regression
trendline for each series of data. ...................................................................................78
Figure 4.6 Phenanthrene desorption from contaminated soil (50, 100, 250, 500 mg/kg)
into aqueous solution in the presence of biosurfactant over a 48 hour period............80
Figure 4.7 Total phenanthrene concentration in solution and suspended/dissolved
organic matter in soil slurries containing BHB and/or biosurfactant (0.25, 1, 5 g/L),
salicylate (100mg/L), and glucose (100mg/L) over a 10 day period ..............................84
Figure 4.8 Total remaining phenanthrene in soil slurries after 10 days of bioremediation,
results presented as phenanthrene remaining in mg/kg of dry soil..............................86
Figure 4.9 Live cell counts (cfu/mL) taken from soil slurry solution over 10 days. Results
presented are averages from duplicate or triplicate plate counts. ...............................88
Figure 4.10 Observed breakthrough curves and fitted breakthrough curve models using
CXTFIT inverse parameter estimation for (a) chloride with v = 1.54cm/h (b) chloride
with v = 77.87cm/h and (c) biosurfactant with v = 1.54 cm/h.......................................93
Figure 4.11 Microbubble dispersion breakthrough curve with a conservative tracer in the
liquid fraction. ................................................................................................................96
Figure 4.12 Pressure distribution across the length of the soil column during biosurfactant
(1 g/L) solution pumping. Data presented corresponds to depth in the column with
the highest pressure at the inlet, and the lowest pressure at -17cm assuming the
outlet is the datum.........................................................................................................98
Figure 4.13 Pressure distribution across the length of the soil column during biosurfactant
microfoam (1 g/L) pumping. Data presented corresponds to depth in the column
with the highest pressure at the inlet, and the lowest pressure at -17cm assuming
the outlet is the datum...................................................................................................99
Figure 4.14 Pressure distribution across the length of the soil column during biosurfactant
microfoam ( 5g/L) pumping. Data presented corresponds to depth in the column
with the highest pressure at the inlet, and the lowest pressure at -17cm assuming
the outlet is the datum...................................................................................................99
Figure 4.15 Trial 1 phenanthrene distribution in soil column after 10 days continuous
upflow at 0.2 mL/min , phenanthrene in column effluent over 10 days. Column 1
influent solution biosurfactant 1g/L + salicylate 100 mg/L; Column 2 influent
solution biosurfactant 1 g/L. Soil distribution assuming effluent (0cm) is the top of
the column and influent (-37 cm) is the bottom of the column. .................................100
Figure 4.16 Trial 2 phenanthrene distribution in soil column after 10 days continuous
upflow with BHB broth at 0.5mL/min; phenanthrene in column effluent over 10
days. Column 1 influent pulse solution biosurfactant microfoam 1 g/L + salicylate
100mg/L; Column 2 influent pulse solution biosurfactant 1 g/L. Soil distribution
assuming effluent (0cm) is the top of the column and influent (-37 cm) is the
bottom of the column. .................................................................................................101

8. viii
LIST OF TABLES
Table 2.1 Chemical characteristics of EPA PAH priority pollutants. Adapted from data
sources (Aitken, Stringfellow et al. 1998; World Health Organization 1998; Rogers,
Ong et al. 2002) ..............................................................................................................10
Table 2.2 Advantages and disadvantages for selecting in situ bioremediation. Adapted
from The Interstate Technology & Regulatory Council (2005) ......................................12
Table 3.1 BHB marine salts broth approximate formula per litre of prepared media............47
Table 3.2 Soil properties ..........................................................................................................51
Table 3.3 Soil slurry media constituents..................................................................................60
Table 3.4 Column trial experimental design............................................................................65
Table 4.1 Rate of phenanthrene degradation in liquid cultures expressed as mg of
phenanthrene degraded / hour .....................................................................................77
Table 4.2 Calculated phenanthrene soil partitioning coefficient Kd, and phenanthrene
partitioning onto soil sorbed surfactant coefficient Ks ..................................................81
Table 4.3 Total percentage removal of phenanthrene due to soil flushing and
biodegradation in soil column tests after 10 days continuous flow............................103

9. ix
ABBREVIATIONS
PAHs polycyclic aromatic hydrocarbon/s
SOM soil organic matter
IRZs in situ reactive zones
CMC critical micelle concentration
HOCs hydrophobic organic compounds
BHB Bushnell-Hass marine salts Broth
LB Lysogeny Broth
PV pore volume
HPLC high performance liquid chromatography
TOC total organic carbon
OD600 optical density at 600nm

10. 1
CHAPTER 1 INTRODUCTION AND THESIS OBJECTIVES
1.1 INTRODUCTION
Interest in the bioremediation processes of soil contaminated with Polycyclic Aromatic
Hydrocarbons (PAH) has been growing over the past two decades. This interest stems from
the identification of microbes with the ability to degrade toxic xenobiotic compounds in soil
and water. Although bioremediation primarily relies on the catalytic roles of soil
microorganisms to break down contaminants into innocuous by-products, the
understanding of the microbial communities’ operation and behaviour in complex soil
systems remains limited. Bioremediation occurs in the natural environment where most
organisms are uncharacterized, and each site is unique in terms of its soil, microbes, and
contamination. These variable site characteristics create numerous challenges to
understanding the interactions taking place which actually contribute to the desired
decrease in harmful contamination. The biodegradation process is mostly treated as a
unknown ‘black box’ process where soil amendments are made and desired contaminant
removal is achieved without fully understanding the microbial processes that were
enhanced to bring about contaminant mineralization (Singh and Ward 2004). Recent
research has focused on the biochemical and physiological aspects of the bioremediation
process with an emphasis on determining key parameters that make the process more
efficient and reliable (Samanta, Singh et al. 2002). This includes improving the
bioavailability of the contaminants and understanding the metabolic pathways and the
enzymatic reactions that are used in contaminant breakdown, with the goal of identifying
the rate-limiting steps. Ultimately obtaining this knowledge will enable scientist to engineer
better bioremediation processes. Biotechnology and advanced molecular techniques are

11. 2
now providing researchers with the tools to advance understanding in these areas. There is
tremendous potential for engineered bioremediation to make microorganisms more
effective and efficient in removing contaminants, accelerating the remediation process.
It is difficult to replicate the complexity of real PAH contaminated sites in constructed lab
scale systems. However, there is certainly a need to determine optimal treatment
conditions and degradative capabilities in a single bacteria strain in order to unravel the
underlying interactions. Simple systems, where most variables can be controlled and
monitored, show insight into the microbial response to specific variables that are altered,
and will offer advances for understanding the underlying complexities of in situ
bioremediation (Pieper and Reineke 2000). This research was carried out to determine why
specific soil amendments, including the addition of biosurfactants, have been shown to
increase (or decrease) overall contaminant degradation. Results from these trials provide
further evidence to the processes responsible for successful in situ bioremediation
treatment and contribute to the understanding and capabilities of these processes in PAH
field contaminated sites.

12. 3
1.2 THESIS OBJECTIVES
The focus of this research was to enhance the degradation of phenanthrene in soil by the
microbe Pseudomonas putida ATCC 17484 (P.putida) with amendments that included co-
substrates, electron acceptors, and metabolic pathway inducers to the system.
Amendments were designed to increase contaminant bioavailability, enhance microbial
degrading activity, and increase the amount of contaminant degradation. To fully
understand the interactions between biodegradation, amendments, and soil, each process
was isolated and independently evaluated in order to focus on specific characteristics that
enhanced in situ biodegradation. The experiments were designed in stages to achieve each
specific objective:
OBJECTIVE 1: To study the effect of co-substrates, metabolic pathway inducers, and
inoculant pre-treatment on the degradation of phenanthrene and naphthalene by P.putida
in liquid cultures.
Task A: studied growth characteristics of P.putida in various substrates
Task B: determined changes in phenanthrene solubility in the presence of rhamnolipid
biosurfactant
Task C: determined contaminant degradation rates in liquid cultures with added
biosurfactant, glucose, and salicylate
OBJECTIVE 2: To evaluate the effects and monitor the changes in the degradation of
phenanthrene by P.putida due to various soil amendments to contaminated soil slurry.
Task A: determined soil characteristics and the contaminant desorption characteristics
in the presence of rhamnolipid biosurfactant
Task B: determined contaminant degradation rates in soil slurries with amendments

13. 4
OBJECTIVE 3: To design a continuous-flow bench-scale micro-environment to model in situ
remediation and observe the degradation of phenanthrene by P.putida in a saturated
contaminated soil. This system was used to study the effects and flow of microfoam through
the system, and analyze the substrate transport parameters in a soil column.
Task A: determined transport parameters in soil columns using non reactive tracers
Task B: determined microfoam characteristics and evaluate pressure build up in the
soil during the injection of microfoam for various flow rates and microfoam
qualities
Task C: evaluated the efficiency of microfoam and various soil amendments on the
overall removal of phenanthrene from contaminated soil
1.3 ORGANIZATION OF THE THESIS
This thesis consists of five chapters:
Chapter One gives a brief introduction to bioremediation and gives an overview of research
objectives and thesis setup. Chapter Two defines the nature of the problem and discusses
the most important areas of investigation. It also provides an overview of the topic and
highlights key knowledge gained in similar areas, which are relevant to the work presented
in this thesis. Chapter Three presents all of the methods that were used in this study.
Experimental designs are presented in detail for each of the experiments that were
necessary to complete the three main objectives. Chapter Four summarizes the results
obtained and offers an interpretation and discussion of them. Chapter Five forms the
conclusion, and provides recommendations for future research.

14. 5
CHAPTER 2 LITERATURE REVIEW
2.1 CONTAMINATED SITES IN NEW ZEALAND
As modern economies move to enhance environmental protection, more effective testing
methods, increased legislation, and stricter monitoring guidelines have been developed. The
result of this has been the location and acknowledgement of numerous sites of soil
contamination (Doyle, Muckian et al. 2008). Many strategies have been proposed, including
physical, chemical, and biological methods to restore contaminated soil sites. Polycyclic
Aromatic Hydrocarbons (PAHs) are present in many contaminated soil sites, stemming
primarily from the use of oil and petroleum-derivatives; including potentially hazardous,
carcinogenic, and toxic hydrocarbons. Sites with high PAH concentration can act as sources
because contaminants mobilize and leach offsite posing extra risks to groundwater, soil
fertility, and living systems (Singh and Ward 2004). PAHs significantly accumulate in surface
and subsurface soils and an increased concentration can result in a highly toxic
environmental site, necessitating cleanup. Depending on the site location and the level of
groundwater contamination such contamination can pose serious human health risks. In
New Zealand, surface and subsurface soil contamination has been linked to historical land
uses which include agricultural and horticultural activities, gas works, landfills, petrol
station, dry cleaners, sheep dips, and timber treatment sites. The New Zealand Ministry of
the Environment (2007) reports 1,238 contaminated sites resulting from industrial inputs
deemed as ‘Hazardous Activities and Industries List’ (HAIL). However, this number could be
over 50,000 when sites not currently on the HAIL list are considered. These include a large
number of urban sites contaminated unknowingly by fill materials—and such sites are
slowly being discovered (Auckland City Council 2007).

15. 6
2.2 BIOREMEDIATION
The term bioremediation can be applied to any biological process that uses enzymes in
microorganisms, fungi, or green plants to break down undesired contaminants and
contribute to the restoration of the environment to its original condition. Biodegradation is
defined as the breakdown of organic compounds to less complex metabolites, or the
complete breakdown through mineralization into the inorganic minerals H20, C02, or CH4.
Understanding the bioremediation process requires the examination and interpretation of
both biochemical and physiological aspects. Knowledge of these processes will allow key
parameters to be manipulated and bioremediation optimized (Singh and Ward 2004).
Generally, the environmental conditions in which microbial processes are occurring must be
altered to encourage the desired outcomes. With bioremediation, a variety of factors
(Figure 2.1) can influence microbial growth and bioactivity which ultimately increase the
microbial, physiological and biochemical activity and enhance the biodegradation of
contaminants. Even if these factors are optimized, PAH degradation can remain slow unless
the mass transfer rates and the bioavailability of the PAHs for microbial metabolism are
increased (Cerniglia 1993). To effectively accelerate the removal of PAHs from contaminated
soils a greater understanding of not only the physical processes involved but the
physiology, biochemistry, molecular genetics and microbial ecology of the degrading strains
of microorganisms is required (Chauhan, Fazlurrahman et al. 2008). Consideration of the
biotic factors such as the production of toxic or dead-end metabolites, metabolic repression,
presence of preferred substrates and lack of co-metabolites or inducer substrates are also
important in optimizing the overall efficiency of the bioremediation process (Chauhan,
Fazlurrahman et al. 2008).

16. Figure 2.1 Factors that influence
7
Factors that influence biodegradation systems in bioremediation. Adapted from
Singh and Ward (2004)
in bioremediation. Adapted from

17. 8
2.3 POLYCYCLIC AROMATIC HYDROCARBONS
PAHs are identified as a class of chemicals with two or more fused aromatic rings containing
solely carbon and hydrogen (Figure 2.2). They are formed as products in the incomplete
combustion or pyrolysis of fossil fuels and organic matter (Harvey 1991). PAHs are natural
components of fossil fuels such as petroleum or coal, but leakage and accidental spill of
these products can cause accumulation in the environment (Doyle, Muckian et al. 2008).
The World Health Organization (1998) reports that the PAHs contained in various
environmental wastes, including coal combustion residues, motor vehicle exhaust, used
motor lubricating oil, and tobacco smoke “are mainly responsible for their carcinogenic
potential.” The largest and most important PAH contaminated sites occur near large
industrial sources where individual PAH levels of up to 1g/kg of soil have been found. These
originate from concentrated emissions of combustion residues and storage and handling of
coal, coke, fly ash or liquid petroleum reserves. Soils present around crude-oil refineries,
fuel storage depots, petrol stations, gas works, landfills, incinerators and wood preservation
facilities are the most common areas where significant PAH accumulation has been
detected. PAH sources such as automobile exhaust have been shown to cause
contamination next to busy roadways in the range of 2-5mg/kg of soil, whereas background
levels of PAHs are 5-100µg/kg soil deriving from natural sources of atmospheric deposition
such as forest fires and volcanic eruptions (World Health Organization 1998).
The environmental fate of PAH contaminants is governed by the number of aromatic rings
present, and the nature of the linkage between the rings (Doyle, Muckian et al. 2008). PAHs
with low molecular weight are typically defined as those containing up to three aromatic
rings and tend to be more soluble and volatile. High molecular weight PAHs are generally

18. 9
defined as those containing four or more aromatic rings and tend to be less soluble, less
volatile and have a tendency to accumulate in the environment as they sorb strongly to soil
organic matter (SOM) (World Health Organization 1998).
Figure 2.2 The diagrammatic structures of 16 U.S EPA priory pollutant PAH compounds.
Modified from Rogers et al. (2002)
The Environmental Protection Agency (EPA) Priority Pollutant List of 126 pollutants includes
16 PAH compounds (Figure 2.2), the important details of their chemical properties include
thier aqueous solubility (Table 2.1). The contaminants included on this list are regulated,
and the EPA has developed analytical testing methods for accurate detection in the
environment (U.S Environmental Protection Agency 2008). This frequently referenced list
originated from the 1972 Clean Water Act and the 1977 Clean Water Act Amendment, and

19. 10
the only modifications to this list were the removal of 3 compounds from the list in 1981
showing the long recognition of PAHs as toxic compounds (Hendricks 2006; U.S
Environmental Protection Agency 2008).
Table 2.1 Chemical characteristics of EPA PAH priority pollutants. Adapted from data
sources (Aitken, Stringfellow et al. 1998; World Health Organization 1998; Rogers, Ong et
al. 2002)
U.S EPA PAH priority
compound
Chemical
Formula
Aqueous
solubility
Csat
(mg/L)
Molecular
weight
(g/mol)
n-Octanol
water
partition
coefficient
(log Kow)
Organic
carbon
partition
coefficient
(log Koc)
Naphthalene C10H8 31 128.17 3.37 3.11
Acenaphthylene C12H8 3.4 152.19 4.0 3.4
Acenaphthene C12H10 3.8 154.21 3.94 3.65
Fluorene C13H10 1.9 166.22 4.18 3.86
Phenanthrene C14H10 1.1 178.23 4.57 4.15
Anthracene C14H10 0.045 178.23 4.54 4.15
Fluoranthene C16H10 0.26 202.25 5.22 4.58
Pyrene C16H10 0.132 202.25 5.18 4.58
Chrysene C18H12 0.002 228.29 5.65 5.3
Benz[a]anthracene C18H12 0.011 228.29 5.91 6.14
Benzo[b]fluoranthene C20H12 0.0015 252.31 5.8 5.74
Benzo[k]fluoranthene C20H12 0.0008 252.31 6.0 5.74
Benzo[a]pyrene C20H12 0.0038 252.31 6.04 6.74
Dibenz[a,h]anthracene C22H14 0.0006 278.35 6.75 6.52
Indeno[1,2,3-c,d]pyrene C22H12 0.062 276.33 7.66 6.2
Benzo[g,h,i]perylene C22H12 0.00026 276.33 6.5 6.2

20. 11
2.4 IN SITU BIOREMEDIATION
Conventional remediation techniques, including ex situ treatment, and removing the soil for
treatment and disposal at another, safer location are difficult to apply to many PAH
contaminated sites. The slow mobilization of contaminants, and leaching into ground water
tables have lead to situations which contaminated ground water supplies and contaminant
plumes migrate below city infrastructure and developed areas, make ex situ treatment and
soil removal impossible. Sites of PAH contaminations are varied, including: developed urban
and industrial areas, parks and natural environments, and spread over large geographical
areas. Many contaminated sites were discovered after industrial activities ceased, as PAHs
are essentially recalcitrant and persistent (Gómez, Alcántara et al. In Press). The increasing
number and widespread distribution of contaminated sites has encouraged the
development of in situ technologies such as heat based injection, air sparging, soil vapour
extraction, soil washing or flushing, bioremediation, enhanced bioremediation, microbial
filters, and others (Warith, Fernandes et al. 1999). In situ technologies offer an advantage
over ex situ techniques as they are designed to be implemented in place, without
transporting or disturbing the soil (Error! Reference source not found.). Furthermore, in situ
technologies have now been proven to be much cheaper alternatives to traditional
methods, saving time and resulting in a less invasive remediation design that can
complement the natural attenuation process (Suthersan and Payne 2005). In situ
bioremediation is a technology based on stimulating the growth of indigenous or introduced
microorganisms to improve the degradation of contaminants without excavating or
transporting the soil to other locations for treatment. In situ bioremediation can include
augmenting the natural microbial population with the addition of specific microbes that can
metabolize and grow on specific compounds.

21. 12
Table 2.2 Advantages and disadvantages for selecting in situ bioremediation. Adapted
from The Interstate Technology & Regulatory Council (2005)
In situ reactive zones (IRZs) are designed to manipulate oxidation and reduction reactions,
and other biogeochemical processes to affect the mobility, transport, and fate of inorganic
and organic contaminants in the subsurface. Successful designs of IRZs need to account for
all the variables presented in Figure 2.1 to optimize the reactions required for the
biodegradation of target contaminants. Each contaminated site has different characteristics
and naturally variable conditions that need to be taken into account in an engineered
remediation solution to create an effective microbial reactive zone.
ADVANTAGES DISADVANTAGES
It is usually less expensive than other
remediation options.
Complete contaminant destruction is not
achieved in some cases, leaving the risk of a
residual toxic intermediate.
It is almost always faster than baseline
pump-and-treat.
Some contaminants are resistant to
biodegradation.
It may be possible to completely destroy
the contaminant, leaving only harmless
metabolic by-products (no ex situ waste
created).
Some contaminants (or their co-
contaminants) are toxic to the
microorganisms and prevent complete
metabolism and site restoration.
It can be designed with minimal
disturbance to the site and facility
operations, and also can be incorporated
into ongoing site development activities.
Biodegradation of organic species can
sometimes cause mobilization of naturally
occurring toxic inorganic species such as
manganese or arsenic.
It is not limited to a fixed area, typical of
chemical flushing or heating technologies,
because it can move with the contaminant
plume.
Alteration of groundwater redox conditions or
substrate supply can reduce the down
gradient effectiveness of natural
bioattenuation processes.
It can treat both dissolved and sorbed
contaminants.
Uncontrolled proliferation of the
microorganism may clog the subsurface.
The processes usually use reagents that are
easily accepted by regulators and the
public.
The hydrogeology of the site may not be
conducive to enhancing the microbial
population.

22. 13
2.5 BACTERIAL DEGRADATION
The biochemical pathways and enzymes responsible for the initial transformation stages are
usually specific to particular contaminants, but bacteria have the capacity to evolve new
catabolic pathways when exposed for long periods of time to specific contaminants. Due to
the complex mixture of low and high molecular weight PAHs present in some contaminated
sites, there tends to be incomplete bioremediation of higher weight PAHs even if aggressive
approaches are used to enhance the process (Singh and Ward 2004). Due to the
recalcitrance of high molecular weight it is difficult for any single microbial organism to use
them as sole energy and carbon growth, but they are more likely to be oxidized in a series of
steps by consortia of microbes (Perry 1979). Cerniglia (1993) stated that “a better
understanding of the metabolism, enzyme mechanisms, and genetics of polycyclic aromatic
hydrocarbon-degrading microorganisms is critical for the optimization of these
bioremediation processes” and this fact holds true 15 years later and remains an effective
motivation for research today.
2.5.1 PHENANTHRENE METABOLISM
Phenanthrene is a three-ring PAH with low aqueous solubility and is commonly used in
laboratory research as an ideal PAH contaminant for the study of various aspects of
microbial metabolism and physiology (Woo, Lee et al. 2004; Labana, Manisha et al. 2007).
Phenanthrene is the smallest PAH which has both a low aqueous solubility and contains an
“L-region” “bay-region,” and a “K-region” which is common in many higher ringed PAHs
(Figure 2.3). Bay-regions are locations where there is a terminal ring on one side of the bay
region (the terminal ring is also termed the A region), K-regions are areas of high electron
density in all resonance structures and L-regions are sites between two ring fusion points

23. 14
(Yan 1985). The understanding of phenanthrene metabolism can be correlated to studies
on higher-ringed PAHs such as benzo[a]pyrene, benzo[a]anthrancene and chrysene. The
metabolism of bay-region and K-region is believed to be important in understanding the
degradation of both higher and lower ringed PAH compounds and phenanthrene serves as
the example (Xiang, Xian-min et al. 2006). The Bay-region dihydrodiol epoxides are believed
to be the main carcinogenic species and in benzo[a]pyrene these metabolites are cytotoxic,
cause DNA strand breaks and are also mutagenic (World Health Organization 1998). The
Bay, K, and L PAH regions (Figure 2.3) are involved in the formation of metabolically active
and highly reactive epoxides. PAH epoxides arise via metabolism of the parent PAH and
occur whenever oxygen atoms are added across double bonds, a process that can be
catalyzed by the action of enzymes or by an uncatalyzed oxidation process (Josephy and
Mannervik 2006).
There are several bacteria strains that are capable of degrading phenanthrene aerobically
and the more commonly identified strains are Pseudomonas sp, Rhodococcus sp.,
Mycobacterium flavescens, Mycobacterium sp., Flavobacterium sp., and Beijerinckia sp.,
which are capable of using phenanthrene as the sole carbon source and growth substrate
(Cerniglia 1993; Samanta, Chakraborti et al. 1999; Chauhan, Fazlurrahman et al. 2008 ).
Figure 2.3 Bay, K and L regions of PAHs involved in the formation of
metabolically active epoxides. Adapted from Chauhan et al. (2008)

24. 15
Phenanthrene has two potential degradation pathways that are established based on the
bacteria present. These pathways take advantage of the biologically and chemically active
bay and K-region epoxides, which can be formed metabolically by enzymes present in
phenanthrene degrading bacteria (Samanta, Chakraborti et al. 1999). Both pathways share
the same common upper route (Figure 2.4) and are initiated by the double hydroxylation of
a phenanthrene ring by a dioxygenase enzyme to yield cis-3,4-dihydroxy-3,4
dihydrophenanthrene, which then undergoes enzymatic dehydrogenation to 3,4-
dihydroxyphenanthrene. From here the diol is cleaved and metabolized, and 1-hydroxy-2-
naphthoic acid remains and is degraded by one of the two routes termed the lower
pathways (Prabhu and Phale 2003).
The lower pathways consist of two separate routes for degradation depending on the
enzymes that are present in the organisms. In route one (Figure 2.5) 1-hydroxy-2-naphthoic
acid is degraded via the naphthalene pathway to salicylate and then further metabolized via
the formation of catechol or gentisic acid, while route two uses the phthalate pathway. Both
naphthalene and phenanthrene share a common upper metabolic pathway and organisms
that degrade phenanthrene via route one have the ability to degrade naphthalene,
salicylate and catechol (Kiyohara, Torigoe et al. 1994; Samanta, Chakraborti et al. 1999;
Prabhu and Phale 2003). Both oxygen and water are consumed during metabolism and H+
ions are produced, which can affect the pH of the environment if enough degradation
activity is occurring. Understanding the metabolic processes that are involved in the
degradation of phenanthrene are important when determining how additions such as
oxygen or salicylate will influence microbial activity, or determining why changes in pH or a
build up of intermediate metabolites is occurring.

25. 16
Figure2.4Illustrationofcommonstepsintheupperpathwayforaerobicmetabolismofphenanthrene
(Samanta,Chakrabortietal.1999;Chauhan,Fazlurrahmanetal.2008;Jun2008)

26. 17
Figure2.5Illustrationofcommonstepsinaerobicmetabolismofnaphthaleneandoneofthelowerpathwaysforaerobic
metabolismofphenanthrene(Samanta,Chakrabortietal.1999;Chauhan,Fazlurrahmanetal.2008;Jun2008)

27. 18
2.6 LIMITING FACTORS AND STRATEGIES FOR BIOREMEDIATION
It is commonly assumed that bacteria can only access PAHs in the aqueous phase, and the
relatively low bioavailability of PAHs in this phase limits their consumption by the microbial
biomass. As this is a limiting factor, increasing the bioavailability of PAHs in the aqueous
phase by increasing mass transfer rates of PAHs from the soil into solution is essential in
furthering research in this area (Wick, Colangelo et al. 2001). Recent review papers
(Cerniglia 1993; Samanta, Chakraborti et al. 1999; Chauhan, Fazlurrahman et al. 2008 )
conclude that virtually all of the PAHs of concern are biodegradable, and that organisms
capable of degrading PAHs are ubiquitous in the natural environment. PAHs also have
strong hydrophobicity and associate with nonaqueous phases in soil and natural organic
matter where they are not bioavailable, meaning they are in a location where they are not
able to be adsorbed and metabolised by microorganisms. Because of these tendencies,
bioremediation of PAHs in the environment is usually incomplete, even when soil
amendments attempt to enhance the system (Singh and Ward 2004).
Looking specifically at the microbial kinetics, there are several methods possible to enhance
the rate of biodegradation of a PAH. The simplest way to determine what factors influence
this rate in a simple solution is to look at Monod growth kinetics (Equation 2.1). The Monod
growth kinetics take the same form as Michaelis-Menten kinetics with the assumption that
a certain number of new cells grow per unit mass of chemical transformed (Hemond and
Fechner-Levy 2000). This equation is the most commonly used in modelling growth kinetics
associated with PAH degradation.

28. 19
Monod Growth Kinetics
࢛ ൌ ࢛࢓ࢇ࢞ ·
࡯
࡯ା ࡷ࢙
Equation 2.1
Where:
• u : specific growth rate [T-1
]
• umax : maximum specific growth rate [T-1
]
• C : concentration of dissolved chemical [M/L3
]
• Ks : half-saturation constant [M/L3
]
The biodegradation rate depends on µmax, Ks and substrate concentration C. Therefore,
increasing µmax, decreasing Ks, increasing microbial cell density, or increasing contaminant
concentration will be sufficient strategies to enhance the biodegradation process. At low
contaminant concentrations, the rate at which bacteria can degrade the substrate can also
depend on the specific affinity for the substrate (Johnsen, Wick et al. 2005). Specific affinity
refers to the ratio of the maximal rate of substrate uptake and the half saturation constant,
and high affinities lead towards efficient contaminant removal at low concentrations due to
steeper concentration gradients and higher transfer rates between the substrate and the
cell (Johnsen, Wick et al. 2005). Enhancements in microbial growth kinetics can only occur if
no chemicals other than the contaminants are limiting the microbial community.
Specifically, oxygen and mineral nutrients must be in excess. Understanding the basic
parameters that influence the degradation rate of contaminants, highlights the importance
of increasing the bioavailability of contaminants for effective in situ bioremediation.

29. 20
2.6.1 BIOAVAILABILITY OF PAH CONTAMINANTS
Significant PAH accumulation in the environment occurs in subsurface organic soil matter
due to the hydrophobicity and low aqueous solubility of PAHs. The majority of PAHs are
difficult to remove because they sorb strongly to soil organic matter. Long term
contamination of soil, which is commonly referred to as aging or weathering, is the result of
chemical oxidation reactions and slow chemical diffusion into small pores, both of which
decrease PAH bioavailability over time (Singh and Ward 2004). The degradation process
therefore involves the transfer of contaminants from the soil to the enzymes in the
microorganisms which begin the mineralization of the contaminant (Noordman 1999).
Contaminant characteristics such as molecular structure, solubility, and the octanal/water
partitioning coefficient (Kow) are relevant for substance sorption in or onto soil and can be
used to indicate the availability of the contaminant to bacteria in the soil.
It is the influence of the dissolution and desorption process of PAHs in soil that are often
cited as the rate limiting step in the degradation process. The slow transport of PAHs from
the soil matrix to bacteria is the slowest process and limits degradation (Mulder, Breure et
al. 2001; Prabhu and Phale 2003; Johnsen, Wick et al. 2005; Doyle, Muckian et al. 2008).
However, contaminant bioavailability can be species-specific, with different bacteria strains
able to access different contaminant pools in the soil-water system. Understanding the
interactions amongst bacteria will also provide further opportunities for enhanced
degradation of bioavailable contaminants (Dean, Jin et al. 2001). For instance, some
organisms have the ability to affect sorption kinetics on their own, through the production
of surface active agents termed biosurfactants. These can increase the apparent solubility of
PAHs in the aqueous phase and concomitantly increase the concentration gradient, allowing

30. 21
improved mass transfer of contaminants from the soil to the aqueous phase (Pignatello and
Donald 1999). Dean et al. (2001) demonstrated that some of the sorbed phase
phenanthrene was bioavailable to certain Pseudomonas bacteria, and called into question
the frequently used assumption that only bulk aqueous phase contaminant is available for
degradation. Woo et al. (2001) included a term for sorbed phase biodegradation of
phenanthrene when modelling the process to account for the rapid degradation that
occurred in soil slurry tests. Kwok and Loh (2003) also proposed that bacteria which have
attached themselves directly to soil particles can utilize the nutrients sorbed at that
location. Several techniques have been developed to effectively enhance the bioavailability
of contaminants.
2.6.2 BIOSTIMULATION AND BIOAUGMENTATION
Enhanced biodegradation is usually accomplished through biostimulation and
bioaugmentation. Biostimulation refers to the modification of the environment via the
addition of oxygen, nutrients, other electron donors or acceptors, and surfactants. These
additions stimulate the existing bacteria and increase the number or rate at which the
organisms are degrading a contaminant. Biostimulation relies on making the natural
environment more favourable to the metabolic capacities of the indigenous microbial
populations, whereas bioaugmentation describes the addition of adapted microorganisms
to the environment that are capable of degrading contaminants that are present.
Depending on the characteristics of the contaminated site, either biostimulation or
bioaugmentation may be needed to achieve the desired outcomes. Ruberto (2006) found
that a combination of both techniques using fish meal for nutrient supply and surfactant
Brij700 with bioaugmentation using a psychrotolerant PAH degrading bacterial consortium

31. 22
caused significant removal (46.6%) of phenanthrene whereas when each technique was
applied separately, insignificant reduction was observed (Ruberto, Vazquez et al. 2006). It is
commonly reported that either the availability of electron acceptors, or nutrient limitations,
are the cause of slow biodegradation processes at contaminated sites (Institute for Ecology
of Industrial Areas 1999). Laboratory studies often report high rates of biodegradation
compared to results actually achieved in the field with similar soil and bacteria types, which
can be due to the optimization of many variables such as temperature, mixing, nutrient
balances and nutrient delivery. These variable are sometimes impossible to replicate in the
field (Institute for Ecology of Industrial Areas 1999).
2.6.3 SURFACTANTS AND BIOSURFACTANTS
Surfactants are used to describe surface-active agents that lower the surface tension of a
liquid (Riser-Roberts 1998). Surfactants have both a hydrophilic group and a hydrophobic
group and can be described as either anionic or cationic depending on whether they release
an anion or a cation when dissociating in water. They are termed non-ionic if no net charge
is dissociated. Therefore an anionic surfactant has an anionic hydrophilic group at its head,
whereas a non-ionic surfactant has no net charge groups at its head. Anionic and non-ionic
surfactants tend to be the best solubilizers and are relatively non-toxic compared to cationic
surfactants (Oostrom, Dane et al. 2006). Jin (2007) ranked the toxicity of the studied
surfactants to bacterial activity in soil and determined the order of toxicity towards bacteria
as follows: non-ionic surfactants (Tween 80, Brij30, 10LE and Brij35) < anionic surfactants
(LAS) < cationic surfactants (TDTMA) (Jin, Jiang et al. 2007).

32. 23
Suitable co-solvents or surfactants must be selected according to solution chemistry, proven
ability to solubilise PAH compounds, and compatibility with the remediation technique. In
addition, they must not be toxic or a threat to human health or the environment (Gómez,
Alcántara et al. In Press). The presence of surfactants in the bulk phase causes an increase in
the free energy of the system. In order to lower the free energy, the surfactant molecules or
monomers are concentrated at the surface and interface, and the surface tension is lowered
increasing the solubility of hydrophobic contaminants (Myers 1988). The surface tension will
decrease to a given value, known as the critical micelle concentration (CMC) beyond which
point it will remain constant (Figure 2.6). Once the concentration of surfactants is above the
CMC, the surfactants begin to aggregate to form micelles, vesicles, and lamellae. Surfactant
micelles increase the apparent aqueous solubility of hydrophobic particles by reducing the
interfacial tension between the oil phase and the aqueous phase. In contaminated systems
this results in PAHs partitioning within the hydrophobic micellar core of the micelles. This
creates higher apparent aqueous solubility as PAHs are dissolved both in aqueous solution
and inside surfactant micells which are present in the bulk aqueous phase (Error! Reference
source not found.) (Noordman, Ji et al. 1998; Cameotra and Bollag 2003; Makkar and
Rockne 2003).
Surfactants can also be produced by bacteria or yeasts from growth on various substrates
including sugars, oils, hydrocarbons and agricultural wastes. These are termed
biosurfactants (Lin 1996). In terms of surface activity, heat and pH stability, many
biosurfactants are comparable to synthetic surfactants (Lin 1996). Biosurfactants are
receiving increasing attention as they have lower toxicity and higher biodegradability
compared to their chemical counterparts (Rosenberg and Ron 1999). Specifically,

33. rhamnolipid biosurfactants produced by
extensively as they have excellent emulsifying power with a variety of hydrocarbon
vegetable oils (Wang, Fang et al. 2007)
glycolipidic surface-active molecules that are produced in mixtures of one or two rhamnoses
attached to β–hydroxyalkanoic acid
resulting in lengths of 8, 10, 12 and 14 carbons
Fang et al. 2007). The in situ
them potentially more cost effective while also using natural resources instead of chemical
inputs.
Figure 2.6 Schematic diagram of physical changes that occur due to surfactant addition
above the CMC. Adapted from
Surfactant
monomer
24
rhamnolipid biosurfactants produced by Pseudomonas aeruginosa have been studied
excellent emulsifying power with a variety of hydrocarbon
(Wang, Fang et al. 2007). Rhamnolipid (Figure 2.7) is the name given to the
active molecules that are produced in mixtures of one or two rhamnoses
hydroxyalkanoic acid. The length of the fatty acid chains can va
12 and 14 carbons (Soberón-Chávez, Lépine et al. 2005; Wang,
in situ production of biosurfactants at contaminated sites renders
ntially more cost effective while also using natural resources instead of chemical
Schematic diagram of physical changes that occur due to surfactant addition
above the CMC. Adapted from (Mulligan, Yong et al. 2001)
Surfactant
Micelle
have been studied
excellent emulsifying power with a variety of hydrocarbons and
the name given to the
active molecules that are produced in mixtures of one or two rhamnoses
an vary significantly,
Chávez, Lépine et al. 2005; Wang,
production of biosurfactants at contaminated sites renders
ntially more cost effective while also using natural resources instead of chemical
Schematic diagram of physical changes that occur due to surfactant addition
g et al. 2001)

34. 25
Figure 2.7 Examples of typical glycolipid biosurfactants produced by Pseudomonas
aeruginosa
Research into the addition of surfactants and biosurfactants have produced mixed results,
from greatly enhanced rates of PAH degradation to the inhibition of PAH degradation
(Pieper and Reineke 2000; Makkar and Rockne 2003; Avramova, Sotirova et al. 2008). There
are several hypotheses explaining the mixed results. Beneficial results may be due to the
facilitation of bioremediation through increases in desorption, solubilisation, and dissolution
of PAHs from soil sorbed or solid phase contaminant into the aqueous phase, which results
in increased bioavailability of PAHs for microbial metabolism (Mulligan, Yong et al. 2001;
Makkar and Rockne 2003; Shin, Kim et al. 2004; Avramova, Sotirova et al. 2008). Negative
result led to an assortment of conclusions including:
• the preferential use of surfactants as a growth substrate by degrading
microorganisms;
• the toxicity of the applied surfactants preventing increased microbial growth;

35. 26
• the toxicity of the PAHs resulting from the increased bioavailability that is caused by
the surfactant solubilization of PAHs;
• the reduction of PAH bioavailability due to the uptake into surfactant micelle which
then could not be available for bacteria;
• the sorption of surfactant into the soil blocking access to PAHs that could have been
further absorbed into the soil or causing PAH sorption into soil sorbed surfactants
(Garcia-Junco, Gomez-Lahoz et al. 2003; Shin, Kim et al. 2004; Avramova, Sotirova et
al. 2008).
An example of these mixed results was meaningfully demonstrated by Allen et al. (1999)
with the use of titron X-100 with Pseudomonas sp. strain 9816/11 and Sphingomonas
yanoikuyae B8/36. Triton X-100 increased the rate of oxidation of phenanthrene with strain
9816/11. Conversely, the surfactant inhibited the biotransformation of both naphthalene
and phenanthrene with strain B8/36 under the same conditions (Allen, Boyd et al. 1999).
These observations show an important knowledge gap in how surfactants truly alter the
biodegradation process and interact with bacteria. Considering that a non-ionic surfactant
could have contrasting effects on the ability to degrade PAHs by different bacteria, there is a
requirement for additional research relating to surfactants, including all stages of soil-water-
surfactant-bacteria interactions.
There is a recurring assumption that the remediation of PAHs in soil or soil-water systems
depends strongly on the desorption rates of the PAHs from the soil into the aqueous phase
(Jin, Jiang et al. 2007). It is assumed that once PAHs are in the bulk aqueous phase, it is
possible to use engineering treatment steps to enhance the remediation process and create

36. 27
an effective bioremediation strategy. However, there are an increasing number of studies
that have demonstrated that bacteria can attach to soil particles and use the nutrients
sorbed to the soil surface (Dean, Jin et al. 2001; Wick, Colangelo et al. 2001). This could
explain why the addition of surfactants to some systems does not predictably enhance
contaminant biodegradation. As the natural role of biosurfactant is to increase the
bioavailability of contaminants by decreasing surface tension, there can be a reduction in
direct adhesion of bacteria to the desired contaminants of interest due to the decrease in
surface tension (Pieper and Reineke 2000).
The mixed effects of surfactant on biodegradation show the complex interactions between
the PAH, surfactant, microorganism, soil, and water in the environment. Due to variable that
are important in the bioremediation process all researchers have to provide caveats in the
conclusions section to isolate results to the unique system of bacteria, soil type,
contaminant, and test conditions that was studied.
2.6.4 SORPTION AND DESORPTION
The partitioning and transport processes (sorption, desorption, and dissolution) between
the soil and water phases of both contaminants and surfactants affect the overall
degradation of contaminants (Schlebaum, Schraa et al. 1999; Kraaij, Ciarelli et al. 2001;
Mulder, Breure et al. 2001; Zhou and Zhu 2005; Zhou and Zhu 2007; Wang and Keller 2008;
Zhu and Zhou 2008; Laha, Tansel et al. In Press). Soil organic matter and natural organic
matter is not homogeneous and PAHs strongly absorb to soot carbon, and more slowly
partition into humic matter (Jonsson, Persson et al. 2007). As PAHs adsorb onto the surface

37. 28
of soil organic mater they slowly begin to penetrate further into cavities and diffuse into the
organic fraction over time. Landrum et al (1992) observed a continuous increase in the
partition coefficient of phenanthrene and pyrene into soil over a period of six months, after
in-lab contamination of the soil. The length of this process makes it impractical for the
determination of single sorption or desorption coefficients to model the process over the
long term. Schlebaum et al (1999) successfully modelled the sorption of hydrophobic
organic compounds (HOCs) from the soil matrix with a kinetic model using two separate
compartments. A Freundlich isotherm represented high affinity sites, and a linear sorption
isotherm and first order kinetics represented low affinity sites. Even if the amount of
organic matter is low, PAHs can still become trapped in pores and voids and these variables
will affect the efficiency and success of any remediation process.
It is not just the average aqueous concentration of the target contaminant that determines
its availability. The rate of mass transfer to microbial cells relative to the intrinsic substrate
utilization capacity of the microbial cells must also be considered because it determines the
bioavailability of the contaminant (Wick, Colangelo et al. 2001). As a result, limited
bioavailability occurs when the environment is unable to deliver the substrate at the rate
consumable by the microbial biomass. The biodegradation rate in the subsurface is often
reported as first-order even when total contaminant concentrations are high. Wick et al.
(2001) provided an explanation for these observations by considering that the mass
transport processes are slow for hydrophobic organic soil pollutants which cause the same
degradation rates to be obtained even when the substrate concentration has changed.

38. 29
When enhancing bioremediation, it is important to consider the effect surfactants can have
on the desorption and dissolution of contaminants from the soil. When surfactants exceed
their CMC , it is well established that there is an increase in desorption of PAHs from the soil
(Noordman 1999). However, when surfactants adsorb to soil they increase the overall
organic content of the soil and provide additional sorption capacity. This can enhance
sorption of hydrophobic compounds onto soil sorbed surfactants. This influences the
amount of PAHs present in the aqueous phase, accessible for biodegradation (Edwards,
Adeel et al. 1994). Conversely, the micelles present in the bulk aqueous phase can greatly
enhance the solubilisation of the PAHs, causing increased desorption from soil. The
efficiency of surfactants at enhancing PAH desorption shows a strong dependence on the
soil composition, surfactant structure and concentration, and PAH properties as concluded
by Zhou and Zhu (2005).
The process of surfactant adsorption to the soil has been described as a three stage process
by Torrens et al.(1998). The first stage is controlled by electrostatic attraction between
surfactants and the soil surface. As the surfactant concentration increases, there is a
tendency for self-association of surfactant ions due to the electrostatic and hydrophobic
forces. This is analogous to the micelle formation but it leads to the formation of
hemimicelles which is the second stage. This stage is more rapid than the first stage, and
results in neutralization of the particle surface, causing the sorption process to slow. After
the second stage, micelle formation begins, which results in the reversal of the surface
charge. This greatly reduces surfactant sorption due to charge repulsion. The third stage is a
plateau region, and additional surfactant will be present in solution (Torrens, Herman et al.
1998). Figure 2.8 encapsulates the interactions that are believed to occur between the soil-

39. 30
water-surfactant systems. The stages of sorption also result in the wetting of soil grains
which enables the washing out of the hydrophobic substances from the soil pores. For most
hydrophobic contaminants they can be assumed to be un-wetted with water, and
surfactants increase the wettablity of the hydrophobic surfaces through attachment and
sorption to the soil surface (Pastewski, Hallmann et al. 2006).
Figure 2.8 Schematic diagram of abiotic processes in a soil-aqueous-surfactant system
containing a non-ionic surfactant, phenanthrene and soil organic matter. Adapted from
Edwards et al. (1994).

40. 31
The distribution of contaminants between the soil fraction and the aqueous phase is
generally described by the partitioning coefficient Kd. Kd refers to the ratio of the
concentration of contaminant in the soil fraction to the concentration in the aqueous phase.
In the simplest form, the equation Cs = KdCw where Cs is the concentration of the
contaminant sorbed by the soil fraction and Cw in the concentration of the contaminant in
the aqueous phase respectively. Kmc is another commonly used partitioning coefficient,
which defines the amount distributed between the aqueous phase and the surfactant
micelle phase. The total amount is commonly referred to as the apparent aqueous solubility
as there is more contaminant in the aqueous phase, although it is located inside the
surfactant micelle. There are many different theoretical models that are used to determine
the partitioning coefficient Kd, taking into account surfactant adsorption modelled by the
Langmuir isotherm, Kow, and the fraction of organic carbon, and PAH sorption (Huang and
Cha 2001). The following equation appears to be the most commonly used to describe PAH
partitioning within a soil-water-surfactant system (Zhu, Chen et al. 2003; Zhou and Zhu
2007; Wang and Keller 2008; Zhu and Zhou 2008).
ࡷࢊ
‫כ‬
ൌ
ࡷࢊା ࡽ࢙ࡷ࢙
૚ା ࡷ࢓࢔ ࢄ࢓࢔ା ࡷ࢓ࢉ ࢄ࢓ࢉ
Equation 2.2
Where:
• ‫ܭ‬ௗ
‫כ‬
: ratio of sorbed PAH to mobile PAH in the aqueous solution (L/kg);
• Kd : PAH sorption coefficient with the soil in the absence of surfactant (L/kg);
• Qs : quantity of surfactant sorbed to the soil;
• Ks : solute distribution coefficient with the soil-sorbed surfactant (L/kg);
• Xmm and Xmc : surfactant monomer and micellar concentration in water (g/L);
• Kmm and Kmx : PAH partitioning coefficients with the surfactant monomer and
micellar phases (L/kg).

41. 32
The overall factors that effect ‫ܭ‬ௗ
‫כ‬
are the partitioning of PAH to soil due to the presence of
sorbed surfactants (terms in the numerator in equation 2.2), and decreased PAH
partitioning to soil by the enhanced aqueous solubility of the PAH in the presence of
surfactant monomers and micelles (denominator in equation 2.2).
Depending on the quantity of surfactant added to the system, the majority may be in the
soil sorbed-phase (Laha, Tansel et al. In Press). The result of this is increased partitioning of
PAHs onto soil until the solubilisation by micellar phase surfactant is at a high enough
concentration to compete with the increased PAH sorption on the surfactant sorbed soil
(Laha, Tansel et al. In Press). However, the cation exchange capacity of the soil can
significantly affect the sorption of surfactants (Ks), and well as the ionic strength or pH of the
system.

42. 33
2.6.5 IONIC STRENGTH AND PH EFFECTS ON BIOSURFACTANTS
Anionic surfactants are strongly affected by the presence of electrolytes in solution as they
can influence the solubilization capacity, cause precipitation of the surfactant from the
aqueous phase, and increase the adsorption to subsurface porous media (Stellner and
Scamehorn 1989; Jafvert and Heath 1991; Guiyun, Brusseau et al. 1998). Torrens et al (1998)
saw 67% rhamnolipid sorption to soil at low K+ concentrations (10mM) but this increased to
98% in the presence of 20mM K+ in solution. The ionic strength and presence of cations in
solution has been shown to further enhance the solubility of hydrophobic organic
contaminants in rhamnolipid solution. Guivun (1998) reported that both Na+
and Mg2+
enhanced the solubility of PAHs as there was an increase in the interior volume of
rhamnolipid micelles in the presence of cations, and Mg2+
, being a divalent cation, had a
stronger affection on reducing the repulsion forces between anionic head groups (Guiyun,
Brusseau et al. 1998). However, Ca2+
had little affect on solubility, due to competing effects
between rhamnolipid precipitation and enhanced contaminant solubility. The presence of
cations also reduced the interfacial tension between rhamnolipid solutions and hexadecane
from 2.2 to 0.89 dyn cm-1
(Guiyun, Brusseau et al. 1998). A decrease in pH from 7 to 6 was
seen to have the same qualitative effect to the interfacial tension as the increase in Na+
concentration. The carboxyl group in the rhamnolipid head group has a pKa of 5.6, causing it
to become more protonated as the pH decreases, thereby reducing repulsion between the
head groups. A similar effect was seen by Shin (2004) as the apparent solubility of
phenanthrene was 3.8 times greater at a pH of 5.5 when compared with a pH of 7 in the
presence of 240 mg/L rhamnolipid. In another study, more rhamnolipid molecules were lost
by sorption to sand particles at a pH 4 than at both higher and lower pH values, explaining
why a dramatic decrease in apparent aqueous solubility of phenanthrene was seen at that

43. 34
pH (Shin, Kim et al. 2008). These findings are particularly important in soil remediation as
subsurface matrix solutions contain electrolytes such as Ca2+
, Mg2+
, Na+
, K+
, and Al3+
which
can have affect the surfactant performance (Guiyun, Brusseau et al. 1998).
2.6.6 BIOSURFACTANT MICROBUBBLE DISPERSIONS
Microbubble dispersions, also known as colloidal gas aphrons (CGA) or microfoam, are a
series of micro-bubbles that were first investigated by Sebba (1971). Microfoam displays
colloidal properties because of its micron-sized bubbles (typically 0.7-100µm) and its unique
bubble structure which consists of multiple layers of surfactant monomers surrounding the
surface of the microbubble. In contrast, standard foam consists of just one layer of
surfactant monomers (Jauregi and Varley 1999; Wan, Veerapaneni et al. 2001; Larmignat,
Vanderpool et al. 2008). Microbubble dispersions can flow like water, and can be pumped
easily without collapse (Jauregi and Varley 1999). Surfactant microfoam technology is a
relatively new approach for enhancing in situ bioremediation, showing promising
advantages over air sparging or surfactant solution application. Foam can flow in a plug flow
manner, delivering oxygen or air uniformly (Wang and Mulligan 2004). Microbubble
dispersion flow is also capable of overcoming heterogeneity in porous media, enhancing
bacterial transport, and delivering oxygen and nutrients to the subsurface (Wan,
Veerapaneni et al. 2001; Choi, Park et al. 2008; Park, Choi et al. In Press). Foam and
microfoam technology is designed either to remove contaminants and/or act
simultaneously as an augmentation for existing technologies such as pump-and-treat and
bioremediation. It is designed to enhance the process and improve removal efficiencies and
cost effectiveness (Wang and Mulligan 2004). Foam stability reflects the ability of the

44. 35
suspension to resist bubble collapse, and is typically measured as the time required for half
of the foam to collapse. The half-life for microfoam can range from minutes to days,
depending on the generation method, surfactant, and additions such as nutrients, bacteria
or soil particles.
Microbubble dispersions can facilitate mobilisation and transport of contaminants trapped
in porous media, and can take less pore volumes to achieve high contaminant removal when
compared to surfactant solutions (Wang and Mulligan 2004). Couto et al (In Press) saw 96%
removal in sandy soils using microfoam in soil flushing to remove diesel oil, versus 88%
removal with regular foam and 35% removal with surfactant solution. Park et al. (In Press)
saw a 2.2-fold increase in phenanthrene degradation when 3 pore volumes of microbubbles
were injected instead of 1 pore volume. There are several studies demonstrating the ability
of conventional foam to enhanced remediation of PAH contaminated soils, and the
beneficial transport mechanisms of foam (Chowdiah, Misra et al. 1998; Rothmel, Peters et
al. 1998). Microfoam appears to have an added advantage over conventional foam as
dispersion can be generated that contain less gas (60-70% versus up to 99% with
conventional foam) in smaller sized bubbles, making them easier to pump through the
subsurface (Roy, Kommalapati et al. 1995; Jauregi and Varley 1999).
Microbubble injection systems have been shown to be efficient oxygen delivery systems in
pilot scale tests that used microbubble generators that were encapsulated in pressurized
chambers that contained oxygen and biosurfactant solution. Leigh et al. (1997)
demonstrated that microbubbles generated using this method persistent in the subsurface
for longer periods of time and have different migration characteristics compared to air

45. 36
bubbles injected in by typical air sparging. Using this generation method and a mixture of
anionic and non ionic surfactants, Wan et al. (2001) was able to generate microbubbles that
were still present in solution up to six weeks after generation.
Subsurface foam and microfoam flow is typically accompanied by a pressure drop due to the
flow characteristics. Higher-viscosity foams flow forward and fill up larger channels and pore
spaces. When the pressure drop builds up in the channel, the foam flows into less accessible
spill areas. This pressure dependent “clogging” process means that channelling, or poor
sweep, should not occur with the microbubble scouring as compared with surfactant
flushing (Riser-Roberts 1998). However, applications could be limited by the pressure drop
required to pump microbubbles into soil with low permeability (Riser-Roberts 1998; Choi,
Park et al. 2008; Park, Choi et al. In Press).

46. 37
2.6.7 METABOLIC PATHWAY INDUCERS
Another biostimulation strategy that can enhance the intrinsic biodegradation rate of target
compounds is the addition of one or more known pathway intermediate catabolite. These
are usually produced by the bacteria when mineralizing a contaminant and they stimulate
growth, enzymatic expression, and ultimately increase the biodegradation of PAHs
(Ogunseitan and Olson 1993; Cho, Seung et al. 2006). This process is defined as co-
metabolism, where bacteria may co-utilize various substrates that compete with the
structurally similar primary substrate for the enzyme’s active sites (Mohan, Kisa et al. 2006).
The introduction of carbon sources that are metabolic pathway inducers into the soil can
enhance in situ bioremediation by stimulating the growth of specific indigenous micrograms
that are capable of degrading organic contaminants. Unfortunately, additional carbon
sources can also be used preferentially by soil bacteria causing diauxic growth which can
have a negative effect on the degradation process (Lee, Park et al. 2003).
A number of studies have used salicylate as a pathway inducer to enhance initial rates of
naphthalene and phenanthrene removal (Chen and Aitken 1999; Lee, Park et al. 2003; Woo,
Jeon et al. 2004; Lee, Lee et al. 2005; Powell, Singleton et al. 2008; Basu, Das et al. In Press).
Salicylate is the third intermediate formed in the degradation of naphthalene (Figure 2.5)
and it is also an intermediate formed in the degradation of phenanthrene for bacteria that
degrade phenanthrene via the naphthalene pathway. Most information about PAH
metabolism has been derived from the study of naphthalene catabolic plasmids in
Pseudomonas putida G7 (Yen and Serdar 1988). In the plasmid there are genes which
encode the pathway for naphthalene degradation (Figure 2.9) In the first operon, there are
genes which encode the pathway for conversion of naphthalene to salicylate, and in the

47. 38
second operon are the genes which code for the conversion of salicylate via catechol meta-
cleavage to acetaldehyde and pyruvate (Eaton and Chapman 1992; Platt, Shingler et al.
1995). The regulatory mechanism for both operons is encoded in a third operon which acts
as the regulatory protein and positively regulates the two operons by the increased
presence of salicylate (Schell and Wender 1986; Atlas and Philip 2005). The principle
mechanism for the aerobic bacterial metabolism of naphthalene is via the oxidative action
of the naphthalene dioxygenase enzyme; that introduces molecular oxygen into the
aromatic ring. The naphthalene (upper pathway) and salicylate (lower pathway) degradation
genes located in the NAH7 catabolic plasmid from Pseudomonas sp. are regulated by
salicylate induction to both operons (Figure 2.9). Chen and Aitken (1999) showed that
salicylate greatly enhanced removal of fluoranthene, pyrene, benz[a]anthracene, chrysene,
and benzo[a]pyrene, all of which are high molecular weight PAHs which the strain
Pseudomonas saccharophila P15 could not use as a sole carbon for growth. This showed
that high-molecular weight PAH metabolism by this organism is induced by salicylate. Lee et
al. (2005) saw phenanthrene degradation rates 3.5-fold higher with Burkholderia cepacia
PM07 compared to the rates achieved without salicylate addition in aqueous solutions. They
also saw a decrease in phenanthrene removal with the addition of glucose (Lee, Lee et al.
2005). Basu et al. (In Press) determined Pseudomonas Putida CSV86 preferentially utilized
aromatics over glucose and co-metabolized them with organic acids, indicating that
intermediate metabolites enhance the mineralization rate of PAHs more effectively than
additional carbon sources (Basu, Das et al. In Press).
In other studies Cho et al. (2006) saw up to 12 times increase in the degradation of target
chemicals per equivalent cell mass with the addition of various intermediate metabolites

48. 39
into solution. In this experiment all phenanthrene was soluble due to the addition of 1%wt
Triton X-100. Woo et al. (2004) saw up to a 3-fold increase in phenanthrene degradation
using salicylate in soil water systems, however addition of triton X-100 saw inhibitory effects
towards total phenanthrene mineralization. Other substances such as 1-hudroxy-2-
naphthoate, catechol, and pyruvate have also shown their potential as effective pathway
inducers to enhance in situ bioremediation (Cho, Seung et al. 2006; Basu, Das et al. In Press).
Chemotaxis is another strategy that can be used to enhance the degradation of
contaminants in the environment. Chemotaxis is “a complex process [in] which bacterial
cells detect temporal changes in the concentrations of specific chemicals, respond
behaviourally to theses changes and then adapt to the new concentration of the chemical
stimuli” (Samanta, Singh et al. 2002). It is not clear if it is the metabolism of the substrate or
if it is the binding of the substrates to the chemoreceptors that is the crucial inducer of
chemotaxic behaviour. The NAH7 plasmid in Pseudomonas putida, which encodes the
enzymes for the degradation of naphthalene and salicylate, also encodes the chemotaxis
towards these compounds (Samanta, Singh et al. 2002). This chemotaxis in Pseudomonas
putida was found to be homologous to chemotaxis, flagellar and mobility genes from other
known E.coli bacteria. The ability to foster chemotaxis phenomenon via metabolic
influences could be important to enhance in situ bioremediation.

49. 40
Figure2.9Plasmid-encodednaphthalene(upperpathway)andsalicylate(lowerpathway)degradationgenesofNAH7catabolicplasmid
forPseudomonassp.GenesnahA-Dencodetheupperpathwayoperonwhichencodesenzymesforthedegradationofnaphthaleneto
salicylateandgenesnahG-Mencodethelowerpathwayoperon,wheresalicylateisfurtherdegradedtopyruvateand
acetylaldehyde.TheproductfromnahR(atrans-actingpositivecontrolregulator)isthepositiveregulatorforbothoperonsandis
inducedbysalicylate.Thelocationofeachrespectiveoperonpromoterisshownandlocationsofgenesencodingthenaphthalene
dioxygenasecomplexareindicated.

50. 41
2.7 DETERMINING TRANSPORT PARAMATERS
Laboratory tracer experiments are useful for flow characteristics in soil. The fundamental
mass balance of the system uses:
Inputs + Production – Outputs – Losses = Accumulation Equation 2.3
All contaminant transport and biodegradations models use this fundamental principal to
derive equations when approximating parameters such as convection and dispersion. These
assumptions allow the derivation of the partial differential equation referred to as the
convection-dispersion (Equation 2.4). The CDE is a mathematical model used in
quantitatively simulating the transport of solutes in porous media. The CDE is derived by
assuming the change in chemical flux into and out of a control volume is controlled by an
advection component (which is controlled by the velocity of the chemical), and dispersion
(which can be through of as mimicking diffusion in the sense that the dispersive flux appears
to be driven by concentration gradients) (Toride, Leij et al. 1995). The CDE for one-
dimensional transport of reactive solutes subject to adsorption, first-order degradation, and
zero-order production, in homogenous soil is given as: (Toride, Leij et al. 1993):
ࡾ
ࣔ࡯
࢚ࣔ
ൌ ࡰ ቀ
ࣔ૛࡯
ࣔ࢞૛ቁ െ ࢜ ቀ
ࣔ࡯
ࣔ࢞
ቁ െ μ࡯ ൅ ઻ሺ࢞ሻ Equation 2.4
The initial boundary conditions used to solve this equation assume that there is a fixed
known concentration of solute added to the system. This is expressed as the following:
•
డ஼
డ௫
ሺ∞, ‫ݐ‬ሻ ൌ 0 (exit condition where the concentration = 0)
• C(x,t) = 0 for x = 0
• C(x,0) = Ci (where the concentration of influent tracer is constant)
• C(0,t) = ቄ
‫ܥ‬଴
0
଴ ழ௧ ஸ ௧బ
௧ வ ௧బ
(tracer on/off after time t)

51. 42
Where:
• C : dissolved aqueous chemical concentration;
• x and t : dimensionless space and time variable respectively;
• ܴ ൌ 1 ൅
ఘ್ ௄೏
௡
(Where ߩ௕ is the bulk density; ‫ܭ‬ௗ is a partitining coefficient; and n is
the porosity);
• D : dispersion coefficient;
• ‫ݒ‬ : pore water velocity;
• ߤ : first order decay rate constant;
• ߛ : zero order production rate constant.
When decay appears in such a system it can be due to strong sorption, chemical and
biological activity, or other physicochemical interactions of the solutes in the porous media.
The inclusion of decay represents the change in dispersive flux out and assumes that decay
affects the mass inside the controlled volume. Another common form of equation 2.4 uses
a term called the Peclet number (P) which is a dimensionless number relating the amount of
advection to dispersion (P=v/D).
The transport of solutes in soil and groundwater systems includes a large number of
complicated physical, chemical, and microbiological processes (Toride, Leij et al. 1993).
Variations to the standard CDE presented in equation 2.3 have been added to account for
the simultaneous effect of sorption (including zero and first order), convective transport,
molecular diffusion, hydrodynamic dispersion, zero-order production, and first order decay
(Chen, Wang et al. 2006). Equilibrium transport processes refer to exchange reactions that
are perceived as instantaneous and are commonly described by equilibrium isotherms
including linear, Freundlich or Langmuir type. However, these equilibrium models appear to
fail in situations where chemical transport processes are not at equilibrium (Nielsen, Van
Genuchten et al. 1986). This has led to the development of non-equilibrium transport

52. 43
models that incorporate first order reactions and various chemical, kinetic and diffusion
limited rate laws to describe the non equilibrium transport. A familiar chemical non-
equilibrium model (Equations 2.5 & 2.6) includes one-site and two-site sorption. This means
sorption onto one site can be considered to be instantaneous (equilibrium) while sorption
onto the second site will be rate limited by first order kinetics (non-equilibrium) (Toride, Leij
et al. 1993). The two-site model presented is also used for physical non-equilibrium and is
termed a two-region (dual-porosity) type formulation which contains two distinct liquid
regions, one being mobile (flowing) and the other being immobile and the rate constants
refer to the mass transfer between the two regions which is modelled as a first-order
process.
ࢼࡾ
ࣔ࡯૚
࢚ࣔ
ൌ
૚
ࡼ
ቀ
ࣔ૛࡯૚
ࣔ࢞૛ ቁ െ ቀ
ࣔ࡯૚
ࣔ࢞
ቁ െ ૑ሺ۱૚ െ ۱૛ሻ െ μ૚࡯૚ ൅ ઻૚ሺ࢞ሻ Equation 2.5
ሺ૚ െ ࢼሻࡾ
ࣔ࡯૛
࢚ࣔ
ൌ ૑ሺ۱૚ െ ۱૛ሻ െ μ૛࡯૛ ൅ ઻૛ሺ࢞ሻ Equation 2.6
Where:
• subscripts 1 and 2 refer to equilibrium and non equilibrium sites respectively;
• β : partitioning coefficient of adsorption sites that equilibrates with instantaneous
and kinetic adsorption sites or mobile and immobile liquid phase;
• ω : dimensionless mass transfer coefficient (Toride, Leij et al. 1995).
The two-site equilibrium and non-equilibrium equation is incorporated into a software
package called CXTFIT developed by Torride et al. (1995), which permits one to fit a variety
of analytical solutions to the concentration distributions observed in laboratory and field
tracer studies as a function of time and/or distance. The transport of PAHs in the presence
of surfactants (Linear Alkylbenzene Sulfonate) has been successfully modelled using the
two-site model presented in equations 2.5 and 2.6 using this CXTFIT software (Chen, Wang

53. 44
et al. 2006). Noordman et al. (1998) were also successful in predicting the removal of
phenanthrene from soil using a rhamnolipid biosurfactant using a similar two site
formulation which accounted for both micellar solubilisation and admicellar sorption due to
the presence of the biosurfactant.

54. 45
CHAPTER 3 MATERIALS AND METHODS
3.1 LABORATORY SUPPLIES AND MICROOGANIM STORAGE
The important characteristics and handling procedures for the chemicals, biosurfactants,
microorganisms, and nutrient broth media purchased for this study are detailed below.
3.1.1 CHEMICALS
HPLC grade acetonitrile, hexane, acetone, and dichloromethane were purchased through
Biolab New Zealand and supplied by Mallinckrodt Baker. Naphthalene technical crystals
were from B.D.H., London, England and from Ajax Finechem, Auckland, New Zealand.
Salicylic acid crystals were from B.D.H., London, England. Phenanthrene crystals were >96%
purity HPLC grade and were supplied by Sigma Aldrich.
3.1.2 BIOSURFACTANT
Biosurfactant was purchased from Jeneil Biosurfactant Co., LLC Saukville, Wisconsin, USA.
The biosurfactant is a glycolipid produced by Pseudomonas aeruginosa with the trademark
name JBR 425. Biosurfactant stock solution contained a 25% solution of Rhamnolipids Rha-
C10-C10 (termed R1 or RLL) with the molecular formula C26H48O9 and Rha-Rha-C10-C10 (termed
R2 or RRLL) with the molecular formula C32H58O13. The rhamnolipids are an anionic
surfactant with a pKa=5.6.
3.1.3 MICROORGANISMS
The microorganisms used in this study were Pseudomonas putida ATCC 17484 (P.putida),
obtained from the American Type Culture Collection and purchased through Cryosite
Distribution in Australia. This isolate is chemoheterotropic, and is from biotype B which is
cited to degrade naphthalene. P.putida are a gram-negative rod-shaped flagellated

55. 46
bacterium which stains a pink colour when a gram stain test is performed for identification
under a microscope. P.putida are aerobic bacteria with an optimum growth temperature
between 25-30°C in a neutral pH environment. They are easily isolated from environmental
samples, and are found in most aerobic soil environments which makes it a good
representative isolate of the soil microbial consortium. For shorter storage periods, nutrient
agar plates inoculated with bacteria cultures were kept at 4°C and were re-streaked onto
fresh agar every two weeks from a single colony. For long term storage a stock of frozen
isolates, consisting of 0.5 mL of fresh overnight culture (approximately 1 x 108
– 1 x 109
cfu/mL) added to an equal volume of sterile 50% glycerol in a 1 mL sterile plastic tube, was
constructed and stored at -80˚C.
3.1.4 MEDIA AND NUTRIENT SUPPLY
All experiments, unless otherwise noted, were carried out using DifcoTM
Bushnell-Hass Broth
(BHB) as the nutrient supply. BHB is designed for study of microbial utilization of
hydrocarbons. It contains no carbon source, but provides all the trace elements necessary
for bacterial growth. It provides the monopotassium and diammonium hydrogen phosphate
to buffer the growth media, in an initial pH of 7.0 at 25°C. BHB was mixed at 3.27g/L and
autoclaved for 15 minutes at 121°C according to manufacturer’s instructions in 0.5 or 1L
increments (Table 3.1). Nutrient agar plates and Pseudomonas isolation agar plates, along
with a gram stain set with stabilized gram iodine, were purchased from Fort Richard
Laboratories, Auckland, New Zealand. Nutrient agar plates were also made in house by
adding 1.5% DifcoTM
agar to Lysogeny broth (LB). LB was purchased from USB corporation,
Cleveland, USA, and was made in a 20g/L solution which contains 10g/L casein peptone,
5g/L yeast extract, and 5g/L of sodium chloride. BHB and Agar were purchased from Becton

56. 47
Dickinson and Company, Sparks, USA. Other solutes that were used include glucose (Glucosa
1-hidrato) and sodium chloride reagent grade, both supplied by Panreac and Scharlav,
Spain.
Table 3.1 BHB marine salts broth approximate formula per litre of prepared media
Approximate Formula Chemical Formula Concentration (g/L)
Magnesium Sulfate MgSO4 0.2
Calcium Chloride CaCl2 0.02
Monopotassium Phosphate KH2PO4 1.0
Diammonium Hydrogen Phosphate (NH4)2HPO4 1.0
Potassium Nitrate KNO3 1.0
Ferric Chloride F2Cl3 0.05

57. 48
3.2 CELL CULTURING
This section details the bacteria storage and inoculant growth methods. Bacteria were
stored on agar plates, and single colonies were used to grow inoculants used in
experiments. Serial dilution plate counts and optical density at 600nm (OD600) methods are
presented to quantify the bacteria used in this study.
3.2.1 AGAR PLATES
Initially, bacteria were streaked out from a freeze-dried culture and allowed to grow at 28°C
until colonies were visible. For shorter storage periods, nutrient agar plates were inoculated
with bacteria cultures and re-streaked from a single colony onto fresh agar every two
weeks. They were grown for 24-48 hours at 28°C until colonies were large, and then stored
at 4°C to be used as inoculant for the liquid cultures. All agar plates were checked to ensure
colonies were of uniform shape, size, colour, and consistency. Periodic tests using
Pseudomonas selective agar and the gram-stain tests were employed to ensure everything
was aseptic and the only culture was indeed a gram-negative rod shaped Pseudomonas.
3.2.2 INOCULANT PREPARATION AND HARVESTING
The liquid cultures used to inoculate all aqueous phases, soil slurries, and soil column tests,
were prepared by transferring one loop from a single bacteria colony from a previously
cultivated plate, to 125 mL or 250 mL Erlenmeyer flasks with 50-100 mL of sterilized
medium. Flasks were then placed on a rotary shaker at 200 rpm and maintained at 25-30°C
overnight (12-20 hours) until the bacteria reached an OD600 of 1.0 to 2.0. This indicates that
they have reached their late exponential growth phase. Standard inoculant growth media

58. 49
(used for all tests other than those outlined in section 3.4.1.1) were a BHB broth with
glucose as per minimal media of 2g/L. After overnight growth, bacteria were transferred to
sterile centrifuge tubes and spun at 4000g for 5 minutes, the supernatant was then poured
off, and cells were re-suspended in 0.85% saline (w/v) at room temperature. This process
was repeated before trials were performed to remove residual broth or carbon sources.
Finally, cells were concentrated to an OD600 of 1.5 to 2.0 depending on the aqueous phase
or soil slurry phase experiment and used as the inoculant for the degradation trials.
3.2.3 PLATE COUNTS
The plate count method was used routinely in all experiments as a means of enumeration of
the viable bacteria present. This method involved performing serial dilutions and plating the
dilution series to obtain a dilution of 10 – 100 colony forming units (cfu), which then can be
accurately enumerated visually. Dilutions were done in either 1 mL culture tubes or in 96-
well micro plates, depending on the number of samples necessary. Serial dilutions in 96-well
plates were performed by adding 10 µL of neat solution to 90 µL of saline solution, while
avoiding immersing the pipette tips in the saline solution. New sterile pipette tips were used
to mix the contents in the well and add 10 µL of the mixture to the next well, until a 1x10-7
dilution series was complete. In the 1 mL culture tubes, the same technique was used,
however 100 µL of neat solution was added to 900 µL of saline solution and 100 µL was
transferred for each dilution. Ten microlitres from each dilution were then transferred onto
nutrient agar plates, and placed in a 28°C incubator overnight, or until colonies were large
enough to enumerate (Figure 3.1). All serial dilutions were done in duplicate or triplicate.
The amount of 10 µL was chosen as the transfer volume to the plate because it evaporated
and soaked into the agar within a few minutes and did not allow bacteria to become

59. 50
detached from the growing colonies. This method also allowed an entire dilution series to
be plated on a single agar plate. Final calculations for cfu/mL follow the formula:
cfu/mL = (colonies on the plate) * 10(dilution# + 2)
Equation 3.1
3.2.4 OPTICAL DENSITY
Optical density (OD) at 600nm was also used as an indicator of cell density in solution. A
standard curve for OD600 and the density of cells per mL was constructed to calibrate the
absorbance value and obtain comparisons between readings. OD was measured by
transferring 1 mL of solution to a disposable UV-cuvette. The same spectrophotometer. The
spectrophotometer was used for all readings, and was zeroed by using the sample matrix
from a sterile stock solution.
CFU/ML ??
100/10 µL 100/10 µL 100/10 µL 100/10 µL 100/10 µL 100/10 µL 100/10
90/900 µL saline
1 2 3 4 5 6 7
x2 duplicate
10 µL of sample
Incubated the plate at 28ºC until colonies
visible and counted dilutions with
number of colonies between 10 and 100
cfu/mL = (colonies on the plate) * 10 (dilution# + 2)
Figure 3.1 Schematic diagram for serial dilutions to determine cfu/mL of solution

60. 51
3.3 SOIL METHODS
Soil was obtained from a laboratory supply (original location unknown) of mixed sand and
silt, and the properties are outlined in the following section. The artificial soil contamination
and contaminant extraction techniques are also presented, along with the efficiency of the
contaminant extraction method.
3.3.1 SOIL PROPERTIES
The particle size distribution of the soil used in all experiments was determined using
American Society for Testing and Materials (ASTM) D6913 standard testing methods for
gradation of soils, using sieve analysis (Figure 3.2). Any material larger the 2mm opening
was discarded to obtain a homogeneous mixture that would be appropriate for bench-scale
testing. According to the ASTM standards, sand is 2mm to 0.05mm; silt is 0.05mm to
0.002mm; and clay is less than 0.002mm. The soil used can therefore be classified as loamy
sand using the soil texture triangle. The loss on ignition test method ASTM D2974-87 was
used to determine the organic content of the soil (Table 3.2). Soil pH was determined in a
1:1 soil slurry with distilled water. Porosity and density were determined gravimetrically
when soil column were packed (Table 3.2 Soil properties
Table 3.2 Soil properties
Soil Parameter Symbol/Units Value
Organic Content % 2.15
pH --- 5.4
Dry Density ρd /(g/cm3
) 1.75
Unit Weight γd (kN/m3
) 17.2
Specific Gravity Gs 2.8
Average porosity n 0.38

61. 52
Figure 3.2 Particle size distribution
3.3.2 SOIL CONTAMINATION
Soil was sterilized by autoclaving it at 120°C in 100g increments three times, after which the
soil samples were plated on nutrient agar plates to ensure continued sterility. Sterile dry soil
was placed in 1 L shot bottles and spiked with phenanthrene dissolved in acetone. It was
then shaken vigorously for 5 minutes to promote homogeneous distribution of
phenanthrene in the soil. The amount of acetone added was sufficient to completely
saturate the soil, without producing excess liquid after shaking. Acetone was then
evaporated by allowing the sample to rest for 3 days at 30°C under a fume hood. The
contaminated soil was aged between 2 weeks and 1 month before each experiment. After
contamination, soil was re-autoclaved (as it was proven not to be sterile) and the
0.00
20.00
40.00
60.00
80.00
100.00
0.010.1110
PercentPassing
Particle Size (µm)