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Evaluation of Bacterial Strategies to Degrade Non-desorbable
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Wateoueindd-pJS

EVALUATION OF BACTERIAL STRATEGIES TO DEGRADE NON-
DESORBABLE NAPHTHALENE IN NATURAL AND ARTIFICIAL SORBENTS
By

Mohammad Sajjad

A DISSERTATION

Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY
Department of Civil and Environmental Engineering

2005

ABSTRACT

EVALUATION OF BACTERIAL STRATEGIES TO DEGRADE NON-
DESORBABLE NAPHTHALENE IN NATURAL AND ARTIFICIAL SORBENT

By
Mohammad Sajjad

Recently, several studies demonstrated that bacteria might be able to degrade
sorbed contaminants without prior desorption in water. Mechanisms including extra-
cellular biomaterial production, chemotaxis/motility, and adhesion were hypothesized to
be responsible for the enhanced bioavailability of these compounds. However,
experimental evidence to confirm this bacterial potential in the soil system has not been
verified. In this study we evaluated the role of the selected bacterial phenotypes in
naphthalene degradation in a sorbent-slurry system. The rate and extent of sorbed-phase
naphthalene was determined in the sorbent-slurry systems inoculated separately with the
wild-type Psuedomonas putida G7 and the mutant strains (with one or other missing
phenotypes), and the results were compared for evaluating the role of these phenotypes.

The experimental results showed that wild-type P. putida G7 degraded sorbed
naphthalene at a rate and extent greater than what can be removed from both Tenax beads
and serial extraction methods. Activity of extra-cellular enzyme was ruled out during the
biodegradation study since the experiments with bacterial filtrates did not show any
degradation of observable free naphthalene during the incubation period. The possible
effect of biomaterial on the sorption/desorption parameters was also observed to be
negligible. The amount of the biomaterial produced during the bio-availability studies

was small enough to have caused any significant effect on mass transfer parameters.

This study demonstrates that the effect of chemotaxis/motility phenotypes could
not be observed in natural soils under both stirred and quiescent conditions. In a stirred
system, the mixing had masked the role of chemotaxis/motility due to maintenance of
isotropic concentration and mixing-mediated adhesion of cells to sorbent particles. Under
quiescent conditions, wild-type P. putida G7 and mutant strains were not able to degrade
the non—desorbable fraction of naphthalene because of pore size exclusion and retardation
of bacterial cell in natural soil. However, the role of chemotaxis/motility was apparent in
the system containing activated carbon, where wild-type degraded the non-desorbable
naphthalene fraction at a greater rate and extent as compared with the system containing
mutant strains. The difference in uptake rates between wild-type and mutant strains was
due to the presence of a significant concentration gradient at the soil/water interface as a
result of a large pool of non-desorbable naphthalene fraction in the activated carbon.

This study also attempts to examine factors affecting bacterial attachment that,
consequently, enhances bio-degradation. The wild type P. putida G7 and the mutant
strains, deficient in adhesion, were observed to have equally accessed non-desorbable
fraction of naphthalene and degraded the fraction at an equal rate and extent. Three
orders of decrease in molar concentration of the buffer medium did not affect the rate and
extent of sorbed phase naphthalene degradation by P. putida G7. Important attachment
phenotypes of P. putida G7 did not play any significant role in enhanced bioavailability
of sorbed phase naphthalene. The results of this study suggest that factors promoting
bacterial attachment (i.e. removal of exo-polysaccharides, presence of flagella and
reduction of buffer medium concentration) do not contribute to biodegradation of non-

desorbable fraction of soil-sorbed naphthalene in a mixed soil-slurry system.

DEDICATION

To my courageous mother, Bacha Bibi

ACKNOWLEDGMENTS

I am genuinely grateful to my major professor, Dr. Thomas C. Voice, who gave me
biggest breakthroughs in my life. He first introduced me to the teaching profession by
selecting me for the position of lecturer at National University of Sciences and
technology, Pakistan. He later gave me the Opportunity to pursue my PhD at the
Michigan State University. Through out my PhD, he supported my financial needs by
providing me Research/Teaching assistantships. Dr. Voice had guided me in formulating
my research ideas and gave me the confidence to overcome the obstacles during the
endeavor. I will always be grateful to him for what he did for me and, for reposing
confidence and trust in me. I also appreciate other committee’s members: Dr. Stephen
Boyd, Dr. Susan Masten and, Dr. Phanikumar Mantha for their help, guidance,
encouragement and, advices throughout the research.

I am indebted to Dr. Jeong-Hun Park who introduced me to the subject of bioavailability,
and helped me with initial laboratory work. I benefited a great deal from the useful
discussions I had with visiting assistant professors: Dr. Xianda Zhao and , Dr. Michael
Dybas and fellow graduate students: Chris Saffron, and Irfan Aslam. I am indeed grateful
to Dr. Tamara Tsoi, who taught me the right way of doing microbiology.

Last but not least, I owe a great debt to my wife, Attiya Sajjad, who took good care of my
emotional and physical well-being. I am also indebted to my sweet daughters: Dania
Sajjad and Zoha Sajjad, for their cooperation and understanding. Also, I thank my

parents, Amir khan and Bacha Bibi, who always loved and prayed for me when I need it

most.

TABLE OF CONTENTS

Chapter 1. Introduction .......................................................................... 1
Introduction ......................................................................................... l
Hypotheses .......................................................................................... 7
References ......................................................................................... 10

Chapter 2. Literature Review .................................................................. 13
Introduction ........................................................................................ 13
Affect of Sorption/Desorption kinetics on Bio-availability .................................. 14

Sorption Process ................................................................................ 15
Desorption Process ............................................................................. 17
Influence of Microbial Kinetics on Bio-availability ......................................... l9
Biodegradation by growing organisms ...................................................... 20
Biodegradation by non-growing organisms ................................................ 21
Bioavailability Models ........................................................................ 22
Bacterial Physiological Consideration ......................................................... 24
Bacterial Chemotaxis and Enhance Bioavailability ......................................... 25
Bacterial Adhesion ............................................................................. 28
Structure of Gram-negative cell surfaces ..................................................... 29
Mechanism of Adhesion ...................................................................... 33
References ......................................................................................... 37

Chapter 3. Enhanced Bioavailability of Sorbed phase Naphthalene: A comparison of

biodegradation and desorption kinetics ........................................ 42
Abstract .......................................................................................... 42
Introduction ..................................................................................... 43
Material and Methods ......................................................................... 46

vi

Data Analysis ..................................................................................... 51

Tenax Model .................................................................................... 51
Coupled Desorption/Biodegradation Model (DBM model) .............................. 52
Results ............................................................................................. 53
Sorption and Desorption ...................................................................... 53
Removal of Naphthalene in soil-slurry system by Tenax ................................. 56
Biodegradation of Naphthalene by PpG7 in soil-slurry .................................... 59
Effect of Extracellular Biomaterial on Sorption/Desorption .............................. 59
Discussion ......................................................................................... 65
Comparison of biodegradation rates and abiotic extraction rate ........................ 65
Extracellular biomaterial and enhance bio-availability .................................... 69
References ......................................................................................... 72

Chapter 4. Affect of Chemotaxis and Motility on the Biodegradation on N on-desorbable

Naphthalene in natural and synthetic sorbent ....................................... 74
Abstract ............................................................................................ 74
Introduction ....................................................................................... 75
Material and Methods ........................................................................... 79
Data Analysis: Desorption-Sorption Model (DBM) .......................................... 88
Results and Discussion ........................................................................... 91

Sorption of Naphthalene ...................................................................... 91
Desorption of Naphthalene ................................................................... 93
Determination of Non-Desorbable Fractions ............................................... 95
Characterization of Strains .................................................................... 98
Naphthalene degradation kinetics in Soil Extract ........................................ 100
Bioavailablity of Sorbed Naphthalene in Natural soil under mixed condition ....... 100
Bioavailabilty of sorbed Naphthalene in Natural soil under static condition ......... 105

Bioavailablity of sorbed Naphthalene in Activated Carbon under mixed system ...109
Discussion ....................................................................................... 1 12

Conclusions ....................................................................................... l 18

vii

References ....................................................................................... 120

Chapter 5. Influence of bacterial Adhesion on Enhanced Bioavailability of

Naphthalene in natural soil .................................................... 122

Abstract ........................................................................................... 122
Introduction ....................................................................................... 123
Bacterial Attachment and Enhance Bio-availability ....................................... 123
Factors affecting bacterial attachments .................................................... 125
Objectives of the study ....................................................................... 128
Material and Methods ........................................................................... 129
Data Analysis: Desorption-Bioavailability Model (DBM) ................................. 135
Results ......................................................................................... 138
Sorption, Desorption, and Serial Dilution Experiments ................................. 138
Cell surface properties ....................................................................... 141
Bioavailability experiments at different Molar Concentration of Buffer medium... 142
Isolation of adhesion deficient mutants ..................................................... 146
Bioavailability Experiment with Wild-Type and Mutant Strains ....................... 147
Discussion ....................................................................................... 148
References ....................................................................................... 156
Chapter 6. Summary and Conclusions .................................................... 159
Summary ......................................................................................... 159
Conclusions ...................................................................................... 162

viii

Table 3-1.

Table 3-2.

Table 3-3.

Table 3-4.

Table 3-5.

Table 3-6.

Table 3—7.

Table 4-1.

Table 4-2.

Table 4-3.

Table 4-4.

Table 4-5.

Table 4-6.

Table 4—7.

Table 5-1.

Table 5-2.

Table 5-3.

Table 5-4.

LIST OF TABLES

Selected properties of soils used in the study ..................................... 47
Parameters for sorption of naphthalene by the A horizon of selected soils. . .55
Desorption parameters evaluated by a three site-desorption model ............ 58

Rate co-efficient of naphthalene removal by PpG7 and Tenax in soil-slurry
slurry system ........................................................................... 60

Naphthalene concentration before and after addition of cell filterate .......... 61

Distribution co-efficient of naphthalene to colwood at different amount

of biomaterial ........................................................................... 64
Non-desorbable fraction from Tenax and serial dilution methods .............. 64
Selected properties of soils used in the study ...................................... 81

Parameters for soption of Napthahlene by the A hoizon of selected soil and
granular activated carbon ............................................................ 92

Desorption parameters evaluated by a three-site desorption model... ... .......94
Fraction of desorbable naphthalene by three methods in natural soil and

granular activated carbon ............................................................ 96
Biokinetic parameters evaluated for three strains of PpG7 ..................... 106

Estimated bio-kinetic co—efficient for aqueous fraction (keq) for wild-type
and non-motile strains in CapacA and Muck soil under static condition. . ..l 10

Estimated bio-kinetic co-efficients for desorbable phase (keq) and non-
desorbable phase (knd) for wild type and non-motile strain in GAC.........111

Selected properties of soils used in the study .................................... 130
Parameters for sorption of naphthalene by the A hoizon of selected soils... 139
Desorption parameters evaluated by a three—site desorption model ........... 140

Cell Surface and hydrophobicity of P. putida G7 ............................... 142

ix

Table 5-5. Biokinetic parameters evaluated for pseudomonas putida G7-WT under
varying molar concentration of chemotactic buffer solution (CB) ............ 144

Table 5-5. Biokinetic parameters evaluated for non-adhesion mutant along with three
other strains of PpG7 ................................................................ 149

LIST OF FIGURES

Figure 2-1. Description of Desorption-Biodegradation model (DBM) ..................... 23

Figure 2-2. Adhesion as a bacterial strategy to directly access sorbed chemical at solid
surfaces ............................................................................... 29

Figure 2-3. Schematic of outer membrane of gram-negative bacteria bacteria along with

its associated LPS layer ............................................................ 30

Figure 3-1. Sorption iSOptherm of naphthalene in natural sorbents ......................... 55
I

Figure 3-2. Kinetic desorption data for naphthalene form three soils ...................... 57

Figure 3-3. Sequential desorption data for the removal of naphthalene from three soil..57
Figure 3-4. Desorption of naphthalene from three natural soils .............................. 58

Figure 3-5. Extraction of Naphthalene by Tenax beads in presence and absence of soils

........................................................................................... 60
Figure 3-6. Degradation of naphthalene by PpG7 in mixed soil-slurry system ............ 61
Figure 3-7. Production of Biomaterial (TOC) by PpG7 at different naphthalene

concentration .......................................................................... 63
Figure 3-8. Decrease in partition co-efficient of naphthalene with increase biomaterial

in soil-slurry system ................................................................... 63

Figure 3-9. Removal of naphthalene by PpG7, serial dilution, and Tenax beads in Muck

soil ...................................................................................... 68
Figure 3-10. Removal of naphthalene by PpG7, serial dilution, and Tenax beads in

Capac A soil .......................................................................... 68
Figure 4-1. Configuration for nitrogen-sparging of naphthalene in liquid phase .......... 89
Figure 4-2. Different compartments of desorption-bioavailability model(DBM) ......... 89
Figure 4-3. Sorption isotherms of naphthalene in natural and synthetic sorbent ........... 92
Figure 4-4. Desorption of naphthalene from natural and synthetic sorbent ................ 94
Figure 4-5. Serial dilution desorption in three natural soil and activated carbon... ........97

xi

Figure 4-6. TEM for PpG7(WT) on left, and non-motile mutant on right ................. 99
Figure 4-7. Naphthalene degradation of three strains in liquid medium ................... 99
Figure 4-8. Naphthalene degradation by PpG7(WT) in three natural soils ............... 102

Figure 49 Naphthalene degradation by PpG7 (non-chemotactic mutant) in three
natural soils .......................................................................... 103

Figure 4-10. Naphthalene degradation by PpG7(non-motile mutant) in natural soil. . 104

Figure 4-11. Naphthalene degradation by three strains in muck soil under static

conditions ........................................................................... 108
I

Figure 4.12. Naphthalene degradation by three strains in Capac-A soil under static

condition ............................................................................ 108
Figure 4.13. Degradation of naphthalene by wild-type and non-motile strain in activated

carbon (GAC) ........................................................................ 11 1
Figure 5.1. Sorption isotherm of naphthalene in natural sorbents ......................... 139

Figure 5-2. Desorption of naphthalene from three natural soil and activated carbon... 140

Figure 5-3. Percent of bacteria adsorbed to hexadecane in BATH test .................... 141
Figure 5-4a. Naphthalene degradation by PpG7(WT) in IOOmM CB solution ........... 144
Figure 5-4b. Naphthalene degradation by PpG7(WT) in 1.0 mM CB solution ........... 145
Figure 5-4c. Naphthalene degradation by PpG7(WT) in 0.1 mM CB solution ........... 145
Figure 5-5. Cell attachment to colwood soil at different time intervals ................... 147

Figure 5-6. Naphthalene degradation in soil-slurry solution by PpG7(WT) and its
three mutants: non-chemotactic, non-motile, and non-adhesion ............. 149

xii

CHAPTER 1. INTRODUCTION

INTRODUCTION

Recalcitrant organic compounds such as polycyclic aromatic hydrocarbons
(PAHs) are ubiquitous in the biosphere due to anthropogenic activities. PAHs are a class
of very stable organic molecules with a fused-ring chemical structure. Some of these
compounds, with intermediate to high molecular weight e.g. bezop[a]pyene and
bez[a]anthracene, are suspected carcinogen or mutagen, and are recognized by the
USEPA as priority pollutants (keith and Telliard 1979). The main source of these
compounds is fossil fuels, and these compounds may enter into the environment via point
( e.g., automobile and refinery exhausts) or non-point( e.g., oil spill and urban run-off)
sources. PAHs persist in the natural environment because of their hydrophobic nature
and high molecular weight, resulting in low water solubility and low volatility. As a
result, these compounds selectively partition into soil or sediment that in turn serves as a
long-term source of PAHs contamination to ground and surface water.

Several treatment options such as pump and treat, air-sparging, air-stripping, and
activated carbon adsorption have been used to remove the organic contaminants.
However, these technologies failed to deliver due to the associated high cost to remediate
large volumes of contaminated sites, incomplete removal of contaminants to meet
regulatory standards, and large waste by-products that need to be disposed of (EPA.
1995). Bio-remediation has emerged over‘the last two decades as a favored and attractive
alternative, which has the potential to successfully clean up the contaminated site without

having the disadvantages associated with traditional physiochemical methods (Head

1998). The initial success of bio-remediation was, however, limited to easily
biodegradable, low molecular weight or volatile organic compounds and soon it was
realized that several issues have to be addressed before this technology can successfully
be applied to soils containing recalcitrant pollutants such as PAHs (Mueller et a1. 1996;
Wilson and Jones 1993)

Two major factors are attributed to incomplete remediation of PAHs
contaminated sites. First, high molecular weight PAHs (for e.g., PAHs with five or more
fused rings like benzo-pyrene) are recalcitrant to biodegradation; and, only low to
intermediate molecular weight PAHs degrading bacteria have been isolated from
contaminated sites (Heitkamp et a1. 1987; Mueller et a1. 1996). Second, due to the
hydrophobic nature of these compounds they have a high tendency to sorb to soil and
sediment’s matrix. As a result, sorbed molecules may commonly be unavailable for the
direct uptake of soil bacteria. For sorbed molecules to be available for bacterial uptake,
they have to diffuse out within the nano- and micropores to soil-water interface and,
finally, diffuse through the aqueous phase to the bacterium.

In dealing with biodegradable compounds, the success of bioremediation,
therefore, depends on the interplay of two steps 1) the pollutant release rate in the soil
matrix and 2) the bacterial uptake rate. These two steps occur sequentially, and the
overall bioremediation rate can be controlled and limited by any one of these two steps.
The term bioavailability is often used to describe the rate of bio-remediation, and is
defined as a ratio of rate of transfer of compound from the soil to the rate of uptake by the
bacteria. Bioremediation of soils contaminated with HOCs is often limited by mass

transfer from soil matrix to bacterial membrane and is referred to as limted bio-

availability. Several laboratory and field studies have demonstrated that slow release of
biodegradable PAHs from soil matrix to aqueous phase is often the rate limiting step in
complete biodegradation of these compounds (Beck et al. 1995; Luthy et al. 1994).

Bio-availability of PAHs has been identified as one of the major research
priorities for bioremediation technologies (US-EPA 1991). Recent research has suggested
that even unsuccessful bioremediation of high molecular weight PAHs is mainly due to
low bio-availability of these compounds, rather than compounds being recalcitrant to
biodegradation (Mihelcic et al. 1993). In the last two decades, extensive research has
been conducted to identify factors affecting bio-availability of recalcitrant organic
compounds to microbial degradation, and a review of the literature was recently
published (Baveye 1999). These studies concluded that bioavailability essentially
depends on interaction between the soil matrix, the contaminant, and the micro-organism;
further, these interactions are in turn dependent on composition and structure of soil
matrix, physico-cherrrical nature and residence time of contaminant, and the type and
state of the micro-organisms.

Until recently, release of sorbed molecules from soil matrix to solution phase has
been considered a pre-requisite for complete bio-remediation (Alvarez-Cohen and Speitel
2001; Ogram et al. 1985; Scow and Alexander 1992). Biodegradation in soils and
sediments was considered de-sorption or diffusion rate limited. It was assumed that de-
sorption rate (release rate from soil matrix) depends only on the interaction between
composition/structure of soil matrix and the physiochemical nature and residence time of
the contaminant. Whereas considerable effort has been undertaken to study the

characteristics of soil-chemical interactions in soil, little was known about the pollutant

mass transfer-enhancing strategies of the bacteria. The role of bacterial phenotypes
(physiology, adaptation etc) on rate of de—sorption or direct uptake of sorbed phase has
been neglected or ignored. The bio-remediation technology is still based on this
widespread and common paradigm and, consequently, bioremediation cleanup endpoint
is based on a high concentration of PAHs that is sequestered in soil matrix and hence un-
available for degradation.

Recent studies have challenged this paradigm based on the evidences that bacteria
may be able to degrade sorbed contaminant without prior desorption in water (Bestiaens
et al. 2000; Calvillo and Alexander 1996; Guerin and Boyd 1992; Tang et al. 1998).
These previous studies suggested that while some bacteria can only uptake contaminant
in an aqueous phase, others might be capable of accessing contaminant in the sorbed
phase as well. Thus the researchers who conducted theses studies concluded that there
exists two physiologically distinct strains of PAH degrading bacteria: one able to degrade
PAHs in solution only, and the other capable of degrading PAHs sorbed to soil. This line
of evidence clearly shows that bioavailability of a contaminant also depends on the type
of PAHs degrading bacteria in the system. This new find has important ramifications for
the interpretation of contaminant bioavailability and remediation endpoints that are still
based on kinetic studies of isolate capable of degrading PAHs in solution, only.

All this led to a concept of enhanced bioavailability, which is based on the fact
that bacteria can also access directly to the sorbed phase, and thus the degradation rate of
contaminant can exceed those predicated from aqueous phase contaminant only.
Mechanisms responsible for the enhanced bio-availability of these compounds in soil

systems is still not clear (Dean et al. 2001; Harms and Zehnder 1995). Recent studies

hypothesized that specific physiological characteristics of the PAH degrading bacteria
might contribute to enhance bioavailability. The mechanisms promoting enhanced bio-
availability include: (1) production of extra-cellular biomaterials (enzymes and polymers)
that either directly degrade inaccessible substrate or facilitate enhanced desorption of
organic chemicals (Dean et al. 2001; Liu 2001; Mata-Sandoval et al. 2000; Vandyke et al.
1993); (2) Chemotaxis trait that guides bacteria to swim using flagella to locate substrate
concentration in soil; (3) reduction of the distance between cells and substrate by means
of adhesion-promoting bacterial-surface or structures such as flagella and lipo-
polysaccharide (Bastiaens et al. 2000; Harms and Zehnder 1995).

The first mechanism stipulates that the biomaterial produces by bacterial enhances
mobilization and release of soil bound organic hydrophobic contaminants (Dean et al.,
2001; Liu, 2001; Mata-Sandoval et al., 2000; Schippers et al., 2000; Vandyke et al.,
1993). The enhancement was attributed mainly to increases in apparent solubility of
contaminants due to dissolved biomaterial. The water soluble biomaterials bind
contaminant in solution, decreasing the solution concentration of the free, unbound
contaminant molecule. This would have resulted in an increase in driving force for
desorption of sorbed contaminant. It was expected that due to better solubalization
properties of excreted biomaterial, an increase in biavailabilty of sorbed organic
contaminants would result.

The second mechanism suggests that bacteria with chemotactic and motile
capability actively seek target chemo-attractants. Chemotaxis ability may bring motile
bacteria into close physical contact with soil-sorbed chemoattractants when soluble

substrates are depleted. Several bacterial strains have been reported to be chemotactic

towards environmental pollutants such as benzene, toluene, trichloroethylene, biphenyl,
naphthalene, p-nitrophenol and carbontetrachloride (Grimm and Harwood, 1997b;
Parales et al., 2000; Witt et al., 1999; Samanta et al., 2000). Chemotactic Pseudomonas
putida G7 can degrade naphthalene more rapidly than a mutant deficient in Chemotaxis
towards naphthalene and a motility-deficient in capillary tube (Marx and Aitken, 2000).
It has been shown that flagella-mediated motility is important for the cells to make
contact with a surface (Lawrence et al., 1987; O’Toole and Kolter, 1998b). Motility
facilitates attachment by bringing the cell close to the surface to overcome repulsive
electrostatic forces between the bacterium and the surface. Although it has been studied
intensively in liquid media, there is evidence that Chemotaxis may guide bacteria to seek
localized concentration gradient in soil matrix (Wall and Kaiser, 1999).

The third mechanism suggests that bacteria with a strong tendency for adhesion in
the soil may have a selective advantage to access sorbed compounds (Harms and
Zehnder, 1995; Calvillo and Alexander, 1996; Tang et al., 1998). Surfaces containing
sorbed chemicals create a microenvironment (e.g., a steeper concentration gradient and a
shorter diffusion distance) that is different from the surrounding bulk liquid. A higher
chemical concentration near the sorbent surface may render the chemical more readily
available to attached bacteria than to bacteria present in the aqueous phase. Bacterial
characteristics contributing to cell adhesive properties include cell surface hydrophobicity
and charge, the presence of extra-cellular polymers, and the presence of cellular
appendages (i.e., flagella and pili). For instance, members of Coryneform bacteria exhibit
increasing cell surface hydrophobicity due to the presence and increasing chain length of

mycolic acids in their cell wall, which in turn is related to increased adhesion to Teflon

surfaces (Bendinger et al. 1993). Special cell surface structures (e.g., pili or extracellular
polymers) help form strong links between cell and solid surface as well as micro-colonies
or biofilm development. Type IV pili (thin filaments that extend from the pole of a
bacterial cell) found in some soil bacteria play an important role in microcolony
formation (cell to cell interaction) and mediate surface-associated movement known as
social gliding and twitching motility (O’Toole and Kolter 1998b; Wall and Kaiser 1999).

This study was undertaken to research the role of specific physiological traits of
bacteria on enhance bioavailability of soil-sorbed organic contaminant. All of the
identified mechanisms were tested in a systematic way in stirred soil-slurry reactors using
a representative PAH, naphthalene. A wild-type naphthalene degrading bacteria
(Pseudomonas Putida G7) along with its mutants strains (non-chemotectic, non-motile,
and non-adhesive) were used for studying these mechanisms. The pure wild type P.
putida G7 has been reported previously to uptake sorbed phase naphthalene directly in
soil-slurry. Furthermore, it has been characterized as being motile and, chemotectic
strain.

HYPOTHESES

The underlying hypotheses for this research work are as follows:
1)Pseudomonas putida G7 (PpG7) uptakes naphthalene intra-cellularly and does not
produce extra-cellular catabolic enzyme in response to substrate limiting condition. PpG7
also excretes biomaterial in suspending solution that affect sorption/desorption
coefficient of naphthalene to soil and hence affect bioavailability of naphthalene.
2) Chemotaxis and motililty are important traits that enable wild-type P. putida G7 to

access localized concentration of pollutant in soil. Alternatively, chemotaxis and motility

deficient mutants of P. putida G7 cannot utilize sorbed naphthalene at the same rate as
the wild-type.

3) Attachment phenotype enables wild-type P. putida G7 to access localized and strongly
sorbed concentration of pollutant at the particle-water interface. A wild-type P. putida G7

degrade sorbed-phase molecules at greater rate than its adhesion-deficient mutant.

To address these hypotheses, this research work involved generating, isolating,
and characterizing mutants with one of the missing phenotypes (motility, chemotaxi,
adhesion). Sorption isotherms and single and series dilution desorption experiments were
conducted to evaluate distribution coefficient, de-sorption rate, and amount of non-
desorbable naphthalene. The initial degradation rate was measured for wild type as well
as for all mutants in soil extract solution. All the studies were carried out assuming that
initial degradation kinetics of all mutants remained same as that of wild type strain.
Bioavailability experiments involved first establishing sorption equilibrium, inoculating
the systems with bacteria, and measuring naphthalene concentrations in both sorbed and
dissolved phases over a period of time. Rate of naphthalene degradation by each culture
in stirred batch systems were fitted with non-linear regression model (DBM model) to
determine sorbed phase degradation rate coefficients. Bio-availability data were
compared for each type of organism for extent and rate of sorbed phase degradation to
evaluate role of each phenotype.

The outcome of this research is expected to demonstrate the role of
microorganism’s physiological traits in enhancing bio—availability of sorbed contaminants

in contaminated soil. The finding of this research work will help to improve

bioremediation technologies by incorporating microorganism equipped with one or more
characteristics that are needed for microorganism to have direct access to sorbed
contaminants. Consequently, the research finding may be instrumental in understanding
enhance bioavailability, which play a major role in clean-up time, cost and end-points of

the bio-remediation efforts.

REFERENCES

Alvarez-Cohen, L., and Speitel, G. E. (2001). "Kinetics of aerobic cometabolism of
chlorinated solvents." Biodegradation, 12(2), 105-126.

Bastiaens, L., Springael, D., Wattiau, P., Harms, H., deWachter, R., Verachtert, H., and
Diels, L. (2000). "Isolation of adherent polycyclic aromatic hydrocarbon (PAH)-
degrading bacteria using PAH-sorbing carriers." Applied and Environmental
Microbiology, 66(5), 1834-1843.

Baveye, P. (1999). Bioavailablity of organic Xenobiotics in the environment: practical
consequences for the environment, Kluwer Acedemics Publisher.

Beck, A. J ., Wilson, S. C., Alcock, R. E., and Jones, K. C. (1995). "Kinetic Constraints
On the Loss Of Organic-Chemicals From Contaminated Soils - Implications For
Soil-Quality Limits." Crit. Rev. Environ. Sci. Technol., 25(1), 1-43.

Bestiaens, L., Springael, D., Waittiau, P., and Diels, L. (2000). "Isolation of adherent
polycyclic aromatic hydrocarbon (PAH)- degrading bacteria using PAH-sorbing
carriers." Applied and Environmental Microbiology, 66, 1834-1843.

Calvillo, Y. M., and Alexander, M. (1996). "Mechanism of microbial utilization of
biphenyl sorbed to polyacrylic beads." Appl. Microbiol. Biotechnol., 45(3), 383-
390.

Dean, S. M., Jin, Y., Cha, D. K., Wilson, S. V., and Radosevich, M. (2001).
"Phenanthrene degradation in soils co-inoculated with phenanthrene-degrading

and biosurfactant-producing bacteria." Journal of Environmental Quality, 30(4),
1 126-1 133.

EPA., U. (1995). "How to Evaluate alternative cleanup technologies for underground
storage Tank sites: A guide for corrective Action plan Reviewers." A guide for
corrective Action plan Reviewers, (EPA 510-B-95-007).

Guerin, W. F., and Boyd, S. A. (1992). "Differential Bioavailability of Soil-Sorbed
Naphthalene to Two Bacterial Species." Appl. Environ. Microbiol, 58(4), 1142-
1152.

Harms, H., and Zehnder, A. J. B. (1995). "Bioavailability of Sorbed 3-
Chlorodibenzofuran." Appl. Environ. Microbiol, 61(1), 27-33.

Head, I. M. (1998). "Bioremediation: towards a credible technology." Microbiology-UK,
144(3-4), 599-608.

10

Heitkamp, M. A., Freeman, J. P., and Cemiglia, C. E. (1987). "Naphthalene
Biodegradation in Environmental Microcosms: Estimates of Degradation Rates
and Characterization of Metabolites." Appl. Environ. Microbiol., 53(1), 129-136.

Keith, L. H., and Telliard, W. A. (1979). "Priority pollutants 1- a perspective view."
Environ. Sci. &. Technol, 13, 416-423.

Liu, A. (2001). "Phenanthrene Desorption from soil in the presence of Bacterial
extracellular Polymer: Observation and Model Predictions and Dynamic
behavior." Water Research, 35(3), 835-843.

Luthy, R. G., Dzombak, D. A., Peters, C. A., Roy, S. B., Ramaswarrri, A., Nakles, D. V.,
and Nott, B. R. (1994). "Remediationg Tar-Contaminated Soils at Manufactured
Gas Plant Sites." Environ. Sci. Technol, 28(6), 266A-276A.

Mata-Sandoval, J. C., Kams, J ., and Torrents, A. (2000). "Effect of rhamuolipids
produced by Pseudomonas aeruginosa UG2 on the solubilization of pesticides."
Environmental Science & Technology, 34(23), 4923-4930.

Mihelcic, J. R., Lueking, D. R., Mitzell, R. J ., and Stapleton, J. M. (1993).
"Bioavailability of sorbed- and separate-phase chemicals." Biodegradation, 4,
141-153.

Mueller, J. G., Cemiglia, C. E., and Pritchard, P. H. (1996). "Bioremediation of
environments contaminated by polycyclic aromatic hydrocarbons."
Bioremediation: Principles and Applications, R. L. Crawford and D. L. Crawford,
eds. ; Cambridge University press, -(-), 125 - 194.

Ogram, A. V., Jessup, R. E., Ou, L. T., and Rao, P. S. C. (1985). "Effects of Sorption on
Biological Degradation Rates of (2,4-Dichlorophenoxy)acetic Acid in Soils."
Appl. Environ. Microbiol, 49(3), 582-587.

Scow, K. M., and Alexander, M. (1992). "Effect of Diffusion on the Kinetics of
Biodegradation." Soil Sci. Soc. Ame. J, 56(1), 128-134.

Tang, W. C., White, J. C., and Alexander, M. (1998). "Utilization of sorbed compounds
by microorganisms specifically isolated for that purpose." Appl. Microbiol.
Biotechnol., 49(-), 117-121.

US-EPA. (1991). "Summary and Report High-Priority Research on Bioremediation."
Bioremediation Research Needs Workshop, Washington, DC.

Vandyke, M. 1., Couture, P., Brauer, M., Lee, H., and Trevors, J. T. (1993).
"Pseudomonas-Aeruginosa Ug2 Rhamnolipid Biosurfactants - Structural
Characterization and Their Use in Removing Hydrophobic Compounds from
Soil." Canadian Journal of Microbiology, 39(11), 1071-1078.

11

Wilson, S. C., and Jones, K. C. (1993). "Bioremediation of soil contaminated with
polynuclear aromatic hydrocarbons (PAHs): A review." Environ. Poll, 81(-), 229-
249.

12

CHAPTER 2. LITERATURE REVIEW

INTRODUCTION

Bioremediation of soil contaminated with polycyclic aromatic hydrocarbons
(PAHs) and other hydrophobic organic compounds (HOCs) is believed to be limited by
slow transfer of these compounds from the soil matrix to the aqueous phase (Harms and
Bosma 1997; Ogram et al. 1985; US-EPA 1991). Many organic compounds that are
easily biodegradable in solution have been found to persist when sorbed to a soil or
sediment matrix. The (bio) availability of sorbed compound to bacteria via the aqueous
phase is, therefore, considered a limiting factor for successful bio-remediation.
Bioavailability is considered a dynamic process dependent on the interplay between
physicochemical and microbial processes (Luthy et al. 1994; Pignatello 1999). Any
change in either one of these processes will affect bioavailability and, hence, the rate and
extent of bioremediation.

The physicochemical processes include sorption, desorption, intra-aggregate or intra-
particle diffusion, and sequestration. These processes, in turn, are dependent on the
characteristics of the chemical, the soil matrix, and the residence time of the chemical.
Whereas considerable effort has been undertaken to study the characteristics of soil-
chemical interactions in soil, little was known about the uptake or mass transfer-
enhancing strategies of the bacteria. However, the roles of bacterial phenotypes
(physiology, adaptation, etc.) on the rate of desorption or direct uptake of the sorbed
phase is rapidly gaining attention.

Evidence supporting the role of specific phenotypes (chemotaxis, mobility,

adhesion, ability to produce bio-surfactants, etc.) is slowly accumulating(Guerin and

13

Boyd 1993; Harms and Zehnder 1995; Tang et al. 1998). The bioavailability is likely to
be a function of both physiological properties of bacteria (chemotaxis, motility, and
attachment) as well as soil and contaminant processes (sorption, desorption, etc.). Any
change in either microbial physiological properties or sorption-desorption characteristics
is expected to affect the contaminant availability to the microbial uptake. It is expected
that bioavailability of the sorbed phase naphthalene varies with strains having different
physiological properties as well as with sorbents with different sorption/desorption
properties. In the following paragraphs, a brief account of important physio-cherrrical and

biological processes in controlling bioavailability will be discussed.

AFFECT OF SORPTION/DESORPTION KINETICS ON BIOAVAILABLITY
For solute to be available for microbial uptake, it has to be accessible by being
present either in the aqueous phase (dissolved) or on the surface of the solid phase
(partition). Limited bioavailability can occur when organic compounds are either poorly
water-soluble, strongly sorbed to a sorbent, or spatially separated from the
microorganism (Mihelcic et al. 1993). The pollutants become inaccessible to
microorganisms when they are absorbed into natural organic matter or adsorbed into the
walls of submicron-sized pores inside soil particles. The adsorbed compound can only be
taken up by a bacterium, if it diffuses out/desorbed of the pores of sorbent particles to the
solid-water interface and is then partitioned into the water phase. These general
observations support the theory that desorption and sorption processes are the limiting
factors during bioremediation of hydrophobic organic compounds in a contaminated soil-
water system. Therefore, it is important to include terms for sorption and desorption in

the bioavailability model; ignoring the kinetics of these processes may lead to false

14

predictions about bioavailability of compounds. In the following paragraph a brief
overview of sorption and desorption, their mechanism, and a conceptual model will be

provided.

Sorption Process

The process of sorption involves partitioning of hydrophobic compounds, initially
dissolved in water phase only, between water phase and the sorbent and results in a
considerable removal of the organic compound from the water phase. Sorption processes
have been considered to be of two types: partition and absorption (Chiou 2002). While,
the partition is a surface phenomenon in which the solute accumulates at the surface or
interface between water and sorbent phase; the absorption, on other hand, refers to
contaminant sorption to the wall of pores inside soil particles. Sorption in soil and
sediments is mainly attributed to the soil organic matter content. Sorption of HOCs has
been shown to increase with hydro-phobicity of the chemical and the content of soil
organic matter. The process of sorption is considered, since early 1980, as a reasonably
fast process on a practical time scale ranging from less than a minute to a few days
(Pignatello and Xing 1996a). When sorption is fast, and completely reversible, a linear

isotherm equation can adequately describe the process as follow:

Cs = Kd . Ce [1]
Where Cs and Cc represent the concentration of the compound in water and the solid
phase, respectively; and K1 is the soil-water partition co-efficient. An example of a

system that shows linear isotherm is the partition of HOCs into natural Organic matter of

soil (also called soft SOM).

15

Numerous laboratory and field studies in the early 1990s noted that sorptive
equilibrium may take weeks to months depending on the compound, and ideal linear
relationships no longer remain valid (Ball and Roberts 1991; Young and Weber 1995).
The simplest equilibrium model could not account for the non-ideal sorption behavior of
HOCs as observed in heterogeneous soils or soils containing hard SOM ( kerogen -like
SOM e.g., soot ). The sorption is, therefore, most commonly represented by the

Frendulich equation as given by Equation 2:
n
CS = Kf . Ce [2]

Where it is an empirical constant indicating non-linearity of the isotherm, and Kf is the
Frendulich capacity factor.

The long-term non-equilibrium sorption shows a bi-phasic pattern in which there
is an initial rapid sorption portion followed by a much slower component of sorption.
Several mechanisms have been proposed for non-equilibrium sorption, including: 1)
intra-particulate diffusion (pore-filling), 2) the presence of a small amount of hi gh-
surface area carbonaceous material (HSACM) , and 3) the distributed reactivity model or
dual mode sorption model. The first mechanism assumes that organic chemical, as taken
up rapidly due to sorption by external sorption sites, can move slowly into internal sites
and is retarded by very slow diffusion into tortuous pores of either an organic matrix or
micro-pores of the mineral domain of soil (Wu and Gschwend 1986). This mechanism is
supported by evidence that larger particles had a slow sorption rate, and dis-aggregation
of larger particles into smaller particles improved the sorption rate (by reducing diffusive

path the length).

16

The second proposed mechanism stipulates the presence of small amounts of
high-affinity adsorption sites (e.g., charcoal like materials) besides significant amounts of
SOM in the soil. The sorption on charcoal like material is largely an adsorption process
(competitive) and gives non-linearity to the overall isotherm shape (Chen and Wagenet
1995; Chiou et al. 1998; Chiou et al. 1979). The distributed reactivity model (Weber and
Huang 1996) hypothesizes that soil organic matter consists of two sorption domains: 3
rubbery or soft carbon domain (domain I) and a glassy or hard carbon domain (domain
II), both domains contributing to the overall sorption process. Domain I exhibits a fast,
reversible, and linear isotherm; domain 11, on other hand, shows a non-linear, and
adsorption (competitive) type of sorption. The overall sorption of a representative PAH is

the sum of the fast and slower components that takes the following mathematical form:

CS=fS.K0c.Cw+fh.Kf.CWn [3]

Where fs and fh represent the weight fraction of rubbery or soft carbon and glassy or hard

carbon, respectively.

Desorption Process

The desorption process is the opposite of the sorption process in which sorbed
phase molecules diffuse out from surfacial and internal sites of the soil particles into the
soil water. Desorption results in the removal of the chemicals from the solid phase and
the concentration of the chemicals in the soil water. Desorption of HOCs from soils and
sediments have also been observed to take place in two stages: a rapid stage followed by
a stage of much slower release. Also, a number of studies have shown that a fraction of

sorbed contaminants is difficult to remove, and this fraction has been called the

17

irreversible fraction (Kan et al. 1997; Khan 1991; Pignatello and Xing 1996a; Pignatello
and Xing 1996b). Several researchers observed that fractions of slower release and
irreversible components of desorption increase with the increasing contact time of
organic compounds in soils; this phenomenon is known as the aging of organic
compounds (Huang and Weber 1997; White et a1. 1997). The most common mechanisms
responsible for sequestration are the entrapment of HOCs within SOM matrices, the
formation of HOCs-soil organic matter (SOM) complexes, and chemical degradation.
Several models have been developed to describe the kinetics of de-sorption
including the two-site non-equilibrium model, the intra-aggregrate diffusion model, and
the stochastic-Gamma model. None of them, however, incorporate the irreversible
component of the desorption. In this study, I used a three-site desorption model (Park et
al. 2001) to analyze the desorption data for rate of release of selected compound. The
underlying assumptions for the models are : 1) the overall desorption sites consist of fast,
slow, and irreversible components 2) sorption is a relatively fast process and
approximated by a linear isotherm 3) the rate of non-equilibrium release rate is
proportional to the difference between the concentration in liquid and non-equilibrium

sorbent sites. The overall desorption site is theoretically represented by:
S = Seq + Sneq + Snd [4]
Where S(p.g/Kg) is the total sorbed phase concentration, and Seq, Sneq, and Snd (pg/Kg)

represent respectively equilibrium, kinetic, and irreversible site fractions of desorption.
The equilibrium process is assumed very fast and the corresponding site (Seq) can be

represented by a linear partitioning model of

seq = feq . Kd .c [5]

18

Where C (pg/L) represents the liquid-phase concentration, feq is the equilibrium site

fraction, and Kd(L/Kg) is the sorption distribution co-efficient. The rate of contaminant

release from non-equilibrium site is assumed to follow the following first-order equation:
andSneq/dt = (I ( Sneq — neq Kd C) [6]
Where fneq is non-equilibrium site fraction, and 0t (min-1) is the desorption rate

coefficient. The term Kd is determined from sorption isotherm. Desorption site fractions

feq, fneq and desorption rate or are estimated by fitting non-linear regression equation to

desorption data, with the constraint that sum of all site fractions is equal to unity. The
non-desorbable fraction is, however, determined from the plateau of desorption data

profile.

INFLUENCE OF MICROBIAL KINETICS ON BIOAVAILABILITY

In soil systems, the degradation of organic compounds may be monitored by

measuring with time either the parent compound or the product formation (such as C02).

An understanding of the kinetics of degradation is important to predict (through
mathematical models) the persistence of organic substances in the environment. The
kinetics of degradation defines the relationship between the concentration of a compound
of interest and the rate at which microorganism can degrade it (Grady et al. 1996). The
natural environment contains different amounts of organic compounds, and as a result

there exist different microbial population with defined kinetic properties. Several

l9

different models have been pr0posed to describe the microbial degradation in soils, and

the most commonly used models are described as follows:

Bio-degradation by Growing Micro-organisms

The empirical Monod model is commonly used to describe microbial growth
kinetics under nutrient limited conditions. The concentration of the substrate is assumed
to be in excess to support microbial growth. For pure culture, the Monod equation can be

written as
1* = (”max - S)/(Ks +8) [7]
Where It is the specific growth rate, ”max is the maximum specific growth rate, S is the

organic substrate concentration, and KS is the half-saturation constant or affinity constant
and is taken as the substrate concentration at which the rate of growth is half the
maximum growth rate. The two kinetic constants (umax, Ks ) of the Monod equation have
been widely used to differentiate between two groups of microbes: l) the slow grower or

oligotroph which exhibits an extremely low affinity constant value KS, and 2) the fast

grower or copiotroph which displays a high growth rate and a very high Ks value. The

oligotroph, with low maintenance requirement as compared to copiotrophes, can
potentially degrade contaminants to low endpoints in a diffusion-limited

microenvironment. Several other growth models (e.g., logarithmic, Logistic, etc.) were
proposed that may be applicable under different substrate concentration and bacterial KS

value.

20

Biodegradation by Non-growing Microorganisms

If the substrate concentration is low (rate limiting chemical) as compared to the
cell density, the microbial growth assumed is either very small or negligible. Under this
condition, the degradation kinetic is best described with the Michelis-Menten equation.

According to this equation, the flux of the chemical through the cell membrane Rbio is

Rbio = (Vmax . C) /( Km + C) [8]
Where Vmax is the maximum flux that a cell can get, C is the chemical concentration and
Km is the Michael-Menten constant (half-saturation constant or substrate affinity) which

is the chemical concentration corresponding to 1/2 Vmax. Integrating Equation 8, and re-
arranging its term, results in the following equation:
t: i-(Km.1n<C/Co>+C-Co>/Vmax1Mo ‘ [91

Where C0 is the initial substrate concentration, M0 is the microbial population density,

and t is the time. The two parameters Km, and Vmax can be obtained by fitting the

substrate depletion data with the integrated Michalis-Menten Equation (9)

When the concentration of the compound being degraded by non-growin g

bacteria is low, relative to the biological activity (C<<Km), the flux through the cell

membrane follows the first order kinetics, which is written as:

Rbio = R: . C [10]
Upon integrating and re-arranging equation 2) takes the form ln(C/Co) =-k1 . t,

and can be solved graphically to determine the rate constant k1. The first order expression

21

is often used to describe biodegradation of contaminants, which is low in concentration

due to mass transfer limitation, in the soil (Bosma et al. 1997).

Bioavailability Models

In a slurry bio-reactor, the intracellular degradation of HOCs depends on their
availability to the degrading organisms. Traditionally, it has been assumed that sorption
reduces bioavailability and only readily desorbed and aqueous-phase substrate are
available to microorgansms. However, there are some studies that suggest sorbed
compounds are available to microorganisms without prior desorption (Guerin and Byod
1992 and others). Recently, Park et al. (2001) demonstrated that naphthalene degrading
bacteria (Pseudomonas putida G7) could access to soil-sorbed naphthalene and,
consequently, the observed degradation rate in soil slurries exceeded those predicted from
the independently determined rate of de-sorption. In light of the above, several
approaches can be taken to model bioavailablity depending on whether the sorbed phase
is available or unavailable to the rrricroorganism. In this study, it has been assumed that
the selected microorganism (P. putida G7) can have direct access to the sorbed phase
substrate.
The bioavailability model to be used incorporated a first-order three-site desorption
model, linear sorption, and biodegradation. The graphical representation of the model is

shown in Figure 2-1.

22

 

 

 

 

 

+n++w+<—>—N

 

 

A = Dissolved and Equilibrium site
B = Non-Equilibrium site
C = Irreversible phase

J n = mass flux from non-equilibrium site

Figure 2-1. Description of Desorption-Biodegradation model (DBM)

The model is based on the following assumptions: a) the desorption parameters as
measured in abiotic condition are applicable in the bio-availability assays; b) degradation
rates of the dissolved naphthalene remain the same in the presence or absence of the
sorbent c) both the attached and suspended cells can utilize the dissolved naphthalene
equally. The overall degradation of naphthalene in a batch slurry reactor using mass

balance on naphthalene can be written as:

-(v . dC/dt +m . dS/dt) =v . Rbio + m . ( keq. seq + 1(an . sneq + knd. snd) [11]

23

Where In (kg) is the total mass of the soil, V(L) represents the volume of the liquid, and

keq, kneq, and knd are the first-order biodegradation coefficients of sorbed naphthalene in

equilibrium, non-equilibrium, and non-desorption sites, respectively. The term Rbio is the

rate of biodegradation of naphthalene in the liquid phase only and is defined in Equation-
8 in the above section. To track the sorbed-phase naphthalene concentration in non-

equilibrium and irreversible sites, the following mass balance equations are required:

dSneq/dt =Jn _kneq . Sneq [12]
dSnd/dt = ‘knd . Snd [13]

The different rate constants ( keq, kneq, and knd) are determined by a nonlinear regression

analysis of the bio-availability data.

BACTERIAL PHYSIOLOGICAL CONSIDERATIONS

Recently, several studies reported evidence that there exist significant diversity in
the abilities of bacteria to degrade soil-sorbed HOCs (Guerin and Boyd 1993, Tang et al.
1998, Calvillo and Alexander 1996). The possible physiological explanation for such
diversity includes the production of biomaterial, variation in attachment to solid surfaces,
and chemotaxis/motility traits. For example, two naphthalene-degrading bacteria
(Alcaligenes sp. NP-Alk and Pseudomonas putida ATCC 17484) have been shown to
possess differential capabilities to degrade soil-sorbed naphthalene (Guerin and Boyd
1993). It was hypothesized that the attachment capability of isolate ATCC 17484 to solid
surfaces caused enhanced degradation due to 1) the exposure of cells to a steeper intra-

sorbent concentration and, 2) cells act as competitive sorption media to a soil-sorbed

24

naphthalene. Other studies have shown the potential of chemotaxis in enhancing the
biodegradation of organic compounds such as naphthalene (Marx and Aitken 1999),
carbon tetrachloride (Witt ME 1999), BTEX compounds (Parales et al. 2000), and
pesticides (Hawkins 2002) in laboratory scale microcosms. The following paragraphs
deal with a brief description of important bacterial strategies to enhance the degradation

of toxic compounds in a soil water system.

Bacterial Chemotaxis and Enhanced Bioavailability

Bacterial chemo-taxis is defined as the movement of motile cells toward (in case
of attractants such as nutrient sources), or away from (in case of repellent such as toxic
compounds) the region of elevated chemical concentration gradient in the media
(Harwood et al. 1994). Movement toward the chemo-attractant is called positive
chemotaxis, whereas movement away from the chemo-repellent is termed as negative
chemotaxis. In the absence of a chemical gradient, the bacteria moves in a random
fashion with no directionally biased movement. The random movement consists of a run,
a brief forward movement, followed by a tumble, when the cell stops and bounces in a
random direction. After each tumble, the bacteria re-orients randomly for the next run.
Thus bacteria move randomly in an environment by means of runs and tumbles without
any net directional movement. In the presence of a chemical gradient, these random
movements become biased. The run length increases in proportion to increasing chemical
concentration and is accompanied by less frequent. The net result of this behavior moves

the bacteria in the direction of higher concentration of the attractant (Parales et al. 2000).

25

The mechanism controlling bacterial movement involves a signal transudation
system, which transmits signals to the cell about the presence of concentration gradient.
The signal transudation system consists of two sets of sensory proteins: 1) sensor protein
located in the membrane and 2) a cytoplasmic protein. The sensor protein has a very
high affinity for the attractant or repellants. As the concentration of attractants or
repellants increases over a period of time, more and more binding of these molecules on
sensor protein takes place. This binding sets in play a series of interactions with the
cytoplasmic protein that eventually affects flageller rotation to increase or decrease the
bacteria run length.

Chemotaxis may play an important role in enhancing bio-availability and ,
consequently, accelerates biodegradation. Chemotaxis allows cells to move in the
direction of higher concentration, and thus bringing them in contact with the chemicals to
be degraded. In soil contaminated with hydrophobic compounds, chemotaxis may enable
microbes to access zones of concentrated chemical for degradation that would otherwise
be inaccessible due to mass transfer limitations, low solubility or sequestration of
chemicals to or inside a solid matrix. This is all the more important in heterogeneous
soils, where there exist large variations in sorbed concentrations across different regions,
and chemotaxis may help cells locate contaminants in these different regions (Jain 2002).

Most prokaryotes are motile, and many are chemotactic to various toxic organic
compounds such as PAHs. Pseudomonas putida G7 and Pseudomonas sp. NCIB 9816-4
have recently been shown to be chemotactic to naphthalene. The chemotactic response is
due to catabolic plasmid present within these organisms. A mutant derivative of P. putida

G7 (G7.C1) which is devoid of plasmid (NAH7) was shown to be non-chemotact to

26

naphthalene. It has also been shown that the chemoreceptor gene ( NahY) encoded by the
NAH7 plasmid is involved in naphthalene chemotaxis (Grimm and Harwood 1997;
Grimm and Harwood 1999).

Limited studies have been carried out to demonstrate the potential role of
chemotaxis in rapid biodegradation of organic compounds in aqueous and porous media.
In studies involving an aqueous medium alone, the concentration gradient was
established by supplying naphthalene either from a glass capillary tube (Marx and Aitken
2000) or a NPAL-water interface (Law and Aitken 2003). Both studies clearly
demonstrated that naphthalene was degraded at a higher rate by the wild-type strain than
either the non-motile strain or the non-chemotactic strain. It has been shown that the
wild-type strain could overcome mass transfer limitations by accumulating in regions of

higher concentration that led to a rapid biodegradation rate in both studies.

Experiments with a porous medium, packed in sand column, involved detecting
changes in the bacterial density distribution with time in response to concentration
gradients. The concentration gradient establishes as bacteria continue to uptake
compound that was present at an initial constant concentration. The experiments with
porous medium, however, either failed to observe chemotaxis (Barton and Ford 1995) or
provided indirect evidences of chemotaxis (Pedit et al. 2002; Witt ME et al. 1999). The
presence of the porous medium has been shown to cause significant reduction in the
bacterial chemotactic and motility responses. Also, if the initial isotropic concentration of
the compound exceeds, the saturation of the chemoreceptor would lead to the suppression

of chemotactic responses (Barton and Ford 1995 and 1997).

27

Witt et al. (1999) demonstrated the movement of Pseudomonas stutzeri KC in
columns packed with aquifer sediments. In a continuous flow column, the cells migrated
through the column at a velocity exceeding the average linear velocity of flow. The
enhanced movement was attributed to the creation of a nitrogen gradient as a result of the
rapid consumption of nitrates by the bacteria. Experimental results with no flow column
showed that the mean velocity of KC cells due to chemotactic movement towards
increasing concentration of nitrates was 5 cm/day. However, the study was not able to
distinguish between chemotactic and random motility.

Padit et al. (2001) clearly demonstrated the role of chemotaxis in increased
biodegradation of naphthalene in a porous medium. They used an experimental system
similar to Marx et al. (1999) with the exception that both the micro-capillary and the
reservoir were filled with glass beads. It was shown that chemotaxis can occur in a
porous medium; however, the presence of solids had significantly reduced the chemotaxis

responses as compared to responses measured in a solid-free system.

Bacterial Adhesion

Adhesion is shown to be an important mechanism through which microorganisms
are involved in the direct utilization of sorbed chemicals (Calvillo and Alexander 1996;
Harms and Zehnder 1995). The phenomenon of adhesion suggests that microorganisms
possess some special cell walls and/ or bonding properties that help them to adhere to
hydrophobic surfaces containing sorbed chemicals. Bacteria, adhering to solid surfaces
containing sorbed chemicals, experience a microenvironment that is different from the

surrounding bulk liquid. The microenvironment, which lies within a thick static boundary

28

layer at the solid-water interface, provides a steep concentration gradient to bacteria that
results in the degradation of the sorbed substrate. The role of adhesion in providing

bacteria direct access to the sorbed phase organic pollutant is illustrated in Figure 2-2

 

 

 

 

Water level 0

33’.

E
Static boundary layer “s fiT '5 \\

° \

g \\:\\
Solid Surface 5 ‘j

 

 

Cs, concentration at
the solid surface

Figure 2-2. Adhesion as a bacterial strategy to directly access sorbed chemicals at solid

The attachment to the solid surface exposes bacteria to a steep concentration gradient by
shortening the diffusion distance. To understand the mechanisms responsible for
bacterial adhesion to solid surfaces, it is important to know the structure and properties of

bacterial cell surfaces.

Structure of the Gram-negative Cell Surface

The bacterial cell surface is a three dimensional, multi-layered, and complex
structure. The surface of most bacteria consists of three layers: the outer membrane with
associated structures, the cell wall, and the inner cytoplasmic membrane. Of these three
layers, the outer membrane is mainly responsible for cell surface properties. The

schematic of the outer membrane along with its associated lipopolysaccharides layer

29

(LPS) is shown in Fig 2-3. The outer membrane is a coating of protein macromolecules
that provides an additional layer on top of the cell wall. The outer membrane serves to
protect the cell and also gives the cell shape and rigidity. It contains pores that facilitates
the selective transport of water and nutrient molecules and avoid the entry of toxic

substances. On top of the outer membrane there exist additional structures made up of

 

Rod Shaped r

Bacterium L

 

 

 

O—antigen

1% (2-40 nm; 1-50

'5 (repeating units)

E E

2 .3 Core Polysaccharide

:3 (23 nm)

8 Outside of cell
Fatty acids in ---- -- Hut"--.
outer
membrane

 

 

 

 

 

 

Figure 2-3: Schematic of the outer membrane of gram negative bacteria along with its

associated LPS layer

protein and polysaccharides molecules; they form what is often called lipo-
polysaccharide layer or simply LPS. The LPS is composed of three components and is
anchored into the hydrophobic region of the outer membrane by its lipid-A component.

The lipid-A component is then attached to a second component of LPS, called core

30

polysaccharide, which consists of a unique eight-carbon sugar called 2-keto-3-de-
oxyoctonic acid (KDO). The core polysaccharide region of the LPS is negatively charged
due to the presence of phosphates and carboxylic groups. The outer part of the LPS is the
O-antigen, which consists of 15 to 20 repeating units of three to five sugars and can
extend out from the cell surface upto 50 nm into the cell surroundings (Brock 2000).

A variety of different other structures can also be found on top of the outer
membrane; they include flagellae, fimbriae, and fabrils. They are appendages of different
length and diameters and are mainly proteinaceous in chemical composition. Many
bacterial strains also produce polymeric substances that are referred to as extra-cellular
polymeric substances (EPS). EPS may be secreted as a loose layer of carbohydrates
called the slime layer or a more rigid layer of proteins known as the capsule (Fletcher
1985; van Loosdrecht et a1. 1990).

The adhesion of bacteria to a solid surface has been described as a four-step
process (van Loosdrecht et al. 1990). The first step involves the transport of bacteria
toward the solid surface by one of three different modes: diffusion due to Brownian
motion, advection due to liquid flow, or active movement of the cell due to chemotactic
response to the concentration gradient using flagellated structures. The diffusion process
is the slowest of all (at the rate of 40 um/hr) and occurs due to the random interaction of
cells with surfaces. The advective transport is, on other hand, a very fast process
depending on the velocity or mixing of water. Finally, bacteria respond to the existence
of the chemical gradient in a heterogeneous medium and move with the help of flagella

toward a region of higher concentration (van Loosdrecht et al. 1990).

31

Once bacteria come in contact with a solid surface, initial adhesion takes place.
The initial adhesion may be reversible or irreversible depending on the mechanism
involved. During a reversible adhesion, there is a continuous exchange between the free
and adhered cells. The initial adhesion of bacteria can best be described by the DLVO
(Derjaguin, Landau, Verwey, Overbeek) theory. The DLVO theory envisages adhesion as
an interplay of long ranging electrostatic and van der Waals interactions; and initial
adhesion occurs when attractive forces overcome repulsive forces. The main factors
influencing these interactions are surface charges and hydro-phobicity. The electrostatic
forces arise due to the fact that both the bacterial surface and solid surfaces are negatively
charged. Therefore, when bacteria approach solid surfaces, they experience electric
double layer repulsion. The intensity of the double layer repulsion depends on the
concentration and valency of the ions in the culture medium. The electrolyte repulsion
can be decreased or eliminated by increased electrolyte concentration of the culture
medium (van Loosdrecht et al. 1989).

Figure 2-3 illustrates how the interaction of long ranging attractive and repulsive
forces governs initial adhesion at various separation distances between bacterium and
solid surface. When cells approach very close to surfaces, and this occurs at high ionic
strength, the attractive forces become very strong, creating a primary minimum. In this
situation, the primary attractive forces are hydrogen bonding and ion pair formation. As
the bacterium moves slightly away from the surface, repulsive forces grow quickly and
cause the adhesion to cease. As cells move further away from the surface, another
shallow minimum exists called the secondary minimum. The reversible adhesion is

thought to exist at the secondary minimum and both cells and surface are not in actual

32

contact at this distance of separation. The reversible part of initial adhesion is very weak,
and cells can easily be removed from surfaces due to slight shear force (Rijnaarts et al.
1993; van Loosdrecht et al. 1989).

There are, however, experimental evidences that indicate that other factors such
as outer membrane proteins, composition of LPS, flagella, and fambriae may be required
for the initial adhesion to surfaces (DeFlaun and Fletcher 1999; Fletcher 1985). There are
many examples of bacteria adhering to a solid surface due to hydrophobic interaction,
thus overcoming any water barrier to bacteria attachment (Stenstrom 1989). After the
initial adhesion, irreversible attachment of cells can occur due to the adsorption of
polymer to surfaces. The LPS polymers are long enough to bridge the distance between
the cells and surfaces, even if no DLVO attraction exists. In irreversible attachment, cells
form mono-layer on the surface. During the final stages of adhesion, the attached cells
grow in micro-colonies and ultimately form a multi-layer growth called bio-film (van

Loosdrecht et al. 1990).

Mechanism of Adhesion

Mechanisms of attachment to inanimate surfaces is still controversial due to the
wide variation of cell surface properties among bacteria and also because bacteria cell
surfaces are dynamic in nature due to varying environmental conditions. It is well known
that bacteria have different abilities to attach to surfaces. Even for a given population of
bacteria, some cells attach, while others remain suspended, and there can be a significant

difference in adhesion properties even among the clones of a single organism (Fletcher

1996).

33

The mechanism of bacterial adhesion is complex and unlikely to be explained by
one or other property of the cell surface (Bos et al. 1999). There are a number of factors
that contribute to the adhesiveness of the bacterial surface that include cell surface
charge, bacterial hydrophobicity, the presence of extra-cellular polymers, and flagella
(Deflaun 1999). Besides cell surface properties, bacterial adhesion is also influenced by
the chemical composition of the solid surface and the liquid medium.

Numerous studies have demonstrated that cell adherence increases with more cell
hydrophobicity (as measured by contact angle), and less negative charge (as measured by
electropherotic mobility) (e. g., van loosdrecht et al 1987, Stenstrom, 1989). Mueller et al.
(1992), found that bacteria with low hydrophobicity ( P. flauresences) attached at far
lower rate (five times) on to a glass surface than bacteriam with high cell hydrophobicity
(P. aeruginosa). Bestiaens et al. (2000) isolated adherent PAH-degrading bacteria by an
enrichment technique using hydrophobic membranes containing sorbed PAHs. The
isolated bacterium was shown to be very hydrophobic and adherent to different surfaces.
Microbial surface charge can also influence adhesion, but to a smaller degree as
compared to cell hydrophibicity. For example, adhesion of the cells increased, when the
surface charge of the cells was chemically modified from a negative to a positive charge
(Klotz et al. 1985). Cells are defined to be hydrophobic when the measure of water
contact angle is above 30. Also, the experiment using Gram-negative indicated that
adhesion with polystyrene and rice roots becomes zero, when the value of zeta-potenial
becomes more negative than -25 V (Bos et al 1999).

LPS and EPS have also been shown to play an important role in cell adhesions,

and are especially useful for strengthening cell binding to surfaces (Rijnaarts, 1995;

34

Fletcher, 1996). However, the characteristics of specific polymers (e.g., compostion, size,
charge, etc.) in turn play an important role in adhesion to the surfaces (Fletcher 1996).
Razatos et al. (2000) used and compared interactions of the wild-type strain (with longer
LPS chain) and its mutant (with truncated LPS chain) with glass surfaces. The authors
reported that while the wild-type was able to adhere to the glass surface, the mutant was
not able to adhere. They hypothesized that as the mutant lost its core polysaccharide
component of LPS, the exposed negatively charged KDO molecules experienced a
repulsive interaction with negatively charge glass surface. The wild-type on the other
hand had experienced an attractive interaction due to the shielding of the negatively
charged KDO by longer LPS molecules.

Deflaun et al.(1999) also investigated the role of cell surface-polymer in adhesion
of B. cepacia G4 to sand surfaces. The authors isolated the adhesion-deficient mutant of
Burkholderia cepacia G4 in a sand column assay and compared its adhesion phenotypes
to that of the wild-type strain. The mutant was reported to have lost its adhesion
phenotype due to the attenuated O-anti gen on the LPS. The loss of O-antigen part of LPS
caused a reduced shielding of the charged group near the membrane (such as the
phosphate group in the core-lipid A region) and, consequently, there was a net repulsive
interaction between the mutant and the negatively charged sand surfaces.

In some studies the truncation of the LPS, however, resulted in a more
hydrophobic and, hence, more adherent mutant than that of the wild-type (Fletcher et al.
1996). The authors generated and isolated mutants of pseudomonas fluorescens that were
altered in adhesion ability. The mutants were demonstrated to have increased adhesion to

a variety of surfaces (quartz sand, hydrophobic polystrene ) when compared to the

35

adhesion of the wild-type strain. The mutants showed an increased attachment to hydro-
phobic surfaces and a decreased attachment to the more hydrophilic sub-straturm. The
authors attributed the enhanced adhesion property of the mutant to attenuation of the 0-
anti gen on the LPS. The lack of the O-anti gen exposed lipid moiety and, hence, rendered

the cell surface more hydrophobic.

36

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40

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41

CHAPTER 3. ENHANCE BIOAVAILABILITY OF SORBED PHASE
NAPHTHALENE: A CONIPARISON OF BIODEGRADATION AND

DESORPTION RATES IN NATURAL SOIL

ABSTRACT

The removal of naphthalene in a soil-slurry system was studied using the bacterial strain
Pseudomonas putida G7 (Pp G7), Tenax beads extraction, and serial extraction with a 0.1
M buffer solution alone. The overall removal of naphthalene by Tenax beads was
observed to be greater than that obtained with serial desorption. At high concentration
both Tenax and serial dilution showed similar trends for naphthalene removal rate.
However, Tenax extracted more sorbed phase naphthalene than the serial dilution
method. A comparison of the results of Pp G7 biodegradation and Tenax extraction
showed that the rate and extent of naphthalene removal by biodegradation was at all
times significantly greater than the corresponding Tenax extraction of sorbed phase. It
seems that naphthalene does not have to be desorbed in to the aqueous phase for
biodegradation by Pp G7. It appears that Pp G7 was able to overcome the mass-transfer
limitation and directly take-up sorbed-phase naphthalene. The experimental results
showed that Pp G7 degraded sorbed naphthalene at a rate and extent greater than what
can be removed by both Tenax beads and serial extraction methods. Also, explored was
the effect of excreted biomaterial on enhancing bioavailability. The activity of extra-
cellular enzymes was eliminated during the biodegradation study, as the experiment with

the bacterial filtrate did not result in any degradation of observable free naphthalene

42

during a 3-days incubation period. It was observed that the partition coefficient (Kd) of
naphthalene to Colwood soil was reduced as the concentration of biomaterial increased.
However, the reduction in K. was observed to be statistically insignificant at a
biomaterial concentration as obtained at naphthalene concentration less than 10 mg/L. It
is, therefore, expected that any sorption experiment carried out with biomaterial obtained
at less than 10 mg/L of naphthalene will not significantly affect the partition coefficient
of naphthalene. All bioavailability experiments were carried out at 2 mg/L of
naphthalene; thus, the possible effect of corresponding biomaterial on partition co-

efficient (and consequently mass transfer) of naphthalene is highly unlikely.

INTRODUCTION

Sorbed phase HOCs have been reported to be available for biodegradation
without prior desorption into the aqueous phase(Guerin and Boyd 1992; Guerin and Boyd
1993; Park et al. 2001; Tang et al. 1998). This finding has led to the concept of enhanced
bio-availability, which has been defined as the availability of organic compounds for
bacterial uptake at a rate greater than abiotic desorption rate from the sorbent. The
evidence that non-desorbable fraction of sorbed phase PAHs can be available for
bacterial uptake has significant implications for remediation endpoint and risk
assessment. Enhanced bioavailability of soil-sorbed contaminants may render
bioremediation a viable technology that will have the potential to decrease residual
soil/sediment contaminant concentration to an environmentally acceptable endpoint that

meets regulatory requirements.

43

The previous studies mostly put forward indirect evidence of enhanced bio-
availability (Harms and Zehnder 1995; Park et al. 2001). They compared independently
measured rate and extent of desorption with intrinsic activity of the test bacteria to define
enhance bioavailability. For example, Harms and Zehnder (1995) reported that attached
bacteria degraded 3-cholorodibenzofuran from porous Teflon beads at a rate greater than
abiotic desorption rate that was measured by sequential replacement of the aqueous
phase. No study has, however, measured desorption kinetics during bio-remediation
treatment.

The approach of independently measuring abiotic desorption rate has some
potential limitation in determining enhance bio-availability. Some investigators have
assessed that rate and extent of desorption measured under abiotic condition does not
represent rate of desorption in presence of degrading bacteria. The desorption rate
changes due to utilization of contaminants by degrading bacteria in aqueous phase,
leading to the desorption of more contaminant from the sorbent to the aqueous phase.
Furthermore, water alone seems to be a poor choice of extractant for determining non-
desorbable fraction of HOCs because: 1) HOCs are sparingly soluble in water and
strongly sorbed and, 2) water cannot access labile fraction of the contaminant in sorbent
matrix. This view has been has been supported by earlier researcher that concluded that
desorption rate alone may under-predict biodegradation potential of bacteria (Comelissen
et al. 1997b).

Furthermore, previous experimental studies also examined role of bacteria and
suggested that the utilization of non-desorbable PAHs requires bio-availability—enhancing

strategies of the bacteria (Calvillo and Alexander 1996; Guerin and Boyd 1992). One of

the strategy that bacteria can employ to enhance bioavailability is to excrete compound
that can either enhance desorption rate or degradation of sorbed phase naphthalene.
These compounds are extra-cellular in nature and include catabolic enzymes and
polysaccharides substances.

While the concept of enhance bioavailability is appealing, the actual mechanism
responsible for the phenomenon is still not clear. The main objective of this study was to
observe phenomenon of enhanced bio-availability by directly comparing abiotic
desorption in presence of Tenax and biodegradation rates for naphthalene at various times
during the treatment period. In the absence of a technique to measure desorption rate in
the presence of bacteria, we used equivalent amount of Tenax beads to represent
degrading bacteria. The Tenax beads have been shown to be a suitable adsorbant to
assess desorption rate and extant during biodegradation treatment. When place into
contact with a soil-water slurry, Tenax maximizes the diffusion gradient by acting as an
infinite sink for desorbing chemicals. Thus, the contaminant mass transfer rate sensed by
degrading bacteria cells could potentially be represented by those measured in the soil-
slurry using Tenax beads. Possible role of extra-cellular biomaterial (such as enzymes
and polymer) was also explored to explain enhance bioavailability. Two hypotheses were
tested: 1) Pp G7 does not produces extra-cellular enzymes during uptake of naphthalene
and, 2) PAHs adsorb to Extra-cellular polymeric substances (EPS) from soil phase and

thereby become more available to the degrading bacteria.

45

MATERIAL AND METHODS
Low;

The three soils selected for this are CapacA, Colwood, and Houghton Muck with
low, intermediate and high soil organic content. . The soil was prepared as described
elsewhere (Park et al. 2001), and their characteristics are summarized in Table 3-1.
Chemicfls and Culture Medium

Both ”C-labeled and unlabeled naphthalene were used in this study. The ”c-
labelled naphthalene (sigma, 0.1 mCi, >98% pure) stock solution was prepared by
supplementing it with unlabelled naphthalene and was used for sorption, desorption,
serial dilution studies. The un-labelled naphthalene (sigma, > 98%) stock solutions were
prepared in methanol and stored in 5 ml screw cap vials (capped with Teflon septa) at
4°C before use. All biological assays were carried out with unlabelled naphthalene stock
solutions. Microbiological media were prepared as described previously. The
experiments were canied out with chemotaxis buffer (pH =6.8) as described elsewhere
(Harwood et al.)

Bactenafirmns

Wild-type Pseudomonas putida G7 was used in this study. Wild-type PpG7 (ATCC
17485) was obtained from Dr. Caroline Harwood (University of Iowa) and has been
reported previously to uptake dissolved and sorbed phase naphthalene in soil slurry (Park
et al. 2001). Furthermore, it has been characterized as being motile, and chemotactic

(Grimm and Harwood 1997; Grimm and Harwood 1999).

46

Table 3-1. Selected properties of soils to be used in this study

 

 

Soil % O.C. % Sand % Silt % Clay PH CEC
[cmol(+)/kg]

Capac 3.3 55 24 21 6.8 24

Colwood 7.8 64 21 15 6.0 43

Houghton Muck 38.3 ND ND ND 5.1 156

 

ND: not determined yet.
O.C.: organic carbon content.

CEC: cation exchange capacity.

47

Sorption, Desorption, and Serial dilution Experiments

Sorption, desorption, and serial dilution experiments were carried out as described
previously (Park et al. 2001). Briefly, all experiments were carried out with initial
equilibrium period of two days. The ratio of sorbent to liquid was kept the same in all the
experiments: 1:15 for Capac-A soil, 1:65 for Colwood soil, and 1:240 for muck. Sorption
experiment was canied out at initial liquid-phase naphthalene concentrations, ranging
from 0-2500 ug/l in triplicate for each type of soil. The desorption and sequential
extraction experiments were both carried out at an initial naphthalene concentration of 2
mg/l. Desorption and sequential extraction experiments were initiated, at the end of 2
days sorption equilibrium period, by adding naphthalene free soil extract to make up the
decanted volume. In all these experiments, vials were tumbled at 9 rpm in the dark. After
mixing, each vial was centrifuged and supernatant sampled. The remaining supernatant
was then decanted and replaced with methanol to extract soil-sorbed naphthalene. The
concentration of naphthalene was then determined by liquid scintillation counting (LSC)
and verified by high-pressure liquid chromatography (HPLC) for some samples.
Bioavailability and Biodegradation rate experiments

Bioavailability assay was adopted from the method described by Park et a1.
(2001). Briefly, soil extract controls and soil slurries were set up in 5-ml vials sealed by
screw caps with Teflon-lined septa. The soil to water ratio was kept same as for
sorption/desorption experiments. Unlabelled naphthalene was injected in all vials at
initial liquid phase concentration of ~ 2000 ug/L. After 2 days of initial equilibrium
period on tumbler, soil slurries and soil-free extracts (to determine degradation rate in

liquid only) were inoculated with cells harvested at early stationary phase. Each vial

48

received bacterial cell suspension to a final cell density of ~ 5x106 CFU/ml. At precise
time intervals, selected vials were centrifuged and l-ml of supernatant was transferred to
a HPLC vial containing 0.01 ml of 10 N NaOH to stop biodegradation of naphthalene.
The soil-sorbed naphthalene was then extracted by methanol after removing left over
supernatant. Appropriate control vials were included, with no bacteria, to check for any
abiotic losses.
Chemical Analysis

Naphthalene was analyzed by HPLC using a C13 reverse-phase column, with 80%
acetonitrile/20% water mobile phase, and fluorescence detection (280 nm excitation, 340
nm emission). Radioactivity was determined by LSC. The analytical detection limits of
naphthalene solution were 5 ppb and less than lppb for HPLC and LSC, respectively.
Extrafilluljar Enzyme Activity Assay

Extra-cellular enzyme activity was measured, if there was any, by first growing
bacteria (PpG7) in a mineral salt medium containing succinate (100 mg/L) for 36 hours
and the cells were removed by filtration through 0.2 um Millipore filter. The culture
filtrate (~ 2 mL) was then injected into each of the triplicate 20 ml screw-cap vials
containing 16 mL of Chemotactic buffer (CB). The solution was then supplemented with
Naphthalene at a final concentration of 1000 ug/L. The control vials did not receive
filtrate and account for any abiotic losses due to volatilization and wall sorption. The
vials were then mixed at 9 rpm and samples were removed at predetermined time

intervals for HPLC analysis.

49

Extra-cellulg Biomaterial Extraction

To obtain extra-cellular polymer, Pseudomonas putida G7 were first grown in
mineral salt medium supplemented with different concentration of naphthalene solution
(10-30 mg/L) in a 50 mL flask with stopper. The solution was aerated by placing
magnetic bar in each flask. The loss of naphthalene by volatilization was compensated
based on control flask. The growth of bacteria was monitored over the period of time by
standard plating. After 4 days of incubation, naphthalene was degraded to a non-
detectable level. The bacterial suspension was then filtered through a 0.22-um Millipore
filter to separate cells and dissolved biomaterial. The filtrate was supplemented with
sodium azide at a final concentration of 400 mg/L to prevent any microbial activity. The
filtrate was then stored at 4 °C before being used for sorption/desorption experiments.
The organic content (TOC) of the filtrate was determined by the TOC analyzer.
Abiotic De-sorption Assavfi Using Tenax Beads

The abiotic rate of naphthalene release from soil slurries was measured using
procedure as described elsewhere (Comelissen et al. 1997a) The initial step for
desorption assay, utilizing Tenax-TA beads, was same as in the sorption experiment. The
solid-to-solution ratio was kept same as in the sorption experiment. Desorption was
initiated by placing Tenax-TA beads (0.05 grams), enclosed in sealed dialysis bags
(MW CO =13000-15000), in the sediment/water slurry bottles. A Teflon-lined cap was
tightly screwed onto each vial that was then tumbled at 9 rpm. During the mixing, the
Tenax beads absorb instantaneously any naphthalene that are present in liquid phase or
released into the aqueous phase from the soil particles by desorption. At the

predetermined time interval, selected bottles were removed and centrifuged for phase

50

separation. The dialysis bags, containing Tenax TA beads, were removed and liquid
phase was sampled. The remaining supernatant was decanted and the sorbed phase was
extracted with methanol. The concentration of naphthalene was determined by LSCE and
verified by HPLC.

Rate of extraction and maximum sorptive capacity of Tenax beads were measured
in separate experiment without any sediment. In this experiment, the CB volume and
naphthalene concentration were kept comparable to those used in the desorption
experiments. Extractions were carried out with 0.05 grams of Tenax for extended period
of time. At predetermined time intervals, water phase was sampled for residual
naphthalene concentrations. The extraction rate was determined from the plot of time vs.

residual concentration by fitting first order kinetic equation to the data.

DATA ANALYSIS
Tenax Model
The Tenax data were analyzed according to Comelissen et al. (1997). The model
is described with one water compartment, and two sediment compartments. The amount

of naphthalene left at a given time is determined by equation —1.

S _. _ _
§L=(Faqe keXtr't'l'Fslow'C kSIOW°t'1'Fv-SIOW°C kv-slowt) [1]

0

Where, St and So (jig) are the naphthalene contact time t (h) and at the beginning of the
experiment, respectively; Faq, F810“, and Fv-slow are the fractions in the liquid phase,

slow fraction in sediment, and very slow fraction in sediment, respectively; and kem,

51

kslow, and kHIOW are the corresponding rate constant of liquid and sediment. The

extraction rate (kem) in aqueous was determined in aqueous phase with out sediments.

Coupled Desorption-Biodegradation Model (DBM model)
The overall degradation of naphthalene in a batch slurry reactor using mass

balance on naphthalene can be represented by equation-2.

—(Vl-£+m--C-i—S—)=(Rbio-V+m-knd-Snd) [2]
(it (it

Where m (kg) is the total mass of the soil, C (ug/L) represent naphthalene concentration

in solution phase, S(p.g/Kg) represent total sorbed-phase naphthalene, V (L) represents

volume of the liquid, and knd is the first-order biodegradation coeffcient for non-

desorbable fraction of total sorbed contaminant (S). Rbio is the liquid-phase

biodegradation rate expression, which can be expressed by following equations:

R bio = M for Michaelis Menten kinetics, [3]
km+C
Rbio = kl' C for first-order kinetics [4]

Where, Vmax is the maximum flux that a cell can get and, Km is the Michaelis-Menten .
constant, and k1 first-order degradation rate co-efficient. The total sorbed phase (S) was

assumed to consist of equilibrium (Seq), non-equilibrium (Sneq) and non-desorbable(Snd)

sites. The equilibrium site represents fraction that release spontaneously during

desorption experiment and is described by a linear partitioning model. The non-

52

equilibrium site fraction allows slower release of contaminant and can be represented by
first order expression. The non-desorption site is defined by Park et al. (2001) to be
containing contaminant that cannot be released to aqueous solution during the
experimental desorption period. Mathematically, these three-sorbed phase fractions are

described as follow:

3“,:th . Kd. c [5]
Snd = fnd che [6]
dSneq/dt = 0!. ( fneq Kd C — Sneq) [7]

Where Kd is the sorption co-efficient, CC, is the liquid phase concentration in sorption
equilibrium, t is desorption time in minutes, Otis the first-order desorption rate coefficient
(per minute). The three remaining terms feq, fneq, and fnd represent, respectively,

equilibrium site fraction, non-equilibrium site fration, and non-desorption site fraction.

The three site fractions were measured from desorption experiment: fnd corresponds to

the plateau of the desorption rate profile; and feq, fneq and or were estimated by non-linear
regression analysis of desorption data with the constraint that:

feq+fneq+ fnd =1 [8]

RESULTS
Sorption and Desorption
Figure 3-1 shows sorption isotherms for naphthalene on the selected natural soils.

The isotherms were found to be linear over the evaluated concentration range of

53

naphthalene. The sorption coefficients (Kd =slope of isotherms) ranged from 15 to 240
m1] g that increases as organic content of the soil increases (Table 2-2). Figures 2-3
shows naphthalene release during desorption and sequential extraction experiments,
respectively. Whereas de-sorption experiment involved single extraction cycle, the
sequential extraction experiment consist of seven extraction cycles. In both type of
experiments, release of naphthalene from the soil was expressed in percent of the total
sorbed naphthalene desorbed at the time of measurement.

The desorption rate profiles showed initial instantaneous release of naphthalene followed
by a period of rate-limited de-sorption (Figure 2-2). The rate-limited de-sorption was
significant within first 6 hrs; however, it becomes negligible beyond 6 hours. In all three
soils, the maximum amount of desorbed naphthalene was observed to be less than 50%.
The attainment of de-sorption plateau after 6 hrs is probably due to apparent de-sorption
equilibrium. Further release of naphthalene was achieved by sequential extraction
experiment (Figure 2-3). The subsequent steps with fresh naphthalene-free CB solution
extracted more naphthalene, but amount of desorbed naphthalene decreases
(exponentially) with each steps. This shows that repeated replacement with fresh CB
solution cause greater total release than a single desorption experiment carried out for a
longer period of time. At the end of seventh extraction cycle, no further release of
naphthalene occurred indicating attainment of extent of desorption endpoint. The fraction
of non-desorbable naphthalene was then determined by extracting soil with methanol at

the end of last extraction cycle (Table 3-2).

54

Table 3-2. Parameters for sorption of naphthalene by the A horizon of selected soils

 

Non-desorbable

 

 

 

 

 

 

 

 

 

 

     

 

 

 

 

Soil K, R2 K...
fraction, fnd
CapacA 15 0.96 457 0.21 (6x103)
Colwood 68 0.98 872 0.19 (2x103)
Muck 240 0.997 626 0.18 (1x102)
3.5e+5
g 3.0e+5 -
a
o- 2.59+5 -
C
8 O Muck
a) 2.0e+5 _ O Colwood
: V Capac-A
2
CO
5 1.Se+5 -
.2
CL
(U
C 1.0e+5 -
'0
a)
g 5.0e+4 -
m ___
0.0 7 7. T I .
0 500 1000 1500 2000

Aqueous naphthalene cone. (ug/L)

Figure 3-1. Sorption isotherm of naphthalene in natural sorbents

55

The three site desorption model (Park et al. 2001) was used to fit de-sorption data
(Figure 3-4) and to estimate de-sorption rate coefficient (or) along with each site fractions
(Table 3-3). The de-sorption data shown in Figure 3-4 were expressed in percent of the
total amount expected assuming complete reversibility. In all the experiments, the overall

naphthalene mass recovery was found to be 94-97%, indicating no degradation.

Removal of Naphthalne in Soil-slurry System by Tenax

In Figure 3-5, adsorptions of naphthalene to Tenax in presence and absence of soil
are plotted as a function of time. The rate of extraction of naphthalene by Tenax in a
dialysis bag, in absence of any soil, can be modeled as a first order kinetics; the value for
km was calculated as 0.97 /hr. The added amount of 0.05 gram of Tenax was sufficient to
extract more than 97% of naphthalene in the system. For an equivalent system but in
presence of soil, however, the rate and extent of naphthalene adsorption to Tenax was
significantly retarded due to slow de-sorption of naphthalene from the soil. After a rapid
removal of dissolved and desorbable naphthalene (as shown by initial steep curve), the
adsorption of further naphthalene was mass transfer/desorpiton limited. The data were
then fitted with two sediment compartments and a water compartment (Comelissen et al.
1997) model and the rate constants for rapid and slow desorption ( Km, and Kslow) are
presented in Table 3-4. The rate of removal naphthalene in muck soil was observed to be
slower than the Capac-A soil. This was probably due to higher content of organic

material in muck.

56

60

 

 

‘6 50 . a
a w I I . I I
r

0 40 ~

3 30

g 0 Cach

E 20 4 I Colwood

3' 10 - MUC"

Z

°\° 0 I T I I I

0 20 40 60 80 100

Contact time (hour)

Figure 3-2. Kinetic disorption data for naphthalene from three soils

90.0 -
80.0 -
70.0 -
60.0 -
50.0 a
40.0 -
30.0 -
20.0 -
10.0 -

Naphthalene extracted (%)

 

0.0

O!
.‘
.1

o Capac-A
I Colwood
Mick

 

I I I l I l l

2 3 4 5 6 7 s
Extraction cycle

Figure 3-3. Sequential desorption data for the removal of Naphthalene from three soils

57

Table 3-3. Desorption parametersa evaluated by a three-site desorption modelb

Soil

 

 

feq fneq fnd a R2
(min'l)
Capac 0.74(0.042) 0.1 0.16(0.009) 0.0025 (0.0008) 0.94

Colwood 0.62(0.023) 0.23 0.15(0.010) 0.0016(0.0004) 0.94

Houghton Muck 0.58(0.034) 0.27 0.15(0.005) 0.005(0.0017) 0.97

 

3‘ feq fneq fnd: equilibrium-, non-equilibrium-, non-desorption-site fractions; 0:: first
order desorption rate coefficient for non-equilibrium sites.

bNumbers in parentheses are standard deviations of the evaluated values.

 

 

 

   

 

 

 

100
E: O Colwood
C 0 Muck
.9 v CapacA
4—0
33
a) 40 -
as
D

20 —

0 I I I I I

0 1000 2000 3000 4000 5000 6000

Desorption time (min)

Figure 3-4. De-sorption of naphthalene from three natural soils

58

Biodegradation of Naphthalene by PpG7 in soil-slurry

Figure 3-6 shows removal of naphthalene in soil-slurry system by Pp G7 for
Capac-A and Muck soil, respectively. Pp G7 degraded naphthalene in a biphasic curve.
The first, rapid decline of naphthalene (~80%) was completed with in first 2 hours and
was followed by a much slower decline. The rapid decline occurred in liquid phase,
which had contained naphthalene either in dissolved phase or desorbable phase ; whereas,
the slower decline was mostly occurred at solid-water interface containing non-
desorbable fraction of naphthalene. The wild-type Pp G7 was able to decrease
naphthalene to a level below the non-desorbable fraction as shown by the dashed
horizontal line. The fitted model (DBM model), as represented by solid lines, reasonably
described degradation of naphthalene in soil-slurry system; the estimated parameters are

shown in Table 3-4.

Effect of Extra-cellular biomaterial on Bioavailability

Experiments were performed to determine role of extra-cellular biomaterial on
enhancing bioavailability either by degradation of non-desorbable fraction due to extra-
cellular enzymes activity or enhancement of rate and extent of de-sorption of naphthalene

due to presence of solubalized bacterial polymers.

Extra-cellular Enzyme Activity

The data showing naphthalene concentration with and with out culture filtrate

(from 0.2 um filter) are shown in Table 3-5. Comparing data between samples receiving

59

Table 3-4. Rate efficient of naphthalene removal by PpG7 and Tenax in soil-slurry

 

 

 

 

 

 

 

 

 

 

 

system
Mixed PpG7 Mixed- Tenax
Soil
I(In Ks Knd cht Krap Kslow
(ug/L) (ug/L-hr) (1/hr) (1/hr) (l/hr) ( 1/hr)
0.05
Capac-A 980 300 (0.0017) 0.7 0.3 0.02
300 0.054
H. Muck 950 (0.0016) 0.65 0.05 0.001
10
O No soil
0 Capac-A Soil
’6) V Muck Soil
3
o'
c
O
O
G)
c
.9
(U
5
.c
O.
(D
Z
““‘L ............................
o s— -------- 3
4D 60 8‘0 100

 

Time (hours)

Figure 3-5. Extraction of Naphthalene by Tenax beads in presence and absence of soils

60

Table 3-5. Naphthalene concentration before and after addition of cell filtrate (0.2 um filter)

 

 

 

 

 

 

 

 

 

 

Sample Initial Naphthalene conc. Control Filtrate addition
no. (ppb) (ppb) (ppb)
1 1000 928 920
2 1000 918 901
3 1000 965 923
Avg. i std. Dev. 937i25 915i12
9
8
3 7
m . Capac- A soil
3 5 0 Muck soil
E 5
o
c
2 4 ’
E
.c 3 '
O.
(U 2 -
Z A...
1 F km,
0 I 1 r r
0 20 40 60 80

Time (hours)

Figure 3-6. Degradation of naphthalene by PpG7 in mixed soil-slurry system

61

culture filtrate and control (with no filtrate) show no hydrolyzing activity of extra-cellular
enzymes. Abiotic losses due to volatilization were observed in both samples and control,

and were similar in amount.

Sorption/Desorption Experiment in presence of Excreted biomJatelial

Figure 3-7 shows amount of excreted biomaterial in filtrate solution, obtained by
growing Pp G7 at a range of naphthalene concentration (10-30 ppm) in CB solution. The
quantity of biomaterial was measured by determining total organic carbon (TOC) of
filtrate solution. It is evident from the data that as bacterial mass increase, in response to
increasing naphthalene concentration, there is proportional increase in excreted
biomaterial in CB solution.

The data obtained from the sorption experiments in presence of different
quantities of biopolymer are shown in Figure 3-8. The data were fitted with linear
isotherm for estimating distribution coefficient of naphthalene to soil (Kd) as shown in
Table 3-6. The K, value decreases as biomaterial concentration increases. The decrease in
K; value A trail de-sorption experiment was carried out in presence of biomaterial
corresponding to 10 mg/L naphthalene. The presence of biomaterial did not alter the de-
sorption kinetics of naphthalene in Colwood soil (data not shown). Since, our bio-
availability experiments were carried out at 2 ppm, therefore, we did not see the need to
further explore the effect of biomaterial on de-sorption and consequently bio-availability

of sorbed phase naphthalene.

62

2000 .
1800 <
1600 .
1400 «
1200 ~
1000 a
800 -
600 ~
4001
200+ O

0

TOC (mg/L)

 

F? = 0.9921

 

O

10

T

20 30 40

Aqueous naphthalene conc. (mg/L)

Figure 3-7. Production of Biomaterial (TOC) by PpG7 at different naphthalene

 

concentration
ie+5
' O No-Biomaterial /0 v
0 Biomaterial at 10 ppm naphthalene .-
V Biomaterial at 20 ppm naphthalene //'€7
89+4 . v

i

N
O
+
b
r

Sorbed naphthalene cone. (ug/Kg)

 

Biomaterial at 30 ppm naphthalene / -

 

l

600

 

I I I I I

800 1 000 1 200 1 400 1 600 1 800

Aqueous naphthalene conc.(ug/L)

Figure 343- Decrease in partition co-efficient of naphthalene with increasing
biomaterial in soil-slurry system

63

Table 3- 6. Distribution co-efficient of naphthalene to Colwood at different amount

 

 

 

 

 

 

of biomaterial
Soil Biomaterial at naphthalene Kd (L/Kg) R2
concentration of (mg/L)
0 65.5 0.996
Colwood 10 61.5 0.995
20 57.0 0.997
30 55.5 0.993

 

Table 3-7. Non-desorbable fraction from Tenax and serial extraction methods

 

 

 

 

 

 

Non-exhaustive Exhaustive
Soil
TA extraction method Serial dilution method
Non-desorbable fraction Non-desorbable fraction
fnd fnd
CapacA 0.12 0.21
Muck 0.14 0.18

 

DISCUSSION

Recently, several researchers have observed and reported enhanced bio-
availability of soil-sorbed PAHs. Their studies suggested that biodegradation of sorbed
PAHs is not exclusively dependent on abiotic mass transfer of the compound into the
water phase. Several arguments supported the observations of enhance bio-availability:
First, it has been shown that coupled biodegradation-desorption model (DBM) that
incorporate degradation of non-desorbable fraction could only be successfully used to
either curve-fit or predict the observed biodegradation data. Secondly, the observation
that contaminant degradation kinetics exceeds independently measure desorption rate and
extent has led to the conclusion that bacteria can uptake from sorbed phase, directly. In
most of these studies, the de-sorpton rate was measured by sequential replacement of the
aqueous phase. This study explores the validity of these observations by directly
comparing biodegradation kinetics with abiotic extraction rate using Tenax and also to

investigate role of bacterial extra-cellular biomaterial on enhance bioavailability.

Comparison of Biodegradation and Abiotic Removal rate in soil-slurry system
The main objective of this research was to compare kinetics of naphthalene
removal in soil slurry system in presence of Tenax adsorbent as a third phase infinite
sink, with biodegradation kinetics by Pp G7 performed in separate experiments. Tenax
has been reported to be a better extractant for determining the size of the rapidly and
slowly desorbable fractions of total sorbed phase HOCs as compared to using salt
solution, alone. (Reid et al. 2000 a). Tenax has also been shown to be a suitable

adsorbent for HOCs removal from liquid phase and can mimics the bio-degradation

65

capacity of PAH-degrading bacteria(Comelissen et al. 1998; Theodora et al. 2003).When
placed into contact with soil-water slurry, it tends to adsorb HOCs rapidly by acting as an
infinite sink (Kp =104 IJKg). By adding fresh tenax to soil-slurry, a near zero naphthalene
concentration was maintained that resulted in a complete extraction of desorbable
fraction of sorbed naphthalene. Thus, it is expected that desorption with tenax represents
biodegradation condition and is more representative of to evaluate the rate and extent of
naphthalene removal from the soil.

Figures 3-6 and 3-7 show naphthalene removal by biodegradation and adsorption
to Tenax beads in Capac-A and Muck soil-slurry systems, respectively. All the
experiments began with equal mass of naphthalene in both liquid and sorbed phases at the
end of the sorption period. During initial period of 1 hour, Pp G7 and Tenax bead
removed the liquid phase naphthalene at approximately equal rates. After this initial
period, the Pp G7 continued to degrade remaining naphthalene in sorbed phase at same
rate until less than 5% of total naphthalene remained in soil after 8 hrs of treatment. The
Tenax beads, on other hand, removed naphthalene at a slower rate in sorbed phase with
biphasic pattern: an initial relatively rapid phase followed by much slower phase. The
fraction of sorbed phase that is released rapidly (qu) was removed at a relatively faster
rate (Kmpid), whereas slowly desorbing fraction (fneq) was removed at much slower
removal rate (Ks|ow). The overall removal of naphthalene by Tenax beads slightly
exceeded total desorbable fraction, as determined by serial extraction method, and shown

by horizontal dashed line.

66

Also plotted on Figures 3-9 and 3-10 results of serial de-sorption (dashed lines
with asterisk symbols) with 0.1 M buffer solution. During serial de-sorption experiment,
it was observed that desorption plateau during first 4 hours of each step (Figure 3-2).
Therefore, an effective time of 4 hours was selected for each step during which de-
sorption complete. Amount of naphthalene left at end of each step was taken and plotted
against corresponding 4 hrs periods.

Both serial extraction and Tenax extraction shows similar trend. Either method
showed presence of non-desorbable fraction of sorbed naphthalene. The amount of non-
desorbable fraction was, however, greater in serial extraction than Tenax extraction
(Table 3-7). The amount of non-desorbable fraction is often assumed to be non-available
for biodegradation, and this underlies the importance of having method that accurately
measure the fraction in solid media. The result shows that Tenax extraction provides a
better estimate of desorbable fraction (and hence bioavailable) and can be used as a rapid
and straightforward alternative to conventional serial extraction.

Comparison of the results of Tenax extraction and Pp G7 biodegradation shows
that the rate and extent of naphthalene removal by bio-degradation was at all time
significantly greater than the abiotic removal rate and extent in the sorbed phase. It seems
that naphthalene do not have to be desorbed in the aqueous phase for bio-degradation by
Pp G7. Apparently, Pp G7 were able to overcome the mass-transfer limitation by

unknown mechanism and could able to directly uptake sorbed phase naphthalene.

67

Muck Soil

    

 

 

 

 

 

 

a
a . 0 PpG7 degradation
3 5 ? o Tenax Extraction
(u l * Serial dilution
E 5 -
‘5. 4
(0
E 3 -
19.
O 2 .
'— . .................................. 6,322"? *
- ~~~~~ Q-~__
1 - -------- a
0 ' 1‘ ' -‘ - e .
0 20 40 60 80
Time (hours)
Figure 3-9. Removal of naphthalene by PpG7, Serial dilution, and Tenax
beads in Muck soil
9 Capac-A Soil
A
U)
a.
:3 0 PpG7 degradation
3 o Tenax Extraction
_ * Serial dilution
.1:
O.
(0
2
E
O
'- _____ 4 el-
--------------------------- o
O '0'
. 40 60 80
Time (hours)

Figure 3-10. Removal of Naphthalene by PpG7, Serial dilution, and Tenax
beads in Capac-A soil

68

Thus, this study lend credence to the concept of enhance bio-availablity, which
has been observed previous works. Two mechanisms had often been cited for the
observed enhance bio-availability: 1) direct uptake by bacteria and, 2) micro-organism
mediated increase in mass transfer(Harms and Zehnder 1995; Rosenberg et al. 1989). Our
tenax data support the former mechanism because tenax adsorb HOCs rapidly in soil-
slurry system by acting as an infinite sink. Thus, tenax can represent all possible bacterial
strategies ( i.e. production of bio-surfactant or EPS) that can result in increase mass-
transfer of sorbed phase naphthalene from soil to water.

It is plausible that when Pp G7 degrade dissolved concentration to zero; passive
uptake by bacteria becomes negligible. Consequently, Pp G7 can uptake further
naphthalene from non-desorabable pool by direct uptake of naphthalene at solid-water
interface. Bacteria have been known to attach soil particles containing organic molecules

and thus degrade these molecules directly.

Extra-cellular Biomaterial and Enhanced Bioavailability

Bacteria can employ a variety of strategies to access sorbed-phase HOCs. Among
these, bacterial production of extra-cellular biomaterial (including enzymes and polymer)
can influence bio-degradation of sorbed molecules either directly or indirectly. In this
study, no catabolic/hydrolyzing activity of extra-cellular enzymes was observed.
Naphthalene metabolism has been observed to be intracellular in Pseudomonas bacteria
(Bugg et al. 2000; Gray and Bugg 2001; Grimm and Harwood 1999; Whitman et al.
1998). The activity of extra-cellular enzymes in response to pollutant uptake is still

unknown. Micro-organism, mostly fungi, are reported to excrete extra-cellular enzymes

69

in response to large (complex polymer) and inaccessible substrate (very low solubility).
The excreted enzymes acts upon large and complex molecules and convert them into
smaller, and more soluble or diffusible compounds (Faber 1979). In soils, micro-
organism including bacteria do not have direct access to contaminant molecules that
becomes sorbed or entrapped with nano- and micro-pores (Alexander et al. 1995).
Microorganism can, however, degrade these trapped molecules by excreting enzymes
molecules, which are small enough to enter these inaccessible molecules.

Pseudomonas putida G7 is known to possess both extra-cellular polymeric
substances (EPS) as well as lipopoly-saccharide (Kachlany et al. 2001). The bacteria (Pp
G7) excrete these polymer under changing environmental condition (low C/N ratio,
starvation condition, etc) (Neu and Marshall 1990). These polymers have high molecular
weight and, consist of polysaccharides, proteins, lipo-polysaccharides, lipoproteins or
complex mixture of these polymers. They are, however, less effective in lowering surface
and interfacial tension in solvent, but can increase compound solubility due to unknown
mechanism(Ron and Rosenberg 2002).

Presence of bacterial polymer in soil-slurry can influence the fate of naphthalene
by affecting it sorption/desorption rate and extent. In this study, partition co-efficient
(Kd) of naphthalene to colwood soil reduced from 6- 15% as biomaterial concentration
increases. This enhancement corresponds to biomaterials obtained at naphthalene
concentration of 10-30%. It is, therefore, expected that any sorption experiment carried
out with biomaterial obtained at less than 10 ppm of naphthalene will insignificantly
affect partition co-efficient of naphthalene. Furthermore, our all bioavailability

experiments were carried out at 2 mg/L of naphthalene, the possible affect of naphthalene

70

on partition co-efficient is highly unlikely. This was further substantiated by desorption
experiment carried out at biomaterial obtained at naphthalene concentration of 10 mg/L,

where we did not see any change in rate or extent of naphthalene desorption.

71

REFERENCES

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Calvillo, Y. M., and Alexander, M. (1996). "Mechanism of microbial utilization of
biphenyl sorbed to polyacrylic beads." Appl. Microbial. Biatechnal., 45(3), 383-
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Comelissen, G., Noort, P. C. M. v., and Govers, H. A. J. (1997a). "Desorption kinetics of
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Comelissen, G., Rigterink, H., Ferdinandy, M. M. A., and Van Noort, P. C. M. (1998).
"Rapidly desorbing fractions of PAHs in contaminated sediments as a predictor of

the extent of bioremediation." Environmental Science & Technology, 32(7), 966-
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Comelissen, G., Rigterink, H., Vrind, B. A., tenHulscher, T. E. M., Ferdinandy, M. M.
A., and vanNoort, P. C. M. (1997b). "Two-stage desorption kinetics and in situ

partitioning of hexachlorobenzene and dichlorobenzenes in a contaminated
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Faber, M. D. (1979). "Microbial-Degradation of Recalcitrant Compounds and Synthetic
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Gray, M. R., and Bugg, T. (2001). "Selective biocatalysis in bacteria controlled by active
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Grimm, A. C., and Harwood, C. S. (1997). "Chemotaxis of Pseudomonas spp. to the
Polyaromatic Hydrocarbon Naphthalene." Appl. Environ. Microbiol., 63(10),
41 11-41 15.

Grimm, A. C., and Harwood, C. S. (1999). "NahY, a Catabolic Plasmid-Encoded
Receptor Required for Chemotaxis of Pseudomonas putida to the Aromatic
Hydrocarbon Naphthalene." J. Bacterial. , 181(10), 3310-3316.

Guerin, W. F., and Boyd, S. A. (1992). "Differential Bioavailability of Soil-Sorbed
Naphthalene to Two Bacterial Species." Appl. Environ. Microbial, 58(4), 1142-
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Guerin, W. F., and Boyd, S. A. (1993). "Bioavailability of Sorbed Naphthalene to
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Sorption and Degradation of Pesticides and Organic Chemicals in Soil. SSSA
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Harms, H., and Zehnder, A. J. B. (1995). "Bioavailability of Sorbed 3-
Chlorodibenzofuran." Appl. Environ. Microbiol, 61(1), 27-33.

Kachlany, S., Levery, B., Kim, J., Reuhs, B., and Lion, L. (2001). "Structure and
carbohydrate analysis of the exoplysaccharide capsule of Pseudomonas putida
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Neu, T. R., and Marshall, K. C. (1990). "Bacterial polymers: Physico-chemical aspects of
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Park, J.-H., Zhao, X., and Voice, T. C. (2001). "Biodegradation of Non-desorbable
Naphthalene in Soils." Environ. Sci. Technol., 35(13), 2734-2740.

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Rosenberg, E., Rosenberg, M., and shoham, M. (1989). "Adhesion and desorption during
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Tang, W. C., White, J. C., and Alexander, M. (1998). "Utilization of sorbed compounds
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Theodora, E. M., Hulscher, T., Postma, J ., and Besten, P. J. (2003). "Tenax Extraction
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73

CHAPTER 4. AFFECT OF CHEMOTAXIS AND MOTILITY ON THE
BIODEGRADATION OF NON-DESORBABLE NAPHTHALENE IN NATURAL

AND SYNTHETIC SORBENT

ABSTRACT

The purpose of this study was to evaluate the role of chemotaxis/motility in soil-
sorbed naphthalene degradation by wild-type Pseudomonas putida G7 in a sorbent-slurry
system under stirred and quiescent conditions. Naphthalene degradation by the wild —type
strain was compared to that of a non-motile strain and a mutant strain deficient in
naphthalene chemotaxis. Any difference in naphthalene degradation rate and extent
between wild-type and mutant strains was considered to come from chemotaxis/motility
phenotypes. In natural soils (with low to high organic content), under both stirred and
static conditions, the rate and extent of naphthalene degradation was observed to be
similar for all the strains. Whereas in stirred system all the strains had an equal access to
a non-desorbable fraction of naphthalene; in the static system, however, none of the
tested strain had gained access to the non-desorbable fraction of naphthalene. The effect
of mixing was clearly seen to have caused the difference of bacterial uptake in mixed and
static systems. The well-mixed batch system is expected to have caused uniform
concentration gradient, greater cell to soil particles interaction, and shorter transfer
distances that contributing to the enhanced bioavailablity. In a well-mixed system,
however, bacterial chemotaxis/motility is a much slower process as compared with other
processes and therefore chemotaxis/moitility effects was not isolated. In a static system,

however, numbers of sites available for bacterial attachment for direct uptake of sorbed

74

phase napthalene were substantially reduced due to the pore size exclusion and
retardation (due to cell adosption to soild surfaces) of bacterial cells. The
chemotactc/motility movements ( and hence access to high concentration sites) were also
impeded due to a lack of mixing in organic soils, bacterial size exclusion, and adsorption
adhesion effects.

Experiments carried out with activated carbon, under stirred conditions, showed
that both WT and mutant strains had accessed to the non-desorbable fraction of
naphthalene. However, WT degraded non-desorbable phase naphthalene at a significantly
greater rate as compared to its mutant strains. The difference of uptake rate between WT
and mutant strains could be explained due to the presence of significant concentration
gradient at soild/water interface as a result of large pool of non-desorbable naphthalene.
The concentration gradient attracted greater number of WT cells, due to its
chemotaxis/motility phenotypes, that resulted in a greater rate of non-desorbable

naphthalene degradation as compared with system inoculated with mutant strains.

INTRODUCTION
Limited bio-availability of chemicals trapped in soil micro-pores pose great
challenges the remediation of contaminated soil and sediments(Head 1998).
Bioavailability of soil sorbed contaminants is considered to be an interplay between
desorption/mass transfer and microbial processes. Any change in either mass transfer
from the soil microenvironment to aqueous phase or bacterial characteristics can affect
the bioavailability of contaminants. During the initial step of bio-degradation, the mass

transfer increases as contaminant is removed from the aqueous phase. As rapid and slow

75

de-sorption sites are exhausted, the bio-degradation of a hydrophobic compound has been
observed to be limited by the extremely slow rate of chemical transfer from non-
desorbable compartments of desorption sites (Bosma et al. 1997; Rijnaarts et al. 1990).
The mass transfer to a microorganism can be, however, improved if the degrading
organism swims toward localized concentration gradient at the soil-water interface. It is,
therefore, quite comprehensible that bacteria capable of sensing and responding to
localized soil gradients can increase net mass transfer rates.

It has been recognized that bacteria can sense and respond to the presence of
aromatic compounds in the environment (Grimm and Harwood 1997; Harwood et al.
1994). Bacterial chemotaxis and motility are two physiological functions that have been
shown to confer advantages to microorganisms to increase the rate of degradation of
aromatic hydrocarbon(Marx and Aitken 1999). Chemotaxis allows cells to move in
direction of higher concentration, thus bringing them in contact with the chemicals to be
degraded. In soil contaminated with hydrophobic compounds, chemotaxis may enable
microbes to access zones of concentrated chemicals for degradation that would otherwise
be inaccessible due to mass transfer limitations, low solubility, or sequestration of
chemicals to or inside a solid matrix. This is all the more important in heterogeneous
soils, where there exist large variations in sorbed concentrations across different regions
and chemotaxis may help cells locate contaminants in these different regions.

Most prokaryotes are motile, and many are chemotactic to various toxic organic
compounds such as PAHs (Jain 2002; Madigan et al. 2002). Pseudomonas putida G7 and
Pseudomonas sp. NCIB 9816-4 are recently shown to be chemotactic to naphthalene

(Grimm and Harwood 1997; Grimm and Harwood 1999). The chemotactic response is

76

due to catabolic plasmid present within these organisms. A mutant derivative of P. putida
G7 (G7.C1) that is devoid of plasmid (NAH7) was shown to be non-chemotact to
Naphthalene. It has also been shown that chemoreceptor gene (NahY) encoded by the
NAH7 plasmid is involved in naphthalene chemotaxis.

Limited studies have been carried out to demonstrate the potential role of
chemotaxis in rapid biodegradation of organic compounds in aqueous and porous media.
In studies involving aqueous medium alone, the concentration gradient was established
either by supplying naphthalene from a glass capillary tube (Marx and Aitken 2000) or
NPAL-water interface (Law and Aitken 2003). Both studies clearly demonstrated that
naphthalene was degraded at a higher rate by the wild-type strain than either non-motile
strain or non-chemotatic strain. It has been shown that the wild-type strain could
overcome the mass transfer limitation by accumulating in region of higher concentration
that led to rapid biodegradation rate in both studies.

Experiments with porous medium involved detecting temporal changes in
bacterial density distribution in packed sand column in response to concentration
gradients. The concentration gradient established as bacteria continued to uptake a
compound that was present at an initial constant concentration. However, the experiments
with porous medium either failed to observe chemotaxis (Barton and Ford 1995), or
provided indirect evidences of chemotaxis (Pedit et al. 2002; Witt ME et al. 1999). The
presence of the porous medium has been shown to contribute significantly to the
reduction in the bacterial chemotactic and motility responses compared to the measured

in aqueous system. Also, if the initial isotropic concentration of the compound exceeds

77

saturation capacity of the chemoreceptor, it would lead to suppression of chemotactic
responses (Barton and Ford 1995 and 1997).

Witt et al. (1999) demonstrated movement of Pseudomonas stutzeri KC in
columns packed with aquifer sediments. In continuous flow columns, cells migrated
through the column at a velocity exceeding the average linear velocity of flow. The
enhanced movement was attributed to creation of a nitrogen gradient that resulted from
the rapid consumption of nitrate by bacteria. Experimental results with no flow column
showed that the mean velocity of KC cells due to chemotactic movement towards
increasing concentration of nitrate was 5 cm/day. However, the study was not able to
distinguish between chemotactic and random motility.

Padit et al.(2001) clearly demonstrated the role of chemotaxis in increase
biodegradation of Naphthalene in porous medium. They used an experimental system
similar to Marx et al.(1999) with the exception that both micro-capillary and reservoir
were filled with glass beads. It was shown that chemotaxis can occur in a porous
medium; however, the presence of solids had significantly reduced chemotaxis responses
as compared to a solid-free system.

All of the previous studies related to bacterial chemotaxis and motility in
biodegradation of organic compounds were carried out either in a homogenous solid
(glass beads or clean sand) or liquid medium. Furthermore, in experiments with solid
mediums, the rate of degradation by wild type was not compared to mutants deficient in
chemotaxis and motility trait to clearly identify the role of the bacterial trait. In earlier

studies with columns packed with porous medium, the loss of substrate was speculated to

78

be either due to chemotactic or random motility. The possible role of chemotaxis in the
bioavailablity of organic compounds in a soil matrix has so far remained unexplored.
The aim of this study, therefore, was to explore the role of chemotaxis and
motility on the rate and extent of biodegradation of both desorbable and non-desorbable
fractions of soil-sorbed organic compound such as Naphthalene in natural and synthetic
sorbents. The rate of sorbed phase naphthalene was monitored in both mixed and static
slurry system with Wild-type P. putida G7 and two mutant strains, one of which was
motile but not chemo-tactic toward naphthalene, and the other of which was non-motile.
Comparing degradation rate and extent of the wild type to the mutant strains allowed us
to distinguish the effect of chemotaxis and motility on biodegradation of sorbed phase

naphthalene in the test systems.

MATERIAL AND METHODS

Geo-sorberm Soil ExtraLts

The three sorbents selected for this study comes from A-horizon of agricultural
soils found in Michigan. The selected soils spanned a range of organic matter content.
These included the Houghton muck soil, which is essentially devoid of mineral matter;
and two mineral soils of differing organic matter contents (Capac-A, and Colwood),
whose mineralogy is predominately illite and vermiculite. The selected characteristics of
these geo-sorbents are summarized in Table 4-1.

Before use, all the soil samples were air-dried, and passed through 2 mm sieve.
The soil were then filled in several 5-ml screw cap bottles and sent them to University of

Michigan Ford Nuclear reactor for y—irradiation. The soil samples were sterilized by

79

gamma irradiation (5 Mrad) and remained sealed in S-ml bottles until use. The soil
extracts were prepared by mixing soil and chemotaxis buffer solution at a ratio of 1:15
for Capac-A, 1:60 for Colwood, and 1:220 for Houghton muck soils. The suspension
were mixed, in sterile plastic bottles, and tumbled on an end-over-end shaker at 9 rpm.
After 2 days of incubation, the suspensions were centrifuged for separating soils particles
from the supernatant. The separated supematants were then filtered through two layers of
Whatman no.2 filter paper before use for desorption and bioavailablity experiments.
Chemical

Naphthalene was selected as the test contaminant because it has often been used as a
model compound for polycyclic aromatic hydrocarbons (PAHs); this class of contaminant
also poses a critical bioavailablity issue at numerous petroleum contamination sites. Both
14C-labeled and unlabeled naphthalene were used in this study. The l4C-labelled
naphthalene (sigma, 0.1 mCi, >98% pure) stock solution was prepared by supplementing
it with unlabelled naphthalene and was used for sorption, desorption, serial dilution
studies. The un-labelled naphthalene (sigma, > 98%) stock solutions were prepared in
methanol and stored in 5 ml screw cap vials (capped with Teflon septa) at 4°C before use.

All biological assays were carried out with unlabelled naphthalene stock solutions.

80

Table 4-1. Selected properties of soils to be used in this study

 

 

Soil % O.C. % Sand % Silt % Clay PH CEC
[cmol(+)/kg]

Capac 3.3 55 24 21 6.8 24

Colwood 7.8 64 21 15 6.0 43

Houghton Muck 38.3 ND ND ND 5.1 156

 

ND: not determined yet.
O.C.: organic carbon content.

CEC: cation exchange capacity.

81

_B_acterial strain and culture preparation

Wild-type Pseudomonas putida G7 along with its two mutant strains were used in
this study. Wild-type PpG7 (ATCC 17485) and a mutant strain [Pseudomonas putida
G7.C1(pHG100)], which is non-chemotactic to naphthalene, were obtained from Dr.
Caroline Harwood (University of Iowa). The non-flagellated (non-motile) mutant strain
of P.putida G7 was obtained from Dr. Michael Aitken (University of North Carolina); the
non-motile strain was spontaneously generated and isolated as described elsewhere (Marx
and Aitken 2000). The pure wild type, pseudomonas putida G7, has been reported
previously to uptake sorbed phase naphthalene in soil slurry. Furthermore, it has been
characterized as being motile, and chemotectic (Grimm and Harwood 1997; Grimm and
Harwood 1999).

All four strains were stored at (-80 °C) in cryo-vials containing aliquots of
overnight culture supplemented with 75 % glycerol. Bacterial culture was grown by
lightly scraping a frozen stock with a sterile inoculating loop and, transferred it to
tryptone broth. For non-chemotectic mutant the broth was supplemented with a final
concentration of 50 ug/l of tetracycline (sigma) to select it for the desired mutants. The
cultures were allowed to incubate at 25 °C overnight on a shaker. The overnight cells
were then centrifuged for 2 min., and subsequently re-suspended in the chemotactic
buffer. To obtain pure colonies, the washed cultures were streaked on minimal media
plates, containing tetracycline for non-chemotactic mutants. Naphthalene was provided as
sole carbon source by placing its crystal in the lids of the plate. The plates were incubated

at room temperature for 3—4 days for colonies to grow. The seed cultures were then stored

at 5°C.

82

For each bioassay, fresh bacterial inoculums was prepared by placing colonies
from seed culture plates (using sterilized inoculating loop) into 250 ml of flask with 50
ml of mineral salt media (Harwood et al. 1994). The pH of the medium was kept at 7.0.
Naphthalene was added as a concentrated stock solution (in methanol) to the sterilized
salt media to a final concentration of 200 mg/L. The liquid cultures were incubated at
room temperature with stirring and growth was monitored by measuring the absorbance
at 600 nm. Cells in the early stationary phase were harvested by centrifugation, and re-
suspended in CB solution(Harwood et al. 1994). This procedure was repeated three times
to ensure the removal of any remaining naphthalene from the cell growth medium. The
number of viable cells was determined by standard plate count. The CB contains 13.6 g/L
of potassium phosphate and 7.4 mg/L of EDTA, and adjusted by 10 N NaOH to pH.
Tetracycline was added at a final concentration of 50 jig/ml to the mineral-salt medium
while growing non-cheomotactic mutants
Sorption. Desorptionband Serial dilution Experiments

Batch sorption experiments were performed for all the test soils. An aliquot of
sterile soil and 4.2 ml of chemotactic buffer was taken in each of the series of 5-ml vials
and spiked with a known amount of U-MC-labeled radioactive naphthalene (7000 —
60000 dpm). The serum vials were sealed with Teflon-lined caps. The ratio of sorbent to
liquid was kept the same as the ratio in the bio-availability assay; the solid-to-solution
ratio was set as 1:15 for Capac-A soil, 1:65 for Colwood soil, and 1:240 for muck soil.
Seven initial liquid-phase naphthalene concentrations (ranging from 0-2500) were used

(in triplicate) for each type of soil. The headspace was less than lml in volume. Control

83

vials without sorbents were prepared in triplicate. Vials were tumbled at 9 rpm in the
dark for 2 days. After mixing, each vial was centrifuged for 5 min at 1200 x g to separate
liquid from the sorbent, and 1 ml of the liquid phase was sampled in duplicate. The
concentration of naphthalene in the supernatant was determined by measurement of
radioactivity using liquid scintillation counter (LSC). LSC results were verified by high
performance liquid chromatograph (HPLC) for representative samples. Solid-phase
concentrations were calculated by subtracting equilibrium liquid-phase concentrations
from initial liquid phase concentration, and the results were confirmed by direct analysis
of sorbed-phase naphthalene concentration in selected samples for each type of soil. The
direct analysis entailed removal of supernatant from the vials and replacing it with
methanol. The extracted amount of naphthalene in methanol was then determined by LSC
and HPLC to verify the mass balance.

Desorption assays were conducted to measure desorption rates. The initial step
for desorption assay was the same as in the sorption experiments utilizing batch soil
slurries. The only two exceptions were that one 50 ml vial was used instead of 7, 5 ml
vials for each soil, and the aqueous phase was spiked with an initial naphthalene
concentration of 2 mg/l. Each 50-ml centrifuge-tube, containing 45 ml of chemotactic
buffer, was spiked with initial naphthalene concentration of 2 mg/l. The solid-to-solution
ratio remained the same as in the sorption experiment. The tubes were capped with
Teflon-lined Minnert valves and screw-sealed. A control tube with no soil was included
in the experiment to keep track of naphthalene losses due to volatilization and wall
sorption, if any. The tubes were tumbled at 9 rpm for two days to reach apparent

equilibrium, as determined in the sorption kinetics assay. At the end of sorption

84

equilibrium, each tube was centrifuged for 20 min at 1200g to separate soil, and the
supernatant was sampled. The concentration of naphthalene in the supernatant was
determined as mentioned in sorption assay. After sampling, the supernatant was decanted
to the maximum extent possible. Preliminary studies showed that the residual
supernatant was 0.27 ml with a standard deviation of 0.03 ml. Desorption was initiated
by adding naphthalene free soil extract to make up the decanted volume. The vials were
tumbled again at 9 rpm, and at pre-deterrnined time intervals they were removed and
centrifuged for liquid phase sampling. The concentration of naphthalene was determined
by LSCE and verified by HPLC. After final desorption samples were taken, the
remaining supernatant was decanted and the sorbed phase was extracted with methanol.
The extracted naphthalene amount was used to verify mass balance. The desorption data
obtained from this assay was fitted with a three-site desorption model to evaluate
desorption parameters and site fractions. These parameters and site fraction information
were utilized in the bioavailability model to quantitatively account for the effects of
desorbed naphthalene on bioavailability.

Serial-dilution desorption assay was conducted to measure the amount of non-
desorbable naphthalene in soil samples. The measurement was verified by an air-
sparging method. The serial dilution experiment was carried out in 5 ml sacrificial vials
containing sorbent slurries (with the same sorbentzsolution ratio as in the sorption study)
being spiked with an initial naphthalene concentration of 2 mg/L in the solution phase.
For each soil triplicate vials were prepared for each sampling period. The slurries were
tumbled at 9 rpm for 2 days, the supernatant decanted, sampled and analyzed for

naphthalene, and the bottles refilled with naphthalene-free liquid to the original volume.

85

This procedure was repeated for a total of 6 successive desorption periods. The non-
desorbable naphthalene was calculated by difference, with confirmation of mass balance
by solvent extraction of the sorbent following this series dilution procedure. The
calculated non-desorbable naphthalene amount was compared to the estimated value by
the three-site desorption model from desorption data.

The air-sparging method involved sample preparation similar to the serial dilution
method as mentioned above. At the end of 2 days sorption equilibrium, the vials were
connected to each other in a serial configuration via steel tubing (as shown in Figure 4-1).
The first vial was connected to a nitrogen supply source and the last vial attached to a
carbon trap through a steel tube. The continuous injection of nitrogen gas below the water
surface of each vial effectively purged dissolved naphthalene into overhead space. The
purging operation was stopped when dissolved concentration of naphthalene was reduced
to a non-detectable limit. The vials were then taken out of the assembly, decanted, and
the non-desorbable fraction was extracted with methanol.

Motility apd chemotaxis w

To determine the motility of the naphthalene-degrading bacteria, cells were grown
in stabs of semisolid media with tetrazolium salt (nutrient broth containing 0.4% agar) for
24 hours at 28°C. Motile bacteria were migrated from the line of inoculation to form
diffuse red-turbidity in the surrounding medium; nonmotile bacteria grew only along the

line of inoculation (Krieg and Gerhardt 1994; Leboffe and Pierce 1999). Chemotaxis was

86

tested with a soft agar swarm plate assay. For the swarm plate assay, cells were
inoculated with a needle stabbing into the center of the minimal medium that contains
naphthalene and is solidified with 0.4% Noble agar. Chemotactic bacteria migrated

outward while consuming naphthalene in the medium.

Bioavailability assay

To evaluate the availability of sorbed naphthalene to bacteria, a bio-availability
assay was carried out as described previously (Park et al. 2001). Briefly, the assay was
carried out in a set-up of 5-ml vials sealed by screw caps with Teflon-lined septa. The
soil to water ratio was kept the same as for sorption/desorption assays. Aliquot of
naphthalene stock solution was injected in each vial at an initial liquid-phase
concentration of ~2000 jig/L. The vials were tumbled on an end-over-end shaker at 9
rpm for 2 days in dark. Sterility of the soil slurries was checked by plating out the
suspension from selected vials on nutrient agar plates. After 2 days, each vial received an
aliquot of microbial cell suspension harvested during the early stationary phase, to initial
biodegradation of naphthalene. The aliquot of cell suspension corresponded to a desired
cell density of 5x106 CFU/ml in serum vials. Inoculated vials were tumbled on the end-
over—end shaker at 9 rpm, and at predetermined time intervals vials were removed and
centrifuged at 3000 rpm for 5 minutes to separate the soil from solution. One milliliter of
supernatant was transferred to a HPLC vial containing 0.01 ml of 10 N NaOH to stop
biodegradation of naphthalene. The remaining supernatant from the vial was decanted

and replace with an equal volume of methanol to extract remaining soil sorbed

87

naphthalene. Both supernatant and methanol extracted naphthalene concentration was
measured by HPLC. In bio-availability experiment two types of control vials, without test
bacteria, were used as a check for losses: one contained only buffer solution, and the
other contained slurry solution.

Biodegradation rates were measured in sorbent-extract solutions to be
representative of those expected in systems with sorbent present. A sorbent-extract
solution was prepared following the method outlined in the previous section. Kinetic
data was obtained using the same procedure as in the bioavailability assay, except that no
sorbent will be present in the vials. Both sterile-CB and soil-extract controls were used

to monitor for abiotic losses of naphthalene during the sorption and degradation periods.

DATA ANALYSIS: DESORPTION-BIOVAIABILITY MODEL (DBM)

A coupled biodegradation-desorption model (Figure 4-2) that incorporates
degradation of a non-desorbable fraction was used to track total as well as liquid phase
and sorbed phase naphthalene in a batch slurry reactor. The overall degradation of
naphthalene in a batch slurry reactor using mass balance on naphthalene can be

represented by equation 4-1.

—(Vl°£+m'-d-S-)=(Rbio'V'i'm'knd'Snd) [1]
dt (it

where m(kg) is the total mass of the soil, C(ug/L) represent naphthalene

concentration in solution phase, S(p.g/Kg) represent total sorbed-phase naphthalene, V(L)

represents volume of the liquid, and knd is the first-order biodegradation co-efficient for

88

0.2 um filter

 

Naphthalene

Figure 4-1. Configuration for Nitrogen-Sparging of Naphthalene in liquid phase

A = Dissolved and Equilibrium site
B = Non-Equilibrium site
C = Irreversible phase

In = mass flux from non-equilibrium

 

’\. ‘_Rb10

 

in
|

 

 

’\. *— Knd Snd

 

+o>t+w+>4>|

 

Figure 4-2. Different Compartments of Desorption-Bioavailability Model (DBM)

89

non-desorbable fraction(Snd) of total sorbed contaminant (S). Rbio is the liquid-phase

biodegradation rate expression, which can be expressed by following equations:

 

Rbio = Vm" ° C for Michaelis Menten kinetics, [2]
m

or

Rbio = K 1' C for first-order kinetics [3]

Where, Vmax is the maximum flux that a cell can obtain, Km is the Michael-

Menten constant, and K1 the first-order degradation rate co-efficient. The non-desorbable

fraction is tracked by following mass balance equation

—" = -kndsnd [4]

The total sorbed phase (S) was assumed to consist of equilibrium (Sal), non-

equilibrium (Sneq) and non-desorbable (Snd) sites. The equilibrium site represents the

fraction that releases spontaneously during the desorption experiment and is described by
a linear partitioning model. The non-equilibrium site fraction allows slower release of
contaminant and can be represented by a first order expression. The non-desorption site is
defined by Park et al. (2001) to be containing contaminant that cannot be released to
aqueous solution during the experimental desorption period. Mathematically, these three-

sorbed phase fractions are described as follows:

Seq=feq.Kd.C [5]
sud = fnd . Kd . ce [6]
dSncq/dt = (X . ( fneq . Kd . C — Sneq) [7]

90

Where Kd is the sorption coefficient, C6 is the liquid phase concentration in

sorption equilibrium, t is desorption time in minutes, or is the first-order desorption rate
coefficient (per minute). The three remaining terms feq, fneq, and fnd represent,

respectively, equilibrium site fraction, non-equilibrium site fraction, and non-desorption

site fraction. The three site fractions were measured from the desorption experiment: fnd

corresponds to the plateau of the desorption rate profile; and feq, fneq and or were

estimated by non-linear regression analysis of desorption data with the constraint that

feq‘l'fneq+ fnd =1 18]

RESULTS AND DISCUSSION
Sorption of Naphthalene
Figure 4-3 shows sorption isotherms for naphthalene on the selected natural soils.
The isotherms on all the three soils were nearly linear over the evaluated concentration

range, with the regression coefficient (n) ranging from 0.94 to 0.996. The sorption co-
efficient (Kd = slope of isotherms) ranged from 15 to 240 ml/g (Table 4-2) that increases
as organic content of the soil increases. To compare the sorption of natural sorbent with
synthetic sorbent, activated carbon was tested using the same experimental procedure as

used for natural soils. Sorption of naphthalene by granular activated carbon displayed a

curvilinear isotherm, and the data were adequately fitted with a Freundlich’s

91

 

 

 

 

 

 

Ge+5
’6 Muck
it 564-5 - Colwood
‘f Capac-A
0
g 4e+5 4 GAC
0
o
C
2 394-5 ~
I!
5
ii
a Ze+5 4
C
'0
§ 1e+5
o
m ._..._-
0 .
0 500 1000 1500 2000

Aqueous naphthalene conc. (ugiL)

Figure 4-3. Sorption Isotherms of Naphthalene in natural soils and
Granulated Activated carbon (GAC)

Table 4-2. Parameters for sorption of Naphthalene by the A horizon of selected soils and

Granular Activated Carbon

 

 

 

Freundlich equation Linear equation
Soil KF n R2 Kd R2 Kac
((ug/kgflwg/Li) (11kg) (L/kg)
Capac
-- -- -- 15 0.94 457
C l ood
0w -- -- -- 68 0.99 872
H t M k
ough 0" "C -- -- —- 240 0.996 626
Granular Activated
15858 0.53 0 99 -- -- --
Carbon

 

92

model (Figure 4-3). Values of the Freundlich’s parameters (Kp, and n) are listed in
Table-4-2. Mass balances in sorption equilibrium studies ranged from 93 to 102 percent

of the initial naphthalene added to the aqueous phase of the system.

De-sorption of Naphthalene

De-sorption of naphthalene from all the selected sorbents was expressed in
percent of the total sorbed naphthalene desorbed during the time of measurement (Figure
4-4). The three site desorption model was used to fit de-sorption data and to estimate
desorption rate coefficient along with each site fraction (equilibrium, non-equilibrium,
and non-desorbable) for all sorbents (Table 4-3). The three-site mathematical model fit
well to all the sorbents, with R2 values ranging from 0.92 to 0.96. The equilibrium site
fraction, which desorbed instantaneously, was observed to decrease as the soil-carbon
content increased ranging from 0.74 for low carbon soil (Capac-A) to 0.013 for activated
carbon.

Desorption from non-equilibrium site fractions of all natural soil was very
significant within the first 1000 minutes; however, further desorption beyond 1000
minutes was negligible. Desorption from the activated carbon was almost completed
within the first 250 rrrinutes. The desorption rate coefficient (or) is generally observed to
increase as carbon content increases, ranging from 0.0016 for Colwood soil to 0.0075 for
activated carbon. The extent of desorption is same for all natural soil (~85% within de-
sorption time of 3 days); however, activated carbon desorbed only about 14% of the total
sorbed naphthalene. A large fraction of the sorbed naphthalene was observed to be
resistant to desorption in activated carbon, as compared to natural soil. This may be due

to the surface adsorption of naphthalene on hard/black carbon of activated carbon

93

Desorption (%)

Table 4-3. Desorption parametersa evaluated by a three-site desorption modelb

Soil

 

 

feq fneq fnd a R2
(min'l)

Capac

0.74(0.042) 0.1 0.16(0.009) 0.0025(0.0008) 0.94
Colwood

0.62(0.023) 0.23 0.15(0.010) 0.0016(0.0004) 0.94
Houghton Muck

0.58(0.034) 0.27 0.15(0.005) 0.005(0.0017) 0.97
Activated Carbon

0.013(0.004) 0.127 0.86 (0.01) 0.0079(0.0006) 0.98
a feq fneq fnd: equilibrium-, non-equilibrium-, non-desorption-site fractions; 6!: first

order desorption rate coefficient for non-equilibrium sites.

bNumbers in parentheses are standard deviations of the evaluated values.

100

 

 

 

 
  

 

 

 

O Colwood

O Muck
v CapacA
O Granulated Activated Carbon (GAC)

40 ~

20 a

O I I I l I l
0 1000 2000 3000 4000 5000 6000

Desorption time (min)

Figure 4-4. Desorption of Naphthalene from natural soils and Activated Carbon

94

as compared to natural soil. This may due to the surface adsorption of naphthalene on
hard/black carbon of activated carbon. The surface adsorption is a competitive
phenomenon and is expected to be stronger as compared with partitioning into natural
soil organic matter.
Determination of Non-Desorbable fractions

The serial dilution and nitrogen-sparging experiments were carried out to
determine non-desorbale fraction of naphthalene for all the sorbents. Table 4-4 presents a
summary of the non-desorbable fraction for each of the sorbent using three different
methods. Figure 4-5 shows desorption isotherms for all the sorbents using serial dilution
experiment. The desorption isotherm for all the sorbents lie significantly above the
sorption isotherm, indicating clear existence of non-desorption sites . The non-desorption
site fraction was then calculated from the intercept of the desortion isotherm and the
initial sorbed-phase concentration. The calculated values were further confirmed by
extraction of the soil with methanol after the last desorption step. For activated carbon,
the non-desorabable fraction was extracted sequentially in three steps. In the sparging
experiment, the non-desorbable naphthatlene was extracted from sorbents when liquid
phase napthalene in soil-water-slurry effectively reduced to a non-detectable
concentration. The fraction was then calculated by dividing the mass of napthathlene in
non-desorbable fraction to initial sorbed-phase.

For all sorbents, the serial dilution method showed greater non-desorbable
fraction than nitrogen the sparging method. This difference could be explained by the fact

that water alone cannot extract a non-labile fraction of naphthalene that has been shown

95

Table 4-4. Fraction of desorbable naphthalene by three methods in natural soil and
and Activated Carbon.

 

Serial dilution Spargin g Non-linear regression

Soil Analysis

 

 

Non-desobable (fnd) Non-desorbable (fnd) Non-desorbable(fnd)

Capac-A 0.21(6x10‘3) 0.13(1.4x102) 0.16(9x103)
Colwood 0.19 (2x103) 0.09(1x103) 0.15(1x102)
Muck 0.17(1x102) 0.1(7x103) 0.15(5x103)
GAC 0.69 (1x102) 0.75(2x102) 0.86 (1x102)

 

96

Colwood CapacA

 

 

 

  

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

A C 18000
a, 70000 .0 16000 4 ,0
E} R2 = 0.9986 ” E 14000 , R2 = 0,9954 ”
V 1 I E 12000~ /’
'8 c 500004 I or A
o I o
E '3 40000. ,’ g
0 I
a) ‘E’ 30000 g
8 20000 - .8
C
0 10000 0
0 0 v . . U)
0 500 1000 1500 1000
Liquid concentration (ug/L) Liquid concentration (ug/L)
. M k
g Granular Activated Carbon (GAC) 250000 uc
68+5 A
or
3
‘z’ . 200000 - 2
g 5°+5 ' g Fl = 0.9975 IX!
8 «+53%, fififififi a. 8 8 150000-
0 I—
f! 33.5 i '2 E 100000 -
2 (D E
E 2e+5 8 50000-
a} 0 Sorption data c
z 1e+5 A step-desorption data 8 o r
'3 0 500 1000
e O r r r l
(3 0 200 400 600 800 1000 Liquid concentration (ug/L)

Aqueous Naphthalene conc. (ug/L)

Figure 4-5. Serial dilution desorption in three natural soils and
Granulated Activated carbon (GAC)

97

to be present in the solid phase. However, the sparged-nitrogen may access a larger pool
of labile fraction of the naphthalene in the soil, perhaps representing a better method of
determining this fraction. In all the abotic experiments, the HPLC analysis indicated that
naphthalene was not transformed during the sorption and de-sorption periods. The total

recovery of naphthalene was higher than 93% for all sorbents used.

Characterization of strains

To clearly observe the role of chemo-taxis and motility in naphthalene
degradation, the wild type PpG7 and its two mutant strains were first checked for these
phenotypes. Each strain was first tested for chemotaxis and motility with a soft—agar
swarm plate assay. Only wild-type cells had a positive chomotactic response to succinate
and naphthalene. The wild type was observed to move outward from the center of plate as
they metabolized the carbon source, forming a sharp ring of growth. Contrary to the wild-
type, both non-chemotactic and non-motile mutants did not respond to the gradient of
carbon source that was created as they metabolized the carbon source at the center of
plate. All three strains were further observed under transmission electron microscope
(TEM) for flagellated structures: the wild type and non-chemotactic strains possessed
polar flagellated structure, while the non-motile strain was devoid of a flagellated
structure (Fig 4-7). These two tests confirmed that the wild type possessed both motility
and chemotactic phenotypes, whereas non-chemotactic and non-motile strains did not
have the capability to sense and move due to missing flagella and chemo-receptor genes,

respectively.

98

 

Figure 4-6. TEM for Pseudomonas putida-G7(wild-type) on Left, and

Non-motile mutant on right

 

 

   
  

 

 

 

 

 

   

 

1.60E-06 3 90°
0.
.9 1.40E-06 — I 3 80° "
E g C 700i
0
C 1.20E-06 — z:
.o A S 600 _j 0 Measured
a C o—r
to 'E 1.00E-06 — c
E 3' 800507 _ g 500 i _.Flegression
8: LL ' 0 Wild Type 8 400 -u
‘9 Q 6.00E-07 - o L
'8. é, I Non- 3 30° ‘
<0 3’ 4-00'5'07 ' Chemotactic E 200 4-
E 2.00E'07 _ ' Non'motlle E 100 L
a a ‘
E 0.00E+00 l l I Z 0 l l - . v e
0 500 1000 1500 2000 0.00 0.50 1.00 1.50 2.00
Naphthalene, Ppb Time, hrs

Figure 4-7. Naphthalene degradation of three strains in Liquid medium

Naphthalene Degradation Kinetics in Soil Extract

Initial studies with wild-type Pseudomonas putida G7, non-motile and non-
chemotactic strains were conducted to determine if these variants biodegrade dissolved
naphthalene at the same rate. The kinetic studies were carried out at initial naphthalene
concentrations of 1 ppm and 2 ppm, with an initial cell concentration of ~2x106 CFU/ml.
The kinetic studies were carried out in soil extract, instead of chemotaxis buffer, since
dissolved soil organic matter can potentially affect naphthalene degradation kinetics.
Figure 4-7 shows initial naphthalene degradation rates for each strain. The three strains
showed similar rates of naphthalene degradation at both 1 ppm and 2 ppm of naphthalene
under a well-mixed system. This confirms that all three strains possess the same enzyme
system that is responsible for naphthalene degradation. Therefore, any difference in
naphthalene degradation rate or extent among these strains in a soil-slurry system would
be due to the respective phenotype (chemotaxis/motility). Also shown in Figure 4-7 is
degradation of naphthalene by PpG7 over an extended period of time. The degradation of
naphthalene in the soil-extract solution followed the Michaelis-Menten equation, and
kinetic parameters were estimated by fitting the Michaelis-Menten equation to
degradation data for three strains. The maximum specific utilization rate co-efficient,
Vmax, and the half-saturation constant, Km, were estimated as 500 ug/L-hr and 301 ug/L,

respectively.

Bio-availability of Sorbed Naphthalene in Natural Soils Under Mixed Conditions

Bio-availability assays for wild-type, non-motile mutant, and non-chemotactic

strains were performed using three natural soils with different sorption co-efficients. Data

100

from the bioavailability assays for all combination of bacteria and natural soils are shown
in Figures 4—8 to 4-10. In all cases, liquid phase naphthalene degraded rapidly and
become negligible within the first 4 hrs for all combinations of soil and bacterial strains.
Sorbed phase fraction, on other hand, degraded in a biphasic manner. The initial steep
curve represents rapid loss of desorbable naphthalene fraction followed by a much milder
curve that represents loss of non-desorabable fraction. The degradation of the desorbable
fraction was completed within the first 5 hours, coinciding with complete degradation of
the liquid phase fraction. At the end of the desorbable fraction, the degradation of the
non-desorabable fraction took place at a slower rate and last over an extended period of
time (> 40 hrs). The later part of the sorbed phase degradation profile clearly indicates
decreased bio-degradation rate due to the presence of the non-desorbable fraction of
naphthalene.

Clearly, in all cases, the initial sorbed-phase naphthalene concentration decreased
to levels below the non-desorbable concentration, as shown by the two horizontal lines.
The upper dotted lines and lower dashed lines represent extent of non-desorbable
fractions as obtained from serial dilution and Nitrogen-sparging experiments,
respectively. It is to be noted that the inflection point of the sorbed phase degradation
profile coincided with lower dashed lines, indicating that nitrogen-sparging method
provided better estimates of the non-desorbable fraction. The decrease of sorbed phase
naphthalene below the dashed lines, which represents a harsher extraction method,
indicates that all strains were able to degrade and remove more sorbed phase naphthalene
than could be accomplished in abiotic experiments. The bio-availability data, which

included both liquid phase and sorbed phase data, were then fitted with the coupled

101

Colwood soil with PpG7-WT CapacA soil with PpG7-WT

A
O

   
  

 

 

.3
O

 

 

 

 

 

 

 

 

 

 

 

 

0 Total Naphthalene 0 Total Naphthalene
9 4:» Solidphase A 9 : mm”
A . | ' a .1 . I 388
g Total Naphthalene W' a 8 'Napm‘a'em
z: w v '
7 «a
co to
as as
o
O 5 c 5 .
5 Sorbed phase Naphthalene % 4
:3 fa-
.: 3 Liquid phase Naphthalene 8' 3 ..
gzl / Z 2‘
1.9: _ _- l :
. o
0o 10 40 _ 40
Time (hours) Time (hours)

Muck soil with PpG7-WT

 

0 Total Naphthalene
+ Solid phase
0 liquid phase

 

 

 

Naphthalene mass (ug)

 

 

 

Time (hours)

Figure 4-8. Naphthalene degradation due to PpG7-Wild type in three natural soils

102

CapacA soil with PpG7-nonchemotactic mutant
10

Naphthalene mass (ug)

D

 

 

.5
O

 

 

 

 

 

 

 

 

 

 

 

 

o TotalNaphthdene 0 Tasmania
4} Solidphase 9 1h Solidphase
O liquidphaee a O liquidphase
a
7 T IN .8 km=0035100014 i} 7 km=0.041i0.003
ota aphtha ne m
lNapNhalene
g 5
2 5
0
El 4 SorbedphaseNaphthalene
a 3
to ' 'd seN hthalene
Z 2 )uqu pha 30
1W
00 1o 20 30
Time (hours) Time (hours)

ngck soil with PpG7-nonchemotactic mutant

 

      
  
  

 

 

 

0 Total Naphthdene
9 4r Solidphase
A O limidphase
0’ 8
a k M = 0.0409i0.0018
a 7 Total Naphthalene
as
E 6
2 5
0
To 4 rbed phase Naphthalene
3. 7mm phse Napflhalene
Z

 

 

Figure 4-9. Naphthalene degradation by Pp—G7( Non-chemotactic mutant) in three natural soils.

103

 

C1aopacA soil with PpG7—nonmotile mutant quwood soil with PpG7-nonmotile mutant
1 .

 

  
  
 

 

 

 

 

 

 

 
 
 
 
 

  
  

 

2 mw 0 Total Napl'lthdene
A - mam A 9’ T We”
3 Total Naphthalene 3 “l ‘ “WW”
3 km=°-°35‘-‘°-°°°9 3 7 T al Naphthalene know-0430003
a to
E 5 E 6
g 5 o ,., Phase Naphthalene 2 5"
o o Sorbedphase Naphthalene
E 4 'a
a .. g
29; 23 £3 you'd Phase Naphthalene

 

 

 
  

 

 

 

 

40 0 10 20 40

Time (hours) Time (hours)

Muck soil with PpG7-nonmotile mutant

 

 

 

 

10' <> TotalNaphthdene
9. It Solidphase

I liquidphase
3‘ k M=0.0401i0.002
<>T Napflhalene

 

7
6
5
4

Naphthalene mass (ug)

 

 

 

20 40

Time (hours)

Figure 4-10. Naphthalene degradation by Pp-G7( Non-motile mutant) in three natural soils.

104

desorption-biodegradation model (DBM). The simplified DBM model provided a
reasonable description of both liquid— and sorbed phase data for all combinations of
strains and soils over the entire experimental period. The DBM model estimated two sets
of degradation rate coefficient: one for easily degradable fraction (liquid and desorbable
fraction) and other for difficult to degrade fraction (non-desorbable fraction). The
liquid/desorbable fraction of naphthalene was assumed to be degraded following Michalis
Menten kinetics with coefficient (Km and Vm). The non-desorbable fraction was, on
other hand, assumed to follow lst order kinetics with coefficient represented by Knd.
These estimated degradation rate coefficients from all of the combinations are shown in

Table 4—5.

Bio-availability of Sorbed Naphthalene in Natural Soil Under Static Condition

To illustrate more clearly the role of chemotaxis and motility in overcoming
mass-transfer limitation, bioavailability experiments were repeated without mixing. All
the other conditions were kept similar as in the mixed system, except the whole content
was under quiescent conditions. The three strains were tested both in low organic (Capac-
A) and high organic (Muck soil) soil-slurry systems. It was expected that only wild-type
(which is motile and chemotactic) could access to localize zones of naphthalene, which is
created due to lack of mixing, as well as to the non-desorbable/sequestrated fraction of
sorbed naphthalene. Thus, comparing the rate and extent of wild type with two mutant
strains would clearly demonstrate role of chemotaxis and motility phenotypes on

biodegradation of soil-sorbed naphthalene.

105

Table 4-5. Bio-kinetic parametersal evaluated for three strains of Pseudomonas putida G7

 

 

 

 

 

 

 

Wild Type Non-motile Non-chemotactic
50,. km Vm kno km vmax knd km Vmax knd
“g’L llhr l/hr lug/L) (l/hr) (l/hr) (ug/L) (l/hr) (l/hr)
cap“ A 980 300 30932) 1050 325 $093251) 890 260 $3157)
C°1w°°d 980 250 $9353) 1080 280 30%;» 1200 300 $3222)
Mm" 950 300 33324) 950 310 3693216) 1050 280 3033123)

 

a km and Vma, are Michaelis-Menten degradation parameters for liquid and desorbable sorbed phases ;

knd = first order degradation rate for non-desorbable fraction. Standard deviations are listed in parentheses.

106

Data from the bioavailability assays, under no-rnix conditions, for all combinations of
selected bacteria and natural soils are shown in Figures 4-11 and 4-12. In all cases,
bacteria could only degrade naphthalene in liquid and desorbable fraction of the
naphthalene; and, were unable to degrade mass-transfer limited and non-desorbable
fractions of the sorbed phase. The degradation profiles for all three strains in each soil
were observed to be similar, with initial rapid degradation phase (represented by the line
with steeper slope) followed by no-degradation phase (represented by the horizontal line)
due to mass transfer limitation of naphthalene from solid phase. None of the profile from
all combinations dropped to a level below the non-desorbable naphthalene concentration,
shown by horizontal dashed lines. At the end of the treatment period, the amount of
residual naphthalene was observed to be greater in Capac A soil than Muck soil. This can
be attributed to the fact that in the case of Muck soil, greater surface area was available
for mass-transfer of naphthalene into the overlying water column as compared to Capac-
A soil. Thus it can be expected that, under quiescent conditions, greater surface area
would offer less resistance to mass transfer of naphthalene from the soil-water interface
and hence increased degradation. The bioavailability data were then fitted with the DBM
model to estimate the degradation rate coefficient for all combinations. By incorporating
reduced desorption fraction, due to static conditions, DBM provides a reasonable
description of total naphthalene degradation data for all combinations of strains and soils
over the entire experimental period. The DBM model estimated only keq, which represent
both liquid and desorbable fraction of the sorbed phase. The degradation of the non-
desorbable fraction of the sorbed phase was estimated as zero. The estimated degradation

rate coefficient (keq) from all combinations is shown in Table 4-6.

107

Naphthalene mass (ug)

.5
N

_L
O

 

  

 

 

 

 

‘wnd rype'(G7)_ M
5o 46 go 810
Tune (hours)

Figure 4-11. Naphthalene degradation by three strains in Muck soil under static condition

12

10

Naphthalene mass (ug)
or

 

 

 

Wild-Type G7
2.0 40 60 80 100 150
Time (hours)

Figure 4-12. Naphthalene degradation by three strains in Capac-A soil under
static condition

108

Bioavailability of sorbed naphthalene in Activated Carbon under mixed system

The bioavailability of naphthalene was also compared for wild type and non-
motile strains in a mixed slurry system containing activated carbon. The objective of the
experiment was to test whether pseudomonas putida G7 could differentially utilize GAC-
sorbed naphthalene, and also to determine its accessibility to the non-desorbable pool of
naphthalene due to its motility and chemotactic phenotypes.

The degradation rate data for both wild-type and non-motile strains in a mixed
slurry system containing activated carbon is shown in Figure 4-13. The profile of
naphthalene degradation for both strains indicates an initial rapid degradation followed
by an extended period of slow degradation. The initial period of degradation, which
mostly occurred in liquid and desorbable fraction, lasts for the first 3 hours. During the
subsequent period, the degradation of naphthalene occurred in a non-desorbable pool of
naphthalene at a relatively slower rate.

It is clear from Figure 4-13 that both strains could degrade non-desorabable
fraction of naphthalene. This has been indicated by the profile of total napthalene mass
dropped to a level below the non-desorbable fraction, shown by the dashed horizontal
lines. However, wild-type strain could able to degrade non-desorbable fraction at a
relatively greater rate, as indicated by slope of degradation profile (Figure 4-13, compare
diamond symbols for wild-type and asterisk symbols for non-motile strain). For
quantitative comparison, the data were fitted with the DBM model to estimate biokinetic

parameters (Table 4-7).

109

Table 4-6. Estimated bio-kinetic co-efficient for liquid/desorbable fraction (Keq) for Wild-
type and non-motile strains in CapacA and Muck soils under static conditions

 

 

 

 

 

Wild Type Non-chemotactic Non-motile mutant
Soil mutant

Keq (h‘W Keqar‘) Keq (h-l)
CapacA 0.50 (6x103) 0.42 (2.7x10‘5 0.41(4x10°2)
Muck 0.52 (2x102) 0.352 (1.57x10'2) 0.352 (5x103)

 

llO

Table 4-7. Estimated bio-kinetic co-efficients for desorbable phase (Keq) and
non-desorbable phase (knd) for wild-type in Activated Carobon (GAC)

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Wild Type Non-motile mutant
Keq (l/hr) Knd (l/hr Keq(1/hr) Knd (l/hr)
1.6 0.004 1.48 0.0029
(0.22) (3.8 x104) (0.3) (5x10“‘)

12

‘ Non-Motile
10 0 Wild Type (G7)
8

 

Non-Motile mutant

 
     

’6:

3

'U

(D

C

'3

g

h 6

(D

(D

E ‘ ~-~ .5 i
0 4’ w/ 1‘
8

E ildType(G7)
5 2~

.C

Q.

(U

2 o

 

 

0 20 40 60 80 100 120
Time (hours)

Figure 4-13. Degradation of Naphthalene by Wild type and non-motile strain in
Activated carbon

11]

The de-gradation coefficient of the non-desorbable fraction was determined to be greater
for the wild type as compared to the non-motile strain, indicating a possible role of

motility and chemotaxis phenotypes.

DISCUSSION
Chemotaxis/motility can provide a selective advantage to bacteria in an environment
where bioavailability limits biodegradation of contaminants. In a soil-slurry system,
containing hydrophobic compound, there always exists heterogeneity in contaminant
distribution due to the presence of hi gh-sorptive capacity constituents and a localized lack
of mixing. Chemotaxis can guide bacteria to swim, using flagella, and bring them into a
close physical contact with soil-sorbed contaminants when desorbable substrate is
depleted. As bacteria accumulate in zones of concentrated chemical gradient, they
degrade the contaminant at a faster rate. In addition, chemotaxis/motility might be
required for the bacteria to move through inter/intra particle pore spaces in response to
micro-scale chemical gradients. Thus even small increases in cell density surrounding
chemical concentration gradients can lead to enhance bioavailablity as observed in earlier
studies (Park et al 2001, Tang et al. 1998).

The role of chemotaxis/motility in enhanced PAH degradation has only recently
been verified in porous media. For example, Pedit et al. (2002) reported an enhanced rate
of naphthalene biodegradation from a capillary tube, filled with glass spheres, in systems
inoculated with wild-type PpG7 as compared with systems inoculated with non-
chemotactic mutant strains. The enhanced degradation was observed corresponding to

increased number of chemotactic bacteria having moved into capillary tubes. Another

112

study found increased number of PAH degraders in rhizosphere, in response to PAHs, as
compared with unplanted bulk soil controls. The author estimated chemotactic motion of
the soil bacteria toward PAHs in the order of 1 mm/min (Ortega-Calvo et a]. 2003).

In the present study, experiments were carried out in both natural and synthetic
sorbents, with low to high non-desorbable fractions, to observe the effect of chemotaxis
and motility on enhanced bio-availability under a mixed and quiescent experimental
condition. In a mixed system containing natural soil, all the strains were observed to have
similar access to a pool of both liquid/desorbable fraction as well as non-desorbable
fraction of soil sorbed naphthalene (Figures 4-9 to 4-11). The evaluated degradation rate
constants for all three strains were found to be similar at a 95% confidence interval for all
three soils (as shown in Table 4-5). In a similar system containing GAC, the wild type
degraded non-desorbable fraction at a significantly (at a 95% CI) greater rate as
compared to two mutant strains. However, as in natural soil, both wild type and non-
motile strains had equally degraded liquid/desorbable fraction (Table 4-7).

Several plausible explanations can be offered for different responses of strains in
activated carbon and natural soils under mixed conditions. In both natural and GAC
sorbents, biodegradation began with a constant concentration of naphthalene. As bacteria
degrade naphthalene from the aqueous phase, the concentration changes from one
isotropic concentration value to another constant value. The maintenance of isotropic
naphthalene concentration in the liquid phase was due to uniform mixing of naphthalene
that desorbed from the sorbed phase. In the absence of a concentration gradient,
chemotactic movement did not occur, and hence both liquid/desorbable phase

naphthalene was degraded equally by all strains in both natural and GAC. If there was a

113

concentration gradient at the solid-water interface, it was highly unlikely to attract
bacteria due to the presence of a large amount of dissolved naphthalene. Thus, existence
of an isotropic concentration condition during the rapid phase of degradation resulted in
no detectable gradient and consequently no chemotactic movement.

After most of the liquid and readily desorbable fraction of naphthalene was
degraded, there remained non-labile fractions of naphthalene in both natural and GAC.
The non-labile fraction diffuses out naphthalene into surrounding liquid at a very low rate
due to film resistance. For biodegradation to continue, the bacteria had to come closer to
particle surfaces and directly uptake naphthalene from within the stagnant hydrodynamic
boundary layer. Chemotaxis/motility would allow cells to move toward this higher
concentration region located at water/solid interface. The results from the present study
did not discern role of chemotaxis/motility, as all strains had gained equal access to the
non-labile pool of naphthalene in both natural soil and GAC. This may also be attributed
to the mixing affect as mixing increases the probability of random collision between soil
particles and bacteria and hence greater chance of adhesion. Thus mixing had seemingly
masked the role of chemotaxis and motility during initial phase of non-labile naphthalene
degradation. Previous study with chemotaxis-mediated biodegradation illustrated the
importance of relative rates of substrate mass transfer and bacterial movement. In a
mixed or partially mixed system, bacterial chemotaxis and motility was shown to be a
much slower process as compared with movement of naphthalene due to
convection(Marx and Aitken 2000).

However, the role of chemotaxis/motility seems to become apparent at a longer

incubation time in GAC (Figure 4-13). After initial mixing-mediated cell adhesion, more

114

cells become attracted from the liquid phase toward interface due to a chemical gradient.
This resulted in a greater number of cells per unit area of the sorbent particles in a system
inoculated with wild type as compared with non-motile strain. Consequently, the rate and
extent of degradation of non-labile naphthalene was greater with PpG7 than
corresponding non-motile strains. Similar observations were not observed in natural soil
probably likely due to short incubation time and reduction of non-labile naphthalene to a
threshold concentration, below which biodegradation stops.

To remove the effect of mixing in masking bacterial chemotaxis and motility, the
experiments were repeated under quiescent condition. All the strains were introduced
carefully at the surface of the water column located well above the thin layer of soil
deposited at the bottom of the vials. It was expected that as bacteria degraded
naphthalene in the liquid phase, a concentration gradient would be created in a vertical
direction. In response to the naphthalene gradient, wild type PpG7 would migrate in large
numbers from the liquid phase toward the sediment/water interface. The mutant strains,
on other hand, either could not respond to gradient or might move downward in fewer
numbers. It was also expected that the presence of a large number of wild type PpG7 and
a higher concentration of naphthalene available at the water-soil interface would result in
a greater rate and extent of Naphthalene degradation.

Furthermore, the porous structure of soils does not seems to hinder the movement
of bacterial cells, since soil pores have sizes ranging from several to hundred of
micrometers, which corresponds to the length of the straight-line segments of random
bacterial motion (Zaval'skii and Voloshin 2003). Thus, it was expected that only bacteria

with motility and chemotactic phenotypes would be able to move from the liquid-phase

115

into soil pores, as a result of large microscopic chemical gradients. The ability of wild
type PpG7 to move within the porous spaces would further result in greater accessibility
to sorbed phase naphthalene and hence increase degradation rate and extent. This role of
chemotaxis in penetrating porous spaces has been well documented in earlier studies; the
chemotactic migration was found to be faster in the porous medium than in the aqueous
phase (Duffy et al. 2003). The faster movement was attributed due to large rnicro-scopic
chemical gradients in porous medium.

However, the experimental results of this study did not support the expectation of
the role of motility and chemotaxis in enhancing bioavailability under quiescent
conditions. None of the strains, in either low or high K, soil were able to access mass
transfer limited/non-desorbable fraction of naphthalene. The degradation profiles for all
the combinations were observed to be similar, with an initial rapid phase of degradation
followed by a phase of almost negligible degradation as indicated by a horizontal line.
This reduction can be attributed to changes in the mass transfer parameter as a result of
biodegradation. As bacteria degrades naphthalene in solution phase, the mass transfer rate
from the sorbed phase decreased gradually and finally diffusion appeared to control the
movement of naphthalene from soil to overlying water. Both soils retained more
naphthalene mass than non-desorbable fraction, clearly indicating degradation is limited
by mass-transfer due to lack of mixing. Further evidence supporting this fact was the
observation that the fraction of mass retained is observed to be greater for Capac-A soil
than muck soil, due to its smaller surface area to total volume ratio. In conclusion, it
appeared that chemotaxis and motility was not observed to have caused enhanced

bioavailability of mass-transfer limited naphthalene.

116

Let us now consider the reasons for observing similar rates of naphthalene
degradation for all tested strains, and absence of chemotaxis/motility phenotypes in
overcoming mass transfer limited conditions in the quiescent system. Since the cells were
introduced into the experimental system in a vertical direction, we may expect the
potential effect of gravity on bacterial movement. The relative importance of gravity can
be seen by comparing the terminal settling velocity and chemotactic velocity of injected
cells. In the aqueous system wild type P. putida G7 respond to chemical gradient at a
typical chemotactic velocity range from 20 to 40 urn/sec (Marx and Aitken, 2000 b),
which is far greater than its corresponding terminal settling velocity of 0.02
um/seC(Pedit et a]. 2001). Therefore, in the experimental system, with a water column
height of 6 cm above soil layer, the relative affect of gravity can be neglected on overall
movement of bacterial cells due to chemotaxis/motility.

In the absence of a gravity effect, several plausible explanations can be offered for
not observing chemotaxis and motility effects on naphthalene gradation in the quiescent
system. If the initial concentration of the attractant is high enough to saturate bacteria’s
chemoreceptor, then the bacteria cannot sense the concentration gradient. For this study,
the saturation of the chemoreceptors could have occurred, but this assumption cannot be
validated since no data is available on dissociation constant for the PpG7 chemoreceptor.

The inaccessibility of WT strains to non-desorbable pool of naphthalene can be
explained in terms of the complex and heterogeneous nature of the soil matrix. The soil
used in this study contained various amounts of organic and clay matters as shown in
Table-1. The presence of organic and clay content could have restricted bacterial

movement due to: 1) bacterial size exclusion effects at the soil water interface, and 2)

117

adsorption-adhesion effect due to physio-chemical interaction between bacteria and soil
particles. Earlier studies that observed chemotactic movement were mainly canied out in
a saturated sand medium that contained continuous macro-pores with no clay or organic
content. Even with clean sand matrix, the previous researcher (Barton and Ford 1995)
observed marked reduction in chemotactic movement as compared to that measured in
fluid system. The reduction in motility coefficient was shown to be a function of soil

porosity and tortousity as given by equation-9:
8
”eff _ ‘1- [9]

Where, peg rs motility in bulk water, and - represents the ratio of $011 porosrty to
1:

its turtosity. In a recent study, the chemotactic sensitivity coefficient, which represents
key chemotaxis property, was determined to be reduced by 4 times in capillary assay in
presence of glass beads (Marx and Aitken 1999). Thus, chemotactic sensitivity
coefficient was greatly reduced in low porosity; high organic soils like Capac-A and

muck soil, and resulted in low bioavailability of sorbed naphthalene in no-rnix system.

CONCLUSIONS
In this study we evaluated the role of chemotaxis/motility on naphthalene
degradation in sorbent-slurry system under stirred and quiescent conditions. The rate and
extent of sorbed-phase naphthalene was monitored in the systems inoculated separately

with the chemotactic wild-type strain and its mutant strains, and compared the results for

118

evaluating the role of chemotaxis/motility. In natural soils (with low to high organic
content), under both stirred and static conditions, the rate and extent of naphthalene
degradation was observed to be similar for all the strains. Whereas in stirred system all
the strains had an equal access to a non-desorbable fraction of naphthalene; in the static
system, however, none of the tested strain had gained access to the non-desorbable
fraction of naphthalene.

Experiments carried out with activated carbon, under stirred conditions, showed
that both WT and mutant strains had accessed to the non-desorbable fraction of
naphthalene. However, WT degraded non-desorbable phase naphthalene at a significantly
greater rate as compared to its mutant strains. The difference of uptake rate between WT
and mutant strains could be explained due to the presence of significant concentration
gradient at soild/water interface as a result of large pool of non-desorbable naphthalene.
The concentration gradient attracted greater number of WT cells, due to its
chemotaxis/motility phenotypes, that resulted in a greater rate of non-desorbable
naphthalene degradation as compared with system inoculated with mutant strains.

This study suggest that in a heterogeneous soil containing medium to high content
of clay and organic content, the role mixing is very important for successful
bioremediation. In the absence of mixing both passive and active uptakes of naphthalene
were retarded. Lack of mixing reduced effective diffusion of molecules in soils due to
solid phases, dead-end pores and high tortousity of the system. Also, lack of mixing
reduced the number of sites available for bacterial attachment which would result in the

direct uptake due to pore sizes exclusion and retardation effect.

119

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Barton, J. W., and Ford, R. M. (1995). "Determination of effective transport coefficients

for bacterial migration in sand columns." Applied and Environmental
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Bosma, T. N. P., Middeldorp, P. J. M., Schraa, G., and Zehnder, A. J. B. (1997). "Mass
Transfer Limitation of Biotransformation: Quantifying Bioavailability." Environ.
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Duffy, K., FOrd, R., and Cumming, D. (2003). "Residence time calculation for
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Grimm, A. C., and Harwood, C. S. (1997). "Chemotaxis of Pseudomonas spp. to the
Polyaromatic Hydrocarbon Naphthalene." Appl. Environ. Microbiol., 63(10),
41 1 1-41 15.

Grimm, A. C., and Harwood, C. S. (1999). "NahY, a Catabolic Plasmid-Encoded
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Hydrocarbon Naphthalene." J. Bacteriol, 181(10), 3310-3316.

Harwood, C. S., Nichols, N. N., Kim, M. K., Kitty, J. L., and Parales, R. E. (1994).
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Involvement in Chemotaxis, Biodegradation, and Transport of 4-
Hydroxybenzoate." J. Bacteriol, 176(21), 6479-6488.

Head, 1. M. (1998). "Bioremediation: towards a credible technology." Microbiology- UK,
144(3-4), 599-608.

Jain, G. P. a. R. K. (2002). "Bacterial Chemotaxis toward Environmental Pollutants: Role
of Bioremediation." Applied and Environmental Microbiology, 68(12), 5789-
5795.

Krieg, N. R., and Gerhardt, P. (1994). "Solid, liquid/solid, and semisolid culture."
Methods for general and molecular bacteriology, N. R. Krieg, ed., American

Society for Microbilogy, Washington, DC, 216-223.

Law, A., and Aitken, M. (2003). "Chemotaxis to Naphthalene desorbing from an NAPL."
Applied and Environmental Microbiology, 69, 5968-5973.

Leboffe, M. J ., and Pierce, B. E. (1999). A photographic atlas for the microbiology
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Madigan, M., Martinko, J ., and Parker, J. (2002). Brock Biology of Micro-organism,
Prentice Hall.

Marx, R. B., and Aitken, M. D. (1999). "Quantification of chemotaxis to naphthalene by
Pseudomonas putida G7." Applied and Environmental Microbiology, 65(7), 2847-
2852.

Marx, R. B., and Aitken, M. D. (2000). "Bacterial chemotaxis enhances naphthalene
degradation in a heterogeneous aqueous system." Environmental Science &
Technology, 34(16), 3379-3383.

Park, J .-H., Zhao, X., and Voice, T. C. (2001). "Biodegradation of Non-desorbable
Naphthalene in Soils." Environ. Sci. Technol., 35(13), 2734-2740.

Pedit, J. A., Marx, R. B., Miller, C. T., and Aitken, M. D. (2002). "Quantitative analysis
of experiments on bacterial chemotaxis to naphthalene." Biotechnology and
Bioengineering, 78(6), 626-634.

Rijnaarts, H. H. M., Bachmann, A., Jumelet, J. C., and Zehnder, A. J. B. (1990). "Effect
of desorption and intraparticle mass transfer on the aerobic biomineralization of a-

Hexachlorocyclohexane in a contaminated calcareous soil." Environ. Sci.
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Sharma, P. K., McInemey, M. J ., Knapp, RM. (1993). "In situ growth and activity and

models of penetration of Escherichia coli in unconsolidated porous materials."
Applied and Environmental Microbiology, 59, 3686-3694.

Terracciano, J. S., and E. Canale-Parola. (1984). "Enhancement of chemotaxis in
Spirochaeta aurantia grown under conditions of nutrient limitation." Journal of
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Witt ME, Dybas MJ, Worden RM, and CS, C. (1999). "Motility-enhanced bioremediation
of carbon tetrachloride-contaminated aquifer sediments." Environ. Sci. &
Technol, 33, 2958-2964.

Zaval'skii, L. Y., and Voloshin, A. (2003). "Bacterial chemotaxis in porous medium."
Microbiology, 72(3), 369-372.

Zhang, W. X., Bouwer, E. J ., and Ball, W. P. (1998). "Bioavailability of hydrophobic

organic contaminants-effects and implications of sorption-related mass-transfer
on bioremediation." Ground Water Monit. R., 18(1), 126-138.

121

CHAPTER 5. INFLUENCE OF BACTERIAL ADHESION AND ENHANCE

BIOAVAILABILITY OF NAPHTHALENE IN SOIL

ABSTRACT

Bacterial attachment has the potential to increase the rate of degradation of
organic contaminants, but its influence on degradation of non-desorbable organic
contaminants in a soil-water system has not been examined. This study was an attempt to
examine factors affecting bacterial attachment and consequently enhanced bio-
degradation. The factors studied included molar concentration of buffer medium and the
bacterial cell surface structure. Bio-availability experiments were carried out in a stirred
batch soil-slurry system inoculated with: 1) wild-type Pseudomonas putida G7 strain
under variable molar concentration of buffer medium, and 2) different strains of
Pseudomonas putida G7 (i.e. wild-type, non-motile mutant, and non-adhesive mutant),
possessing variable attachment capabilities, under constant molar concentration of buffer
medium. Three orders of decrease in molar concentration of buffer medium did not affect
rate and extent of sorbed phase naphthalene degradation by P. putida G7. Important
attachment phenotypes (e.g., expo-polysaccharides, flagella) of PpG7-WT were also
observed to have played no significant role in enhanced bioavailability of sorbed phase
naphthalene. The wild type P.putida G7 and the mutant strains were observed to have
equally accessed non-desorbable fraction of naphthalene and degraded the fraction at an
equal rate and extent. The results of this study suggest that factors promoting bacterial

attachment ( i.e. removal of exo—polysaccharides, presence of flagella and reduction of

122

buffer medium concentration) do not contribute to biodegradation of non-desorbable

fraction of soil-sorbed naphthalene in soil-slurry system.

INTRODUCTION
Bacterial Attachment and Enhanced bioavailability

Recently, several researchers observed enhanced bioavailability of soil-sorbed
PAHs (Calvillo and Alexander 1996; Guerin and Boyd 1992; Park et a1. 2001; Tang et a1.
1998). Their studies suggested that biodegradation of sorbed PAHs is not exclusively
dependent on abiotic mass transfer of the compound into the water phase. Enhancement
in observed bioavailablity was attributed either to the direct bacterial access to sorbed
molecules or increase in abiotic mass transfer to the solution phase due to specific
physiological traits of bacteria (e. g. bio-surfactant production, attachment etc). It has
been demonstrated that different bacteria degrade sorbed PAHs at different rates, due to
difference in cell physiological trait (Guerin and Boyd 1992). This finding was
subsequently supported by several other studies(Tang et al. 1998, Calvillo and Alexander
1996).

Several studies have hypothesized the role of bacterial attachment in enhancing
sorbed phase contaminant (Calvillo and Alexander 1996; Guerin and Boyd 1997; Harms
and Zehnder 1995). In soils, the degradation of a non-desorbing compounds have been
found to be essentially carried out by particle-bound bacteria only (Kirchman 1990; Kuhn
1987). (Fletcher 1986) observed that cells attached to solid surface utilized glucose at
faster rate (by a factor of 5 or more) than un-attached cells. The author also examined

role of type of solid surface (polythelene, glass, or polyvinylidene fluoride) on activities

123

of attached cells, and found that composition of solid surface had no affect on utilization
of substrate. It was hypothesized that increase utilization of attached cells was either due
to concentration of nutrient at surfaces due to adsorption or physiological modification of
surface associated cells.

Guerin and Boyd (1992, 1997) used two bacterial species (Pseudomonas putida
strain 17484 and Alcaligenes sp. Strain NP-Alk) to demonstrate different capabilities of
bacterial species toward sorbed phase naphthalene in soil water system. Strain 17484
could degrade sorbed phase at the rate greater than de-sorption rate in all natural sorbents;
strain NP-Alk, on other hand, was able to degrade aqueous phase naphthalene, only. The
authors did not offer mechanism for the differential capabilities; however, they
considered surface chemistry and geometry of cells as one of the major factor for the
difference. The strain 17484 was shown to be more hydrophobic than strain NP-Alk, and
consequently possess greater propensity to attach to soil particles. The attachment
capability of strain 17484 was speculated to cause enhance degradation due to 1)
exposure of cells to a steeper intra-sorbent concentration and, 2) cells acts as competitive
sorption media to a soil-sorbed naphthalene.

Harms and Zehnder (1995) also studied role of bacterial properties in enhancing
rate of degradation of 3-Chrlorodibenzofuran off the teflon granules. They observed that
degradation rate exceeded de-sorption rate in a system containing bacteria, and that the
degradation rate increased with increasing amount of cell attachment to the sorbent. The
enhanced de-sorption rate was proposed to be due to specific affinity of the degrading
bacteria and the tendency of the organisms to attach to the sorbent. The specific affinity

depends on a type of bacteria, and is a measure of efficiency with which a bacterium

124

reduces the concentration of a substrate. The specific affinity increase as cells approaches
closer to sorbent surface. As the cells attached to the sorbent surfaces, they exposed to the
steeper concentration gradient, and consequently increase degradation rate.

Tang et a1. (1998) also observed enhanced biodegradation rate of sorbed phase
phenantherene as compared to abiotic desorption rate in a system containing poly-acrylic
beads or sediments. The enhanced biodegradation was, however, observed, with the
bacterium (PS-2) being obtained by enrichment on sorbed phenanthrene; and the
bacterium that was obtained on non-sorbed phenanthrene could not able to cause enhance
biodegradation rate. The authors hypothesized that direct utilization of sorbed chemicals
was due to attachment of microorganism to solid surfaces.

Wick et al. (2002) investigated physiological response of Mycobacterium sp.
LB501T to poorly water-soluble anthracene (which serves as a sole carbon source)
crystal in a batch culture. To overcome low dissolution flux of anthracene,
Mycrobacterium had attached to the solid crystal via extra-cellular polymeric substances -
(EPS). This indirect attachment of cell favors diffusive mass transfer and, consequently,
dissolution of crystal. To further optimize attachment and consequently anthracene
uptake, the Mycobacterium changes its cells-surface properties. Mycobacterium and
related genera excrete mycolic acids in their cell wall, and these molecules are believed

to stimulate attachment to hydrophobic surfaces.

Factor Affecting Bacterial Attachments

The mechanism of bacterial attachment to inanimate materials is complex and

unlikely to be explained by one or other property of the cell surface (Bos et al. 1999).

125

There are number of factors that contribute to the attachment of bacteria to the surface
that include cell surface charge, bacterial hydrophobicity, the presence of extra-cellular
polymers and, flagella. Besides cell surface properties, bacterial adhesion is also
influenced by chemical composition of solid surface and liquid medium (Bos et a1. 1999).

Numerous studies interpreted bacterial adhesion in terms of hydrophobicity or
surface free charge (Stenstrom 1989; van Loosdrecht et al. 1989). Both parameters
influence cell adhesion to solid particles; however, the relative influence of high cell
hydrophobicity is dominant as compared to surface charge on bacterial adhesion.
(Mueller et al. 1992) found that bacteria with low hydrophobicity ( P. flouresences)
attached at far lower rate (five times) on to a glass surface than bacterium with high cell
hydrophobicity (P. aeruginosa). Bestiaens et al. (2000) isolated adherent PAH-degrading
bacteria by enrichment technique using hydrophobic membranes containing sorbed
PAHs. The isolated bacterium was shown to be very hydrophobic and strongly adherent
to different surfaces.

Microbial surface charge can also influence adhesion, but to a smaller degree as
compared to cell hydrophibicity. For example, the contribution of surface charge to
adhesion increases when it is chemically modified from negative to positive charge
(Klotz et a1. 1985). Surface charge also becomes more influential when bacterial cells
displayed low degree of hydrophobicity (i.e., more hydrophilic), which allows the
bacteria to adhere in the so-called secondary minimum.

Besides hydrophobicity and surface charge, surface components of Gram-
negative bacteria such as lipo-polysaccharides (LPS) on the outer bacterial membrane,

and more loosely associated layer called extra-cellular polymeric substances (EPS) have

126

also been shown to play important role in cell adhesions(Pring1e et al. 1983; Rijnaarts et
al. 1995; Williams and Fletcher 1996). The characteristics of these surface structures are
still not well-known, and therefore their affects on bacterial attachment yields
controversial results.

Razato (2000) used wild type strain (with longer LPS chain) and its mutant (with
truncated LPS chain) to compare their interaction with glass surface. The author observed
that while wild type was able to adhere to the glass surface, the mutant failed to adhere.
They hypothesized that as mutant lost its core polysaccharide component of LPS, the
exposed negatively charged KDO molecules experienced repulsive interaction with
negatively charge glass surface. The Wild type, however, experienced attractive
interaction due to shielding of negative charge KDO by longer LPS molecules.

DeFlaun and Fletcher (1999) also concluded role of cell surface polymer in
adhesion of B. cepacia G4 to sand surfaces. The authors isolated adhesion-deficient
mutant of Burkholderia cepacia G4 in sand column assay and compared its adhesion
phenotypes with wild type strain. The mutant was reported to have lost adhesion
phenotype due to attenuated 0 antigen on the LPS. The loss of 0 antigen part of LPS
caused reduced shielding of charged group near the membrane (such as phosphate group
in the core-lipid A region) and, consequently there was a net repulsive interaction
between mutant and negatively charged sand surfaces.

In some studies the truncation of LPS, however, resulted in a more hydrophobic
and, hence more adherent mutant than wild type (Williams and Fletcher 1996). The
authors generated and isolated mutants of Pseudomonasfluorescens that were altered in

adhesion ability. The mutants were demonstrated to have increased adhesion to a variety

127

of surfaces (quartz sand, hydrophobic polystrene ) than wild type strain. The mutants
showed increased attachment to hydro-phobic surfaces and decreased attachment to the
more hydrophilic sub-straturm. The authors attributed enhanced adhesion property of
mutant due to attenuation of 0 antigen on the LPS. The lack of 0 antigen exposed lipid

moiety and hence rendered the cell surface more hydrophobic.

Objective of the Study
The objective of this study was to investigate the role of bacterial attachment in
degradation of soil-sorbed naphthalene. Previous studies did nOt attempt to determine
mechanism by which bacteria attached with soil particles for accessing the soil-sorbed
contaminant molecules. In this study, a systematic approach was taken to examine factor
affecting bacterial attachment and consequently rate and extent of naphthalene
degradation in a batch soil-slurry system. The factors studied included molar
concentration of buffer medium and the bacterial cell surface structure. Following
hypotheses were tested for observed enhanced bioavailability of sorbed phase
naphthalene in a soil-slurry system inoculated with P. putida G7:
1. Attachment of the P.putida G7 cells with soil surfaces increases as the
molar concentration of buffer medium increases
2. Wild type P. putida G7 adhere to soil particle more efficiently than mutant
strains deficient in expo-polysaccharide layer (EPS) and flagellated structures
Bio-availability experiments were carried out in a stirred batch soil-slurry system
inoculated with: l) wild-type Pseudomonas putida G7 strain under variable molar

concentration of buffer medium, and 2) different strains of Pseudomonas putida G7 (i.e.

128

wild-type, non-motile mutant, and non-adhesive mutant), possessing variable attachment
capabilities, under constant molar concentration of buffer medium. Increase in molar
concentration of buffer medium has been reported to decrease electrostatic repulsion
between bacteria and surfaces and, hence increase in bacterial attachment (Rutter and
Vincent 1984). Several researches have also demonstrated significance of EPS layer on
adhesion of bacterial cells to solid. It was expected that mutant strain would have limited
excess to sorbed phase naphthalene, due to its inability to attach with surface, as

compared to wild type.

MATERIAL AND METHODS

(A
9.

The three soils selected for this are CapacA, Colwood, and muck with low,
intermediate and high soil organic content. The soil was prepared as described elsewhere
(Park et a1. 2001), and their characteristics are summarized in Table 5-1.

Chemical and Media

Both 14C-labeled and unlabeled naphthalene were used in this study. The 14C-
labelled naphthalene (sigma, 0.1 mCi, >98% pure) stock solution was prepared by
supplementing it with unlabelled naphthalene and was used for sorption, desorption,
serial dilution studies. The un-labelled naphthalene (sigma, > 98%) stock solutions were
prepared in methanol and stored in 5 ml screw cap vials (capped with Teflon septa) at
4°C before use. All biological assays were carried out with unlabelled naphthalene stock

solutions.

129

Table 5-1. Selected Properties of Soils used in the Study

 

 

Soil % O.C. % Sand % Silt % Clay PH CEC
[cmol(+)lkg]

Capac 3.3 55 24 21 6.8 24

Colwood 7.8 64 21 15 6.0 43

Houghton Muck 38.3 ND ND ND 5.1 156

 

ND = not determined

130

Microbiological media were prepared as described previously. The experiments were
carried out with chemotaxis buffer (pH =6.8) as described elsewhere (Harwood et al.)
5%

Wild-type Pseudomonas putida G7 along with its two mutant strains were used in
this study. Wild-type PpG7 (ATCC 17485) and a mutant strain [Pseudomonas putida
G7.C1(pHG100)], which is non-chemotactic to naphthalene, were obtained from Dr.
Caroline Harwood (University of Iowa). The non-flagellated (non-motile) mutant strain
of P. putida G7 was obtained from Dr. Michael Aitken (University of North Carolina);
the non-motile strain was spontaneously generated and isolated as described elsewhere
(Marx and Aitken 2000). The pure wild type, Pseudomonas putida G7 , has been reported
previously to uptake sorbed phase naphthalene in soil slurry. Furthermore, it has been
characterized as being motile, and chemotactic (Grimm and Harwood 1997; Grimm and
Harwood 1999).

Sogption/Desorption/Serial Dilution Experiments

Sorption, desorption, and serial dilution desorption experiments were carried out
as described previously (park 2001). Briefly, all experiments were canied out with initial
equilibrium period of two days. The ratio of sorbent to liquid was kept the same in all the
experiments: 1:15 for Capac-A soil, 1:65 for Colwood soil, and 1:240 for muck. Sorption
experiment was carried out at initial liquid-phase naphthalene concentrations, ranging
from 0-2500 ug/L in triplicate for each type of soil. The desorption and serial dilution
were both carried out at an initial naphthalene concentration of 2 mg/L. Desorption and
serial dilution experiments were initiated at the end of 2 days sorption equilibrium period

by adding naphthalene free soil extract to make up the decanted volume. In all these

131

experiments, vials were tumbled at 9 rpm in the dark. After mixing, each vial was
centrifuged and supernatant sampled. The remaining supernatant was then decanted and
replaced with methanol to extract soil-sorbed naphthalene. The concentration of
naphthalene was then determined by liquid scintillation counting (LSC) and verified by
high-pressure liquid chromatography (HPLC) for some samples.
Bioavailability and Biodegradation Rate Experiments

Bioavailability assay was adopted from the method described by Park et al.
(2001). Briefly, soil extract controls and soil slurries were set up in S-mL vials sealed by
screw caps with Teflon-lined septa. The soil to water ratio was kept same as used for
sorption/desorption experiments. Unlabelled naphthalene was injected in all vials at
initial liquid phase concentration of ~ 2000 ug/L. After 2 days of initial equilibrium
period on tumbler, soil slurries and soil-free extracts ( to determine degradation rate in
liquid only) were inoculated with cells harvested at early stationary phase. At precise
time intervals, selected vials were centrifuged and l-ml of supernatant was transferred to
a HPLC vial containing 0.01 ml of 10 N NaOH to stop biodegradation of naphthalene.
The soil-sorbed naphthalene was then extracted by methanol after removing left over
supernatant. Appropriate control vials were included, with no bacteria, to check for any
abiotic losses.
Chemical Analysis

Naphthalene was analyzed by HPLC using a C13 reverse-phase column, with 80%
acetonitrile/20% water mobile phase, and fluorescence detection (280 nm excitation, 340
nm emission). Radioactivity was determined by LSC. The analytical detection limits of

naphthalene solution were 5 mg/L and less than 1 mg/L for HPLC and LSC, respectively.

132

Isolation of Adhesion Deficient Mutant.

An adhesion-deficient mutant was generated by transposon mutagenesis using a
standard procedure. Escherichia coli strain Sl7-l (pUT) was used as the donor strain and
Pseduomonas putida G7 serves as recipient strain. The donor and recipient strains were
grown in Luria-Bertani broth (LB) to late log phase and then mixed at an appropriate
ratio. The mixture was syringed out on a filter paper so that donor E. coli mate with P.
putida G7. Filters were incubated for several hours generating approx. 7000 mutants per
mating. The filters were then transferred to LB broth in 50 ml tubes for incubation at 30
°C for 30 min. After incubation the tubes were vortexed, and the bacterial suspension was
plated on a minimal medium supplemented with Kenamycin (50 [lg/mL) and
naphthalene. The resulting transposon mutants that grew on the selective plates were
then screened for to identify mutant defective in the production of exo-polysaccharides.
Screen for E_PS deficient mutagls

Calcoflour white was used to identify the defective mutants. Mutants were
streaked on LB plates containing 75 ug/mL of Calcoflour; After 48 hrs of incubation, the
florescence of the cells was observed under long wave UV light source and compared
with the wild type strain. Calcoflour binding by wild type cells was clearly visible, while
no Calcoflour binding was detected by Exo' mutant strains. The adhesion deficiency was
further verified by the soil adhesion assays as described elsewhere (Williams and Fletcher
1996).

Screen for adhesion-deficient mutants
Several potential mutants, as isolated by Calcoflour screening process, were first

grown in LB to the logarithmic phase. Equal amount of each mutant was then added in

133

CB solution at a combined titer of approx. 1x108 CFU/mL. An aliquot of 3 ml cell
suspension was then carefully placed on top of 20-mL syringe barrel filled with 12 grams
of sterilized Ottawa sand. The cells were allowed to interact with the column material for
1 hr, and then column was washed with 3 ml of CB solution. The first 3 mL was collected
at the bottom, containing non—adherent cells, and passed it over a second column and
subsequently third column to obtain adhesion deficient Tn5 insertion mutants.
Cell surface characterization

Hydrophobicity of bacterial cells was determined by measuring bacterial adhesion
to hexadecane (Rosenberg et al. 1980). Four milliliter of washed cells (at initial OD4oo
=1) was mixed with 1 milliliter of hexadecane in 10 ml glass test-tube and vortexed for 5
minutes. The mixture was allowed to stand for 30 minutes to separate phases. The
aqueous phase, at the bottom of the tube, was then sampled for measuring final OD4oo
with spectrophotometer. The percentage of adherence to hexadecane was then calculated

by following equation-1:

(Initial OD4oo — final OD4oo)/(Initial OD5oo) x 100 [1]

Adhesion experiment

The cells were grown in MSM with 200 mg/L naphthalene that also included
ring-U-[l4C] naphthalene (sigma, > 96% pure) at 400000 dpm/mL. After 48 hrs of
incubation, cells were harvested in early stationary phase, centrifuged and washed three
times with CB solution. The final inoculum was re-suspended in CB solution before

adhesion assay. The cell density was determined by standard plate count, and the

134

corresponding initial radioactivity was determined by adding inoculums (1 mL) to 10 mL
of scintillation fluid and activity was counted on Packard 1500 Tri-Carb Liquid
Scintillation Analyzer for 5 minutes.

The attachment assay consists of two sets of 5 ml vials; one containing sterile soil.
suspensions and other soil extract. The soil-water ratio and other conditions were similar
to desorption and bioavailability experiments. At the end of 2 days equilibrium period,
each vial was inoculated with approx 1 x107 CFU/mL. The tubes were then mixed at
rotating mixer, and at predetermined time interval, the tubes were centrifuged at 500 rpm
for 3 min. The radioactivity in the supernatant and soil-free control was determined by
taking 1 mL sample into 10 mL scintillation liquid for activity count. The % age of cell

retained in soil was calculated by the following equation-2

(DPM / ml) in soil free control - (DPM / ml) in supernatant
(DPM/ml) in soil free control

 

[2]

DATA ANALYSIS: DESORPTION-BIOVAIABILITY MODEL (DBM)
Coupled biodegradation-desorption model that incorporate degradation of non-
desorbable fraction was used to track total as well as liquid phase and sorbed phase
naphthalene in batch slurry reactor. The overall degradation of naphthalene in a batch

slurry reactor using mass balance on naphthalene can be represented by equation-3.

dC dS
—(Vl'E—+M'E)=(Rbi0'v +M'knd ~Snd)

[3]

135

Where m (kg) is the total mass of the soil, C (pg/L) represent naphthalene

concentration in solution phase, S (pg/Kg) represent total sorbed-phase naphthalene,

V(L) represents volume of the liquid, and knd is the first-order biodegradation co-

efficient for non- desorbable fraction (Snd) of total sorbed contaminant (S). Rbio is the

liquid-phase biodegradation rate expression, which can be expressed by following

equations:

- C
Rbio = m— for Michaelis Menten kinetics, [4]
m
or
Rbio = K1' C for first-order kinetics [5]

Where, Vmax is the maximum flux that a cell can get and, Km is the Michael-

Menten constant, and K1 first-order degradation rate co-efficient. The non-desorbable

fraction is tracked by following mass balance equation

dSnd

dt = *kndSnd [6]

The total sorbed phase (S) was assumed to consist of equilibrium (Seq), non-

equilibrium (Sneq) and non-desorbable (Snd) sites. The equilibrium site represents

fraction that release spontaneously during desorption experiment and is described by a

136

linear partitioning model. The non-equilibrium site fraction allows slower release of
contaminant and can be represented by first order expression. The non-desorption site is
defined by Park et al. (2001) to be containing contaminant that cannot be released to
aqueous solution during the experimental desorption period. Mathematically, these three-

sorbed phase fractions are described as follow:

Seq = eq KdC 171
Snd = fnd KdCe [8]
dSneq/dt = 0L ( fneq Kd C - Sneq) [9]

Where, Kd is the sorption co-efficient, Ce is the liquid phase concentration in

sorption equilibrium, t is desorption time in minutes, or is the first-order desorption rate

coefficient (per minute). The three remaining terms feq, fneq, and fnd represent,

respectively, equilibrium site fraction, non-equilibrium site fraction, and non-desorption

site fraction. The three site fractions were measured from desorption experiment: fnd

corresponds to the plateau of the desorption rate profile; and f , fm:q and or were

estimated by non-linear regression analysis of desorption data with the constraint that:

feq+fneq+ fnd =1 [10]

137

RESULTS
Sorption, Desorption, and Serial dilution Experiments

Sorption of naphthalene was found to be linear for the test soils (Figure 5-1). The

sorption coefficients (Kd =slope of isotherms) ranged from 15 to 240 mL/g (Table 5-2).

All the test soils observed non-desorbable (Sud) fractions as determined by serial dilution

method. The non-deosorbable fraction is calculated from the intercept of the step
desorption isotherm (for total six steps) and the initial sorbed-phase concentration (Table
5-2). The calculated values were further confirmed by extracting from soils non-
desorbable naphthalene at the end of last step desorption. Desorption of naphthalene from

all soils were essentially complete within 15 hours (Figure 5-2). The desorbable fraction

(fd) and desorption rate coefficient (or) were determined by fitting non-linear regression

equation to desorption rate data (Table 5-3). In all the experiments, the overall
naphthalene mass recovery was found to be 94-97%, indicating no degradation.
Biodegradation of naphthalene in soil extract for all strains was well described by

Michaelis-Menten kinetics . The estimated maximum specific utilization rate co-efficient
(Vmax) and the half-saturation constant (Km) were 500 ug/L-hr and 301 ug/L,
respectively. Since dissolved soil organic can potentially affect naphthalene degradation
kinetics, soil extract instead of CB solution was used to determine biodegradation rate of

naphthalene. Furthermore, the rate of naphthalene degradation was observed to be lower

in soil extracts than in chemotaxis buffer solution (CB).

138

Table 5-2. Parameters for sorption of naphthalene by the A horizon of selected soils

 

N on-desorbable

 

 

 

 

 

 

 

 

 

     

 

 

 

 

Soil K., R2 K.,c
fraction, fnd
CapacA 15 0.96 457 0.21 (6x103)
Colwood 68 0.98 872 0.19 (2x103)
Muck 240 0.997 626 0.18 (1x10'2)
3.5e+5
% 3.0e+5 -
3
d 2.Se+5 ~
C
s 3 29...
O O
2 2.09% ‘ v Capac-A
.92
a:
5 1.59+5 -
.C
D.
(U
C 1.0e+5 .
'O
a:
g 5.0e+4 -
(.0 L.
0.0 7 e . 1 l l
0 500 1 000 1500 2000

Aqueous naphthalene conc. (ug/L)

Figure 5-1. Sorption isotherm of naphthalene in natural sorbents

139

Table 5-3. Desorption parametersa evaluated by a three-site desorption moder

 

 

Soil re fne fn a R2
(min-1)

Capac 0.74(0.042) 0.1 0.16(0.009) 0.0025(0.0008) 0.94

Colwood

0.62(0.023) 0.23 0.15(0.010) 0.0016(0.0004) 0.94

Houghton Muck 0.58(0.034) 0.27 0.15(0.005) 0.005(0.0017) 0.97

 

afeqfne» fnd: equilibrium-, non-equilibrium-, non-desorption-site fractions; a: first
order desorption rate coefficient for non-equilibrium sites.

1Numbers in parentheses are standard deviations of the evaluated values.

 

 

   

 

 

 

 

 

 

100
80
°\° o Colwood
‘E’ 60 o Muck
.9 V CapacA
'5. Regression-Colwood
5 —— Regression-Muck
8 40 ~ ——— Regression-CapacA
D
20 -
0 T I I T I

 

 

0 1000 2000 3000 4000 5000 6000
Desorption time (min)

Figure 5-2. Desorption of naphthalene from three natural soils and activated carbon (GAC)

140

Cell Surface Properties

The bacterial surface hydro-phobicity was determined by the BATH assay

(Figure5-3). The cells were grown on naphthalene then harvested, washed, and re-

suspended in CB solution at the early stationary phase (40 hrs). All strains including WT

 

 

 

 

 

 

 

 

 

 

 

 

16-
14- 1-
12- L
10-
8‘, a- I
8= ,_
rt .
1
08
ex 2‘ I
«2 0 . r r .
o\° «a ,
2 310‘ a“
e" 1 . 1'
Q9 We We" 990
Strain

Figure 5-3. Percent of bacteria adsorbed to hexadacane in BATH test

showed less than 15% reduction in cell suspension optical density after mixing with

hexadecane, indicating very low hydrophobicity (Bos et al. 1999). All strains can,

therefore, be classified as hydrophilic because they remained in the aqueous phase and

did no adhere to the hexadecane during BATH test. Alternatively, bacterial surface

hydrophobicity and charge can be determined by measuring water content angle and cell

zeta potential, respectively. The measured values of these parameters for P. putida G7 are

given in Table 5-4. Water contact angle (0) value for P. putida G7 lies between 30 and 35

indicating hydrophilic surface; bacterial surface with 0 value greater than 70 is

considered hydrophobic (Bastianens May 2000, AE&M). The value of zeta potential for

141

P. putida G-7 indicates that it surface is negatively charged at neutral pH. In summary, P.
putida G7 is hydrophilic bacteria carry negative charge, and it is expected that its
attachment properties will vary with ionic strength of chemotaxis buffer.

Table 54 Cell surface and hydrophobicity of P. putida G7 grown on different
carbon sources

 

 

 

Strain (carbon Cell surface Cell Cell zeta
source on which it hydrophobicity electrophoretic potential
was pre-grown) (0w) mobility (U) (C)
PpG7(glucose) 31¢ 5 246: 0.39 -32.0i5.2
PpG7(naphtha1ene) 29i2 -2.52i0.49 -32.7i~6. l

 

Bioavailability Experiments at Different Molar concentration of Buffer medium
Bioavailability and biodegradation rate experiments were carried out at different
molar concentration (100 mM, 10 mM and 1 mM) of chemotactic buffer. Bio-availability
experiments were carried out with Colwood soil and wild type P. putida G7, only. All the
vials containing soil slurry and soil-extract were inoculated with bacterial suspension of 5
x 106 CFU/ml. The lowest Ionic strength (0.1 mM) of suspending solution did not affect
degradation capability of bacteria. We found that the initial rate of naphthalene
degradation and kinetic parameters were similar in all tested CB molar concentration.
Data from the bioavailability assays, carried out at different ionic strength in
Colwood soil, are shown in Figures 5-4 (a-c). The depletion of liquid phase naphthalene
was initiated first, followed by desorbable fraction of sorbed phase, and at last non-
desorbale fraction of naphthalene. In all cases, WT strain could able to degrade sorbed
phase naphthalene to a level below the non-desorbable level, as shown by dashed line.
When bacteria have been added and start to degrade the naphthalene the concentration of

dissolved naphthalene decreases. As bacteria continue to degrade naphthalene in liquid

142

phase, more and more naphthalene transferred from solid phase to liquid phase. Thus,
mass transfer from solid to liquid phase reduces as solid phase naphthalene decreases.
Within first four hours both liquid phase and desorbable fraction of sorbed naphthalene
decreased to zero, indicating mass transfer limited condition. The rate of degradation of
naphthalene is relatively faster during this period, as indicated by steeper and shorter part
of the data. About 85% of the total naphthalene was degraded during this step and is
mostly carried out by the bacteria suspend in liquid. At the end of the fist phase of
degradation, a mass transfer limited degradation phase begun (represented by long and
nearly horizontal part of the data). This phase is marked with non-desorbable fraction of
naphthalene, which is mainly degraded by attached bacteria (micro-graph showing
attached bacteria with soil)

The bio-availability data, which included both liquid phase and sorbed phase data,
were fitted with the coupled desorption-biodegradation model (equation-3). The DBM
model provided a reasonable description of both liquid and sorbed phase data for all three
cases over the entire experimental period. The DBM model estimated degradation rate

co-efficient (knd) for non-desorbable fraction (Table 5-5).

143

Table 5-5. Biokinetic parameters evaluated for Psudomonas putida G7-WT under

varying molar concentration of chemotactic buffer solution (CB)

 

 

Molar First order degradation rate for
Soil concentration non-desorbable fraction
(mM) K... (11")
100 0.0585 i 0.0077
Colwood 1 0.0318 i 0.003
0.1 0.036 t 0.002

 

 

Colwood soil with PpG7-WT in 100 mM CB

       
  
  

 

 

 

0 Total Naphthalene
g 9* Solid phase
a 8 ' O llqu1d phase
3
‘3 Total naph.
(U
S Solid phase Naph.
C
% Liquid phase Naph.
E.
.C
D.
(B
2

Time (hours)

Figure 5-4a. Naphthalene degradation by Pp-G7(WT) in 100 mM CB solution

Colwood soil with PpG7-WT in 1.0 mM CB
10 0 Total Naphthalene

 

      
 

 

 

 

  
    

9 4% Bolid phase
a 8 I llqwd phase
3
‘3 7 Total Naph.
(U
E Solid phase Naph.
C
33 Li uid phase Naph.
(U
.C
.0—0
.5
Q.
{U
2

Time (hours)

Figure 5-4b. Naphthalene degradation by Pp-G7(WT) in 1.0 mM CB solution

' Colwood soil with PpG7-WT in 0.1 mM CB
10)

 

 

 

 

{> Total Naphthalene
9 _ elk Solid phase
0 liquid phase
3<> T
otal naph.

Solid phase naph.

 

Naphthalene mass (ug)

Time (hours)

Figure 5-4c. Naphthalene degradation by Pp-G7(WT) in 0.1 mM CB solution

145

Isolation of Adhesion Deficient Mutants
To delineate possible role of Extra-cellular polymeric substances (EPS) on
attachment, mutant was generated and isolated that had decreased attachment ability. The

decrease in observed attachment ability is expected to be due to altered EPS layer.

Mutagenesis

Out of approx. 900 transposon mutants (Km'Rif’) streaked on plate, containing 75
uglml Calcoflour, only 46 mutants did not florescence under long wave UV light. This
showed that these mutants lost it ability to produce EPS. These selected Exo' mutants
(combined titer of ~1x108 CFU/ml) were then passed through sand columns three times.
The final count of mutant collected at the bottom of third pass was determined as ~ 4x107
CFU/ml. The bacterial suspension obtained at the end of third pass was taken as culture
with low adhesion ability. The culture was then purified and preserved in cryo-vials at —
80 °C before use. The mutants deficient in producing EPS are hereafter denoted as Exo'.
Batchadhesionjmy

Tests for attachment to the test soil (colwood) were performed over a period of 10
hours. The initial cell concentration for each strain was ca 1x107 CFU/ml. The percentage
of bacteria retained in soil varied from average 20, 15, 10, and 10% for non-chemotactic
strain, WT strain, non-motile strain, and Exo' strain. For WT and non-chemotactic
mutant strain, cell attachment remained almost constant over the period of time (Figure 5-

5). However, non-motile and Exo’ mutants showed a steady increase (from 3 to 22%) in

146

In G7
a Non-Chemotact
D Non-motile

 

: I Exo-m utant
o

in

c

o

B a

.E 2

S 2

e r.

.2 9-

L-

2

o

a

.n

°\o 576

Time (in minutes)

Figure 5-5. Cell attachment to Colwood soil at different time interval

attachment over a period of time. At the end of 10 hrs, the % retention for all strain

remained same at around 20%.

Bioavailability Experiments with Wild type and Mutant Strains

The bio-availability assays were repeated for WT P.putida G7, and three mutant
strains. The three mutant strains were non-motile, non-chemotatic, and mutant deficient
in producing EPS layer (Exo'). The cell density was kept low at 5x105 CFU/ml in all
experiments to avoid rapid degradation in liquid phase as was observed in experiment
with relatively higher cell density (1x 107 CFU/ml). Previous research has also
recommended use of low density cells for isolating role of bacterial phenotypes
(chemotaxis, attachment etc) of interest (Marx and Aitken 1999). Data from the
bioavailablity assays for four strains in colwood soil are shown in Figure 5-6. All strains
showed almost similar degradation profile in both liquid phase as well as solid phase. All

strains, including Exo' mutant, could able to degraded sorbed phase naphthalene to levels

147

below he non-desorbable level, as shown by dashed line. In contrast to earlier study with
high cell density, both liquid phase and desorbable portion were degraded very slowly
over an extended period of time (20hrs), because of low cell density. It appeared that
slower degradation rate in liquid phase had resulted in a prolong availability of the
desorbable fraction of sorbed phase naphthalene. After all desorbable naphthalene
degraded, the degradation of non-desorbable fraction started at same time for all the
strains (later part of the data). The non-desorbable fraction was then degraded very
slowly by all strains over a period of 120 hrs.

The bio-availability data were then fitted with the coupled desorption-
biodegradation model (DBM). The model provided a reasonable description of sorbed
phase data for all strains over the entire experimental period. The DBM model estimated
degradation rate coefficient of non-desorbable fraction (knd) and is shown in Table 5-6.
For all strains, the estimated knd values were determined to be statistically similar. Thus,
non-desorbable fraction was apparently degraded at same rate and to same extent,
indicating that all strains possess similar capabilities to degraded non-desorbable

naphthalene.

DISCUSSION
Recently, several researchers observed enhanced bioavailability of sorbed HOCs in
which rate of degradation by bacteria was faster than the rate of abiotic desorption. The
enhance bioavailability, however, requires existence of a mechanism that facilitates

degradation of sorbed chemical. Previous studies have suggested that the phenomenon of

148

Table 5-6. Biokinetic parameters evaluated for non-adhesion mutant along with

 

 

 

 

three other strains of PpG7.
Michaelis-Menten
parameters
Soil Strain Knd
Km Vs
<hr")
(mg/L-hr) (mg/L)
WT 0.025+0.001 90 50
Colwood Non-chemotactic 0.023+0.004 30 10
Non-motile 0.027+0.007 9O 50
Non-adhesion (Exo’) 0.0196+0.001 9O 50

 

Sorbed phase naphthalene(uglkg)

 

O

 

4t) 60 80
Time (hours)

8 .

 

 

Figure 5-6. Naphthalene degradation in soil-slurry-solution by PpG7- WT, and its
three mutants: non-chemotatic. non-motile. and non-adhesion

149

enhances bioavailability is likely dependent on the specific bacterial trait e. g., production
of bio-surfactant or bacterial attachment. It has recently been shown that neither synthetic
nor microbially produced surfactant aided in utilization of sorbed HOCs by bacteria
(Calvillo and Alexander 1996). However, numerous studies cited attachment as a
plausible explanation for enhanced bioavailability of soil-sorbed HOCs.

The previous studies attributed observed difference in rate of sorbed phase
degradation and corresponding abiotic desorption rate to bacterial attachment. However,
the role of bacterial attachment in direct uptake of soil—sorbed HOCs is still not clear. The
effect of attachment on degradation of HOCs has been widely studied yielding
controversial results. For example, Mukherji and Weber (1998) suggested that microbial
attachment could even hinder both mass transfer and biodegradation.

In this study an attempt has been made to examine the factors affecting bacterial
attachment and consequently enhance bio-degradation. The factors studied included
affect of molar concentration of buffer medium and EPS layer of Pseudomonas putida
G7. By varying these factors, we intended to isolate affect of attachment on
bioavailability of sorbed phase naphthalene. Initial studies on Pseudomonas putida G7
were conducted to determine characteristics of its surface properties. Evaluation of
BATH test and other cell surface properties (zeta potential and contact angle) indicated
that wild-type Pseudomonas putida G7 has very low hydro-phobicity with corresponding
low zeta potential. It was, therefore, expected that initial cell attachment would be
controlled by electrostatic interaction, the magnitude of which could be manipulated by

varying molar concentration of buffer medium at constant pH.

150

Bio-availability experiments carried out at varying molar concentration of CB
solution and constant pH values, however, did not yield any distinguishable difference.
Bacteria could equally access and degraded all fractions (i.e., liquid, desorbable, and non-
desorbable) of naphthalene under varying molar concentration of buffer medium. This
finding seems to rule out affect of electrostatic interaction on initial attachment of
bacteria. Both BATH test and soil attachment assay showed 15-20 % of cell attachment,
and this fraction of cells attachment could either due to hydrophobic site on cell surfaces
or electrostatic attraction between negative site on bacterial surface and localized positive
sites on soil clay. This fraction of attached cells were concluded to be unaffected by
electrostatic interaction as was observed from bioavailability experiments at different
ionic strength. The fraction of 15-20% attached bacteria were able to degrade non-
desorbable fraction of sorbed phase naphthalene.

The probable reason for this finding can be explored in term of complex nature of
soil. Previous research with bacterial attachment mostly dealt with clean sand or artificial
surfaces e. g., Teflon and glass. The interaction of bacteria with these uniform surfaces
cannot be reproduced with complex system like soil. Lahlou et al. (2000) demonstrated
that attachment results from glass and Teflon couldn’t be reproduced with soil materials.
For example, the author concluded that one strain showed significant attachment to teflon
and glass but did not show any affinity toward soil components. The soil used in this
experiment (i.e., Colwood) consists of a complex mixture of quartz (75%), clay (15%),
and organic (8%) matter carrying either negative or positive surface charge depending on
solution pH. Since the pH of buffer medium was kept near neutral (pH =65), all soil

components mostly carried negative charge. However, clay minerals may also exhibit

151

fixed positive charge resulting from breakage at the edges to expose trivalent or
tetravalent positive ions or from the exchange of OH groups (Stotzky 1985). Marshall
and coworker showed adsorption of bacterial cells to clay particles resulting from
positive charges on platelets clay edges and breakage at the edges(Marshall 1976).
Another plausible explanation is that decreasing ionic strength might have
reduced electrostatic repulsion between bacterial and negative site, but did not affect
neutral site or even positive site (of clay surfaces due to broken edges). In presence of
positive charges on clay surfaces, lowering of ionic strength even caused opposite effect
i.e., reduction in electrostatic repulsion between bacterial surfaces and clay positive
surface and hence greater attachment. Thus, it is highly probable that due to presence of
both negative and positive charge surfaces in soil, lowering of ionic strength might not
have effected overall attachment of bacteria. Furthermore, at neutral pH, P. putida G7 is
relatively more negatively charged than soil surfaces. Due to large difference in negative
charges between bacterial and particle surface charge, the change in Gibbs free energy of
adsorption due to electrostatic interaction is negligible, regardless of ionic strength.

After verifying that charge at bacterial surface did not affect bioavailability of
sorbed phase naphthalene, we turned toward the possibility of role of specific structure on
the outer membrane (e.g., EPS and LPS layer). Pseudomonas putida G7 is known to
possess both extra-cellular polymeric substances (EPS) as well as smooth and rough lipo-
polysaccaride (LPS)(Kachlany et a1. 2001). EPS may be secreted as a loose layer of
carbohydrates called slime layer or a more rigid layer of protein and is known as capsule.
Several researchers have demonstrated significance of EPS and LPS on adhesion of

bacteria to solid surfaces (Allison and Sutherland 1987; Pringle et al. 1983). The EPS and

152

LPS promote adhesion by adsorption to surfaces; polymers bridge the distance between
cells and the surfaces and may cause adhesion even when the cells do not experience
attraction forces (Figure5- 7).

By selecting for mutants with altered surface characteristics, we can obtain strains
that have significantly different adhesion properties than parent strain. Transposon
mutagensis was used to generate cells with altered surface characteristics. Screening on
Calcoflour plate is a convenient method for isolating non-EPS producing mutant,
especially when wild type strongly fluoresces under UV light. In case of Pseudomonas
putida, we observed a weak fluoresce signals that made it difficult to screen for non-EPS
producing mutants. In absence of alternative screening system, we picked colonies that
were observed being dull and opaque under UV light. Passing sequentially non-EPS
mutants through a sand column, however, allowed us to collect non-adhesive mutants.

Both, Bath test and batch adhesion test revealed that non-adhesive mutant had
lower retention with hexadance and live soil, respectively. The batch test however
indicated that non-adhesive mutant recovered it attachment capability with time. The
plausible explanation for this trend is that non-adhesive mutant re-developed its
attachment phenotype under starvation condition. This trend can be explained base on
earlier observations that bacteria employ different modes by which it can acquire surface
characteristics that influence its attachment and detachment during exogenous energy and
nutrient deprivation (Neu and Marshall 1990). In the beginning of the batch adhesion test
with non—adhesive mutant, substrate was readily available in dissolved as well as

desorbable form. However, as naphthalene depleted with time, bacteria somehow

153

acquired adhesion trait, by unknown mechanism, to get access to otherwise non—
desorbable , difficult to get substrate in the system.

The observation, as described in previous paragraph, can equally be applicable to
our bioavailability experiment, where non-adhesive strains could equally degrade non-
desorbable naphthalene. After depletion of readily available fraction, naphthalene was
present under mass transfer limited condition. This condition created an apparent
starvation condition that forced bacteria to make-up for its inability to attachment by
some other mechanism (e. g. increase in surface hydrophobicity etc). Our results also
substantiated previous research that observed that degradative ability of the adhesion-
deficient and wild type were similar under substrate limiting conditions (Deflaun and
Fletcher 1999).

Even though mutagensis is an invaluable tool in isolating bacterial surface
component, yet it success demands several confirmatory experiments. The confirmation
of successful mutagensis requires large alterations in adhesiveness, multiple replicate
adhesion assay, and statistical comparisons (Williams and Fletcher 1996) . Thus,
screening for correct mutants is very difficult and time-consuming. This exercise of
selecting non-adheisve mutant is even more difficult because attachment can employ
variety of surface polymer (including protein ligand, carbohydrates and lipid on the
bacterial surface) under different substratum chemistry and environmental conditions
(Williams and Fletcher 1996).

In conclusion, the results of this study suggest that factors promoting bacterial
attachment (i.e. removal of exo-polysaccharides, presence of flagella and reduction of

buffer medium concentration) do not contribute to biodegradation of non-desorbable

154

fraction of soil-sorbed naphthalene in soil—slurry system. Further research is clearly
needed in establishing this finding under a well defined system, since bacterial
attachment to soil matrix is complex and unlikely to be explained by one or other
property of the cell surface ( Bos et al. 1999). Different cells employ different polymers
in mediating attachment depending on different substratum chemistries and medium
composition. For example, some polymer (eg pilli and LPS) may mediate in attachment
to hydrophobic surface; whereas, other polymer (EPS) help bacteria to attach with
hydrophilic surfaces. Certain combinations of bacterial, solid surface and liquid medium

properties could result in large deviations from the null hypotheses.

155

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‘adhesion of freshwater bacteria." Journal of general Microbiology, 133, 1319-1327.

Bestiaens, L., Springael, D., Waittiau, P., and Diels, L. (2000). "Isolation of adherent
polycyclic aromatic hydrocarbon (PAH)— degrading bacteria using PAH-sorbing
carriers." Applied and Environmental Microbiology, 66, 1834-1843.

Bos, R., van der Mei, H., and Busscher, H. J. (1999). "Physico-chemistry of initial
microbial adhesive interactions - its mechanisms and methods for study." FEMS
Microbiology Reviews, 23, 179-230.

Calvillo, Y. M., and Alexander, M. (1996). "Mechanism of microbial utilization of
biphenyl sorbed to polyacrylic beads." Appl. Microbial. Biatechnal., 45(3), 383-390.

DeFlaun, M. F., and Fletcher, M. (1999). "Alterations in Adhesion, Transport, and
Membrane Characteristics in an Adhesion-Deficient Pseudomonad." Applied and
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CHAPTER 6. SUMMARY AND CONCLUSIONS

SUMMARY

The phenomenon of the enhanced bioavailability of soil—sorbed organic molecules
is rapidly gaining attention. It is defined as the availability of organic compounds for
bacterial uptake at a rate greater than a-biotic de-sorption rate from the sorbent. The aim
of this study was to investigate the role of specific physiological traits of bacteria in the
enhanced bioavailability of soil—sorbed organic contaminants. The specific bacterial traits
considered included: production of biomaterial, variation in the attachment to solid
surfaces, and chemotaxis/motility. To accomplish the aim, three main objectives were
pursued in sequence. All of the identified traits were tested in a systematic way in stirred
and unstirred soil slurry reactors using a representative PAH, naphthalene. Wild type
naphthalene degrading bacteria along with its mutant cultures (non-chemotectic, non-
flagellated, and non-adhesive) were used for studying the role of these traits. The pure
wild type, Pseudomonas putida G7, has been reported previously to uptake sorbed phase
naphthalene directly in the soil slurry. Furthermore, it has been characterized as being a
motile and chemotectic strain.

The research work involved generating, isolating, and characterizing mutants with
one of the missing phenotypes (motility, chemotaxi, adhesion). Sorption isotherms and
single and series dilution desorption experiments were conducted to evaluate the
distribution coefficient, de-sorption rate, and amount of non-desorbable naphthalene. The
initial degradation rate was measured for wild type as well as for all mutants in the soil

extract solution. All the studies were carried out assuming that initial degradation kinetics

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of all mutants remained the same as that of wild type strain. Bioavailability experiments
involved first establishing sorption equilibrium, inoculating the systems with bacteria,
and measuring naphthalene concentrations in both sorbed and dissolved phases over the
period of time. The rate of naphthalene degradation by each culture in stirred batch
systems was fitted with a non-linear regression model (DBM model) to determine the
sorbed phase degradation rate coefficients. Bioavailability data were compared for each
type of organism for the extent and rate of sorbed phases degradation to evaluate the role

of each phenotype.

The first objective was to observe phenomenon of enhanced bio-availability by
comparing the removal of naphthalene in a soil-slurry system using bacterial strain P.
putida G7, Tenax beads extraction, and serial extraction with the 0.1 M buffer solution
alone. Also explored under this objective was the affect of excreted biomaterial on
enhanced bioavailability. Comparison of the results of the Tenax beads extraction and
serial dilution with PpG7 biodegradation showed that the rate and extent of naphthalene
removal by bio-degradation was at all times significantly greater than the abiotic removal
rate and extent in the sorbed phase. It seemed that naphthalene did not have to be
desorbed in the aqueous phase for bio-degradation by PpG7. Apparently, PpG7 were able
to overcome the mass-transfer limitation by an unknown mechanism and could able to
directly uptake sorbed phase naphthalene. The activity of the extra-cellular enzyme was
ruled out as no significant naphthalene loss was observed with bacterial the filtrate during
3 days of the incubation period. Also, the possible effect of extra—cellular polymer on

sorption/deosprion parameters was minimal. Since all bioavailability experiments were

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carried out at 2 mg/L of naphthalene, the resulting biomaterial production was expected
to be low enough to have caused any affect on partition co-efficient (and consequently
mass transfer) of sorbed naphthalene.

For the second objective, we evaluated the role of chemotaxis/motility on
naphthalene degradation in sorbent-slurry system under stirred and quiescent conditions.
In the stirred system, containing natural soils, the mixing had completely masked the role
of chemotaxis and motility due to the presence of an isotropic concentration and mixing-
mediated adhesion of cells to sorbent particles. However, in the stirred system with GAC
the role of chemotaxis/motility became apparent at a longer incubation time. The
presence of a large pool of non-labile naphthalene, which is released over an extended
period of time, attracted more cells at the activated carbon surface than what could be
accumulated due to mixing alone. The increasing number of cells accumulation due to
chemotaxis/motility caused a greater rate of non-labile naphthalene degradation as
compared with the system inoculated with a mutant. Experiments carried out with soils
under a no-mix condition, however, revealed that none of the strains were able to degrade
the non-desorbable fraction of naphthalene. The present study suggested that in a
heterogeneous matrix ( i.e., soil) , containing medium to high content of clay and organic,
the role of mixing is very important in successful bioremediation. The lack of mixing
reduced the effective diffusion of molecules in soils due to solid phases, dead-end pores
and high tortousity of the system. Also, the lack of mixing reduced the number of sites
available for bacterial attachment for direct uptake due to pore sizes exclusion and

retardation affect (due to sorption).

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The third objective was to examine the factors affecting bacterial attachment and
consequently the rate and extent of biodegradation of naphthalene in a soil-slurry reactor.
The factors studied included effect of ionic strength and EPS layer of pseudomonas
Putida G7. Bio-availability experiments were carried outwith: 1) PpG7-WT strain under
variable ionic strength of suspending solution, and 2) different strains of PpG7 possessing
variable attachment capabilities under constant ionic strength of suspending solution.
Three orders of decrease in molar concentration of buffer medium did not affect rate and
extent of sorbed phase naphthalene degradation by P.putida G7. Likewise, important
phenotypes (e.g., EPS, flagella, attachment) of PpG7-WT apparently did not contributing

toward enhanced bioavailability of sorbed phase naphthalene.

CONCLUSIONS
1. The effect of extra-cellular enzyme and polymers was not observed in enhanced
bioavailablity of soil-sorbed naphthalene.
2. In natural soils, chemotaxis/motility traits do not play any role on enhanced
bioavailability. Chemotaxis/motility contributed in enhance degradation in a system with
granualr activated carbon, which contain greater fraction of non-desorbable naphthalene.
3. Factors promoting bacterial attachment (i.e. removal of exo-polysaccharides, presence
of flagella and reduction of buffer medium concentration) do not contribute to
biodegradation of non-desorbable fraction of soil-sorbed naphthalene in soil-slurry
system.

4. Mixing plays an important role in bioavailablity of sorbed contaminant in natural soils.

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