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APPLICATION AND TRANSGENIC PRODUCTION OF THE
BIOSURFACTANT CYCLODEXTRIN FOR ENHANCED
POLYAROMATIC HYDROCARBON PHYTOREMEDIATION

presented by

Rachada Settavongsin

has been accepted towards fulfillment
of the requirements for the

Doctoral degree ' Crop & Soil Sciences

    

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APPLICATION AND TRANSGENIC PRODUCTION OF THE BIOSURFACTANT
CYCLODEXTRIN FOR ENHANCED POLYAROMATIC HYDROCARBON
PHYTOREMEDIATION

By

Rachada Settavongsin

A DISSERTATION

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

DOCTOR OF PHILOSOPHY
Department of Crop and Soil Sciences

2005

ABSTRACT

APPLICATION AND TRANSGENIC PRODUCTION OF THE BIOSURFACTANT
CYCLODEXTRIN FOR ENHANCED POLYAROMATIC HYDROCARBON
PHYTOREMEDIATION
By
RACHADA SETTAVONGSIN

Phytoremediation is the use of plants to cleanup organic and inorganic
contaminants and is considered a cost effective, aesthetically pleasing, and more
environmentally friendly alternative to engineering-based remediation. Plant roots and
their associated microbes are considered the major mechanism for phytoremediation of
some persistent organic pollutants such as polycyclic aromatic hydrocarbons (PAHs).

PAHs are carcinogenic and mutagenic compounds that are stably sorbed to soil
organic matter, which limits aqueous solubility and consequently biodegradation
effectiveness. Biosurfactants are biologically synthesized detergent-like compounds that
have been demonstrated to improve bioremediation effectiveness for some pollutants.
Cyclodextn'ns (CDs) are sugar-based biosurfactants produced from starch by the bacterial
enzyme cyclodextrin glycosyltransferase (CGTase). CDs have been shown to enhance
solubility of various organic and metal contaminants fi'om polluted soils including
enhanced biodegradation of the PAH compound phenanthrene.

We hypothesized that the presence of biosurfactants (CDs) in the rhizosphere will
improve bioremediation effectiveness by enhanced solubilization of otherwise persistent
organic pollutants and proliferation of PAH biodegrading bacteria. Beta-cyclodextrin

(flCD) amendments were tested in a 4-month greenhouse study of four phytoremediation

plant species grown in PAH-impacted, industrial waste soils. Soil PAH levels and

bacterial community response were evaluated in [3CD treated and untreated control soils.
[3CD amendments were demonstrated to enhance both PAH bioremediation effectiveness
and stimulate total bacterial and biodegrader cell densities in most plant + biosurfactant
treatment combinations. PAH levels were generally observed to be more effectively
reduced in [3CD treated rhizospheres relative to unamended control soils.

As a strategy to biologically generate [3CD in contaminated soils for in situ
biosurfactant-enhanced phytoremediation, we genetically engineered transgenic plants
transformed with the bacterial cyclodextrin glycosyltransferase (cgt) gene. The cgt gene
of the BCD-producing bacterium Paenibacillus illinoisensis (PI) was cloned and
expressed in tobacco (Nicotiana tabacum). Transgenic PI-cgt plants produced and
secreted an extracellular enzyme, which was shown to degrade starch in culture medium.
Preliminary assays of cgt-tobacco in PAH-spiked soils were not conclusive for clear
indication of biodegradation. It is anticipated that cgt-tobacco “phyto”-BCD will enhance

solubization and subsequent biodegradation of otherwise persistent organic pollutants.

ACKNOWLEDGEMENTS

I would like to express my appreciation to my major advisor, Dr. Clayton Rugh,
for giving me an opportunity to work in his lab with this exciting project and for his
suggestions and advice. I would like to thank Dr. Stephen Boyd, Dr. Rebecca Grumet,
Dr. Kenneth Keegstra, and Dr. Thomas Voice for serving on my graduate committee and
for their valuable comments, encouragement and advice during the course of my study
and research at Michigan State University. I also thank Dr. Thomas Voice for allowing
me to use his laboratory equipment at times in my research.

In addition to my committee members, I would like to thank Dr. Lee Jacobs, Dr.
James Tiedje, and Dr. Brian Teppen for technical support and allowing me to utilize their
lab equipment and space over the course of my study. I extend my great appreciation to
Dr. Delbert Mokma for advice with soil sampling and soil classification methodologies.
My special thanks go to Chris Saffron for help with sorption and desorption isotherm
experiments and to Huang Xuewen for statistical analyses of my research data.

Let me also say ‘thank you’ to my lab coworkers for their contributions to soil
preparation, transplanting, harvesting, and sample collection: Cindy Wan, Sarah Kinder,
Endang Susilawati, Chris Saffron, Theresa Wood, Nathan Stewart, Michael Roberts,
Andrew Bender, Kevin Crist, and Danielle Abshagen. I would like to express my thanks
to Dr. Pulla Kaothien for suggestion and ideas for my molecular biology work.

I would like to thank my new friends at MSU for their help and support. Sarah
Kinder, I thank you for being my ‘walking dictionary’ and for helping me with all kinds

of your valuable skills: related computer support, molecular biology techniques, and

iv

chemical analyses. I am deeply indebted to Endang Susilawati for helping me with my
research as well as emotional support during the course of my study. I appreciate support
and suggestion from Cindy Wan. I would also like to thank my Thai friends, Anurat
Saithong, Pawika Boonyapipat, Nalinee and Vareemon Tuntivanich, for their advice,
support, and encouragement. Last but not least, I would like to express my gratitude to
my parents, sisters, and brothers for their love and support.

My dissertation program at MSU was supported by the Royal Thai Government
(2000-2004). For 2005 spring and summer semesters, I was supported by funding from
Ford Motor Company Environmental Quality Office, the Consortium for Plant
Biotechnology Research, and the Michigan State University College of Agriculture and

Natural Resources Dissertation Completion Fellowship.

TABLE OF CONTENTS

PAGE
LIST OF TABLES ............................................................................ viii
LIST OF FIGURES ........................................................................... ix
INTRODUCTION ............................................................................. 1
REVIEW OF LITERATURE ................................................................ 3
Polycyclic aromatic hydrocarbons (PAHs) ....................................... 5
Chemical and biological surfactants ............................................... 8
Phytoremediation ..................................................................... 14
Genetically engineered plants for phytoremediation ............................. 25
Summary and conclusion ............................................................ 29
Literature Cited ........................................................................ 42
CHAPTER 1.APPLICATION OF CYCLODEXTRIN AMENDMENTS FOR
ENHANCED PAH PHYTOREMEDIATION OF COKE-OVEN
SOILS ......................................................................... 43
Introduction ........................................................................... 46
Materials and Methods .............................................................. 51
Results and Discussions ............................................................. 59
Conclusion ............................................................................ 76
Literature Cited ........................................................................ 77

CHAPTER 2. EFFECTS OF EXOGENOUS CYCLODEXTRIN APPLICATION
ON BACTERIAL COMMUNITY STRUCTURE DURING PAH

RHIZOREMEDIATION OF COKE-OVEN SOILS .................... 84
Introduction ............................................................................ 85
Materials and Methods ............................................................... 89
Results .................................................................................. 93
Discussions ........................................................................... 109
Literature Cited ...................................................................... 115

CHAPTER 3. TRANSGENIC TOBACCO PRODUCTION OF THE BACTERIAL
BIOSURFACTANT CYCLODEXTRIN FOR ENVIRONMENT

REMEDIATION ............................................................ 122
Introduction .......................................................................... 123
Materials and Methods .............................................................. 128
Results ................................................................................. 142

vi

PAGE

Discussions ........................................................................... 162
Conclusion ............................................................................ 168
Literature Cited ...................................................................... 169

vii

LIST OF TABLES

PAGE

Table 1.1. Soil agronomic and physical properties for materials used in

phytoremediation growth chamber and greenhouse studies ............... 57
Table 1.2. Soil agronomically available nutrient content from growth

chamber BCD-phytoremediation study ....................................... 72
Table 1.3. Mulberry plant tissue nutrient mineral content from growth

chamber BCD-phytoremediation study ....................................... 74
Table 1.4. Big bluestem plant tissue nutrient mineral content from growth

chamber BCD-phytoremediation study ....................................... 75

Table 2.1. Bacterial cell density of total heterotrophic, phenanthrene degrading,
pyrene degrading, and cyclodextrin producing cells per gram soil
dry weight in uncontaminated (Clean) and PAH-contaminated
(Coking Oven) experimental source soils .................................... 96

Table 2.2 (a-b). Percentage of phenanthrene degrading bacteria (ZFU) relative to
total heterotrophic bacteria (CFU) for various [3CD x planted
treatments in Clean-soil or PAH-soil after 4-months treatment ......... 106

Table 2.3 (a-b). Percentage of pyrene degrading bacteria (ZFUpy) relative to total
heterotrophic bacteria (CF U) for various [3CD x planted treatments in
Clean-soil or PAH-soil after 4-months treatment ......................... 108

Table 3.1. Selective marker segregation analyses of T1 generation
cgt-tobacco seeds .............................................................. 152

Table 3.2. Phenanthrene (phen), fluoranthene (flra), pyrene (pyre),
and summed PAH (tPAH) concentrations in spiked

sand-soil mix (100 ug/ g each at T0) after planted and unplanted
treatments for 30 days (T1) and 45 days (T2) .............................. 164

viii

LIST OF FIGURES

PAGE

Figure 1.1. Sorption isotherm of l4C--phenanthrene for soil with various

cyclodextrin types and solution concentrations ............................. 60
Figure 1.2. Growth Chamber Study: Average soil tPAH concentration .............. 63
Figure 1.3. Greenhouse Study at 2 months sampling time: Average soil

tPAH concentration ............................................................. 64
Figure 1.4. Greenhouse Study at 4 month sampling time: Average soil

tPAH concentration ............................................................. 65
Figure 1.5 (a-c). Plant tissue biomass ....................................................... 68

Figure 1.6. Greenhouse study: Shoot biomass of plants after 4 months treatment. . ..70

Figure 2.1. Reduction from initial soil tPAH concentration by planted X BCD
amendment treatment combinations after 4 months under greenhouse
conditions ......................................................................... 94

Figure 2.2 (a-b). Comparison of total heterotrophic bacterial cell densities between

[3CD amendment concentrations for each plant species in treated
Clean-soil and PAH-soil ........................................................ 97

Figure 2.3 (a-b). Comparison of heterotrophic bacterial cell densities between plant

species for each [3CD amendment concentration in treated Clean-soil and
PAH-soil .......................................................................... 99

Figure 2.4 (a-b). Comparison of phenanthrene degrader cell densities between [3CD
amendment concentrations for each plant species in treated Clean-soil
and PAH-soil ..................................................................... 101

Figure 2.5 (a-b). Comparison of phenanthrene degrader cell densities between

plant species for each [3CD amendment concentration in treated
Clean-soil and PAH-soil ........................................................ 102

Figure 2.6 (a-b). Comparison of pyrene degrader cell densities between BCD
amendment concentrations for each plant species in treated
Clean-soil and PAH-soil ....................................................... 103

Figure 2.7 (a-b). Comparison of pyrene degrader cell densities between plant

species for each [3CD amendment concentration in treated
Clean-soil and PAH-soil ...................................................... 105

ix

PAGE

Figure 3.1 (a-b). Diagram of the pAPC9K and pCAMBIA 1300 cloning
and plant gene expression vectors .......................................... 135

Figure 3.2. Diagram of the expression cassette in plant transformation
vector pE1778 ................................................................ 136

Figure 3.3. Cyclodextrin glycosyltransferase (CGTase) indicator assay of soil
bacterial cultures ............................................................. 143

Figure 3.4. Thin-layer chromatograph of CGTase enzyme assay of supematants
of putative CD-expressing bacterial isolate cultures .................... 145

Figure 3.5. Cyclodextrin production by bacterial isolate culture supernatants. . .. 146

Figure 3.6. Paenibacillus illinoisensis C36 isolate cyclodextrin glycosyltransferase
Gene (cgt) nucleotide coding sequence .................................... 147

Figure 3.7. Paenibacillus illinoisensis strain C36 CGTase amino acid sequence,
calculated molecular mass of 78.2 kDa derived from the nucleotide
sequence ........................................................................ 149

Figure 3.8 (a-c) LB, starch degradation assay, and BCG CD-indicator assay of
Paenibacillus illinoisensis, E. coli strain DHch and transformed
cgt-E. coli DHSa ............................................................... 150

Figure 3.9 (a—b). Average leaf areas of the (a) 3rd leaf and (b) 4m leaf
(counted from the base of stem) of wild type and cgt—tobacco

plant lines 45 days after planting ............................................ 153

Figure 3.10. Average plant height of wild type and cgt plant lines 45 days
after planting .................................................................. 154

Figure 3.11. Average number of days from planting to flower of wild type
and cgt plant lines ............................................................ 155

Figure 3.12 (a—b). Growth and development comparison of wild type and

transgenic cgt-tobacco lines ................................................ 156
Figure 3.13. Genomic PCR analysis of putative cgt-tobacco transfonnants ........ 157
Figure 3.14. RT-PCR analysis of cgt tobacco leaf extracts ........................... 159

Figure 3.15 (a-b). Native-PAGE and zymogram analysis of CGTase enzyme
from cgt-tobacco leaves ..................................................... 160

PAGE

Figure 3.16 (a-b). Starch degradation analysis of wild type and
cgt-tobacco plants ............................................................ 161

Figure 3.17. tPAH concentration in spiked sand-soil mix after

planted and unplanted treatments for 30 days (T1)
and 45 days (T2) .............................................................. 163

xi

INTRODUCTION

Polyaromatic hydrocarbons (PAHs) are world wide environmental hazards with
carcinogenic and mutagenic properties. PAHs are relatively insoluble in water and
partition to soil organic matter, which limits their bioavailability and contributes to their
long-term persistence in the environment. Conventional engineering-based approaches
such as extraction and containment strategies have been routinely used to treat
contaminated sites. Biologically based remediation may offer the advantages of reduced
incidental mobilization of pollutants, minimal site disturbance, and more rapid restoration
of the rehabilitated habitat. Microbial PAH biodegradation has been extensively studied
at the biochemical and genetic levels. Phytoremediation is the use of plants and their
associated microbes to clean up environmental pollution. Some plant species are capable
of direct degradation of organic, or carbon-based, contaminants. Other plants release
phenolic compounds from their roots, which can serve as metabolic carbon substrates and
stimulate cometabolic microbial biodegradation of PAHs. Transgenic approaches have
been used to confer novel phytoremediation capabilities to engineered plants for
enhanced removal and detoxification of various inorganic and organic pollutants.

For many pollutants, the rate-limiting characteristic is their stable associate with
soil constituents. Chemical or biologically produced surfactants, or detergent-like
compounds, have been used to enhance solubilization of persistent contaminants for
accelerated removal and/or biodegradation. Cyclodextrins (CDs) are a group of well-
studied biosurfactants derived from starch via enzymatic reaction of bacterial

cyclodextrin glycosyltransferases (CGT). CDs are torus-shaped molecules with a

hydrophilic exterior surface and a nonpolar interior cavity, which allows complexation
with poorly soluble molecules and subsequent aqueous solubilization.

The basic hypothesis of this dissertation project is that application or production
of cyclodextrins in the rhizosphere will improve the biodegradation rate of otherwise
persistent organic pollutants, e. g. PAHs. Our overall goal is development of transgenic
plants engineered to facilitate rhizosphere biosynthesis of CD biosurfactants to increase
hydrophobic compounds bioavailability and bioremediation efficiency.

The specific objectives of this dissertation research study are:

(1) Evaluate the effectiveness of applied cyclodextrins for enhanced
phytoremediation of hydrophobic contaminants.

(II) Evaluate the effect of applied cyclodextrins on biodegrader and biosurfactant-
producing bacteria in contaminated soils.

(III) Develop transgenic plants that secrete biosynthetic enzymes for cyclodextrin
production into the soil matrix.

(IV) Determine the effectiveness of cyclodextrin producing plants for enhanced
bioavailability and biodegradation of target compounds.

In this project, we tested cyclodextrin amendments on PAH contaminated soils
planted with species reported to be effective for phytoremediation of organic pollutants:
alfalfa (Medicago sativa), mulberry (Moms rubra), monkey flower (Mimulus ringens),
and little bluestem (Andropogon scoparius). The greenhouse-scale study was conducted
over a 4-month period using weekly applications of different CD amendment
concentrations with periodic sampling and analysis of soil PAH concentration, soil

microbial cell density, and plant growth and vigor. The results from this work will serve

as a proof-of-concept trial for our stated primary goal and further our understanding of
processes involved in rhizosphere-cyclodextrin interactions during PAH-
phytoremediation. In addition, this study will allow development of guidelines for future
applications of biosurfactant-assisted phytoremediation of persistent organic pollutants.
Finally, we genetically engineered tobacco to express bacterial gene constructs
encoding the CD biosynthetic enzyme cyclodextrin glycosyltransferase (CGTase).
Transgenic tobacco was tested for cgt expression and CGTase function in experimental
biochemical and molecular genetics assays. Confirmed cgt-tobacco plants were tested in
pilot-scale grth chamber trials for their effect on PAH biodegradation in planted,
spiked soils. We propose that transgenic cgt-plants would be superior to application of
either biosurfactant-producing bacteria or purified biosurfactant, due to easier system

management and less risk of contaminant leaching.

REVIEW OF LITERATURE

Environmental Pollution

Environmental contamination of air, soil and water with hazardous materials is a
worldwide concern due to health risks to humans and wildlife, degraded natural
resources, and impacted ecosystem stability. Environmental pollution is usually
anthropogenic, typically occurring from industrial or military activities. In 2005, there
were 1,247 Superfund National Priorities List sites in the United States mandated by the
US Environmental Protection Agency Superfund Program to reduce pollutant levels
(EPA, 2005). Additionally, thousands of properties formerly operated by the US.
Department of Defense may contain hazardous toxic compounds including explosives
and radioactive wastes in soils, waters, or underground storage tanks (GAO, 2001).

Terrestrial pollutants are generally divided into two major groups: elemental
pollutants include heavy metals (e. g. aluminum, zinc, cadmium, lead) and radioactive
elements or radionuclides (e. g. cesium-137, strontium-90); and carbon-based, or organic,
chemicals (Schaffner et al., 2002). The major sources of metal pollution are mining,
industrial processing, manufacturing, and weathering of minerals and soils, disposal of
industrial and domestic wastes, fertilizers and pesticides. Sources of radionuclides are
nuclear testing, accidental release, or nuclear energy production. Inorganic pollutants are
transported into atmospheric, aquatic, and terrestrial environment as solutes or
particulates, which may accumulate to high concentrations causing risks to public health.
Organic pollutants include petroleum hydrocarbons, polycyclic aromatic hydrocarbons
(PAHs), polychlorinated biphenyls (PCBs), explosives [e. g. 2,4,6-trinitrotoluene (TNT),

hexahydro-l,3,5-trinitro-l,3,5-triazine (RDX)], pesticides, herbicides, industrial solvents

(e. g. trichloroethylene, toluene, xylene), and dry cleaning solutions. Many of these
compounds are resistant to biodegradation and persistent in the environment and are so-
called persistent Qrganic pollutants (POPS). POPS are chemical substances that do not
readily undergo biogeochemical reactions, persist in the environment, and tend to
biomagnify in lipid tissues of exposed animals. Examples of POPS include the pesticide
l,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (DDT), PCBS, and some PAHS. POPS have
low aqueous solubility and strongly associate with soil organic matter (SOM), making
them resistant to abiotic decomposition and biodegradation (Eduljee, 2001). With
increased soil or sediment residence time, POPS become more tightly bound to the soil
matrix and less available for biodegradation.

Polycyclic aromatic hydrocarbons (PAHS)

PAHS can be produced from either natural and anthropogenic processes. Unlike
synthetic pesticides and PCBS, PAHS are formed naturally in the environment when
organic materials are burned with insufficient oxygen, such as during the burning of
vegetation in forest and bush fires. PAHS are also formed by industrial or common
human activities, including the burning of fossil fuels, coal processing, petroleum
refining, and wood-preserving treatments (Wilson and Jones, 1993). Some PAHS are
manufactured for use in coal tar, roofing tar, creosote, medicines, dyes, pesticides and
plastics. PAHS are volatilized and dispersed from incomplete combustion and can either
remain as vapor or sorb to dust particles, returning to the earth’s surface through rainfall
or particle settling. PAHS discharged from industrial and wastewater treatment plants

adhere to particles and settle to river or lake bottoms. According to the US.

Environmental Protection Agency (EPA), PAHS have been found above threshold levels
in 600 of 1,408 National Priority List Sites (EPA, 2005).

PAHS are of ecotoxicological concern because several PAH compounds are
potentially carcinogenic and mutagenic, e. g. benzo[a]pyrene and benzo[a]anthracene,
causing tumors and reproductive abnormalities in laboratory animals and cancer in
chronically exposed humans (ATSDR, 1995). The metabolic conversion of PAHS by
liver cytochrome produces diol-epoxide intermediates, which have been identified as the
mutagenic and carcinogenic agents (Sims and Overcash, 1983). Some lower molecular
weight PAHS, e. g. naphthalene, phenanthrene, fluorene, pyrene, were shown to reduce
survival rates of soil-dwelling Springtail (Folsomiafimetaria) (Sverdrup et al., 2002). In
phytotoxicological assays, PAHS were demonstrated to decrease the germination rate of
garden cress (Lepidium sativum) by 20 and 50% in soils contaminated with 50 and 1000
ppm PAHS respectively (Maila and Cloete, 2002).

Microbial Metabolism of PAHS

PAHS are degraded both biologically and abiotically. Abiotic degradation occurs
via photolysis or chemical oxidation. Volatilization and leaching occurs to some extent
for two- and three-ring PAHS, though these are negligible dissipation processes for higher
molecular weight PAHS, which have very low vapor pressures and are strongly sorbed to
soil organic matter (Reilley et al., 1996). PAHS are biologically degraded under aerobic
conditions by soil and aquatic microorganisms, typically bacteria and fungi.
Pseudomonas, Alcaligenes, Rhodococcus, Beijerinckia, Mycobacterium, Staphylococcus,
and Arthrobacter are examples of soil bacteria with the ability to degrade or co-

metabolize higher molecular weight PAHS such as pyrene, fluorene and

benzo(a)anthracene (Cerniglia, 1993; Cheung and Kinkle, 2001). PAH biodegrading
bacteria use dioxygenase enzyme activity to incorporate two oxygen atoms into the PAH
ring structure to produce dioxetanes, cis-dihydrodiols and catechol, which are further
degraded to carbon dioxide and water (Cerniglia, 1993; Wilson and Jones, 1993). In
contrast to bacteria, fungi utilize monooxygenases to incorporate one oxygen atom into
the PAH compound to form arene oxides (epoxides) followed by the enzymatic addition
of a water molecule to form trans-dihydrodiols and phenols. Trans-dihydrodiol are
oxidized to catechols, which may be further degraded to acetaldehyde and succinic,
fumaric, pyruvic, and acetic acids, which are used as carbon sources by microorganisms.
The white rot fungus (Phanerochaete chrysosporium) produces lignin peroxidases, the
extracellular enzymes that catalyze free radical attack on PAHS by a Single electron
transfer to form quinones (Cerniglia, 1993).

PAHS are hydrophobic, have low water solubility, and high octanol partitioning
coefficients (logKow), resulting in high sorption to SOM and environmental persistence.
Microbial biodegradation of PAHS is considered the prevalent process for PAH
dissipation in contaminated soils. Microorganisms capable of degrading organic
pollutants are typically abundant in PAH contaminated soils. Freshly added PAHS are
initially degraded very rapidly with little or no additional loss of the compound detected
after several years (Chung and Alexander, 1998; White and Alexander, 1996). Increasing
the exposure time between chemicals and soil or sediment decreases the rapid desorbed
fraction as well as biodegradation (Comelissen et al., 1998). Several physical-chemical
processes such as sorption/desorption, dissolution, and diffusion govern bioavailability of

specific compounds. Two main factors involved in microbial bioremediation are the rate

of contaminant transfer to the cell and the rate of uptake and metabolism (Cuypers et al.,
2002). The slow mass transfer of hydrophobic compounds from a sorbed phase to the
degrading microorganisms reduces the removal rates of the contaminants. Sustained
bioremediation becomes limited due to the irreversible sorption of PAHS to SOM, low
water solubility, and low bioavailability (Kc et al., 2003). It is generally believed that
microorganisms can only uptake and degrade aqueous solubilized PAHS, and therefore
the slow release of the substrates from soils or sediments to interstitial water is
considered the rate limited step for biodegradation.
Chemical and Biological Surfactants

To overcome the limited aqueous solubility of POP compounds, synthetic and
biosynthesized surfactant chemicals have been proposed to solubilize contaminants for
extraction and subsequent treatment. Surfactants usually consist of a hydrophilic head,
which may be constituted from a wide variety of chemical structures, and a hydrophobic
tail, which is frequently a hydrocarbon chain. Above a certain threshold concentration
called the critical micelle concentration (CMC), surfactant molecules cluster together
forming aggregates called micelles. The hydrophilic portion makes surfactants soluble in
water, while the hydrophobic moieties orient toward the center of the micelles, forming a
hydrophobic core (Volkering et al., 1997). Surfactants enhance rates of mass transfer
from a sorbed phase by increasing the solubization and dissolution of non-polar
contaminants by partitioning them into the hydrophobic micelle core (Volkering et al.,
1998)

Chemical surfactants have been demonstrated to enhance desorption and

solubilization of sorbed contaminants in soils (Deitsch and Smith, 1995; Guha eta1.,

1998; Sun and Boyd, 1993; Thangamani and Shreve, 1994; Thibault et al., 1996), thus
making the substrates more readily available for microbial metabolism. A fraction of
phenanthrene partitioned into micellar phase of nonionic surfactants was found to be
readily bioavailable to microbes (Guha and J affe, 1996; Roch and Alexander, 1995). The
nonionic surfactant alcohol ethoxylate was observed to enhance biodegradation rates of
phenanthrene, hexadecane (Macur and Inskeep, 1999) and biphenyl in soil (Aronstein et
al., 1991). Applied surfactant concentrations above the critical micelle concentration are
typically most effective for contaminant solubilization (DiVincenzo and Dentel, 1996;
Guha and J affe, 1996), though may be prohibitively expensive or inhibitory to microbial
activity. Several of the surfactants were toxic to the test bacteria and prevented
biodegradation of biphenyl and phenanthrene at concentrations below and above CMC
due to an injurious influence of the surfactant to bacterial membrane proteins (Laha and
Luthy, 1991; Roch and Alexander, 1995).

Certain bacteria, yeast, or fungi have been demonstrated to enhance hydrophobic
pollutant bioavailability and degradation rates by biosynthesis of high surface-active and
emulsifying biosurfactants that increase solubilization of non-polar molecules in the
aqueous phase and emulsify hydrocarbon-aqueous phase mixtures (Kanga et al., 1997).
The biosurfactant produced by Rhodococcus H-13A was Shown to increase the solubility
of naphthalene (Kanga et al., 1997), as well as three- and four-ring PAHS (Page et al.,
1999). Rhamnolipids of Pseudomonas aeruginosa were demonstrated to enhance the
solubility and biodegradation rate of phenanthrene (Zhang et al., 1997). PAH degrading
bacterial isolates from a soil contaminated with petroleum waste were able to produce

their own surfactant, subsequently increasing dissolution of naphthalene and

phenanthrene (Deziel et al., 1996). Biosurfactants can be glycolipids, lipopeptides,
polysaccharide protein complexes, phospholipids, fatty acids or neutral lipids (Hommel,
1990; Maier, 2003). Biosurfactants are typically composed of a hydrocarbon tail of one
or more fatty acids, which may be unsaturated, saturated, hydroxylated, or branched;
which is linked by a glycosidic ester or amide bond to the hydrophilic moiety of the
molecule (Hommel, 1990). Unlike many synthetic surfactants, biosurfactants are usually
readily biodegraded, are of lower cellular toxicity than chemical surfactants, and retain

effectiveness at wide ranges in pH and temperature (Georgiou et al., 1992).

Cyclodextrins (CDs) are circular oligosaccharide biosurfactants with a hydroxyl-
rich, hydrophilic external surface and a hydrophobic, interior cavity (Figure 1) (Szejtli,
1991). Cyclodextrins are formed by enzymatic degradation of starch and circularization
of the fragment via a ct-l-4 linkage. The CD biosynthetic enzyme, cyclodextrin
glycosyltransferase, is produced by a large number of Bacillus Species (Larsen et al.,
1998), some strains of Brevibacillus brevis (Kim et al., 1998), and Klebsiella pneumonia
strain M5A1 (Binder et al., 1986). CD3 are designated a-, 0-, or 'Y-CDS having 6, 7, or 8
glucose units, respectively, which determines the cavity size conferring the specificity
and range of complexation with organic compound “guest” molecules (Matioli et al.,

2000; Wang and Brusseau, 1993). Cyclodextrins and modified cyclodextrins have been

demonstrated to enhance solubility and extractability of various organic and metal
contaminants from polluted or spiked soils (Sheremata and Hawari, 2000; Wang and
Brusseau, 1995). Biodegradation and dechlorination of the PCB mixture Aroclor 1221
was Significantly enhanced by addition of hydroxypropyl-B-cyclodextrin to shaken liquid

cultures of Pseudomonas sp. strain CPEl (Fava et al., 1998). Unlike micelle-forming

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surfactants, cyclodextrins have no critical micelle concentration requirement and are thus
effective at any concentration, are considered physicochemically stable, and do not sorb
to soil or organic matter (Ko et al., 1999). Also, in contrast to many synthetic surfactants
and some biosurfactants, CDS are not harmful to microbes, plants, or macrofauna except

at extremely high concentrations (Bar and Ulitzur, 1994; Szejtli, 1988).

Pollution Remediation Technologies

Conventional engineering-based approaches over the last three decades include
excavation and landfill disposal, solidification, carbon adsorption, UV oxidation,
chemical precipitation, soil washing, incineration and vapor extraction (Cunningham and
Berti, 1993; FRTR, 2003). However, the choice of remediation technique depends on the
nature of the contaminants. Metals and radionuclides tend to associate with soil organic
matter and soil minerals, forming insoluble complexes with phosphates or hydroxides
(Dragun, 1998). Therefore, they are relatively immobile and considered difficult to
remediate. Soils contaminated with metal are usually excavated and land filled, or treated
by acid washing or electrochemical separation. Soils contaminated with low
concentrations of radionuclides are removed from the site and treated with various
dispersing and chelating chemicals. Disposal of large volumes of liquid chemical wastes
from decontaminated radionuclide-polluted soils may be very expensive. It was estimated
that treatment of all radionuclide-polluted soils in the US would cost about $200-300
billion (Entry et al., 2001).

Soils contaminated with organic pollutants are usually treated by excavation and
thermal desorption, vapor stripping, washing, incineration, or landfilling. Excavation is

environmentally disruptive to the site, may generate toxic air emissions, and, if buried in

12

landfills, is not considered a permanent solution. In addition, finding new secure and
hazardous waste landfill Operations is becoming more difficult (Cunningham and Berti,
1993), so alternatives to burial are needed. Biologically based remediation technologies
such as bioremediation and phytoremediation may be possible alternative approaches for
cost effective and permanent environmental cleanup (EPA, 2001).
Bioremediation

Bioremediation is defined as the use of living organisms, typically microbes, to
degrade, detoxify, or transform pollutants into non-toxic or less harmful compounds. In
bioremediation, indigenous or supplemented microbial species or complex community
interactions are exploited for contaminated site remediation. Bioremediation
effectiveness is dependent on microbial community structure and density as well as
abiotic factors, such as nutrient availability, aerobic conditions, and soil pH, temperature,
and moisture. Microbes transform or degrade organic compounds via direct or co-
metabolic degradation. In direct metabolism, microbes are capable of utilizing the target
organic pollutant as a carbon or energy source for growth. For co-metabolism, the
microorganism cannot utilize the target contaminant as a carbon or energy source, so the
pollutant may be degraded only during grth on another carbon or energy source. For
instance, microbes utilize the structurally related chemical biphenyl as the growth
substrate to degrade PCBS (Focht, 1995).

Successful bioremediation has been demonstrated for numerous organic
compounds. In a bioremediation field study, addition of Slow release fertilizer or fish
composts increased microbial population and accelerated oil residue biodegradation rates

in treated soil relative to unamended soils (Delille et al., 2002). Indigenous aerobic

13

microorganisms were capable of biodegradation of lightly chlorinated PCBS present in
Hudson River sediments, which was enhanced by addition of inorganic nutrients,
biphenyl, and oxygenation (Harkness et al., 1993). A mesocosm study of crude oil
contaminated soil showed that petroleum hydrocarbons decreased 70-90 % (Cl 1-C22
fraction), 40-60 % (C23-C32 fraction) and 35-60 % (C35-C44 fraction) after 11 months
of remediation (Salanitro et al., 1997). Mineralization of atrazine in aquifer sediments
was enhanced 75 % by addition of Pseudomonas sp. strain ADP, an effective atrazine
mineralizing bacterium, relative to non-bioaugmented sediments under laboratory
conditions (Shapir et al., 1998). In addition to biodegradation, bioremediation has also
been demonstrated to effectively reduce biotoxicity of hydrocarbon polluted substrates.
Bioremediated crude oil contaminated soil displayed decreased toxicity to earthworms
and reduced inhibition of seed germination after 5 months treatment (Salanitro et al.,
1997; Wang et al., 1990). In a laboratory study, bioremediation was Shown to
significantly reduce the concentration of carcinogenic PAHS in coal tar contaminated soil
after 180 days treatment.
Phytoremediation

Phytoremediation is defined as the use of green plants to clean up contaminated
soil and water containing organic or inorganic wastes. Plants can detoxify some organic
contaminants by direct accumulation and enzymatic degradation or by stimulation of
microbial degradation in the root zone, or rhizosphere. Some plants take up inorganic
contaminants, such as metals and radioactive substances, from soil or water and
concentrate them in their tissues (Cunningham and Berti, 1993; Schnoor et al., 1995).

Excavation and processing of metal-contaminated soil may cost about a quarter million

14

dollars per acre, while phytoremediation costs range from 2 to 4 orders of magnitude less
than engineered approaches (Cunningham and Ow, 1996). For example, the ash resulting
from incinerated, metal-laden plant biomass may be only 20-30 tons compared to 5,000
tons of the contaminated source soil (Watanabe, 1997). In addition to the reduced costs
that may be realized with phytoremediation, plants reduce site erosion, minimize seepage
of contaminated leachate, and Simultaneously restore soil fertility. There are several
categories of phytoremediation treatment technologies: phytoextraction,
phytostabilization, phytodegradation, and phytostimulation. Which phytoremediation
application to use for a particular site treatment is dependent upon various factors, such
as pollutant chemistry, i.e. inorganic or organic, and site conditions, i.e. aquatic or
terrestrial (Cunningham and Ow, 1996).

Phytoextraction. also called phytoaccumulation, refers to the uptake and
translocation of metal contaminants into the aboveground parts of plants (Salt et al.,
1995). Plants normally extract Significant quantities of macronutrients, e. g. Ca, Mg, and
Fe, and modest levels of micronutrients, e. g. Zn, Cu, and Se, from soils. However,
nonessential elements such as Cd, Pb, Cr, Hg, and As, may also be accumulated to
variable degrees depending on soil parameters and the plants’ physiological and
morphological capabilities (McGrath et al., 2001; Negri and Hinchman, 1996). Certain
plants, known as hyperaccumulators, have the ability to acquire heavy metals in
aboveground parts to levels one hundred or more times the typical shoot tissue
concentration (Baker and Brooks, 1989). The latex of the tropical tree species Sebertia
acuminata growing on nickel-rich soil was observed to accumulate more than 20% nickel

on a dry weight basis (N egri and Hinchman, 1996). Brassicajuncea (Indian mustard) has

15

been demonstrated to accumulate and tolerate extremely elevated levels of Pb, Cd, Cr
(VI), Ni, Zn and Cu (Salt et al., 1995). Poplars were used as bio-pumps to reduce
drainage and extract water contaminated with boron from timber industry waste
(Robinson et al., 2003). Commercially phytoextraction of nickel by the Ni
hyperaccumulator genus, Alyssum, was recently developed. Improved genetic selection,
cultural practices, and soil management practices led to development of cost effective
methods of Ni recovery and recycling for commercial phytoextraction applications (Li et
aL,2003)

High levels of metal uptake in hyperaccumulators like T hlaspi caerulescens may
be due to high expression of metal transporter proteins located at the plasma membrane
(Lombi et al., 2001; Pence et al., 2000). Hyperaccumulators may further facilitate uptake
and tolerance of metals by producing metal chelators like histidine, which is involved in
Ni translocation in xylem of Alyssum lesbiacum (Kramer et al., 1997). Phytochelatins
(PCS), metal-induced chelator peptides, are believed to play a major role in heavy metal
detoxification in plants. PCs bind free metal ions with subsequent transfer to vacuoles
thereby protecting metal-sensitive cytoplasmic functions. PCs are synthesized by PC
synthase from glutathione (GSH) afier induction by the presence of heavy metal ions (Ha
et al., 1999). The cadmium-sensitive mutant of Arabidopsis thaliana, cad 1-1, possessed
15 to 30 % glutathione and 10% PC content of wild type plants and was observed to be
highly sensitive to cadmium exposure (Howden et al., 1995). Another class of heavy
metal binding proteins is the metallothioneins (MTS), cysteine-rich metal binding

proteins. Transgenic tobacco expressing yeast metallothioneins (CUPl) were shown to

16

accumulate 2-3 times the level of copper than non-transgenic wildtype tobacco plants
(Thomas et al., 2003).

Synthetic chemicals such as nitrilotriacetic acid (NTA), ethylene-tetraacetic acid
(EDTA), diethyltriamine-pentaacetic acids (DTPA) have been demonstrated to facilitate
higher metal uptake by hyperaccumulator plants. Other studies have the utility of
potentially less harmful organic acids such as citric acid (Turgut et al., 2004; Zhou et al.,
2003); the biodegradable, synthetic organic chelate ethylenediaminedisuccinic acid
(EDDS) (Kos and Lestan, 2003); and the biosurfactant compounds rhamnolipids (Maier
et al., 2001) and cyclodextrins (Brusseau et al., 1997) to enhance metal uptake in plants.

There exist some limitations and concerns for phytoextraction applications. Some
of the published phytoextraction research was performed by growing plants in
hydroponic solutions, which may have a tendency to overestimate its effectiveness.
Phytoextraction efficiency in the field might be less than in the laboratory due to limited
bioavailability of metals in soil media. Effective phytoextraction may be dependent on
the root architecture, particularly with regard to depth, and Shoot biomass, for
accumulation and storage of significant mass of extracted contaminant. Most naturally
occurring hyperaccumulator plants have small biomass, slow grth rates, and require
specialized growth conditions leading to uncertain effectiveness or very long treatment
times. The use of synthetic chelator amendments in field applications for facilitated metal
phytoextraction may also be questionable. NTA is a Class II carcinogen, DTPA is a
potential carcinogen, and EDTA and DTPA are toxic to invertebrate model organisms
and possess limited biodegradability. Leaching of these compounds and their metal

complexes is of concerns (Grcman et al., 2001; Sun et al., 2001; Vassil et al., 1998).

17

Phytostabilization is the use of plants to immobilize metals in soils by
accumulation into roots, adsorption onto roots, precipitation due to valence changes in the
root zone, or binding to organic matter. Soil management strategies, such as addition of
soil amendments, e. g. organic matter, phosphates, alkalizing agents, can further decrease
solubility of metals and minimize leaching to groundwater (Chaney et al., 1997). Metal
tolerant vegetation cover reduces contaminant loss from water or wind erosion,
dispersion due to run off, or leaching. In the root zone, metals are stabilized by
transformation or complexation of soluble forms to insoluble chemically modified states.
Indian mustard (Brassica juncea) roots were demonstrated to reduce toxic and soluble Cr
(VI) to insoluble and less toxic Cr (III) (Salt et al., 1995). Lead in soil may be
bioavailable and absorbed if ingested, though lead-phosphate or lead chlorophosphate
(pyromorphite) complexes are highly insoluble and nonbioavailable (Chaney et al.,
1997). Roots of colonial bent grass (Agrostis capillaris) growing in highly lead and zinc-
contaminated soil are able to form pyromorphite from soil lead and phosphate, greatly
reducing biological exposure (Chaney et al., 1997). Colonial bentgrass and red fescue
(F estuca rubra) were combined with beringite, an industrial byproduct with strong metal
immobilization capacity, and compost, to revegetate highly metal polluted acid sandy soil
After five years treatment, the beringite-amended phytostabilization mixture reduced the
water-extractable metal fraction of the treated soil 70-fold relative to untreated control
soil (Vangronsveld et al., 1996). However, a limited number of plants are able to grow in
soil with adverse pH, low organic matter, poor water holding capacity, and high

concentrations of heavy metals. In addition, large volumes of soil amendments, sustained

18

agronomic inputs, and extended time periods may be required to produce a persistent
vegetative cover for effective, longterrn phytostabilization.

Phflodegradation, also called phytotransfonnation, is defined as the breakdown of
contaminants taken up by plants through metabolic process within the plants. Plants may
degrade contaminants by release of enzymes into the soil or by taking up and
transforming the chemicals via lignification, volatilization, or metabolism to carbon
dioxide and water. Plant dehalogenase, nitrogenase, peroxidases, laccase, and nitrilase
enzymes have been proposed to be involved in phytodegradation processes (Schnoor et
al., 1995). In plant cells, organic contaminants may be converted by oxidation, reduction
or hydrolysis by P450 monooxygenases or carboxylesterases, which introduce functional
groups such as —OH, -NH2 or SH (Schuler, 1996). Organic pollutants may also be
conjugated with glutathione, sugars or amino acids by glutathione S-transferase (GST),
O-and N-glucosyltransferase, or malonyltransferase enzymes as a means of
detoxification. The conjugated forms may be further converted to other conjugates,
deposited into vacuoles or integrated into cell wall components (Dietz and Schnoor,
2001; Edwards et al., 2000).

The aquatic plants duckweed, parrot feather, and elodea were able to uptake 0, p ’-
DDT and p,p’-DDT from aqueous culture medium, and subsequently degrade or
incorporate them into plant materials (Gao et al., 2000). Callus cultures of crop plants,
soybean and wheat, and non-crop field plants, foxglove and jimsonweed, were Shown to
incorporate and transform pyrene (Huckelhoven et al., 1997). Cell suspension cultures of
soybean (Glycine max), white clover (T rifolium repens), and rose (Rosaceae) were

reported to metabolize a variety of di-, tri- and tetrachlorinated PCB congeners (W ilken

l9

et al., 1995). Trees of the Salicaceae family include poplars and willows and are referred
to as “phreatophytes”, or water loving plants, due their high volume water uptake.
Phreatophytes have been shown to be effective for containment of organic contaminant
plumes in soils and groundwater (Newman et al., 1997; Newman et al., 1999; Trapp et
al., 2001). In addition to hydraulic containment, poplars have been shown to possess
phytodegradation abilities. Hybrid poplars (Populus trichocarpa x Populus deltoides and
P. trichocarpa x P. maximowiczii) have been shown to mineralize the widespread
groundwater contaminant trichloroethylene (TCE) to C02 and less toxic metabolites
through a cytochrome P450 monooxygenase enzyme activity (Schnoor et al., 1995).
Axenic poplar tissue cultures and whole plant hydroponic studies showed metabolism
and incorporation in cell walls of radiolabeled CM-TCE breakdown products (Gordon et
al., 1998). Whole plants and tissue cultures of poplar were demonstrated to degrade
perchlorate in solution cultures by monitoring breakdown products of radiolabeled C136-
perchlorate (Van Aken and Schnoor, 2002). The major plant mechanisms for TCE
detoxification were observed to be formation of epoxides by cytochrome P450 or
conjugation to glutathione by glutathione S-transferase (Dietz and Schnoor, 2001).
Plants infected with the bacterium Agrobacterium rhizogenes, the causative agent
of hairy root syndrome, Show greatly increased proliferation of secondary root tissues,
which has been proposed to have potential advantages for phytoremediation. Carrot hairy
root cultures were observed to remove 90% of phenol and chloro-phenol derivatives from

aqueous medium within 120 hours of exposure with subsequent metabolism by cellular

peroxidases (de Araujo et al., 2002). Solanum nigrum (black nightshade) hairy root

20

culture was demonstrated to transform 60% of Delor 103, a mixture of 59 PCB
congeners, within 30 days of incubation (Mackova et al., 1997).

There are limitations to phytotransforrnation applications. The efficiency of plant
uptake and translocation of organic contaminants is dependent on compound
hydrophobicity, solubility, polarity and molecular weight (Briggs et al., 1982).
Compounds with octanol-water partition coefficients (log Kow) between 1.0 and 3.5, such
as benzene, toluene, ethylbenzene, xylene, chlorinated solvents, and Short chain aliphatic
hydrocarbons, are accumulated by plants. In contrast, more hydrophobic compounds bind
tightly to the root surface and are not efficiently translocated from roots to shoot tissues
(Briggs et al., 1982; Trapp, 1995). However, in some studies, PCBS and high molecular
weight PAHS were reported to be accumulated and transformed by plant cell suspension
cultures (Huckelhoven et al., 1997; Wilken et al., 1995). Members of the plant taxa
Cucurbitae, including zucchini and pumpkin, have been shown to accumulate elevated
levels of the highly hydrophobic pollutants PCBS, DDT and chlordane in their
aboveground tissues when grown on spiked or historically contaminated soils (White,
2002; White et al., 2005). The mechanisms for this atypical phytoaccumulation of
hydrophobic organic compounds is poorly understood, but thought to possibly be due to
Specialized modifications of soil chemical properties and/or cellular membrane function
by roots of these species (White et al., 2003). Despite reports of elevated capabilities for
plant uptake of organic or hydrophobic pollutants, another limitation may be the narrow
range of degradative pathways for these pollutants in plants. For example, hybrid poplars
were shown to uptake the explosives TNT and RDX from aqueous medium into living

tissues, though no chemical transformation of these compounds was detected (Thompson

21

et al., 1999; Thompson et al., 1998). Plant shoot tissue accumulation of organic
contaminants or their metabolites could potentially be considered as vegetation
contamination and pose a hazard to human and wildlife herbivores.

Phytostimulation, or plant-microbe bioremediation, has been shown to increase

 

biodegradation of hazardous organic compounds in planted soils relative to unplanted
sites. Enhanced dissipation of organic contaminants in rhizosphere soils has been
confirmed to be dominated by plant supported microbial activity rather than direct plant
uptake or abiotic degradation. Transformation of l4C-phenanthrene and 14C-
benzo[a]pyrene was monitored in tall fescue (F estuca arundinacea) rhizospheres (Banks
et al., 1999). After 6 months, 44 % of l4C-phenanthrene remained in fescue treated soils,
while 53 % of the starting amount remained in unplanted soils. Benzo[a]pyrene
mineralization was minimal, abiotic loss of the PAHS was only about 2% of starting
concentration, and less than 0.12% was observed to accumulate in plant tissues. Plant-
assisted biodegradation of two common PAH contaminants, anthracene and pyrene, was
investigated in a greenhouse study of four different plant species and unplanted soils
(Reilley et al., 1996). After 24 weeks, rhizosphere soils had 30-44 % lower PAH
concentrations than unplanted soil with leaching, plant uptake, abiotic degradation, and
mineralization shown to be negligible modes of PAH reduction. Various plant species
have been shown to accelerate pollutant biodegradation relative to unplanted soils,
including enhanced n-alkane degradation by ryegrass (Gunther et al., 1996), TCE
mineralization in rhizosphere of a legume (Lespedeza cuneata), loblolly pine (Pinus
taeda), and soybean (Glycine max) (Anderson and Walton, 1995), degradation of

pentachlorophenol by crested wheatgrass (F erro et al., 1994), streambank wheatgrass

22

degradation of 2-chlorobenzoate (Siciliano and Germida, 1997), and more efficient
degradation of herbicide, atrazine, in the Kochia scoparia rhizosphere (Perkovich et al.,
1996). The rate of herbicide 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) biodegradation
in a laboratory observation was higher for field-collected rhizosphere soil than non-
rhizosphere soil (Boyle and Shaun, 1998). In greenhouse study, the disappearance of 100
mg/kg anthracene and pyrene in rhizosphere of fescue, sudangrass, switchgrass and
alfalfa was greater than unplanted soil after 24 weeks (Reilley et al., 1996). In field study
of crude oil contaminated soils, a rotation crop of rye-soybean and St. Augustine grass-
cowpea plot decreased PAHS relative to unplanted plots after 21 months treatment
(Schwab and Banks, 1999). In a PAH phytoremediation field trial using 18 species
Michigan native plants, big bluestem (Andropogon gerardii), little bluestem
(Andropogon scoparius), New England aster (Aster novae-angliae), boneset (Eupatorium
perfoliatum), joe-pye weed (Eupatorium purpureum) were shown to reduce soil total
PAH (tPAH) levels over the course of the first growing season to a greater extent than
unplanted soil treatments (Wan, 2002).

Rhizosphere soils have been Shown to host higher cell numbers and diversity of
microbial populations than unplanted soils, resulting in increased degradation rates of
organic contaminants in vegetated soils. Microorganisms associated with plant roots are
typically oligotrophic, slow-dividing bacteria due to limited carbon resources. Root tips
and emerging lateral roots are zones of relatively high microbial activity resulting from
carbon-rich root exudates or cell lysates (Crowley et al., 1997). The population of
bacteria capable of biodegrading petroleum hydrocarbons in contaminated soil was

increased in the rhizosphere of red clover and tall fescue (Siciliano et al., 2003). The

23

number of microorganisms capable of degrading 2,4-dichlorophenoxyacetic acid (2,4-D)
in sugar cane rhizosphere soil was higher than non-rhizosphere soil (Sandmann and Loos,
1984). Alfalfa, broad bean, and ryegrass rhizospheres in oil-impacted soil displayed 2- to
3-fold greater total petroleum hydrocarbon (TPH) reduction and 10-to 100-fold higher
microbial cell density relative to unplanted soil (Yateem et al., 2000). In a PAH
phytoremediation field study of native Michigan plant species, rhizosphere soils were
generally observed to possess higher total heterotrophic bacteria and PAH biodegrading
bacterial cell densities than unplanted soils (Rugh et al., 2005).

Differences among plant species in root-released nutrients and secondary
compound composition strongly influence microbial growth, community diversity, and
metabolic degradation of xenobiotic pollutants (Crowley etal., 1997). Root exudates
contain a wide variety of phytosynthetic compounds, including aliphatic and aromatic
hydrocarbons, amino acids, and sugars, which may increase bacterial community density
and stimulate degradation of contaminants (Shimp et al., 1993; Siciliano and Germida,
1998). Many root-secreted compounds are chemically similar to certain pollutants and
may act as pollutant analogs or cometabolites. Cometabolites are compounds that may
serve as carbon and/or nitrogen sources enabling biodegradation of target pollutants that
cannot be utilized as sole carbon metabolites. Soils amended with terpene-containing
orange peels and eucalyptus leaves were shown to increase PCB biodegradation rates
compared to unamended soils (Hernandez et al., 1997). Phenolic compounds found in
root leachates of specific plants were demonstrated to stimulate grth and activity of the
PCB-degrading bacterial strains, Alcaligenes eutrophus H850, Corynebacterium sp.

MB], and Pseudomonas putida LB400 (Donnelly et al., 1994). Root exudates of red

24

mulberry (Moms rubra), osage orange (Maclura pomifera), and crabapple (Malusfusca)
contained relatively high levels of phenolic compounds that supported growth of PCB-
degrading bacteria (Fletcher et al., 1995). PCB phytostimulatory root phenolic
compounds might also be able to induce cometabolic grth and activity of PAH-
degrading microbes (Donnelly et al., 1994; Hegde and Fletcher, 1996).

Phytostimulation or rhizodegradation offer the advantages of in situ organic
contaminant destruction, reduced risk of herbivore exposure relative to tissue pollutant
accumulation, with easier site management as harvesting of vegetation is not required.
However, a prolonged time period or agronomic soil amendments may be required to
establish an effective vegetative cover especially in soils with adverse properties, such as
high contaminant concentrations, extreme climates, or Short growing seasons.
Rhizostimulation may also be restricted to shallow depths and subsequently ineffective
for immobile contaminants beyond the root profile. The primary limiting factor for
improvement and utilization of rhizosphere-mediated biodegradation is minimal
understanding of the critical mechanisms in this complex process, necessitating more
intensive and technically advanced experimentation.
Genetically engineered plants for phytoremediation

In addition to broad surveys of natural plant species for potential use as
phytoremediators, extensive efforts are ongoing to develop transgenic plants with
enhanced range and ability for detoxification of hazardous pollutants (Pilon-Smits et al.,
1999; Rugh, 2001; Rugh, 2004). For some contaminants, the plant detoxification or
biodegradation pathway may be known allowing development of plants genetically

engineered to overexpress the rate-limiting enzymes for enhanced phytoremediative

25

capabilities. Alternatively, foreign genes may be isolated from bacteria, mammals, or
other plants and transferred to more suitable, e. g. high biomass or deep rooting, plant
species for development of novel phytoremediation applications.

Metal phytoremediation strategies are based on stabilization, accumulation and
volatilization, each step of which has been the target of efforts for transgenic
improvement (Pilon-Smits and Pilon, 2002). Overexpression of metal transporter genes
may result in elevated metal uptake, translocation, and sequestration. Enhanced
production of metal binding proteins, such as phytochelatins (PCS) or metallothioneins
(MTS), may affect plant metal tolerance, uptake and translocation. PCS, non-translated
polymers of 5 to 7 units of [y-glutamylcysteinyl]glycines, are biosynthesized in plants by
the enzymes cysteine synthase, glutathione synthetase, and phytochelatin synthase. Indian
mustard (Brassicajuncea) was engineered to overexpress E coli glutathione synthetase
and observed to enhance cadmium tolerance and accumulation (Zhu et al., 1999).
Transgenic tobacco (Nicotiana glauca R. Graham) overexpressing a wheat phytochelatin
synthase gene, TaPCSl, displayed increased tolerance to Pb and Cd (Gisbert et al., 2003).
Transfer of TaPCSl gene to the cadmium, mercury, and arsenic sensitive Arabidopsis
mutant cad 1-3 restored wildtype-level metal tolerance and increased shoot uptake of Cd2+
(Gong et al., 2003). Transgenic Arabidopsis thaliana overexpressing yeast YCFI, a
member of ATP-binding cassette (ABC) transporter family, Showed increased transport
of glutathione-conjugated metal to the vacuoles resulting in elevated cadmium and lead
resistance (Song et al., 2003). Transgenic tobacco expressing a modified yeast CUP]

gene encoding metallothionein engineered with an additional Six histidine residues was

26

demonstrated to accumulate 90% more cadmium in aboveground tissues than non-
transgenic control plants (Macek et al., 2002).

There are no known naturally occurring plants capable of mercury
phytoremediation. Mercury bioremediation has been proposed by utilizing bacteria
containing the mercury detoxification pathway (mer), which includes the merA and merB
genes, encoding mercuric reductase and organomercury lyase, respectively (Barkay et al.,
1992). Plants were engineered to express the bacterial mer genes for phytodetoxification
of ionic and organic mercury compounds. The bacterial merA gene encoding mercuric
reductase was modified and expressed in transgenic plants to confer enhanced levels of
ionic mercury resistance and detoxification to Arabidopsis (Rugh et al., 1996), yellow
poplar (Rugh et al., 1998), and tobacco (Heaton et al., 1998). Since the geochemical
conditions for production of the ecotoxic organomercurial compounds are found in
saturated soils, merA genes were transferred to the flood tolerant plant species eastern
cottonwood (Populus deltoides) (Che et al., 2003) and domestic rice (Oryza sativa)
(Heaton et al., 2003) as a potential tool for mercury removal from these habitats. Each of
the merA engineered plant species were shown to possess elevated resistance to otherwise
toxic levels of ionic mercury (Hg2+), enhanced removal of Hg2+ from spiked media, and
achieved conversion of Hg“ to less toxic, less reactive elemental mercury (Hgo). In order
to directly detoxify organomercury compounds, e. g. methyl mercury (MeHg) and
phenylmercury acetate (PMA), plants were engineered with the bacterial merB gene.
MerB-Arabidopsis was shown to possess enhanced tolerance to organomercurial
compounds (Bizily et al., 1999), though plants engineered with both merA and merB

genes were more resistant due to complete detoxification of organomercurials to Hg0

27

(Bizily et al., 2000). Since hydrophobic organic mercury diffusion is the rate-limiting
step for broad-spectrum organomercury detoxification, plants were engineered with a
merB gene construct modified with a signal peptide targeting MerB to the cell wall. Cell
wall localized MerB was shown to increase the methylmercury transformation rate 10- to
70-fold over cytosolic expressing merB plants (Bizily et al., 2003).

Elevated levels of selenium (Se) are harmful due its substitution for sulfur in
biological compounds, most notably as seleno-amino acids (e. g. Se-Cys, Se-Met) in
proteins, thereby interfering with their essential functions. Cellular Se can be detoxified
by conversion to less toxic forms, including methyl-Se-Cys, methyl-Se-Met,
dimethyldiselenide (DMDSe), or Seo. Transgenic Indian mustard overexpressing plastid
ATP sulfurylase displayed increased reduction of selenate organic Se and elevated Se
resistance and enhanced tissue accumulation (Pilon-Smits et al., 1999). Overexpression
of a mouse gene encoding selenocysteine lyase catalyzed the metabolism of Se-Cys to
Sc0 and alanine, enhancing Se tolerance and accumulation in Arabidopsis (Pilon et al.,
2003) and Indian mustard (Garifullina et al., 2003). Transgenic overexpression of the
selenocysteine methyltransferase gene in Arabidopsis and Indian mustard enhanced
transformation of Se-Cys to the non-protein seleno amino acid, MetSeCys, subsequently
reducing misincorporation of Se into proteins with enhanced Se tolerance and
accumulation (LeDuc et al., 2004).

Arsenic exists in soil and water primarily in its oxidized form, arsenate A5043",
which is taken up by plants and translocated to leaves as an analog of phosphate.
Arabidopsis thaliana expressing the E. coli gene encoding arsenate reductase (ArsC) and

y-glutamylcysteine synthetase (y-E CS) displayed enhanced resistance and accumulation

28

of arsenic, due to the activity of arsenate reductase, which reduced A3043' to arsenite
As033', which was consequently bound to the thiol-peptide formed by y-E CS (Dhankher
etaL,2002)

Genetically enhanced plants have also been developed for phytoremediation of
organic, or carbon containing, pollutants. Transgenic tobacco plants expressing a gene
encoding a mammalian cytochrome P450 were Shown to degrade trichloroethylene (TCE)
at 640-fold the rate of non-transgenic control plants (Doty et al., 2000). Plants engineered
with the bacterial gene encoding pentaerythriol tetranitrate reductase gained the novel
capability to degrade explosive nitrate esters and nitroaromatic compounds (French et al.,
1999). Transgenic tobacco plants expressing the bacterial nitroreductase gene from
Enterobacter cloacae were able to tolerate and detoxify TNT (Hannink et al., 2001).
Transgenic Arabidopsis expressing the cotton secretory laccase gene (lacl), which is
involved in plant phenolic metabolism, exhibited elevated resistance to tetrachlorophenol
and its derivative compounds (Wang et al., 2004). Transgenic approaches offer the
potential for development of engineered, habitat-optimized plant species with novel
capabilities for treatment of enhanced ranges of environmental contaminants.

SUMMARY & CONCLUSION

PAHS tend to partition to soil organic matter in soils and sediments and become
unavailable for biodegradation. Phytoremediation is the use of plants to treat organic and
inorganic contaminants and may offer several advantages over conventional remediation.
PAHS in contaminated soils are decreased primarily by microbial metabolism, which has
been demonstrated to be enhanced in rhizosphere environments. Various studies have

revealed that microorganisms principally metabolize only dissolved PAH substrates.

29

Synthetic and biologically generated surfactants have been used to increase
solubilization, mobilization and desorption rates of hydrophobic contaminants, such as
PCBS and PAHS. Cyclodextrins (CDs) are a class of starch-derived biosurfactants shown
to enhance solubilization and biodegradation rates for PAHS and several additional
hydrophobic compounds.

In this research project, we examined the effect of aqueous cyclodextrin
amendments upon PAH contaminated soils in laboratory and greenhouse-scale
phytoremediation trials as a proof-of-concept study for further biotechnological
strategies. Exogenous cyclodextrin-amended soils were analyzed for PAH content
reduction and effects on cell density of total heterotrophic bacteria, PAH biodegrading
bacteria, and cyclodextrin producing bacteria. AS an alternative to application of
exogenous biosurfactant amendments, we developed transgenic plants engineered with
the bacterial cyclodextrin biosynthesis gene (cgt). Development of transgenic plants
expressing the bacterial cyclodextrin biosynthesis genes may provide a novel tool for
treatment of persistent, hydrophobic contaminants. This approach may be more effective
than bioaugrnentation with surfactant-synthesizing microbes, due to the more
environmentally adaptable and stable physiology of plants relative to supplemented
bacterial inocula. In addition, transgenic plant production of cyclodextrin biosurfactants
Should pose less risk than flood-application of surface-active compounds due to the
reduced likelihood of contaminant escape beyond the transpiring plant’s hydraulic control
and outside of the metabolically active rhizosphere. Cgt-plants offer the potential for
accelerated pollutant detoxification and ecological remediation via biophysical

engineering of contaminated soils.

30

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44

CHAPTER I

APPLICATION OF CYCLODEXTRIN AMENDMENTS FOR ENHANCED PAH

PHYTOREMEDIATION OF COKE-OVEN SOILS

45

INTRODUCTION

Polycyclic aromatic hydrocarbons (PAHS) are classified as persistent organic
pollutants due to their resistance to degradation, widespread distribution in the
environment, and suspected carcinogenicity. PAHS consist of two or more fused benzene
rings in linear, angular, or cluster arrangements and sorb strongly to soil organic matter
(SOM) due to their hydrophobicity. Partitioning of nonpolar organic compounds from
water to soil organic matter is the primary process for solid phase sorption of these
compounds (Chiou et al., 1998; Chiou et al., 1984). PAHS are formed from natural
processes during forest and brush fires and by anthropogenic activities, including
combustion of fossil fuels, coal processing during coke production, and manufactured gas
plant operations. The US. Environmental Protection Agency lists PAHS in 600 of 1,408
National Priority List (N PL) Superfimd Sites (ATSDR, 1995) and these pollutants are
additionally prevalent at industrial brownfields (EPA, 2004), dredged sediments (EPA,
2004), and numerous military properties (GAO, 2001).

Various strategies have been developed for the removal of PAHS from
contaminated soils, including physical, chemical, or biological treatment technologies.
Conventional engineering-based approaches such as containment and extraction
strategies have been routinely used to treat contaminated sites (Cunningham and Berti,
1993). Extractive techniques include excavation of hazardous waste materials, soil
washing, vapor extraction, addition of solidifying agents, and burial in secure landfills.
These approaches may be extremely expensive. The estimate cost of PAH incineration
ranges from 8300-440 per ton compared to 825-100 per ton by phytoremediation

treatment (Ensley, 2000; EPA, 20013; EPA, 2001b). Bioremediation is the utilization of

46

biological organisms, commonly microorganisms, or biological processes to detoxify
pollutants. PAH biodegradation has been observed in a wide array of bacterial genera and
various fungal species. PAH biodegrading bacteria use dioxygenase enzyme activity to
incorporate two oxygen atoms into the PAH ring structure to produce cis-dihydrodiols,
which may be ultimately converted to tricarboxylic acid cycle intermediates (Cerniglia,
1993). Biodegrader bacteria may metabolize two- and three-ring PAHS as a carbon
source (Ahn et al., 1999; Daane et al., 2001), though with increasing ring number, the
rate and extent of degradation decreases.

Bioremediation effectiveness is limited by the low aqueous solubility of PAH
compounds (Alexander, 2000; Comelissen et al., 1998), which is also a function of soil
organic matter content (SOM) and soil residence time of the contaminant (i.e.
weathering) (Alexander, 2000; Hatzinger and Alexander, 1995; Sun et al., 2003), and the
Specific PAH generation process (Luthy etal., 1997). PAH residues generated during the
thermal processing of coal during the manufacture of coke for use in smelting may be
particularly stable due to their solidified, refractory soil aggregate structure (Ghosh et al.,
2001; Hong et al., 2003). In spite of the limited solubility and bioavailability of organic
contaminants, some laboratory experiments indicated that microorganisms might access
sorbed contaminants (Guerin and Boyd, 1992; Park et al., 2001; Park et al., 2001).
Additionally, some bacteria were found to utilize sorbed naphthalene and biphenyl by
direct attachment to soil particles more rapidly than predicted by the contaminant
desorption rates from the sediment matrix (Calvillo and Alexander, 1996; Guerin and

Boyd, 1997). While some bacteria appear to utilize solid phase PAH compounds,

47

facilitated solubilization might increase bioavailability to a broader spectrum of
microbes.

A subcategory of bioremediation is phytoremediation, or the use of green plants
and their associated microbial communities to immobilize, sequester or degrade
contaminants in soil and ground water (Cunningham and Ow, 1996). Phytoremediation
has been intensively evaluated against a wide range of hazardous chemicals (EPA,
2001a) and has been applied at an increasing number of field trials and Site installations
(2001a; EPA, 2005). Heavy metal and inorganic contaminants are phytoextracted, or
accumulated to high shoot concentrations by selected plants for harvest and disposal or
recycling (Chaney et al., 1997). By contrast, organic, or carbon-based, pollutants may be
degraded in plant tissues or detoxified in situ in the rhizosphere via root-microbe
symbiotic processes (Schnoor et al., 1995). Several plant Species have been shown to
enhance biodegradation rates of persistent organic pollutants such as PAHS and pesticides
(Crowley et al., 1997; Gunther et al., 1996; Liste and Alexander, 2000; Perkovich et al.,
1996; Reilley et al., 1996). In rhizosphere soils, PAHS may be directly transformed by
biodegraders or by cometabolism among nonspecific heterotrophic bacteria (Corgie et al.,
2003; F ocht, 1995; Nedunuri et al., 2000; Pradhan et al., 1998). Root secretions and dead
or decaying cells provide soluble organic substances such as sugars, amino acids, and
organic acids, which increase microbial growth and activity. Root exudates contain plant
secondary metabolites, such as terpenes or phenolics, which may influence the metabolic
activity of microbes, leading to increased degradation of xenobiotic compounds via
cometabolism (Schwab et al., 1995). Addition of carbon sources such as glucose or

organic acids has been demonstrated to increase both soil microbial cell density and

48

enzymatic activity (Alden et al., 2001; Crane and Novak, 2001). In the presence of
artificial root exudates containing glucose, fructose, sucrose, succinic acid, malic acid,
arginine, serine and cysteine, soil PAH dissipation rates were elevated (Joner et al.,
2002). Corn root exudates were demonstrated to significantly accelerate mineralization of
CM-pyrene (Yoshitomi, 2001) and petroleum hydrocarbons (Chaineau et al., 2000).

PAHS are stably sorbed to soil organic matter, where they are not readily available
for microbial biodegradation and bioremediation. Synthetic surfactants, or detergent-like
compounds, have been demonstrated to enhance desorption and solubilization of sorbed
contaminants in soil, thus making the compounds available for microbial metabolism
(Aronstein et al., 1991; Lowe et al., 1999; Macur and Inskeep, 1999; Sun and Boyd,
1993; Sun etal., 1995). Surfactant enhancement of organic contaminants solubilization is
typically most effective at concentrations above the critical micelle concentration (CMC).
At elevated concentrations, surfactants may be inhibitory to microbial activity, form
highly viscous emulsions that are difficult to remove, or be prohibitively expensive to use
in large-scale applications. Some microbes have been shown to enhance hydrophobic
pollutant bioavailability by synthesizing surface-active biosurfactants that solubilize
hydrophobic molecules in the aqueous phase (Kanga et al., 1997). Biosurfactants can be
glycolipids, lipopeptides, polysaccharide protein complexes, phospholipids, fatty acids or
neutral lipids (Hommel, 1990; Maier, 2003). Cyclodextrins (CDs) are circular
oligosaccharide biosurfactants with hydroxyl-rich, hydrophilic external surface and a
hydrophobic interior cavity. Cyclodextrins are formed by enzymatic degradation of starch
by cyclodextrin glycosyltransferase (CGTase) circularizing the fragment via a a—l, 4

linkage. The common CDS consist of 6, 7, or 8 glucose units and are designated a-, B-,

49

and 'y-cyclodextrins respectively (Szejtli, 1991). The number of glucose units determines
the nonpolar, interior cavity size, which confers the specificity and range of complexation
with organic “guest” molecules, subsequently enhancing their dissolution into the
aqueous phase. Among CDS, B-cyclodextrin (BCD) is the most extensively studied due to
its effectiveness and low price. CD5 and modified CDs have been shown to enhance
solubility and extractability of various organic and metal contaminants from polluted or
Spiked soils (Brusseau et al., 1994; Brusseau et al., 1997; K0 et al., 1999; K0 and Y00,
2003; McCray and Brusseau, 1998; Morillo et al., 2001; Tanada et al., 1999). Moreover,
the presence of cyclodextrin applications were reported to increase biodegradation rate of
the PAH compound phenanthrene (Wang et al., 1998). Unlike other biosurfactants,
cyclodextrins do not require a threshold CMC for activity, are less microbiocidal, and are
more readily biodegraded in environmental settings.

Based on published literature, cyclodextrins appear to offer an effective method to
enhance PAH biodegradation. The purpose of this study was to determine whether
exogenous BCD amendments effectively accelerate PAH biodegradation rates in
rhizosphere soil treatments. Growth chamber and greenhouse trial experiments were
performed to test the effects of [3CD amendments on PAH phytoremediation in impacted
soils planted with plant species previously reported to enhance organic pollutant
biodegradation: alfalfa (Medicago sativa), mulberry (Morus alba), monkey flower
(Mimulus ringens), big bluestem (Andropogon gerardii), and little bluestem (Andropogon
scoparius) (Pradhan et al., 1998; Wan, 2002). The results from this work will advance

our understanding of processes involved in PAH-phytoremediation and test the efficacy

50

of biosurfactant amendments in rhizosphere-assisted bioremediation of persistent organic
pollutants.
MATERIALS AND METHODS

In vitro PAH-Cyclodextrin Sorption/Desorption Analysis
Phenanthrene-Cyclodextrin Sorption Isotherm

Sorption isotherm trials were performed to examine the influence of cyclodextrin
(CD) upon PAH partitioning from an aqueous phase to a soil solid phase system.
Phenanthrene was selected as a representative PAH compound for this study.
Phenanthrene sorption experiments were conducted with a Blount A soil (10% organic
matter, 24% sand, 34% silt, 42% clay, and CEC 24.3mequiv/ 100g of soil). Phenanthrene
adsorption isotherms were obtained using a batch equilibrium technique. l4C-
phenanthrene, unlabeled phenanthrene, and analytical grade or, [3, and yCD stock
chemicals were purchased from Sigma Chemicals (St. Louis, MO) and commercial grade
[3CD (cBCD) was obtained from Wacker Chemicals (Adrian, MI). A serial dilution
mixture of 14C- and unlabeled phenanthrene was prepared in 20ml deionized, distilled
water (ddHZO) or in various concentrations of orCD (5, 10, and 20mM), [3CD (05,10,
and 5.0mM), yCD (05,10, and 5.0mM), or cBCD (5.0mM) in ddHZO and was
equilibrated with 0.06g Blount A soil. Soil-solution mixtures (each combination, N = 4
replicates) were equilibrated at 25°C in Teflon-lined, screw capped, 20ml amber vials for
7 days on a mechanical rotator at IOrprn. Samples were centrifuged at 2000rpm for 5 min
to remove soil particulates. To determine the concentration of l4C-phenanthrene at
equilibrium, 1.0ml aliquot of supernatant solution was mixed with 10ml of scintillation

liquid and its activity was measured on a Packard liquid scintillation counter Model Tri-

51

CARB 1500 (Packard Instrument Co., Downers Grove, Illinois). The amount of
phenanthrene adsorbed was determined by the difference between the amount of 14C in
the initial solution and that remaining in solution after equilibrium with the soil.
Phenanthrene sorption data were fitted to the linearized sorption equation, C5 = KdCe",
where Cs is the amount of sorbed phenanthrene (pg/g); Cc is the equilibrium
concentration of phenanthrene in solution (mg/L), and K., is the partition (or distribution)
coefficient. Kd is defined as the ratio of the mass of the particular compound sorbed to
soil and mass of the compound remaining in solution at equilibrium.
Cyclodextrin-Amended PAH-Desorption Soil Analysis

For PAH desorption analysis, sterile soil was prepared by y-irradiating ~200g
coke oven soil (Rouge Manufacturing Complex, Dearbom, MI) aliquots in 2500c amber
jars at the Phoenix Memorial Emission Source Facility (Ann Arbor, MI). Soil total PAH
content (tPAH) was determined via Accelerated Solvent Extraction (ASE 200, Dionex
Corporation, Sunnyvale, CA) and high-pressure liquid chromatography (HPLC) analysis.
For PAH extraction, glass fiber filters were placed at the bottom of stainless steel ASE
cells. 10g PAH-soil was mixed thoroughly with 10g ammonium sulfate and transferred to
the ASE cell. Ottawa sand was used to fill the remaining volume of the cell, which was
then capped. The ASE 200 was operated through its automatic extraction cycles as per
manufacturer’s instructions. At the end of the extraction, extracts were transferred to pre-
weighed 60ml Amber vials with Teflon liners. Soil extracts were analyzed via high-
pressure liquid chromatography (Perkin-Elmer, Wellesley, Massachusetts) with a Waters
2487-dual wavelength absorbance detector and a Waters 474-scanning fluorescence

detector. Samples were run over a SUPELCOSIL LC-PAH 58229 Col: 2351203 column

52

under the following conditions: flow rate = 1.5ml/min; temperature = 30°C; injection
volume = 10111; mobile phase = acetonitrile2water at a linear gradient from 40:60 to 100:0
during the run time of 75min per sample. Solubilized phenanthrene content was
determined by HPLC-Fluorescence at excitation 249nm and emission wavelength 362nm
(McElmurry and Voice, 2004). Phenanthrene solubilization was monitored as a surrogate
PAH compound due to its relatively high concentration in the irradiated soil (8.72

:0.78ug/ g) and suitable relationship between aqueous solubility (1.10ug/mL) and

instrument detection limit (0.36ug/mL).

For the soil PAH desorption experiment, 1.0g aliquots of sterile, PAH-
contaminated soil were aseptically transferred to sterile 4m] vials and filled completely
with ~3.3mL sterile solutions of 10mM BCD or ddHZO-only control treatments (N = 5
replicates per mixture ratio) and sealed by screw caps with PTFE/silicone liners. After 2
and 4 weeks of mixing via mechanical rotator wheel (10rprn), the vials were removed and
centrifuged at 2000rpm for 7min. 1.0ml of supernatant was transferred aseptically to an
HPLC vial with cap and Teflon liner, the remaining supernatant was discarded, and
reactor vials refilled with the original treatment solution, either pure ddeO or 10mM
BCD, and returned to the rotator. Harvested samples were centrifuged at 3000rpm for
25min and aqueous supematants analyzed for phenanthrene content via HPLC as
previously described.

Growth Chamber-Scale Cyclodextrin-Amended Phytoremediation Trial
Soil Preparation
PAH-contaminated soil was obtained from the coke oven facility site of the

Rouge Manufacturing Complex (Dearbom, MI). The “Rouge-soil” contained 13%

53

organic matter, 81% sand, 11 % silt and 8% clay with a summed total of 14 PAH
compounds (tPAH) average concentration of 718.7 :17.1 (SE) ppm soil DW. PAH-soil
was passed through a stainless steel sieve (2.36mm mesh), to remove debris and ~500g
transferred to lOOOmL glass beakers. To improve bulk soil drainage after watering,
beakers were filled first with 200g of 5mm glass beads and secondly with another 200g
of 3mm glass beads.
Plant Materials

Two plant Species, Mulberry (Morus alba) (field Specimen, East Lansing, MI) and
big bluestem (Andropogon gerardii) (W ildtype Native Plant Nursery, Mason, M1), were
chosen for growth chamber PAH phytoremediation trials. 1-month old big bluestem
seedlings and 1-month old mulberry stem cuttings were transplanted to soil-filled
beakers. For unplanted control treatments, a wooden dowel (8mm diameter, 300m length)
was placed in the center of each filled beaker as an abiotic surrogate plant stem for
similar influence on soil water infiltration. 50ml 5mM [3CD (0.25 mg/kg soil FW) or
ddHZO was applied for per beaker every week. The growth chamber phytoremediation
study was conducted at 22+/-3°C with 16 hr day cycle (60-90 uE.s"m'2). All treatments
were fertilized with 50m] l/2X Murashige and Skoog (MS) media at 2-week intervals over
the course of the study. Watering was provided at 2-4 day intervals with weighing before
and after watering with sufficient water to return each 500g soil treatment pot to
gravimetric field holding capacity (20% soil moisture).
Sample Processing

At 1-, 2- and 4-monthS after planting, three 0.5cm diameter soil cores were

collected from each beaker (N = 4 for each treatment combination) and the cores pooled

54

and mixed. Sampled core holes were filled with sand to maintain normal soil percolation
and gas exchange and marked with toothpicks to avoid resampling at the same location.
Pooled soil core samples were sieved (2.36mm stainless steel mesh) and triplicate
subsamples of 1.0g for each replicate beaker were extracted in 10.0ml dichloromethane
(DCM) and 3.0m] saturated KCl in 20ml amber vials by vortexing 208ec, sonication
10min, and 16hr shaking on a rotating platform at ~150 rpm. The organic phase was
filtered and quantified for PAH content via gas chromatography-flame ionization
detection (GC-F ID). GC-F ID analyses were performed with an Agilent 6890 Gas
Chromatograph (Palo Alto, CA) equipped with an Agilent 3396B/C integrator and
Agilent 7683ALS auto injector under the following conditions: AT-S capillary column
30m x 0.53mm i.d. 1.20pm film thickness (Alltech, Deerfield, IL) with helium carrier
gas, injector temperature 270°C, flame ionization detector at 330°C, column initial
temperature at 100°C for 1min followed by elevation at 100°C/min to 310°C, and an
injected sample volume of 5 ul.

At the final time point, 4 months after planting, all beaker-pots were destructively
harvested and plant materials were removed and separated into roots, leaves and stems
for mulberry and roots and Shoots for big bluestem. Tissue fresh weights (F W) were
recorded and tissues oven-dried at 80°C for >72h for dry weight (DW) determination.
Dried plant tissues and soil samples were analyzed for agronomic nutrient content at
A&L Great Lakes Analytical Labs (Fort Wayne, IN).

Greenhouse-Scale Cyclodextrin-Amended Phytoremediation Study

55

 

Soil Preparation

PAH-contaminated soil (tPAH ~500ppm) was obtained from the Rouge coke
oven area. Uncontaminated soil was obtained from the Michigan State University Plant
Sciences greenhouse facility. Soils were sieved through a 5mm screen to remove debris
and sieved coke oven soil was mixed with uncontaminated soil (“Clean-soil”) at a ratio of
1:2 v/v respectively (“PAH-soil”) for a final concentration of 135.5ppm tPAH (soil DW).
Agronomic and physical properties of the individual and mixed soils were determined by
A&L Great Lakes Analytical Labs (Fort Wayne, IN) (Table 1.1). Clean- and PAH-soils
were transferred to 1.5L plastic pots (9cm top diameter x 24cm tall) (Hummerts, Earth
City, MO). To assist with drainage, the bottom of each pot was lined with a fitted piece of
plastic window screen (2mm mesh), then ~2cm of pea gravel, then a second screen, and
finally overlaid with a piece of paper towel to prevent soil particulate escape or blockage.
1.7kg of soil was added to each pot leaving approximately 1cm from the top unfilled.
Plant Materials

Experimental study soils were planted with species previously observed or
reported to enhance organic compound biodegradation: alfalfa (Medicago sativa)
(obtained from Dr. Sisir Dutta, Howard University), mulberry (Morus alba) (field
specimen cuttings, East Lansing, MI), monkey flower (Mimulus ringens) (Ernst
Conservation Seed, Meadville, PA), and little bluestem (Schizachyrium scoparium or
Andropogon scoparius) (Wildtype Native Plant Nursery, Mason, M1). Alfalfa, monkey
flower, and little bluestem were established from seed and the mulberry clonally
propagated from stem cuttings. Filled pots were planted with single plants of either 1-

month old plantlets of alfalfa, monkey flower or little bluestem or 2—month old rooted

56

 

Table 1.1. Soil agronomic and physical properties for materials used in phytoremediation

growth chamber and greenhouse studies.

 

 

Particle Size Total CEC
, OM P K
Soil (4’) pH N (meq
(W 0 (ppm) (ppm)
sand silt clay (A) [1003)

 

Rouge-soil 78 18 4 4.5 7.5 0.43 118 106 33.6

 

Clean-soil 80 12 8 2.7 7.3 0.10 9 28 14.7

 

PAH-soil 78 14 8 2.4 7.6 0.17 64 43 17.8

 

 

 

 

 

 

 

 

 

 

57

 

cuttings of mulberry. Bacterial bioremediation controls consisted of unplanted pots filled
with PAH-soil mix. The study was conducted under greenhouse conditions at 22+/-3°C
with 16hr supplemental lighting (100-300 uE.s'lm'2 +/-35-75 uES'lm'z). All treatments
were fertilized at 2-week intervals with lOOmL ~325 ppm Peters 20-20-20 N-P-K
solution (The Scotts Company, Marysville, OH) and watered as needed. Clean-soil and
PAH-soil pots were distributed across four greenhouse benches in a completely
randomized block design for all plant species X B-CD amendment combination
treatments.
,BCD application

[3CD (commercial grade BCD, Wacker Chemicals, Adrian, MI) was prepared at
1.25, 5.0 and 10.0mM dilutions in diH20. At weekly intervals over the course of the
study, 100 ml of BCD solution (0, 1.25, 5.0 or 10.0mM) was added to the soil surface of
pro-labeled, designated pots using a plastic graduated cylinder.
Plant and soil sampling

At 2-months and 4-months after planting time points, four replicate pots for each
treatment were harvested by randomly selecting pots from the greenhouse benches. After
harvest, shoots were removed from the pots and stored in individual, pre-labeled paper
sacks. For soil collection, one side of the plastic pot was removed by cutting it away with
a utility knife. Soil was collected from the middle third (~4-8 inch depth) of the soil
column, passed through a stainless steel sieve (2.36mm mesh), and transferred to one
each 20ml and 40m] amber vials. Roots were obtained from the whole pot by manual

collection or physical separation by shaking, sieving and washing off remaining soil.

58

 

Root and shoot fresh weights were obtained immediately upon harvest and bagged
samples were oven dried at 80°C for >72 hr for dry weight determination.
Soil PAH analysis

For soil PAH determination, 6.0g sieved soil was extracted with 20.0ml
dichloromethane (DCM) and 6.0 m1 saturated KCl in 40ml amber vials. Extract mixtures
were vortexed for 205cc, sonicated for 10min, and placed on a rotating shaker (~150rpm)
for 16hr; after which the organic phase was decanted, filtered, and quantified for PAH
concentration by GC-F ID as described previously.
Statistical Analysis

Statistical significance for multiple parameter data (e. g. soil tPAH X plant species
X [3CD levels) was determined using SAS version 8.2 (SAS Institute Inc., Cary, NC) with
one-way AN OVA and least significant difference (LSD) for comparison of treatment
means with p S 0.05. Different treatments (e. g. plant biomass +/- BCD) were statistically
compared by t-test analysis with StatView Version 5.01 (SAS Institute Inc., Cary, NC).

RESULTS AND DISCUSSION

In vitro PAH Sorption/Desorption Analysis
Phenanthrene-Cyclodextrin Soil Sorption Isotherm

Soil-aqueous phase equilibrium isotherms of phenanthrene were determined for
different concentrations of analytical grade or, [3, and 7CD and a bulk-supplier source of
commercial-grade yCD (cyCD). The equilibrium partition coefficient (K.,) of
phenanthrene between soil and water was 622 (mL/ g), which decreased by up to 80% in
the presence of 5mM BCD, yCD, and cyCD (Figure 1.1). By contrast, ctCD was observed

to decrease K; by only 20% at 20mM and no decrease for Kd was observed at lower

59

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aqueous chD concentrations. cBCD was shown to effectively decrease the amount of

phenanthrene sorbed to soil comparably to the analytical grade cyclodextrins. cBCD at
$11/kg is ~100 times cheaper than the analytical grade CDs and therefore used as the
biosurfactant amendment in the phytoremediation studies.
Cyclodextrin-Amended PAH-Soil Desorption Analysis

The tPAH content of the y-radiated Rouge soil used for the desorption study
wasl26ppm with phenanthrene Specifically monitored due to its sufficient soil
concentration (8.7 :0.8ug/ g) to be observable relative to its aqueous solubility (1.1ug/ml)
and HPLC analytical method reporting limits (0.36ug/ml). In contrast to the sorption
study, which examined removal of aqueous phase phenanthrene by soil sorbent, this
experiment used weathered, PAH-impacted soil material to test for transfer to the
aqueous phase. Batch experiments were performed to determine the ability of [3CD to
enhance desorption of PAHS from aged Rouge soil. After 2 and 4 weeks of incubation, no
phenanthrene was observed in 10mM [3CD or ddHZO aqueous phase in the batch reactors.
This result indicates that 10mM [3CD was not capable of desorbing 315% of the sorbed
phenanthrene from the coke oven residue material in this trial. Desorption analysis
demonstrates that PAHS are stably sorbed to the coke oven material and likely
recalcitrant to rapid biodegradation, though the possibility exists that BCD amendments
may still enhance PAH phytoremediation over time. Numerous publications report the
ability of BCD and modified cyclodextrins to enhance desorption and solubilization of
organic contaminants in soils (Brusseau et al., 1994; K0 and Y00, 2003; McCray et al.,
2000; Morillo et al., 2001; Perez-Martinez et al., 2000; Sheremata and Hawari, 2000).

However, additional studies demonstrated irreversible sorption of PAHS to SOM with

61

further decline in aqueous solubility over time (Alexander, 2000; Chung and Alexander,
1998; Ke et al., 2003; White, 1996), which may also make coke-sorbed PAHS recalcitrant
to BCD-enhanced solubilization and limiting for endogenous bacterial or rhizosphere-
assisted biodegradation.
Cyclodextrin-Amended PAH thremediation Experiments

In the growth chamber phytoremediation trial, [3CD amendments were observed
to significantly reduce tPAH content in mulberry rhizosphere soils relative to unplanted
and big bluestem rhizosphere soils after 4 months treatment (Figure 1.2). [3CD amended
mulberry planted soils decreased tPAH concentration 20.2% from the starting level,
relative to a decrease of only 7.2% in unamended mulberry rhizospheres and
approximately 8.5% for both amended and unamended big bluestem treatments. In
contrast, unplanted treatments Showed no reduction over the course of the 4-month
experiment with or without BCD. From this preliminary study, we concluded that [3CD
may enhance PAH phytoremediation in coke oven soils in mulberry planted soils.

Under greenhouse conditions, BCD amendments were demonstrated to enhance
PAH dissipation rates in most treatments (Figures. 1.3 and 1.4). No apparent trend of
tPAH reduction relative to plant species was observed for the planted treatments at either
2-month or 4-month sampling events. Unplanted soil tPAH concentration was unchanged
by ddH20 control applications, however unplanted soils amended with 1.25, 5.0 or
10.0mM [3CD decreased soil tPAH levels 6.2%, 11.1% and 11.1% after 2 months and
12.7%, 12.9%, and 22.1% respectively after 4 months treatment. There was a Significant
reduction of tPAH in unplanted soil at 10mM [3CD amendment level after 4 month,

though this reduction was not statistically different from the starting concentration

62

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65

(ANOVA LSD p S 0.05). Although there was no direct effect of plant species on tPAH
reduction, plant roots may help accelerate tPAH dissipation by improving [3CD
infiltration and diffusion of oxygen through soil. After 4 months of [3CD amended, most
planted treatments showed an apparent decrease in soil tPAH concentrations with a few
species-[3CD combinations displayed statistically significant soil tPAH reduction: alfalfa
+ 10mM BCD, little bluestem + 10mM BCD, unamended monkey flower, and monkey
flower + 5.0mM [3CD amendments. 10mM BCD amendment significantly reduced soil
tPAH from the starting concentration for alfalfa (29.6% reduction; LSD 5 0.05) and little
bluestem (25.8% reduction; LSD 5 0.05) grown soils. It is possible that there may be a
plant specific optimal [3CD application rate, though this can not be determined from this
study due to the uneven PAH reduction results across BCD amendment levels for each
species. BCD-enhanced tPAH reduction in unplanted treatments indicates that
endogenous bacterial biodegradation activity may have increased in response to [3CD
applications.

The soil PAH data in this study indicates that the highly sorptive nature of the
coke oven material limited observable plant-mediated reductions, which was exacerbated
by contaminant heterogeneity in the PAH-Soil mix. Soil PAH volatilization or leaching
are potential sources of loss, the latter of which could possibly be enhanced by
biosurfactant addition. However, the stabilizing effect of longtime weathering (~60+
years) of the coke oven material would strongly minimize PAH loss through either of
these pathways. The disappearance of PAHS was found to decline over time with little or
no loss of the compound detected after several years. Other studies also have been

reported irreversible sorption of PAHS to SOM, and some show that it is chemically

66

incorporated in to SOM (Ke et al., 2003; Alexander, 2000; Chung and Alexander, 1998;
White and Alexander, 1996). As observed in the BCD desorption study, the presence of
the biosurfactant did not detectably enhance phenanthrene transfer into the aqueous phase
indicating leaching is highly unlikely to be a route of soil PAH loss from the open-bottom
pots in the greenhouse study. Moreover, complexation of BCD and gust molecules
protects its complexes from volatilization (Szejtli, 1988). Subsequently, enhanced
biodegradation was observed for some planted and unplanted treatments demonstrating
that cyclodextrin-amendments may enhance biological treatability of coking residue
PAHS.
flCD-amendment influence on plant growth

It is essential to evaluate phytotoxicity or other physiologically stressful effects of
biosurfactant amendments being considered for use in phytoremediation treatments.
Biosurfactants may enhance soil pollutant aqueous solubility or bioavailability, possibly
to phytotoxic levels. It is also possible that biosurfactant compounds may negatively
affect plant health, root-microbe interactions, or overall plant vigor. During the growth
chamber phytoremediation study, [3CD addition was observed to minimally reduce big
bluestem shoot biomass, while mulberry whole plant biomass was significantly reduced,
particularly for shoot growth (Figure 1.5a). These data reflect the negative influence of
the BCD amendment on mulberry and big bluestem leaf development relative to root
growth, the latter of which was only marginally reduced in the presence of the
biosurfactant subsequently lowering mulberry shoot-to-root ratio (Figure 1.5b). [3CD

increased the whole plant dry biomass to fresh biomass ratio (FW/DW) for bluestem

67

Figure 1.5a. Plant tissue biomass

Biomass (g DW)

 

 

 

 

 

 

 

 

Root Shoot Root Shoot
Mulberry Big Bluestem
Figure 1.5b. Shoot / Root ratio (DW) Figurel .5c. Tissue DW / FW ratio
1.6 l ~ 7 7‘27 ,7 v -i 0.4
; ** LEtCB_P:12_..Jl -+CD [3+qu **
1.2 (r 7 V v I 0.3
i l
0.8 ~—~ 77-“ ,7 71 0.2
1 .
0.4 5 7 L} 0.1
I l
0.0 i , *r” l 0.0
Mulberry Big Bluestem Mulberry Big Bluestem

Figure 1.5. Growth chamber study: Mulberry and big bluestem biomass afier 4 months
+/- 0.5mM [3CD amendment in PAH-soil. (1.5a) root and shoot tissue biomass (DW),
(1 .5b) compares ratios of shoot biomass (DW) to root biomass (DW), (1.5c) compares
ratios of whole plant DW / F W. Significant differences between BCD-amended and

ddH20 control treatments indicated by * (p 5 0.10), ** (p 5 0.05).

68

plants though not for mulberry (Figure 1.5c), possibly indicating a different affect on
tissue hydration physiology between monocots (bluestem) and dicots (mulberry).

In the 4-month greenhouse study, [3CD amendments were observed to inhibit little
bluestem and mulberry grth with no effect on monkey flower and alfalfa biomass
(Figure 1.6). Relative to the ddl-120 control treatment, 10mM BCD amendments reduced
final (4mo) shoot biomass (DW) of little bluestem 45% in Clean-soil and 61% in PAH-
soil and mulberry shoot biomass 49% in Clean-soil and only 34% in PAH-soil. Root
biomass among all tested plant species was unaffected by any BCD-soil treatment
combinations. Published reports of [3CD phytotoxicity have given somewhat
contradictory results. Barley seed germination was enhanced 162% in the presence of
BCD, though maize and rye seed germination and plantlet growth were inhibited ~70%
and 60% respectively (Szejtli, 1988). Conversely, cyclodextrin-containing seed dressing
was shown to improve seed germination rate and eventual plant shoot mass in lettuce,
celery, and paprika. No differences in amylase activity, DNA or protein biosynthesis, or
potassium uptake were found between water and cyclodextrin treated seeds in published
reports (Szejtli, 1988).

In a separate study, aBD and [3CD used at low concentrations and in combination
with indolebutyric acid was shown to stimulate seed germination and early root
development in the poorly germinating species jojoba (Simmondsia chinensis) (Apostolo
et al., 2001). Published reports in combination with the experimental results from this
laboratory study indicate that cyclodextrin compounds appear to interact with plants in a

species-specific manner, however the mechanisms for these effects are not understood.

69

Figure 1.6a. Clean-soil

 

 

 

 

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Figure 1.6b. PAH-soil
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(1.6a) Clean-soil or (1.6b) PAH-soil with BCD amendment concentrations as indicated. *

Significant differences relative to OmM (i.e. ddHZO control treatment), p S 0.05.

70

Many chemical pollutants possess phytotoxic qualities, including PAHS as demonstrated
by germination reductions in garden cress (Lepidium sativum) seed assays (Maila and
Cloete, 2002), which would unlikely be the case for weathered, coking-derived materials
due to their stable sorption to the material matrix (Ghosh et al., 2001; Talley et al., 2002).
In the greenhouse phytoremediation study, there was no difference in biomass between
plants grown in Clean-soil versus PAH-soil for any of the species at either 2- or 4-month
treatment times.
,BCD-amendment influence on soil and plant nutrient content

Since cyclodextrins may interact with inorganic as well as organic chemicals, it is
possible that BCD amendments could affect mineral nutrient bioavailability and
subsequent plant biomass and physiological status. Afier the 4-month sampling event of
the growth chamber phytoremediation study, mulberry and big bluestem plants and
planted and unplanted soils were harvested and subsamples analyzed for nutrient content
and soil agronomic nutrient bioavailability. Significant differences were observed in soil
nutrient status for various plant X [3CD treatment combinations (Table 1.2). Bioavailable
soil nitrogen (N) was reduced by the addition of [3CD in each of the planted or unplanted
treatments, significantly for unplanted and mulberry pots. Agronomic N was greatly
reduced (p_<_0.001) in the BCD amended, unplanted pots with reduction of >80% soil N
relative to the ddeO control. Big bluestem planted soils were also reduced in agronomic
N content compared to the unamended, unplanted treatment (102 :1 Sppm soil DW) with
BCD-amended (12 :4ppm) and unamended (18 :7ppm) bluestem soils. Agronomic N
was reduced in mulberry planted soils to less than 6% of the unplanted, ddH20 treated

soil N with significantly less available N in BCD-amended mulberry treatments (2.3

71

 

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i0.3ppm) relative to unamended mulberry pots (5.8 10.9ppm). Agronomic phosphorous
(P) was also decreased 15-20% in all soils receiving BCD relative to the unplanted control
soils, which was much less than the 60-80% decrease observed for N. In contrast to the
reduction in bioavailable N and P, [3CD enhanced bioavailable soil concentrations of
other macro- and micronutrients, including potassium (K), sulfur (S), manganese (Mn),
and iron (Fe). BCD-influenced nutrient differences were more common between the
unplanted soil treatments, due possibly to active uptake of available nutrients by the
growing plants.

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soils, which may be a result of the significantly lower leaf tissue concentrations of N, K,
and zinc (Zn) relative to ddH20 control treatment plants (Table 1.3). Mulberry root tissue
K and Zn content was not significantly (p>0.10) reduced by added BCD. In contrast to
mulberry, BCD amendments altered big bluestem root nutrient content with little effect
on leaf tissue nutrient status (Table 1.4). Nutrient content in big bluestem leaves was
minimally affected by addition of [3CD with only moderate (p50.01) reductions in leaf
calcium (Ca) and Zn. However, [3CD amendments were observed to significantly perturb
big bluestem root nutrient composition with decreased N and Mg and increased Mn and
Fe concentrations relative to ddHZO control treatments.

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in the soil C—N ratio with possible consequences for microbial and plant growth and
metabolism. Utilization of [3CD as a readily available carbon source may have increased
soil microbial biomass and overall metabolic activity resulting in nitrogen consumption

and subsequent N-limited plant growth, particularly as observed for mulberry soil and

73

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23:0 mam: ”2::

.236 w 3 .2. .8: .o w 5 .. .3 3865

00:08:00.: 280:8 20:0 2002080500 :00300 000580.220 0:85:w2m .39 0:32: :0: 23:05::80280 wucwn: :o 2355:5895

o\.. 00 :0>2w 00:23, .5030 :02002200808300A200 0080.20 036% :80 0:00:00 2:00:28 0:053: 0:002: 0:020 80:00:20 w2m2 .V.2 0200.2.

75

leaf tissue nutrient status in [3CD amended soils. To compensate for apparent plant
nutrient deficiencies in the growth chamber experiment, 375ppm Peters 20-20-20
fertilizer was used to increase N, P, and K application rates in the greenhouse study ~5-
fold, 37-fold, and llO-fold respectively, relative to the grth chamber study. In addition
to other species-specific plant-[3CD interactions, mineral nutrition application rates may
also need to be experimentally optimized for each species to enhance biosurfactant-
amended rhizoremediation performance.
CONCLUSIONS

Chemical and biological surfactants have been utilized to enhance hydrophobic,
organic pollutant aqueous solubility and in some instances accelerate biodegradation.
Due to the shorter environmental stability of most biosurfactants, they are thought to be
less ecological harmful than chemical surfactant amendments. We performed a growth
chamber-scale and larger greenhouse study of the effects of the starch-derived
biosurfactant BCD application on PAH rhizosphere-assisted biodegradation for a variety
of plant species. BCD amendments were demonstrated to enhance PAH biodegradation in
rhizosphere soils of mulberry in a growth chamber trial and for alfalfa, monkey flower
and little bluestem in a greenhouse study. Evidence of plant physiological and/or
nutritional stress was observed for some of the BCD amended treatments, thus further
studies will be needed to balance plant vigor, soil-plant nutrient status, and rhizosphere-

assisted degradation effectiveness.

76

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83

CHAPTER II

EFFECTS OF EXOGENOUS CYCLODEXTRIN APPLICATION ON

BACTERIAL COMNIUNITY STRUCTURE DURING PAH

RHIZOREMEDIATION OF COKE-OVEN SOILS

84

INTRODUCTION

Polyaromatic hydrocarbons (PAHS) are a class of persistent soil pollutants with
potential mutagenic or carcinogenic properties. PAHS are highly stabilized in soil
complexes by movement into soil micropores and tendency to sorb to soil organic matter
(Aprill and Sims, 1990; Ke et al., 2003). Biological degradation is thought to be the most
prevalent mechanism for PAH decomposition. PAH biodegradation has been observed in
a wide array of bacterial genera and various fungal species (Ahn et al., 1999; Amellal et
al., 2001; Cemiglia, 1993). PAH biodegrading bacteria use dioxygenase enzyme activity
to incorporate two oxygen atoms into the PAH ring structure to produce cis-dihydrodiols.
The cis-dihydrodiol is oxidized to form catechols, which may be further converted to
tricarboxylic acid cycle intermediates (Cerniglia, 1993). Some bacteria can efficiently
utilize two- or three-ring PAHS as their soil carbon source for grth and multiplication
(Ahn et al., 1999; Daane et al., 2001), though as the number of fused rings increases, the
rate and extent of degradation decreases. For PAHS with more than three fused rings, co-
metabolism is believed to serve as the major mechanism of biodegradation (Cookson,
1994). In one study, the presence of two- or three-ringed PAHS was demonstrated to
enhance the degradation of higher-ringed PAHS by tested bacteria (Schwab et al., 1995).
Fluoranthene degrading bacterial isolates were shown to metabolize pyrene, though were
only able to degrade fluorene, anthracene, benzo[b]fluorene, benzo[a]anthracene or
benzo[a]pyrene if cometabolically induced with the 3-ring PAH phenanthrene (Ho et al.,
2000). The 5-ring PAH, benzo[a]pyrene (BaP), was reported to be co-metabolized by a
number of microorganisms grown on other substrates or in soils containing suitable co-

substrates (Cerniglia, 1979; Gibson et al., 1975). Burkholderia cepacia has been reported

85

to cometabolize dibenzo[a,h]anthracene and BaP in the presence of phenanthrene (Juhasz
et al., 1997). BaP was been reported to persist in otherwise uncontaminated sediments up
to 60 years, while the turnover times could be as short as 3.3 years in oil-contaminated
sediments (Herbes and Schwall, 1978). CM-BaP mineralization in uncontaminated soil
was limited to 9% reduction, while 25% of the added CM-BaP was mineralized in
contaminated soil from a coal gasification site, due possibly to cometabolism supported
by the soil co-contarninants (Carmichael and Pfaender, 1997).

Phytoremediation is the use of plants and their associated microorganisms to
sequester, extract, or detoxify pollutants (Cunningham et al., 1996; Cunningham et al.,
1995; Rugh, 2004). Plant-based bioremediation has been effectively demonstrated for
various heavy metal (inorganic) and carbon-based (organic) pollutants. In addition,
plants roots colonize subsurface soils, facilitate water and air movement, extend
distribution of soil microorganisms, and regenerate soil structure with deciduous and
decaying biomaterials. Plant root exudates stimulate growth and activity of pollutant
degrading microbes, resulting in enhanced organic pollutant metabolism, a process
known as phytostimulation, root-assisted bioremediation, or rhizoremediation (Crowley
et al., 1997; Kuiper et al., 2004; Siciliano and Germida, 1998). Root exudates contain a
wide array of compounds, including sugars, amino acids, and secondary plant
metabolites. Various plant secondary compounds have been to shown to stimulate
microbial cell division and metabolic activity (Anderson et al., 1993; Chang and
Corapcioglu, 1998; Crowley et al., 1997; Curl and Truelove, 1986).

Due to highly stable sorption to soil organic matter, PAH compounds are not

readily available for microbial biodegradation and are consequently environmentally

86

persistent contaminants. Chemical surfactants, or detergent-like compounds, may be used
to increase the aqueous solubility and subsequent bioavailability of hydrophobic, organic
compounds. Synthetic surfactants have been demonstrated to enhance soil desorption and
solubilization of sorbed contaminants, thus making the substrates more readily
biometabolized (Aronstein et al., 1991; Lowe etal., 1999; Sun and Boyd, 1993). Various
bacteria species produce surface-active agents referred to as biosurfactants, which have
the advantage of being more biodegradable and less toxic. Cyclodextrins (CDs) are a
class of biosurfactants produced by bacterial enzymatic conversion of starch to create
cyclic oligosaccharides of six or more glucose units linked by or-l ,4 bonds. CD producing
bacteria are able to both convert starch and metabolize CDs as a utilizable carbon-source
(Feederle et al., 1996). CD8 have a doughnut-like shape with a hydrophobic interior and
hydrophilic exterior to form inclusion complexes with organic compounds, thereby
increasing the aqueous solubility of guest molecules. Numerous studies have shown that
cyclodextrins and modified cyclodextrins enhance aqueous extractability of various
organic and metal contaminants from polluted or spiked soils (Brusseau et al., 1994;
Brusseau et al., 1997; K0 and Yoo, 2003; McCray et al., 2000; McCray and Brusseau,
1998; Morillo et al., 2001; Murai et al., 1998; Sheremata and Hawari, 2000; Wang and
Brusseau, 1995). Studies of the effects of cyclodextrin on overall microbial growth and
activity have produced mixed results. Biodegradation and dechlorination of the PCB
mixture Aroclor 1221 was significantly enhanced by addition of hydroxypropyl-B-
cyclodextrin (HPBCD) in shaken liquid cultures of Pseudomonas sp. strain CPEl (Fava
et al., 1998). Toluene and p-toluic acid biodegradation by Pseudomonas putida was

enhanced in the presence of beta-cyclodextrin (BCD) as a result of decreased toxicity of

87

 

the compounds (Schwartz and Bar, 1995). In contrast, dimethylated BCD (Dimeb)
amendments were reported to be toxic to Mycoplasma and E. coli and inhibit grth of
Rhodococcus erythropolis (Bar and Ulitzur, 1994; Greenberg-Ofrath et al., 1993; J adoun
and Bar, 1993). It is conceivable that cyclodextrin could be sufficiently toxic to sensitive
microorganisms to inhibit rhizoremediation processes at elevated concentrations.

We conducted experiments testing the effect of [3CD amendments in PAH
contaminated rhizospheres of four different plant species: alfalfa (Medicago sativa),
mulberry (Morus alba), monkey flower (Mimulus ringens) and little bluestem
(Schizachyrium scoparium or Andropogon scoparious). Numerous studies have described
the ability of alfalfa to enhance rhizosphere microbial community structure due to its
highly branched root system and soil enrichment by nitrogen fixation and root exudation
(Brophy and Heichel, 1989; Kirk et al., 2004; Muratova et al., 2003). Mulberry has been
reported to contain various phenolic compounds in its root exudates, which were
demonstrated to enrich and accelerate microbial biodegrading activity in its rhizosphere
(Fletcher et al., 1995; Hegde and Fletcher, 1996). In previous experiments in our
laboratory, monkey flower and little bluestem have been demonstrated to enhance PAH
biodegradation and bacterial metabolic activity (Rugh et al., 2005; Susilawati, 2003;
Wan, 2002). This current study was performed to evaluate the effects of exogenous BCD
application on PAH phytoremediation effectiveness and bacterial community response in

coking-impacted soils.

88

MATERIALS AND METHODS

Soil Preparation

PAH-contaminated soil (total PAH concentration, tPAH ~500ppm soil DW) was
obtained from the Rouge coke oven area. Uncontaminated soil was obtained from the
Michigan State University Plant Sciences greenhouse facility. Soils were sieved through
a 5mm screen to remove debris and sieved coke oven soil was mixed with
uncontaminated soil (“Clean-soil”) at a ratio of 1:2 v/v respectively (“PAH-soil”) for a
final concentration of 135.5ppm tPAH (soil DW). Clean- and PAH-soils were transferred
to 1.5L plastic pots (9cm top diameter x 24cm tall) (Hummerts, Earth City, MO). To
assist with drainage, the bottom of each pot was lined with a fitted piece of plastic
window screen (2mm mesh), ~20m layer of pea gravel, a second screen, and finally
overlaid with a piece of paper towel to prevent soil particulate escape or blockage. 1.7kg
of soil was added to each pot leaving approximately lcm from the top unfilled.
Plant Materials

Experimental study soils were planted with species observed or reported to
enhance organic compound biodegradation: alfalfa (Medicago sativa) (Dr. Sisir Dutta,
Howard University), mulberry (Moms alba) (field specimen cuttings, East Lansing, MI),
monkey flower (Mimulus ringens) (Ernst Conservation Seed, Meadville, PA), and little
bluestem (Schizachyrium scoparium or Andropogon scoparius) (W ildtype Native Plant
Nursery, Mason, MI). Alfalfa, monkey flower, and little bluestem were established fi'om
seed and the mulberry clonally propagated from stem cuttings. Filled pots were planted
with single plants of either l-month old plantlets of alfalfa, monkey flower or little

bluestem or 2-month old rooted cuttings of mulberry. Endogenous bioremediation

89

 

controls consisted of unplanted pots filled with PAH-soil mix. The study was conducted
under greenhouse conditions at 22+/-3°C with 16hr supplemental lighting (at bench level:
100-300 |.LE.s'lm'2 +/-35-75 uE.s"m'2). All treatments were fertilized at 2-week intervals
with lOOmL ~325 ppm Peters 20-20—20 N-P-K solution (The Scotts Company,
Marysville, OH) and watered as needed. Clean-soil and PAH-soil pots were distributed
across four greenhouse benches in a completely randomized design for all plant species X
8CD amendment combination treatments. Clean-soil unplanted treatments were not used
in this study.
8CD application

100ml of either 5.0 or 10.0mM commercial grade BCD (Wacker Chemicals,
Adrian, MI) solution, which was equivalent to 334 and 667mg [3CD per kg soil (ppm
FW) respectively, was added at weekly intervals to the soil surface of pre-labeled,
designated pots using a plastic graduated cylinder on each application event. Control
treatments received only dinO.
Plant and soil sampling

At 2-months and 4-months after planting time points, four replicate pots for each
treatment were harvested by randomly selecting pots from the greenhouse benches. After
harvest, shoots were removed from the pots and stored in individual, pre-labeled paper
sacks. For soil collection, one side of the plastic pot was removed by cutting it away with
a utility knife. Soil was collected from the middle third of the soil column (between 4 and
8 inches depth), passed through a stainless steel sieve (2.36mm mesh), and transferred to
one each 20m] and 40m] amber vials.

Soil PAH analysis

90

 

For soil PAH determination, 6.0g sieved soil was extracted with 20.0ml
dichloromethane (DCM) and 6.0 m1 saturated KCl in 40ml amber vials. Extract mixtures
were vortexed for ZOsec, sonicated for 10min, and placed on a rotating shaker (~150rpm)
for 16hr; after which the organic phase was decanted, filtered, and quantified for PAH
concentration via gas chromatography-flame ionization detection (GC-FHD). GC-FID
analyses were performed with an Agilent 6890 Gas Chromatograph (Palo Alto, CA)
equipped with an Agilent 3396B/C integrator and Agilent 7683ALS auto injector under
the following conditions: AT-5 capillary column 30m x 0.53mm i.d. 1.20um film
thickness (Alltech, Deerfield, IL) with helium carrier gas, injector temperature 270°C,
flame ionization detector at 330°C, column initial temperature at 100°C for lrrrin
followed by elevation at 100°C/min to 310°C, and an injected sample volume of Sul. For
soil moisture determination, 5g subsamples were weighed into aluminum weigh-boats as
fresh weight (F W), oven-dried at 80°C for >72h, and re-weighed for dry weight (DW).
Soil Bacteria Metabolic Analysis

Treated soil samples were analyzed for both Clean— and PAH-soils using plate
culture methods to determine soil cell densities for total heterotrophic bacteria, bacterial
biodegraders of phenanthrene and pyrene, and cyclodextrin producing bacteria. Total
heterotrophic bacteria were quantified by standard plate count assay. Soil bacteria were
extracted by mixing 4.0g of soil in 36ml SPP buffer (0.1% tetrasodium pyrophosphate,
pH 8.0) by vortex at full power for 1min, pulse 305ec, and 1min. Serial dilutions of
bacterial extracts were plated on semisolid Yeast Extract-Peptone-Glucose (YEPG)
medium (per liter: 0.05 g yeast extract, 0.5 g polypeptone, 0.25 g D-glucose, 0.05 g

NH4N03, 1.5 % Bactoagar, pH 8.0; all components from Becton Dickenson and

91

Company, Sparks, MD; except NH4N03 from Sigma-Aldrich, St. Louis, MO). Plated soil
extracts were incubated at 25°C for 5 days and colonies counted for quantification of total
heterotrophic colony forming units (CFU).

The PAH spray plate assay (Spray-Assay) was used to quantify phenanthrene-
and pyrene-degrading bacterial cell density in treated soils (Ahn et al., 1999; Kiyohara,
1982). For quantification of phenanthrene degraders, soil extracts were prepared as
described previously and plated at dilutions estimated to achieve ~200 colonies on
Mineral-Bacto Medium (MBM recipe per liter: 2. 13 g NazHPO4, 1.3 g KH2P04, 0.5g
NH4Cl, 0.2g MgSO4.7H20, 15g Bacto Agar (Difco Laboratories), pH 8.0) (Kastner,
1998). Phenanthrene solution (1% w/v in acetone) was sprayed onto freshly plated soil
bacteria extracts on MBM plates using a thin-layer chromatography sprayer flask
assembly pressurized by the exhaust port (~20 psi) of a vacuum pump. Spray plates were
returned to plastic Petri dish bags to prevent desiccation or excessive PAH evaporation,
incubated for 10 days at 25°C, and degrader-positive isolates identified by formation of
clear zones in the cloudy residue around the colonies. Clear zone forming colonies were
scored as Zone Forming Units (ZFUs). The Spray-Assay was also used for quantification
of pyrene degraders, though with some modifications. YPEG medium was supplemented
with glucose (0.25 g/L) as a carbon-source cometabolite. Additionally, colonies were pre-
grown for 5 days prior to spraying with pyrene solution ( 1% w/v in acetone) and
incubated at 25°C for up to 30 days to quantify clear zone forming colonies in the cloudy
pyrene residue (ZFUpy).

Cyclodextrin producing (CDP) bacteria were identified by using selective

medium for colorimetric indication of starch conversion to cyclodextrin: 1% soluble

92

starch, 0.25% yeast extract, 0.25% tryptone, 0.05% KzHPO4, 0.01% MgSO4‘7H20; 0.01
CaClz'HZO, 0.05% (NH4)2SO4, 2% Bacto agar, and alter autoclaving addition of 5ppm
Nystatin (all salts and antibiotics from Sigma-Aldrich, St. Louis, MO) (Larsen et al.,
1998). CDPs are scored positive by clearing of the dark blue dye surrounding the
colonies. Soil samples were extracted in distilled water and appropriate dilutions cultured
on plates for quantification of vegetative cyclodextrin producing colonies. For
enumeration of spore-forming CDPs, aqueous extracts were incubated at 80°C for 10 min
prior to plating. Plates were incubated at 250C for 2 days and 55°C for 1 day before
counting CPD positive colonies.

Bacterial cell counts were adjusted to cell number per soil dry weight (e. g. CF U/ g
soil DW). Soil moisture content was determined by weighing a subsample of each treated
soil (FW), then drying at ~80°C for _>_72 hrs, followed by re-weighing.

Statistical analysis

Statistical significance was determined using SAS version 8.2 (SAS institute,
Cary, NC) with one-way ANOVA and least significant difference (LSD) for comparison
of treatment means with p S 0.05. Different treatments (e. g. plant biomass +/- 8CD) were
statistically compared by t-test analysis with StatView Version 5.01 (SAS Institute Inc.,
Cary, NC).

RESULTS
Soil PAH reduction

After 4 months greenhouse treatment, total soil PAH (tPAH) levels were reduced

from the starting soil concentration (135.6 : 3.9ppm soil DW) by various plant-[3CD

combinations to levels ranging from no change to ~27% reduction (Figure 2.1). The

93

 

 

 

 

 

 

 

 

 

 

. -"!"1-'-‘ - ‘_"..;.:'.,c- -j
._ .. "l""- .. _ .... ..
‘l....b-or~,,',‘... ..

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Reduction in soil tPAH (%)

 

 

 

 

 

 

 

-15 4*—

 

Wunplanted alfalfa _bluestem monkey flower mulberry—C

Figure 2.1. Reduction from initial soil tPAH concentration by planted X BCD amendment
treatment combinations after 4 months under greenhouse conditions. BCD amendment
concentrations: El = OmM BCD, = 5mM BCD, I = 10mM BCD. Values were
calculated relative to PAH-soil starting concentration (135.5 ppm per g soil DW). Error
bars = standard error (SE). Different letters over bars indicate significant differences (or S
0.05) in soil tPAH reduction between BCD application rates within a plant species

treatment.

94

greatest soil tPAH reductions were obtained in 10mM BCD amended unplanted, alfalfa,
and little bluestem treatments. Conversely for monkey flower and mulberry, the 10mM
treatment did not result in tPAH reduction, though displayed substantial variation in
tPAH concentrations among replicated treatment pots (SE = 12.5% and 11.0%
respectively, N = 4). Monkey flower was the only plant species to achieve statistically
greater contaminant reduction for unamended and 5mM BCD treatments relative to the
unamended, unplanted control soil. In general, BCD application rate appeared to correlate
with enhancement of tPAH biodegradation across all treatments with ~10.5, 11.5, and
14.8% tPAH reduction for OmM, 5mM, and 10mM BCD amendment levels respectively.
Despite this apparent trend, wide variation in response was observed among the different
plant species, indicating plant-specific BCD amendment optima or possibly sensitivity to
the biosurfactant for some species at these levels.
Total heterotrophic bacteria (CF U)

There was no apparent difference from soil CFU counts prior to any treatments
(Time 0, Table 2.1) between Clean-soil (To CFU :25 x 106) and PAH-soil (To CFU :14 x
106). BCD amendments were observed to enhance CFU counts for all rhizosphere soils
relative to unamended for both Clean- and PAH-soil (Figures 2.2a and 2.2b). Total soil
bacterial densities among all samples in each BCD treatment level (0, 5mM, and 10mM
BCD) revealed a steady increase relative to amendment concentration with averages of
~7.0, 22.9, and 30.0 CFU for Clean-soil and ~13.5, 37.1, and 43.6 CFU for PAH-soil,
respectively. Higher total bacterial cell counts were observed in bluestem, monkey flower
and mulberry planted soils relative to unamended rhizosphere soils at both 5mM and

10mM BCD application rates (Figures 2.2a and 2.2b). Alfalfa soil CFU cell density was

95

 

Table 2.1. Bacterial cell density of total heterotrophic, phenanthrene degrading, pyrene
degrading, and cyclodextrin producing cells per gram soil dry weight in uncontaminated
(Clean) and PAH-contaminated (Coking Oven) experimental source soils. Bacterial cell
types and cell density counts compared (unpaired T-test) from the respective source soils
were determined prior to treatments. Significantly different values (p < 0.01) between the

two tested soils are indicated by different superscript letters for lower (a) and higher (b)

 

 

 

values.
Untreated Source Soil
Bacterial cell density per g soil DW Clean Coking Oven
b
heterotrophic (CFU x 106) 25-0 14-3 a
phenanthrene degrader (ZFU x 103) 7.0 a 1075.4 b
pyrene degrader (ZFUpy x 103) 10.0 a 55.0 a
cyclodextrin producer (CDP x 104) 47.8 b 3.0 a

 

 

 

 

 

96

Figure 2.2a. Clean-soil

80 221--

l, A,

   

 

CFU per g soil DW (x106)

 

alfalfa little bluestem monkey flower mulberry

Figure 2.2b. PAH-soil
80-

70 ‘

60

50 r

40 a
30 a
20 ‘ -
10 ‘

0 .

 

CFU per g soil DW (x106)

unplanted alfalfa little monkey mulberry
bluestem flower

Figure 2.2. Comparison of total heterotrophic bacterial cell densities between BCD
amendment concentrations for each plant species in treated (2.2a) Clean-soil and (2.2b)
PAH-soil, shown as colony forming units (CFU) per gram of soil (DW). Error bars =
standard errors. Different letters over bars indicate significant differences (or S 0.05)

between [3CD amendment concentrations within a plant species treatment.

97

higher than unamended rhizosphere soil only for the 10mM [3CD treatment level in the
Clean- and PAH-soils. In contrast, BCD was not observed to significantly increase the

CF U number in unplanted PAH-soils (Figure 2.2b). In addition, very minor differences in
total heterotrophic cell counts occurred between plant species in unamended soils (Figure
2.2b). In general, BCD amendments were observed to stimulate higher CF U densities in
most rhizosphere soils, though differently for each plant species with regard to
amendment level and soil type (Figures 2.2a and 2.2b).

When comparing the effect of plant species on CFU counts at specific BCD
application rates, higher heterotrophic cell densities were observed in alfalfa rhizosphere
in both amended and unamended Clean-soil (Figure 2.3a). In 5mM [3CD amended Clean-
soil, CFU counts for mulberry rhizosphere were observed to be particularly high. In
unamended PAH-soil, alfalfa and mulberry CFU counts appeared higher than other
species, though were not statistically different (Figure 2.3b). Addition of 10mM BCD
enhanced CFU counts for both monkey flower and alfalfa rhizosphere soils, however at
5mM BCD amendments, only monkey flower CFU counts were elevated relative to
control treatments.

PAH degr_ading and biosurfactant producing bacteria

Culture plate assays were used to quantify phenanthrene degrading (ZFU), pyrene
cometabolizing (ZFUpy), and cyclodextrin biosynthesizing (CDP) cells in
uncontaminated (Clean-soil) and coking-process impacted (PAH-soil) material after the
4-month greenhouse treatment. Bacterial community responses for these metabolic

capabilities were quantified to analyze the effects of BCD amendments on specific PAH

98

 

Figure 2.3a. Clean-soil

 

80 — - ~ - ,. -7~——
70 L, alfalfa Ilittle bluestem Imonkey flower ilmulberryi#

 

l
l
l

  
  
 
  

50 ,__l.”_

l

 

J)
O

 

ND.)
00

 

CFU per g soil DW (x106)
S

 

OmM BCD 5mM BCD 10mM BCD

Figure 2.3b. PAH-soil

 

0 . 7, 7 . ,7" . , , . .2777 v, ,7
70 l D unplanted alfalfa I little bluestem I monkey flower I mulberry
F7 ' W ' T W” "— "bii ’#

  

60 j 2- 7

CFU per g soil DW (x106)

5mM BCD 10mM BCD

 

Figure 2.3. Comparison of heterotrophic bacterial cell densities between plant species for
each BCD amendment concentration in treated (2.3a) Clean-soil and (2.3b) PAH-soil,
shown as colony forming units (CFU) per gram of soil (DW). Error bars = standard
errors. Different letters over bars indicate significant differences (a s 0.05) between plant

species within a [3CD amendment concentration treatment.

99

rhizoremediation processes. Parallel Clean-soil treatments were also analyzed as an
indication of potential stimulatory effects of the selected plant species or
cyclodextrin amendments on endogenous bacterial dioxygenase or biosurfactant
biosynthetic capabilities in the absence of PAH compound influence.
Phenanthrene degraders

Phenanthrene degrader cell density (ZFU) was far greater in PAH-soils than
Clean-soils in both To samples (~150-fold) and at the completion of the 4-month
greenhouse experiment (~100- to 10,000-fold) for all planted and BCD amended
treatment combinations (Figures 2.4 and 2.5). [3CD amendments did not significantly
alter ZF U cell density relative to unamended controls (Figures 2.4a and 2.4b). Only the
10mM BCD amended monkey flower planted Clean-soils (2.4a) and little bluestem
planted PAH-soils (2.4b) displayed significantly higher ZFU counts. When comparing
plant species ZFU counts at a specific [3CD amendment rate, no significant differences
were observed between species in unamended soils, though alfalfa and monkey flower
displayed higher ZFUs in BCD amended Clean-soils (Figure 2.5a) and bluestem had
higher ZFU counts in PAH-soil amended with 10mM [3CD (2.5b).
Pyrene degraders

Pyrene degrading bacterial cell densities (ZFUpy) were not statistically different
between the initial, pre-treated Clean-soil and PAH-soil mixes (Table 2.1). Generally
speaking, ZFUpy counts were at least 10- to lOO-fold higher in nearly all greenhouse
soils, including the unamended treatments, after the 4-month study relative to the initial
time-zero (To) levels. BCD amendments further enhanced cometabolic ZFUpy numbers

in both unplanted and planted Clean-soils (Figure 2.6a) and PAH-soils (Figure 2.6b).

100

 

Figure 2.4a. Clean-Soil

  

 

 

 

  

 

12%,,_.1_I__Ih_ -WW— — — —— — — —
2 l ClOmM ClSmM

"5310i _________ W

as

a 8i ” r **

a m _ . a

8 I

3 4 -1. ,_____ i

8. I l1

N 0.4:

 

 

alfalfa 7 little bluestem monkey flower mulberry

Figure 2.4b. PAH-Soil

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

700 , ~- W , W — W —
«g ClOmM 035 mM IlOmM b.
i 600 W— — W
3 500 *» ,, _-_-_ -— , --—————— W ~WW
D
= 4001 W— —
8
:0 300 a __ -
8. r -2. a Y
ID 200 F3 a

unplanted alfalfa little monkey mulberry

bluestem flower

Figure 2.4. Comparison of phenanthrene degrader cell densities between BCD
amendment concentrations for each plant species in treated (2.4a) Clean-soil and (2.4b)
PAH-soil, shown as clear zone forming units (ZFU) per gram of soil (DW). Error bars =
standard errors. Different letters over bars indicate significant differences (or S 0.05)
between BCD amendment concentrations within a plant species treatment. Note use of

different Y-axis scales on respective charts.

101

Figure 2.5a. Clean-soil

 

 

20W-WW -W W W

 

 

 

 

 

 

 

 

 

 

 

 

"2; 6 4_ magma f little bluestem Imonkey flower Imulben'y
X
212W~W W W W fi-n—V
E l
DD 8 .0 ._ __ _ . __
E. l b
D 4 (i T
LN“ l a a a a a a a

0 M

OmM BCD 5mM [3CD 10mM BCD

Figure 2.5b. PAH-soil

 

700 -—___ -, --.
600 TU unplanted f alfalfa I little bluestem Imonkey flower I mulberry

 

 

 

 

500
400 +§ W

300 if

200 -—wa iii

100 W—T—l
0 l -

OmM BCD 5mM BCD 10mM BCD

 

 

 
 
 
 

 

 

 

 

 

ZFU per g soil DW (x103)

 

 

 

 

 

 

 

 

Figure 2.5. Comparison of phenanthrene degrader cell densities between plant species for
each BCD amendment concentration in treated (2.5a) Clean-soil and (2.5b) PAH-soil,
shown as clear zone forming units (ZFU) per gram of soil (DW). Error bars = standard
errors. Different letters over bars indicate significant differences (a s 0.05) between plant
species within a [3CD amendment concentration treatment. Note use of different Y-axis

scales on respective charts.

102

Figure 2.6a. Clean-Soil

10000 ,___ W- -— WWW—W ._ _,_ ’__._ ____. _-
J-bClOmM I5 mM I 10mM

 

 

 

 
   

8000 W -

 

 

 

 

 

    

ZFUpy per g soil DW (x103)

 

 

 

 

alfalfa little bluestem monkey flower mulberry

Figure 2.6b. PAH-Soil

6000 7*

l DOmM USmM b
5000 «.— -W W . ,________ _, ,

 

 

 

 

 

3000 T ,_

 

 

 

 
  

 

   

ZFUpy per g soil DW (x103)

 

.-.2 .4}:

unplanted alfalfa little monkey mulberry
bluestem flower

Figure 2.6. Comparison of pyrene degrader cell densities between BCD amendment
concentrations for each plant species in treated (2.6a) Clean-soil and (2.6b) PAH-soil,
shown as clear zone forming units (ZFUpy) per gram of soil (DW). Error bars = standard
errors. Different letters over bars indicate significant differences (a S 0.05) between BCD
amendment concentrations within a plant species treatment. Note use of different Y-axis

scales on respective charts.

103

Particularly notable pyrene biodegrading cell density increases were observed in BCD
amended rhizospheres of alfalfa (up to ~6-fold) and monkey flower (up to ~10-fold)
relative to unamended soils. Statistical comparison between planted treatments further
illustrates the highly significant species-specific [3CD enhancement of ZFUpy
populations in alfalfa and monkey flower rhizospheres in Clean-soil (Figure 2.7a) and
PAH-soils (Figure 2.7b). Of additional note are the observable increases in ZFUpy
numbers in unamended soils planted with alfalfa and monkey flower, indicating the
potential for pyrene-specific rhizostimulatory root exudate constituents by these two
species. Despite the absence of detectable pyrene in the uncontaminated field soil, high
ZFUpy cell numbers were observed again in correlation with BCD application rates and
most notably for alfalfa and monkey flower rhizosphere soils in Clean-soil (Figure 2.7a)
and PAH-soils (Figure 2.7b).
Relative PAH degrader abundance

In contrast to general bioaugmentation (i.e. addition of biodegrader cell inocula)
and biostimulation (i.e. supplementation with nutrients and growth substrates),
phytostimulation is proposed to enrich growth and metabolic activity of specific
biodegrader bacteria against the target pollutants. As a measure of phytostimulation
specificity and effectiveness, we compared the relevant degrader population as a
percentage of the overall culturable bacterial community (CFU) for the various BCD
amended, planted treatment combinations.

The percentage of phenanthrene degrading cells as a component of the total
bacterial community [(ZFU/CFU)* 100] was similar among the various plant X [3CD

amendment treatment combinations within a specific soil type (Table 2.2). As expected,

104

Figure 2.7a. Clean-soil

 

1000 'l W~ _, W W H. , WWWWW .,__ . ___ __zi .
l alfalfa I little bluestem I monkey flower I mulberry Tb
5 ___“ b

 

 

 

800 _. m...“ z- #_-z__

 

 

ZFUpy per 2 soil DW (x103)

 

 

 

Figure 2.7b. PAH-soil

 

 

6000 ~ — ~-~
It] unplanted alfalfa I little bluestem I monkey flower I mulberry
5000 *- 7 - -
I b

4000 {L- - b

 

 

 

ZFUpy per 2 soil DW (x103)

 

OmM BCD 5mM [3CD 10mM BCD

Figure 2.7. Comparison of pyrene degrader cell densities between plant species for each
[3CD amendment concentration in treated (2.7a) Clean-soil and (2.7b) PAH-soil, shown
as clear zone forming units (ZFUpy) per gram of soil (DW). Error bars = standard errors.
Different letters over bars indicate significant differences ((1 S 0.05) between plant
species within a [3CD amendment concentration treatment. Note use of different Y-axis

scales on respective charts.

105

Table 2.2. Percentage of phenanthrene degrading bacteria (ZF U) relative to total

heterotrophic bacteria (CFU) for various BCD x planted treatments in (2.2a) Clean-soil or

(2.2b) PAH-soil afier 4—months treatment. Different letters indicate significantly different

(a S 0.10) percentages of ZFU cells: lower case letters = comparison between BCD

amendment levels (within a column) for the same plant species; upper case letters =

comparison between plant species treatments (within a row) at a given BCD amendment

level.

2.2a. Clean-soil

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

BCD level little monkey
(mM) Implanted alfalfa blues tem flower mulberry
0 _ 0.005 0.006 0.007 0.009
a, A a, A a, A b, A
5 _ 0.010 0.002 0.003 0.001
a, B a, A a, A a, A
10 _ 0.010 0.003 0.035 0.007
a, B a, A b, B b, A
2.2b. PAH-soil
[3CD level little monkey
(mM) Implanted alfalfa blues tem flower mulberry
0 1.0 1.1 1.5 1.7 0.7
a, A b, A b, A b, A a, A
5 0.7 0.9 0.4 0.4 0.4
a, A b, A a, A a, A a, A
10 0.5 0.2 1.1 0.4 0.4
a, AB a, A b, B a, A a, A

 

 

 

 

 

 

 

 

106

relative ZFU abundance was much lower in Clean-soils (5 0.01% ZF U) than observed for
the contaminated PAH-soil mix (~0.5-2% ZF U) for all planted X BCD combinations. In
uncontaminated soils, only 10mM BCD amended monkey flower displayed enrichment
for phenanthrene ZFU, having ~3- to lO-fold greater relative abundance than other Clean-
soil treatments (Table 2.23). In PAH-soils, the relative abundance of phenanthrene
degrading cells displayed an apparent decrease with addition of BCD (Table 2.2b). For all
unplanted and planted soils, the %ZFU was lower in amended soils compared to the
water-only control treatment. Modestly significant differences in relative ZF U abundance
were observed between planted treatments at the 10mM BCD amendment level in both
soil types.

BCD amendments were observed to increase the relative abundance of
cometabolic pyrene degrading bacterial cells [(ZFUpy/CFU)* 100] for some planted
treatments in PAH-soils (Table 2.3b), but not in Clean-soils (Table 2.3a.). Rhizospheres
of alfalfa and monkey flower contained ~10-fold higher percentages of ZFUpy bacteria in
Clean-soils relative to other planted treatments (Table 2.3a), though BCD supplements
did not appear to influence %ZFUpy values for these two plant species. By contrast,
relative ZFUpy abundance appeared to be enhanced to a modest extent for biosurfactant
amended Clean-soils planted with bluestem (~3-fold) or mulberry (~5-fold). In PAH-
soils, however, alfalfa and monkey flower appeared to enrich for ZFUpy cells in relation
to BCD application rates (Table 2.3b). Biosurfactant amendments appeared to enhance
relative pyrene degrader abundance for alfalfa (~2-fold), monkey flower (~3-fold), and

mulberry (~3-fold) relative to unamended PAH-soils.

107

Table 2.3. Percentage of pyrene degrading bacteria (ZFUpy) relative to total

heterotrophic bacteria (CFU) for various BCD x planted treatments in (2.3a) Clean-soil or

(2.3b) PAH-soil after 4-months treatment. Different letters indicate significantly different

(a S 0.10) percentages of ZFUpy cells: lower case letters = comparison between BCD

amendment levels (within a column) for the same plant species; upper case letters =

comparison between plant species treatments (within a row) at a given [3CD amendment

level.

2.3a. Clean-soil

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

BCD level little monkey
(mM) unplanted alfalfa blues tem flower mulberry
0 _ 1.2 0.08 4.4 0.04
a, A a, A b, B a, A
5 _ 2.3 0.11 2.3 0.05
a, B a, A a, B a, A
10 _ 1.6 0.23 2.8 0.19
a, AB a, A ab, B a, A
2.3b. PAH-soil
BCD level little monkey
(mM) unplanted alfalfa blues tem flower mulberry
0 1.6 2.3 3.7 2.2 2.2
a, A a, A b, A a, A a, A
5 1.6 2.7 0.9 2.9 6.4
a, A a, AB a, A b, B b, B
10 2.5 4.2 2.7 6.8 1.9
a, A b, AB b, A b, B a, A

 

 

 

 

 

 

 

 

108

Cyclodextrin producing bacteria (CDP)

There was no significant difference between plant species or [3CD amendment
treatment combinations upon cyclodextrin producing bacteria (CDP) cell density, relative
abundance, or type (spore-forming or vegetative) (data not shown). It was initially
hypothesized that bacteria capable expressing cyclodextrin glycosyltransferase would be
capable of preferential utilization of BCD and consequently display greater cell densities
in amended soils than non-metabolizing bacteria (Feederle et al., 1996). Contrary to
expectations, addition of cyclodextrin did not increase the number of soil bacteria capable
of utilizing BCD.

DISCUSSION

B-cyclodextrin (BCD) amendments increased PAH bioremediation effectiveness
for unplanted soils and some planted treatments as evidenced by improved reduction of
soil PAH concentration and enhancement of biodegrader cell numbers. Plant species and
BCD applications singly and in combination were demonstrated to influence soil bacterial
densities of heterotrophic cells (CFU), phenanthrene degraders (ZFU), pyrene degraders
(ZFUpy), though not cyclodextrin producers (CDP), the latter of which was statistically
similar across all treatments. Different bacterial community responses were observed in
Clean-soil and PAH-soil treatments, which was anticipated for degrader bacteria due to
pre—selection in the coking-impacted soils for utilization of PAH contaminants as carbon
sources. Clean-soil initially hosted higher numbers of heterotrophic and CD-producing
bacteria, though contained lower numbers of PAH biodegrading bacteria than the PAH-

soil. BCD supplements were demonstrated to enhance total heterotrophic bacteria,

109

phenanthrene degraders, and pyrene degraders in both uncontaminated and PAH-
impacted soils.

Soil microbial communities are affected by many factors: physical and chemical
characteristics of the soils (e. g. soil texture, nutrients, pH), climate, and vegetation
(Anderson et al., 1993; Kuske et al., 2003; Marschner et al., 2004). Plants support
microbial grth by releasing root exudates into the rhizosphere, which is defined as the
region of soil in which microbes are influenced by the root system (Hiltner, 1904). In the
root elongation zone located a few centimeters behind root tips, sugars, polysaccharides,
amino acids, organic acids, fatty acids, sterols, and other growth factors are released into
the rhizosphere soil (Kuiper et al., 2004). Along older root parts, the primary substrates
for microbial growth are cellulose and other cell wall materials, including phenolic
compounds such as flavonoids, which form the skeleton backbone of tannins (Marschner
et al., 2004; Singer et al., 2003). Different plant species growing in the same soil release
different quantities and composition of root exudates (Marschner et al., 2001; Rengel,
1997), which influences rhizosphere microbe community structure due to differential
abilities to metabolize and compete for different carbon substrates. Some reports
indicated that different plant species growing in the same soil possess similar rhizosphere
microbial communities (Buyer et al., 1999). Conversely, the same plant species growing
in different soils may be colonized by different bacterial communities (Grayston et al.,
1998; Miethling et al., 2000). During phytoremediation, plants have been shown to
stimulate PAH biodegrading activity by release of root exudates into their rhizosphere
and providing surfaces for more extensive soil colonization (Kirk et al., 2004; Liste and

Alexander, 2000; Macek et al., 2000; Siciliano and Germida, 1998).

110

Combined BCD and planted treatments increased phenanthrene-degrader (ZFU)
cell density relative to their respective single-component controls. However, analysis of
degrader cell percentages did not indicate a specific phytostimulation effect for this PAH
compound, which would require disproportionate increases in phenanthrene biodegrader
cell density, rather than overall bacterial cell number increases. Contrastingly, relative
cell abundance of pyrene degrading bacteria (ZFUpy) was highest in combined plant +
[3CD treatments. [3CD amended rhizospheres displayed enriched %ZFUpy relative
abundance, particularly for alfalfa and monkey flower rhizosphere soils indicating
contaminant-specific phytostimulation by these two plant species.

Enhanced PAH reduction occurred for most of the planted and amended
treatments, though this effect was somewhat obscured by substantial variation in soil
PAH concentration for the planted soil treatments, likely due to the natural heterogeneity
of root colonization and possibly influences on contaminant localization by root surfaces
(Liste and Martin, 2000). Despite the substantial variation, PAH biodegradation was
observed to be significantly enhanced in several combined plant + BCD treatments. In
unplanted treatments, BCD amendments appeared to positively impact bacterial cell
density and PAH biodegradation capability. This benefit was also generally observed in
most rhizospheres, however it was not uniform with regard to PAH dissipation or
bacterial community structure across species and BCD amendment rates. Root-microbe
interactions are strongly influenced by various plant qualities, including plant age and
developmental state, exudate production rate, root epidermal biophysical behavior, and —
in this study - by plant response to biosurfactant “detergent” effects. Given the consistent

increases in rhizosphere bacterial cell numbers relative to unamended or unplanted

111

treatments, it is evident that BCD improved PAH phytostimulatory phenomena in most
instances. It is possible the surface-active properties of BCD caused modifications to the
root epidermal structure, plant secondary compound exudation and/or cometabolite
solubilization, each of which could have improved bacterial colonization efficiency and
PAH degradation effectiveness.

Addition, or biosupplementation, of simple carbon sources like sucrose or glucose
to soil were demonstrated to stimulate induction of enzymes necessary for catabolism of
more complex organic compounds (Grayston et al., 1998) and to increase overall
metabolic activity (Alden et al., 2001). Addition of sodium citrate as a carbon source was
found to effectively stimulate more efficient atrazine biometabolism in contaminated
soils (Mandelbaum et al., 1993). Therefore, it is possible that cyclodextrin supplements
could serve as carbon resources for heterotrophic and PAH biodegrading microbial
populations resulting in increased soil PAH biodegradation rates. As cyclodextrin is
initially metabolized by CD-utilizing strains (Lee et al., 2002; Rajdevi and Yogeeswaran,
2001), other soil bacteria possessing (1.1-4 glycosidic enzymes such as a-amylase, or-
glucosidase, amylopullulanase would be capable of subsequent utilization of CD-
digestion products for heterotrophic grth (van der Maarel et al., 2002). Such a process
may indicate why CD producing/utilizing bacterial abundance was not increased, though
overall microbial cell density was consistently enhanced in [3CD amended soils. Soil
PAH levels were generally observed to be more effectively reduced in BCD treated soils,
which may be explained in part by enhancement of biodegrading cell abundance for the
PAH compound pyrene, though less so for phenanthrene degraders. It is not readily

apparent why pyrene biodegrader abundance would be stimulated, though not

112

phenanthrene. One possibility is that the cometabolic biodegradation pathway for pyrene
might benefit more from BCD effects on plant exudation quantity and availability or
possibly by the use of cyclodextrin and its breakdown products as co-substrates as
discussed previously. A more extensive survey of bacterial degrader abundance for more
of the sixteen US EPA classified PAH compounds would provide a clearer picture of the
extent and specificity of [3CD amendment influence on broad-spectrum PAH
phytostimulatory processes.

It is possible that BCD could negatively affect root-soil biophysical processes
resulting in root stress or toxicity, alterations in nutrient or water transport, or imbalances
in root-microbe nutrient competition. [3CD amendments were shown to reduce the growth
rate of some plant species and soil nutrient status for some elements (Chapter 1). Such
perturbations may influence root-microbe interactions and consequent rhizoremediation
effectiveness. Though fertilizer was added, mulberry appeared stressed and stunted,
though little bluestem, monkey flower and alfalfa were not affected. Plants under
nutritional stress have been observed to alter root exudate amount and composition
causing changes to rhizosphere microbial community (Marschner et al., 2004; Yang and
Crowley, 2000). Excess carbon loading is possible with enhanced root exudation and
cyclodextrin addition, promoting rapid nitrogen assimilation by heterotrophic
microorganisms and subsequent N limited plant stress (Gilliam et al., 2005; Merckx et
al., 1987). With continuous carbon loading, an imbalanced C-N ratio will exacerbate
competition between roots and microbes (Griffiths and Robinson, 1992). It may be
essential to systematically characterize biosurfactant-soil interactions for each plant

species to nutritionally optimize BCD-assisted phytoremediation.

113

In this study, we used coke-oven residues as our PAH-contaminated substrate,
which have been demonstrated to have very strong sorptive characteristics resulting in
minimal PAH bioavailability (Ghosh et al., 2000; Ghosh et al., 2003). Unamended,
unplanted soils achieved relatively limited biodegradation or bacterial biodegrader
response in these soils with only periodic watering and fertilization. By contrast, BCD
amendments effectively enhanced both PAH bioremediation effectiveness and
phytostimulation of bacterial community grth and metabolism in most plant-
biosurfactant treatment combinations. BCD complexation has been shown to effectively
alter the physical and chemical properties of a range of guest molecules by enhancing
aqueous solubility, fixation of volatile compounds, or controlled release of target
compounds (Szejtli, 1988). BCD amendments were demonstrated in this study to enhance
symbiotic processes for rhizosphere-assisted bioremediation. Differences in BCD
amended PAH biodegradation rates and bacterial community response suggests different
mechanisms of action among the tested plant species, which should be further

characterized by additional studies.

114

 

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CHAPTER HI

TRAN SGENIC TOBACCO PRODUCTION OF THE BACTERIAL

BIOSURFACTAN T CYCLODEXTRIN FOR ENVIRONMENT REMEDIATION

122

INTRODUCTION

Polyaromatic hydrocarbons (PAHS) are persistent organic pollutants formed from
natural combustion processes, such as forest or field fires, and by anthropogenic activities
including coal-processing at coking plants, manufactured gas plants, and oil and gas
processing facilities. PAHS are of particular environmental and health concerns due to
their tendency to biomagnify to higher tissue concentrations in the food web and their
carcinogenic and mutagenic properties. The US. Environmental Protection Agency lists
PAHS in 600 of 1,408 National Priority List Superfund sites (ATSDR, 1995). PAHS are
also prevalent at industrial brownfields (EPA, 2004), dredged sediments (EPA, 2004),
and numerous military properties (GAO, 2001).

Current remedial treatments for PAH contaminated soils include both physical
and biological processes. Conventional engineering-based approaches such as extraction
and containment strategies have been routinely used to treat sites contaminated with
PAHS (Cunningham and Berti, 1993). Extractive techniques include excavation of
hazardous waste materials, in situ thermal desorption, and vapor extraction. PAH
contaminated soils may also be solidified and stabilized by adding a cementing material
followed by burial in secure landfills. Biological remediation offers the advantages of
reduced incidental mobilization of pollutants, minimal site disturbance, and reduced
environmental impact relative to intensive engineering-based treatments. Biological
treatments of contaminated soil include natural attenuation, bioremediation and
phytoremediation. Natural attenuation allows hazardous wastes to break down naturally
over time into non-hazardous material and is typically utilized if there is little chance that

the contamination will pose an interim threat to human or wildlife communities.

123

Bioremediation is environmental detoxification mediated by living organisms, generally
bacteria or fungi, to treat soil and water contaminated with organic and elemental
pollutants. Bioreactor techniques and land farming approaches have been developed and
utilized for microbial degradation of organic contaminants (Civilini et al., 1996; Lilja et
al., 1996; Sayles et al., 1999; Sims and Sims, 1999). Bacterial degradation of PAHS via
dioxygenase enzymatic metabolism has been well characterized at the biochemical and
molecular genetic levels (Bumpus, 1985; Cemiglia, 1993; Field et al., 1992).
Phytoremediation is the use of planted systems for treatment of environmental
contaminants and has been shown to enhance organic pollutant biodegradation rates for
natural attenuation, bioremediation and land farming applications (N edunuri et al., 2000;
Pradhan et al., 1998; Reilley et al., 1996). Various plant species have been demonstrated
to enhance biodegradation rates of persistent organic pollutants such as PAHS and
pesticides (Gunther et al., 1996; Liste and Alexander, 2000; Perkovich et al., 1996;
Reilley et al., 1996). Plant-assisted bioremediation of organic pollutants occurs
principally by root-secreted compound stimulation of bacterial biodegrader cell density
and metabolic activity (Corgie et al., 2003; Crowley et al., 1997; Yoshitomi, 2001).
PAH biodegradation efficiency is limited by low aqueous solubility and
bioavailability. PAHS tend to partition to soil organic matter in soils and sediments and
become largely unavailable to biodegrading microorganisms. Synthetic and biologically
generated surfactants have been demonstrated to increase solubilization, mobilization and
desorption rates of various hydrophobic compounds (Deitsch and Smith, 1995; Guha et
al., 1998; Hommel, 1990; Maier, 2003; Sun and Boyd, 1993; Thangamani and Shreve,

1994; Thibault et al., 1996). Cyclodextrins (CDs) are circular oligosaccharide

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biosurfactants with a hydroxyl-rich, hydrophilic external surface and a hydrophobic,
interior cavity due to the orientation of glucose units in the doughnut-shaped cyclodextrin
molecule (Szejtli, 1991). The CD nonpolar interior cavity can form inclusion complexes
with hydrophobic compounds, referred to as “guest molecules”, while the hydrophilic
exterior keeps the complex soluble. CDs have been used in the food, cosmetic,
pharmaceutical, and pesticide industries as emulsifiers, antioxidants, and stabilizing
agents (Lee and Kim, 1991). Biologically produced CDs and synthetic, modified CDs
have been studied for soil and ground water remediation (Brusseau et al., 1997; K0 and
Yoo, 2003; Morillo et al., 2001; Perez-Martinez et al., 2000; Sheremata and Hawari,
2000). CD compounds have been demonstrated to enhance solubility and extractability of
various organic and metal contaminants from polluted or spiked soils (Sheremata and
Hawari, 2000; Wang and Brusseau, 1995).

Cyclodextrins are cyclic, nonreducing oligosaccharides formed from starch by the
bacterial enzyme cyclodextrin glycosyltransferase (CGTase) encoded by the cgt gene.
Common cyclodextrins consist of 6, 7, or 8 glucose units linked together via a a-1-4
linkage and are designated a-, [3-, or y-CDs respectively. CGTases are produced by a
large number of Bacillus species: B. circulans, B. macerans, B. megaterium, B.
stearothermophilus, B. ohbensis, B. alkalophilus (Larsen et al., 1998), and some strains
of Brevibacillus brevis (Kim et al., 1998), Klebsiella oxytoca (Binder et al., 1986), and
T hermoanaerobacter (Wind et al., 1995). CGTases catalyze three different major
transferase reactions: cyclization, coupling, and disproportionation (Szejtli, 1991).
Cyclization, or circularization, is the transfer of the nonreducing-end sugar to another

sugar residue in the same oligosaccharide chain, resulting in formation of cyclic

125

 

compounds (Uitdehaag et al., 1999). Coupling is the combination of a cyclodextrin
molecule and a linear oligosaccharide, resulting in a longer oligosaccharide molecule.
Disproportionation is the transfer of part of a linear oligosaccharide chain (donor) to
another linear oligosaccharide chain (acceptor), resulting in the mixture of smaller and
longer oligosaccharide (Lawson et al., 1994). CGTases and a-amylases are functionally
related in that both enzymes utilize starch as a substrate, the former produce circular
oligosaccharides while the later form linear products (Lawson et al., 1994; Nakamura et
al., 1992). Although the amino acid sequences of CGTases and or-amylases are
dissimilar, their predicted secondary structures have high similarity. CGTases and 0t-
amylases share a three dimensional structural resemblance in an N-terrninal domain of
approximately 400 residues, which fold into (B/a) g-barrel (N akamura et al., 1992).
However, CGTase proteins have two additional C-terminal domains (D and E) that fold
into B-pleated sheets and are absent from a-amylases (Svensson, 1994). Therefore,
CGTases generally have a molecular weight of 70-75 kDa, while a-amylases generally
have a molecular weight of 45-55 kDa (Schmid, 1989). All CGTases produce a number
of cyclodextrins and other products with specificity dependent on the specific bacterial
host and CGTase protein. The primary cyclodextrin formed during the cyclization
reaction depends on the type of CGTase, though as the enzymatic reaction continues,
other cyclodextrin forms are also usually synthesized. B. macerans, B.
stearothermophilus, Klebsiella oxytoca produce mainly orCD, while B. circulans, B.
megaterium, B. ohbensis, B. alkalophilus produce mainly BCD, while 7CD is usually
produced in lower quantities relative to the other cyclodextrins (Schmid, 1989). Various

selective benefits are proposed for bacterial CGTase secretion into its surrounding

126

 

medium (Lawson et al., 1994). CD5 can form inclusion complexes with toxic compounds
to prevent toxicity to bacterial cells or alternatively, CDs may stabilize compounds
needed for bacterial growth. Another possible benefit is conversion of starch to CD
compounds that may serve as selective carbon sources for bacteria able to metabolize
cyclodextrins.

As an alternative to application of exogenous biosurfactant amendments, we have
proposed the development of transgenic plants engineered with the bacterial cyclodextrin
biosynthesis gene (cgt). Most phytoremediation research and field applications have used
naturally occurring plant species, however transgenic plants have been developed with
enhanced abilities for detoxification of hazardous pollutants (Glass, 1998; Pilon-Smits
and Pilon, 2002; Rugh, 2001). Overexpression of enzymes involved in phytochelatin
(poly [y-glutamylcysteinyl]glycines) biosynthesis in transgenic plants resulted in
enhanced metal tolerance and accumulation (Bennett et al., 2003). Engineered Indian
mustard (Brassicajuncea) expressing E coli glutathione synthetase displayed enhanced
cadmium tolerance and accumulation (Zhu et al., 1999). Plant engineering with the
bacterial merA and merB genes encoding mercuric reductase and organomercurial lyase
respectively, conferred elevated ionic mercury and methyl mercury resistance and
detoxification in transgenic plants (Bizily et al., 2000; Rugh et al., 1998; Rugh et al.,
1996). Transgenic tobacco plants expressing a gene encoding a mammalian cytochrome
P450 were shown to degrade trichloroethylene (TCE) at 640-fold the rate of non-
transgenic control plants (Doty et al., 2000). Plants engineered with the bacterial gene
encoding pentaerythriol tetranitrate reductase gained the novel capability to degrade

explosive nitrate esters and nitroaromatic compounds (French et al., 1999). Plant genetic

127

modification for enhanced degradation or metal tolerance capabilities may be particularly
valuable for species with desirable phytoremediation traits such as high biomass or deep
root systems. Development of transgenic plants to secrete CD biosynthetic enzymes into
contaminated rhizosphere soil could represent an effective and novel tool for treatment of
persistent, hydrophobic contaminants.

MATERIALS AND METHODS
Cyclodextrin-Producing Bacteria Screening and Characterization
Isolation of CGTase Expressing Bacteria
Soil samples were collected from PAH contaminated, coal processing facility and a clean
wooded field (Dearbom, MI) and assayed for bacterial CGTase producers. 1.0g soil was
placed in an 18x150mm test tube containing 9m] sterile water. To select for spore
forming Bacillus species, the most common cyclodextrin producers, sample tubes were
heated at 80°C for 10min then allowed to cool 1hr. Tubes were vortexed vigorously and
100ul aliquots of serial dilutions spread on Basic Medium agar plates (2% w/v soluble
starch, 0.5% w/v yeast extract, 0.5% w/v tryptone, 0.1% w/v KzHPO4, 0.02% w/v
MgSO4.7H20, 0.02% w/v CaC12.2H20, 0.1% w/v (N114); S04, pH to 7.0, 3% Bactoagar,
0.03 % bromcresol green and after autoclaving addition of 5mg/l Nystatin; all
components from Becton Dickenson and Company, Sparks, MD; except mineral salts
from Sigma-Aldrich, St. Louis, MO) (Larsen et al., 1998). Plates were inoculated with
soil extracts and incubated 3days at either 37°C or 55°C. CGTase secreting colonies were
identified by clear zone formation in the colored medium due to complexation of the
bromcresol green dye.

TLC Analysis of Cyclodextrins

128

Cyclodextrin production was confirmed by thin-layer chromatography (TLC)
separation and comparison to pure, commercially supplied CD compounds (Jindrich et
al., 1995). Bacterial isolates were grown in 5.0m] liquid Basic Medium at 37°C with
rotary shaking at 200rpm. After 48hr, 1.5m] supernatant was transferred to microfuge
tubes and cleared by centrifugation at ~13000rpm for 5min. 0.5m] supernatant containing
crude bacterially secreted CGTase was incubated in 2m1 1.25% soluble starch (w/v 0.1 M
phosphate buffer, pH 7.0) for 24h at 50°C. 2.0111 reaction medium was spotted on 5x10cm
TLC plates cut from 20 x 20cm silica gel 60/Kieselguhr F254 aluminum TLC sheets (EM
Science, Gibbstown, NJ). a-, B-, and y-CD (Sigma-Aldrich, St. Louis, M0) were spotted
from 1% (w/v) solutions as standards and run adjacent to experimental extracts. The
mobile phase was acetronitrile-water-ammonium hydroxide (6:3: 1) and the assay was run
in a sealed, glass TLC tank. Completed TLC assay plates were processed by brief
immersion of the plate into Vaugh’s solution [1 g of Ce(SO4)2, 24g of (NH4)2MoO4, 50ml
concentrated H2S04, 450ml water] and developed on a hot plate until blue spots became
visible.

Colorimetric Quantification of Cyclodextrins

To determine the type and relative quantity of CDs produced by the CGTase
secreting bacterial isolates, a modified colorimetric assay was performed (Kaneko et al.,
1987). Bacterial strains were grown in Basic Medium broth and supernatant cleared as
described previously. For BCD quantification, 0.1m] supernatant (crude enzyme) was
added to 1.0m] 1.25% soluble starch (w/v 0.1M phosphate buffer, pH 6.0) in 18x150mm
glass test tubes and incubated at 50°C. After 30min the reaction was stopped by addition

of 3.5m] 30mM NaOH, then 0.5m] 0.02% phenolphthalein (w/v 5 mM NazCO3) was

129

 

added, vortexed briefly to mix, and incubated at room temperature for 15 min. Reaction
product color intensity was determined spectrophotometrically (Molecular Devices,
SpectraMax 190 microplate spectrophotometer, Sunnyvale, CA) at 550nm. A 4-point
standard curve was run using a range of 0.05 to 0.3mM BCD in 0.1 M phosphate buffer
pH 6.0. For determination of yCD, 0.1ml supernatant (crude enzyme) was added to 1.0m]
1.25% soluble starch (w/v 0.1M phosphate buffer, pH 6.0) in 18x150mm glass test tubes
and incubated at 50°C. After 120min the reaction was stopped by boiling for 5min, then

0.1ml 5mM bromcresol green and 2.0m] sodium citrate, pH 4.2 added, vortexed briefly,

 

and immediately spectrophotometn'cally analyzed at 630nm. A 4-point standard curve
was run using a range of 0.1 to 0.3mM yCD in 0.1 M phosphate buffer pH 6.0.
Identification of CGTase Positive Soil Bacteria

Selected CGTase positive bacterial isolates were identified by 16S rRNA gene
sequence analysis. Bacterial genomic DNA was isolated from pure cultures using
standard protocols (Ausubel et al., 1987). The 16S rRNA encoding gene was amplified
using universal primers 11f (5’-GTT TGA TCC TGG CTC AG-3’) and 1392r (5’- ACG
GGC GGT GTG T-3’) corresponding to E. coli l6S rRNA gene position 11-27 and 1404-
1392, respectively (Lane, 1991). The PCR reaction contained 0.5ul template DNA,
20mM Tris pH 8.3, 50mM KCl, 2 mM MgC12, 200uM each dNTP, 0.2uM each primer,
0.2ug bovine serum albumin (BSA), and 2.5 units T aq DNA polymerase (Invitrogen,
Carlsbad, California) in a final reaction volume of 25111. 16S rRNA gene sequences were
amplified at 94°C for 3 min followed by 35 cycles at 94°C for 45s, 60°C for 303, 72°C for
305. Afier amplification, the tubes were incubated at 72°C for 10min before cooling to

4°C. The 16S rRNA amplification product was cloned into vector pCR2.1 and

130

transformed into chemically competent E. coli Top10 using the TOPO TA cloning kit per
manufacturer’s instructions (Invitrogen, Carlsbad, CA). Recombinant pCR2.1 plasmids
were screened and confirmed using standard restriction digest and gel electrophoresis
protocols. Plasmids containing fragment sizes expected of the 16S rRN A gene product
were purified using QIAGEN Plasmid Mini Kit according to manufacturer’s protocol
(QIAGEN Inc., Valencia, CA). The 165 rRNA gene insert was partially sequenced using
universal priming sites on the pCR2.1 plasmid at the Michigan State University
Genomics support facility (genomics.msu.edu) and compared to existing ribosomal
sequences in Michigan State University Ribosomal Database (rdp.cme.msu.edu).
Bacterial cgt Gene Cloning

The four bacterial isolates obtained from uncontaminated field soil and coke oven
soil residue that produced mainly BCD were identified as Paenibacillus illinoisensis
(isolates 2-8B, 2-13A, and C36) and Bacillus subtilis (isolate 2-8A) and selected for cgt
gene cloning. The preparation of bacterial isolate genomic DNA was described
previously (Ausubel et al., 1987). Forward primers used to clone cgt were generated from
B. circulans cgt gene sequences, accession numbers: X68326, X78145, and AF 302787,
respectively: Cgt-Fl: 5’-TTCACGAAGGGTGGATTAC-3’,
Cgt-F 2: 5 ’-GGACAAGCCTGGAATTCAA-3 ’, and
C gt-F3 : 5 ’-GAGGAGGTATAGTATGAAA-3 ’.
The reverse primer used to clone cgt was generated from B. circulans cgt sequence,
accession number AF 302787: Cgt-Rl: 5’-ATTAAGGCTGCCAGTTCAC-3’.
The PCR reaction contained 5ul template DNA, 20mM Tris pH 8.3, 50mM KCl, 2.5mM

MgClz, 300uM each dNTP, 0.5uM each primer, 0.2ug bovine serum albumin, and 6.25

131

 

units Taq DNA polymerase (Invitrogen, Carlsbad, California) in a final reaction volume
of 50ul. PCR amplification conditions were 96°C for 3min followed by 30 cycles of 96°C
for 1.5min, 42°C for 1.5min, 72°C for 2.5min. After amplification, the tubes were
incubated at 72°C for 10 min before cooling to 4°C. Amplification products were gel
purified using the QIAquick gel extraction kit (QIAGEN Inc., Valencia, CA) and cloned
into pCR2.1 (Invitrogen, Carlsbad, CA). Transformed colonies were screened using
standard restriction digest and gel electrophoresis protocols and confirmed recombinant
PCR2.l plasmids were purified using QIAGEN Plasmid Mini Kit according to
manufacturer’s protocol (QIAGEN Inc., Valencia, CA). Cloned inserts were sequenced at
the Michigan State University Genomics support facility (genomics.msu.edu), DNA and
amino acid sequences analyzed with the Wisconsin sequence analysis package (GCG),
and compared to published cgt gene sequences using the NCBI Blast database search tool
(www.ncbi.nlm.nih. gov).

The cgt gene from Paenibacillus illinoisensis C36 isolate (PI-cgt) was amplified
for subcloning using the forward primer:
Pngt-F l: 5 ’GAATTCGGCGGCCGCTTAAAGAAGGATTAACAATGTTCAATGG
to introduce EcoRI and NotI sites at the 5’end of the PCR product and reverse primer:
Pngt-Rl: 5’-CTGTACGGATCCGAGCTCATTAAGGCTGCCAGTT
to introduce BamHI and SacI sites at the 3’end of the PCR product. Underlined regions of
the primers indicate the location of introduced restriction sites in the primer sequence
with the remaining nucleotides complementary to the target cgt gene. The PCR reaction

contained Sul template DNA, 20mM Tris pH 8.3, 50mM KC], 2.5mM MgClz, 300uM

each dNTP, 0.5uM each primer, 0.2ug bovine serum albumin, and 6.25 units Taq DNA

132

 

polymerase (Invitrogen, Carlsbad, California) in a final reaction volume of 50p]. PCR
amplification conditions were 96°C for 3min followed by 30 cycles of 96°C for 1.5min,
42°C for 1.5min, 72°C for 2.5min. The cgt PCR product was gel purified and subcloned
into pBluescript II SK- (pBS, Stratagene, La Jolla, CA) using the NotI and BamHI
restriction sites. Screening for successful insertions was performed using standard
restriction enzyme digest and gel electrophoresis methods. pBS-cgt transformed E. coli
colonies were cultured on Basic Medium + starch plates to screen for CGTase production
via clear-halo formation (Larsen et al., 1998).
Paenibacillus illinoisensis CGTase PAGE Analysis

The P. illinoisensis C36 (PI-C36) bacterial isolate was analyzed for CGTase
protein production via denaturing polyacrylamide gel electrophoresis (PAGE). PI-C36
was cultured in liquid Basic Medium and supernatant cleared by centrifugation at
8000rpm for 15min at 4°C . Culture supernatant proteins were precipitated by addition of
3 volumes of ice-cold ethanol (95% v/v), concentrated by centrifugation at 8000rpm for
15min at 4°C, and the pellet resuspended in 0.1M phosphate buffer pH 6.0. Secreted
protein products and protein molecular weight markers were separated via sodium
dodecyl sulfate (SDS)-PAGE and visualized with coomassie blue staining (Ausubel et al.,
1987)
Plant Transformation with PI-cg
Construction of the plant expression vectors

Two plant transformation vectors were constructed for introduction and
expression of the bacterial PI-cgt gene in plants (Figure 3.1). Plasmid pCAMBIA-Plcgt

(pC-PIcgt) was constructed using pAPC9K, which contains the Arabidopsis thaliana

133

 

Actin2 promoter (An et al., 1996) and the PE21 terminator of the Citrus sinensis cv.
Valencia pectinesterase gene (N aim et al., 1998) (Figure 3.1a). pCAMBIA-PIcgt was
constructed by inserting the PI-cgt expression cassette from pAPC9K-PIcgt into the
multiple cloning site of the binary plant expression vector, pCAMBIA1300 (Genbank
accession number AF234296) (Figure 3.1b). The PI-cgt gene was subcloned from
plasmid pBS-PIcgt into the pAPC-9K NotI and SacI restriction sites. The expression
cassette containing the Actin2 promoter, PI-cgt coding sequence, and PE21 terminator
was excised from pAPC9K by SpeI and XmaI restriction digestion and ligated into the
pCAMBIA1300 XbaI and XmaI restriction sites. pCAMBIA1300 contains the hptII
hygromycin resistance gene under the control of the 35S promoter as a transgenic plant
selectable marker.

Plant expression construct pEl778-PIcgt (pE-PIcgt) was assembled by
transferring the PI-cgt gene from pAPC9K into the X7101 and SacI restriction sites of the
pEl778 multiple cloning site (Figure 3.2). pEl778 contains the so-called “super
promoter”, synthetically assembled by combining the noncoding, upstream motifs of the
A grobacterium tumefaciens octopine synthase gene activator and manopine synthase
gene activator promoter (Ni et al., 1995). pEl778 utilizes the nptII gene under the control
of the Nos promoter conferring kanamycin resistance as a transformed plant selective
marker.

For all cloning manipulations, E. coli DHSa was used. Transformants were grown
on LB medium plates containing 100ug/ml ampicillin or 50ug/ml kanamycin. Antibiotic

resistant colonies were analyzed using standard plasmid restriction enzyme digestion and

gel electrophoresis methods.

134

 

Figure 3.1a. pAPC-9K plant expression cassette cloning vector.

Spel Notl xhol Sacl Xmal

 

Actin-Z PE21

pAPC-9K

Figure 3. 1 b. pCAMBIA 1300 plant transformation and gene expression binary vector.

 

 

 

 

 

 

Xba I Xma I
3SS-term
+ {Cansssl am | |-——*
RB "" LB

Figure 3.1. Diagram of the (a) pAPC9K and (b) pCAMBIA 1300 cloning and plant gene
expression vectors. The PI-cgt gene was cloned into the X1101 and Sacl sites of pAPC9K
under control of the Actin2 promoter and PE21 terminator. The Act2-cgt-PE21 cassette
was excised by SpeI and Xmal restriction enzyme digestion and integrated into Xbal and
Xmal restriction sites of pCAMBIA1300. Abbreviations: LB = T-DNA left border; RB =
T-DNA right border; CaMVBSS = cauliflower mosaic virus 358 gene promoter region;

hpt = hygromycin phosphotransferase gene; 35 S-term = CaMV35S terminator region.

135

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Plant transformation and characterization of transgenic plants

The pC-PIcgt and pE-PIcgt binary plant transformation vectors were used to
transform Agrobacterium tumefaciens LBA4404 (Invitrogen, Carlsbad, California) by
electroporation using a Micropulser apparatus according to the manufacturer’s
instructions (Bio-Rad, Hercules, California). Confirmed PIcgt vector containing
Agrobacterium lines were used to transform tobacco (Nicotiana tabacum cv Havana)
using standard leaf disc transformation methods (Horsch et al., 1985). Inoculated leaf

tissues were cultured on semi-solid MS medium (Murashige and Skoog, 1962) containing

25mg/L hygromycin (pC-PIcgt) or 300mg/L kanamycin resistance (pE-PIcgt) at 25°C

 

with 16h light/8h dark regime. Primary transformed shoots (To generation) were rooted
and transferred to soil for seed production under growth chamber or greenhouse
conditions. To determine segregation rates of the antibiotic resistance selective marker
genes as an estimate of T-DNA insertion copy number, T1 seeds were collected from To
regenerants, surface sterilized, and germinated on l/2X MS agar medium containing
25mg/L hygromycin or lOOmg/L kanamycin.

T 1 generation cgt-tobacco and wild type plants were planted in potting soil in
4inch pots and grown to maturity under greenhouse conditions. Leaf area was calculated
from leaf length and width measurements (Walch-Liu et al., 2000) and plant height
determined 45 days after transplanting. The numbers of days to first flower opening for
the different genotypes was recorded for up to ~80 days after potting.

Genomic DNA PCR Analysis of Putative cgt-Tobacco
To confirm the integration of PIch in plant genome, plant genomic DNA was

isolated and purified by using DNeasy Plant Mini Kit (QIAGEN Inc., Valencia, CA)

137

according to the manufacturer’s protocol. For detection of trans genes in regenerated
tobacco, primers specific to the PI-cgt open reading frame (ORF) coding sequence were
used, the forward primer for cgt ORF location nt464-483:

PIcgtl-464f: 5 ’-GATCAATACGGCACATG CCA
and reverse primer for cgt ORF location nt2134-2160:

PIcgtl-r2: 5’-TTA AGGCTGCC AGTTCACTGTAACGGT
to produce an expected PCR partial gene product of 1695bp. The reaction was performed if

in 25ul volume containing ~50ng genomic DNA, 16mM Tris pH 8.3, 40mM KCl,

2.2mM MgClz, lOOuM each dNTP, 0.16uM each primer, 0.016ug bovine serum

 

albumin, and 11.11 of crude-prepped T aq DNA polymerase. PCR reaction conditions were
96°C for 3min; 35 cycles of 96°C 30 sec, 56°C 45 sec, 72°C for 1min; 72°C for 10min
before cooling to 4°C. PCR products were separated on a 0.8% agarose gel, stained with
ethidium bromide, and visualized on a UV transilluminator platform.
RT-PCR analysis of cgt-Tobacco

For reverse transcriptase PCR (RT-PCR), total RNA was extracted from young
leaves of TI plants using the RNeasy Plant Mini Kit (QIAGEN, Valencia, CA). First
strand cDNA synthesis from plant RNA extracts was performed using Superscript II
reverse transcriptase primed with oligo(dT) primer following the manufacturer’s protocol
(Invitrogen, Carlsbad, California). 2.0111 cDNA reaction product was amplified in a total
volume of 25111 containing 3.5mM MgClz, 0.2mM of each dNTPs, 200uM for each of the
forward and reverse primers, and 1.0 unit of T aq DNA polymerase in the buffer supplied
with the polymerase (Invitrogen, Carlsbad, California). For cgt transgene mRNA, RT-

PCR primers were used: PIcgtl-f2 (homologous to cgt ORF bases 1-36):

138

 

5’ATGTTCAAATGGACCAAGCGAATCACTCTCAGTACC
and reverse primer PIcgtl-r2 (complementary to cgt ORF bases 2134 to 2160):

5’-TTAAGGCTGCCAGTTCACTGTAACGGT
PCR amplification conditions were: 96°C for 3min; 96°C 30 sec, 61°C 45 sec, 72°C for
1min for 35 cycles; 72°C for 10 min before cooling to 4°C. The expected fragment size
was 2160 bp. For RT transcript size and specificity confirmation, PCR products were
separated on a 0.8% agarose gel, stained with ethidium bromide, and visualized on a UV
transilluminator platform. Negative control reactions were performed on mRNA extracts
without the 1°'-strand cDNA reaction in parallel RT-PCR assays to test for genomic DNA
contamination artifacts.
Cgt-Tobacco CGTase Analysis

For determination of CGTase activity in tissues 3 different cgt-tobacco lines,
extracts were analyzed via a modified zymograph protocol (Davis, 1964; Kakefuda and
Duke, 1984; Omstein, 1964; Steup and Gerbling, 1983). 10.0g leaf tissue of transgenic
and wild type tobacco plants was homogenized in 30ml 60mM Tris-phosphate, pH 7.0
containing 0.2mM phenylmethylsulfonyl fluoride and 0.1% B-mercaptoethanol (Bio-Rad
Labs, Hercules, CA). Crude extract homogenates were filtered through miracloth
(Calbiochem, La Jolla, CA) and the supernatant cleared by centrifugation at 11000rpm
for 15min. Supernatant protein concentration was determined by the Bradford method
(Bradford, 1976). For PAGE analysis, 150ug protein extract was loaded in each lane of a
non-denaturing gel [50mm x 100mm x 1.5mm thick: 7% acrylamide, 0.184% (w/v)
N,N'-methylenebisacrylamide, 0.28ul/ml TEMED, and 378mM Tris pH 8.9 and 0.07%

(NH4)28203; from Bio-Rad Labs, Hercules, CA]. Electrophoresis was performed at

139

 

15mA/ gel through the stacking gel and 20mA/ gel through the separating gel at 4°C. Each
sample was replicated on the gel so that half the gel could be used for protein staining
(Coomassie Blue) and half for electrophoretic transfer and CGTase activity staining.
Following electrophoresis, the separating gel was subjected to electrophoretic transfer
under non-denaturing conditions into a starch-polyacrylamide gel [starch-PAGE:
15.6mM Trizma-base, 120mM glycine, 1.87mM CaCl2, 0.289u1/ml TEMED, 7% (w/v)
acrylamide, and 0.52% (w/v) soluble starch (Sigma) and polymerized with 0.07%
(NH4)2S203]. Transfer buffer consisted of 25mM Trizma-base (Invitrogen, Carlsbad,
California), l92mM glycine, and 3mM CaCl2 adjusted pH to 8.45. Gels were sandwiched
in direct contact between wetted filter paper and sponge pads and inserted into the
electrophoretic transfer cell. Electrophoretic transfer was at 200mA and 10V at room
temperature for 6h. The starch-PAGE was removed and stained in iodine solution (10mM
KI and 14mM 12).

To determine if PIcgt tobacco plant roots secrete CGTase, T2 seedlings of wild
type and transformed tobacco seedlings were grown in Petri dishes or in GA7 magenta
tissue culture boxes (Sigma-Aldrich, St. Louis, MO) containing semi-solid MS medium
with 0.2% soluble starch (MS+starch) at 25°C with 16h light/8h dark regime. After 4
weeks plant growth, rooted media were stained with iodine solution (10mM KI and
14mM I2). Secreted CGTase was evident by unstained, clear zone around the roots due to

degradation of the starch.

PAH Biodegradation with cg-Tobacco

140

PAH-spiked soil preparation

Clean sand was mixed in equal parts with uncontaminated field soil from the Michigan
State University greenhouse facility (2.7 % soil organic matter, 80% sand, 12% silt, and
8% clay v/v) and sieved through a stainless steel 2.36mm sieve. 2.0kg sand-soil mixture
was spiked with phenanthrene, fluoranthene, and pyrene (0.5% w/v acetone) to achieve
500 mg/kg'1 for each compound. The 2.0kg spiked soil was mixed with 8.0kg clean,
sieved sand to a final concentration for each of the individual PAH of 100 mg/kg".

Wild type and cgt-tobacco (transgenic line PE6) seeds were surface sterilized and
grown for 3 weeks on MS agar medium, plus lOOppm kanamycin for transgenic seeds,
then transferred to MS hydroponics medium without antibiotic for 4 additional weeks. 7-
week old seedlings were transferred to glass test tubes (25x200mm) containing 30g PAH
spiked sand and wrapped with aluminum foil to prevent photodegradation of the PAH
compounds. To provide a bacterial biodegradative consortia, planted and unplanted tubes
were inoculated with 1.0ml bacterial extract of weathered coking substrate at final soil
concentration of 10° total bacterial cells/g soil dry weight (DW) and 104 phenanthrene
degrading cells/ g soil DW (Rugh et al., 2005). The experiment was performed in a
lighted growth chamber at 28+/-3°C with 16h day length (150230 ltla-s"-m'2). Planted
and unplanted control test tubes were placed in racks in a completely randomized design
and harvested in two randomly selected subsets after 30- and 45 -day treatments.

Soil PAH and Plant Biomass Analysis

At the time of harvest, plant and soil was removed from the test tubes. Shoots

were cut and weighed to determine fresh weight (FW). Roots were washed to remove

remaining soil, blotted dry, and weighed. Shoot, root, and soil subsamples were placed in

141

an oven at 80°C for 72h for dry weight (DW) determination. For determination of soil
PAH concentration, 6.0g soil was placed into 40ml amber vials and 20ml
dichloromethane and 6m] saturated KC] added. Sample vials were vortexed for 203ec,
sonicated 10min, and shaken on a rotating platform at 150rpm for 16h. After extraction,
vials were removed from the shaker, and centrifuged for 15min at 1000rpm. The organic
phase was filtered through a glass pipette loosely plugged with glass wool and quantified
for phenanthrene, fluoranthene and pyrene content via gas chromatography-flame
ionization detection (GC-F ID). GC-FID analyses were performed with an Agilent 6890
Gas Chromatograph (Palo Alto, CA) equipped with an Agilent 3396B/C integrator and
Agilent 7683ALS auto injector under the following conditions: ICB-PAH capillary
column 12m x 0.25mm i.d. 0.15pm film thickness (J&K Scientific, Ontario, CAN) with
helium carrier gas, inlet temperature 270°C, flame ionization detector at 330°C. The
initial column temperature was 80°C, the first ramp was at 40°C/min up to 220°C and the
second ramp was at 8°C/min up to 285°C. The injected sample volume was 4ul.
RESULTS

Isolation of CGTase Expressing Bacteria

Bromcresol green (BCG) dye was added to BM-starch plates as a color-shift assay
for CGTase producing colonies from soil extracts. CGTase expressing bacteria were
demonstrated to secrete the CGTase enzyme into the culture media by conversion of
starch to BCD, which subsequently formed a colorless complex with the BCG dye. 41
positive colonies were identified as CGTase expressers by clear zone formation on the
BCG containing agar plate (Figure 3.3). Single colonies of selected isolates were grown

in LB medium overnight and analyzed for CD production via in vitro enzyme assay.

142

 

 

Figure 3.3. Cyclodextrin glycosyltransferase (CGTase) indicator assay of soil bacterial
cultures. CGTase-positive colonies are shown by clear zone formation in Basic Medium
supplemented with starch and bromcresol green. Images in this dissertation are presented

in color.

143

Crude CGTase enzymes were prepared from cleared culture supernatant and incubated
with soluble starch. TLC analysis of the enzymatic reaction medium demonstrated [3CD
production by 16 of the 41 tested isolates (Figure 3.4).

To categorize the CGTase specificity and activity of the CD-positive isolates, a
spectrophotometric method was used to quantify BCD and yCD production based on
complexation with phenolphthalein and bromcresol green, respectively. The relative
amounts of BCD and yCD produced were highly varied among the bacterial isolates
(Figure 3.5).

Paenibacillus illinoisensis cgt Gene Cloning

The 16S rRNA gene of confirmed BCD producing soil-isolated bacteria were PCR
amplified, cloned, and partially sequenced. The 16 different CGTase expressing isolates
were identified as members of Bacillus, Paenibacillus and Brevibacillus genera by
comparison of the partial sequences to the Michigan State University Ribosomal
Database (rdp.cme.msu.edu).

Culture supematants obtained from the soil bacterial isolates 2-8A, 2-8B, 2-13A,
and C36 produced substantially greater amounts of cyclodextrin than the other isolates
and were selected for subsequent cgt gene amplification and cloning efforts. Of the 4
selected CGTase isolates, 2-8A, 2-8B, 2-13A, and C36, only the Paenibacillus
illinoisensis isolate C36 (PI-C36) cgt gene target was successfully PCR amplified using
the designed cgt-specific primers. The full length PI-C36 cgt (PI-cgt) gene was cloned
and sequenced (Figure 3.6). In comparison to cgt sequences in the Genbank database, the
cloned PI-cgt was 77.1%, 76.5%, and 73.2% identical to the Bacillus licheniformis

X15752, B. circulans X68326, and B. sp. X66106 cgt genes, respectively. Computer

144

 

Figure 3.4. Thin-layer chromatograph of CGTase enzyme assay of supematants of
putative CD-expressing bacterial isolate cultures. Isolates 2-8B, 2-13A and C12-B were
observed to be positive for cyclodextrin production. Pure standards of aCD, BCD, and

7CD were run on left side of TLC plate. Images in this dissertation are presented in color.

145

 

 

 

 

(2-6) (2-7) 2-8A 2-88 2-9A 2-11B 2-13A 2-13B G7 B-380

 

Cvclodextrin (uM x102)

 

 

2-15A 2-153 2-16A (2-18) C12A C123 C33 C36 G7 B-380

Bacterial Isolates

Figure 3.5. Cyclodextrin production by bacterial isolate culture supernatants. Stacked

bars show total CD production with specific forms as shown: BCD (I) and yCD (El).

146

ATG ttc aag tgg acc aag cga atc atc ctc agt acc acg ttg agc ttc
agc ttg ctg gcc gga agt gca ctg cca tta ttc ccg gcg gcc tcc gta
ttc gca gat gct gat acg gct gta agc aac aag cag aac ttc agc acc
gat gtt att tat cag gta ttt acc gac cgc ttt ctg gac ggc aat ccc
tcc aac aac cct acc ggg gga gcc tat gac gcc tcc tgt tct aac ctg
aag ctc tat tgt ggg ggc gac tgg caa gga ctg atc aac aag atc aac
gac aac tat ttt tcc gat ctc ggc att acg gca ctc tgg att tcc cag
ccc gtg gag aac att tac tcg ctg att aac tat tcc ggt gtc aac aat
acg gct tat cac ggc tat tgg gca cgt gac ttc aag aag acc aac cct
gct ttc ggt acg atg acc gat ttc cag aat ctg atc aat acg gca cat
gcc aaa ggc att aaa gtc atc att gat ttt gcc ccc aac cat acc tct
cct gcc atg gaa acc gac acc tcg ttc gca gag aat gga aag ttg tac
aat aac ggt acg ctg ctt ggt ggc tac acc aat gat acg aac aaa ctt
ttc cac cac aat ggc ggt tcc gac ttc tcc aca ctg gag aac ggt atc
tac aaa aac ctc tat gat ctc gct gac ctg aat cac aat aac agc acc
att gac acc tac ttc aag gat gcg atc aag ctc tgg ctg gat atg ggt
att gac gga atc cgg gtg gat gcg gtg aag cat atg ccg atg gga tgg
caa aag aac tgg atg tcc tcc atc tat ggc tac aaa ccg gta ttc acg
ttt ggt gaa tgg ttc ctg ggt tca tcc gca tcc gat gcg gac aat acc
aac ttc gcc aac cag tcc ggc atg agt ctg ctt gac ttc cgt ttc aac
aat gaa gta cgt aac gtg ttc cgt gat aac aca tcc acg atg gtt gcc
ctg gat tcc atg atc acc agc act gcc gct gat tat gca cag gtg aat
gac caa gtg act ttt att gat aac cat gac atg gat cgc ttc aaa aca
agt gcc gtt aac aac cgt cgt ttg gaa caa gct ctt gca ttt aca ctt
acc tca cgg ggc gta cca gcc atc tat tat ggc act gag cag tat atg
aca ggt aac ggt gac ccg gat aac cgt gct aaa atg cct tcg ttt tcc
aaa aca aca act gct ttt aac gtc atc agc aag ctg gct cca ttg cgc
aaa acc aat cca gcc atc gcc tat gga aca acc cag cag cgc tgg att
aac aac gat gtg tat gtg tat gaa cgt aaa ttc ggc aac aac gtg gcc
gtc gta gcg gtt aac cgt aat ctt tcg act ccc aca tcc atc agt ggt
ctg acc act tca ctc cct tcg ggt aca tac aac gat gtt ctt gca ggt
gca tta agt ggt aat aac att acc tct aca gga ggt aac gtc gct aac
ttc act ttg gca gca gga gct aca gca gtc tgg cag tat aca gcc aat
acc act aca cca acc atc gga cat gtg ggc cct gta atg ggc aaa gcg
ggt aat acc gtt acc att gat ggt cgt gga ttt ggt act acc aaa ggc
act gtc tat ttc gga acg acc gct gtt acc ggt tca gcc att aca tca
tgg gaa gat act cag atc aaa gtg acc atc cct gca gtt gct gcc ggc
aac tat gca gtc aaa gtc gca gcg agc gga gtc aac agc aac act tat
aac aat ttc acg att ctt agt ggc aac cag gtt tcc gtg cgg ttt gtc
atc aac aat gca tcc acg aca ctc ggt cag aat ctc tat ctg acg ggc
aat gtc gcg gaa ctg ggt aac tgg tcc aca ggc cct ctc gcc att ggt
ccg gcg ttc aat cag gtg atc tat tcc tat cca acc tgg tat tat gac
gtg agt gtg cca gcc gga acg agt cta gaa ttc aag ttt ttc aaa aag
aac ggc tct acg att acg tgg gaa aat ggc aat aat cat act ttc acc
aca cca gcc agt gga acc gca acc gtt aca gtg aac tgg cag cct taa
taa

Figure 3.6. Paenibacillus illinoisensis C36 isolate cyclodextrin glycosyltransferase gene

(cgt) nucleotide coding sequence (2163nt). ATG start codon shown in capital letters.

147

analysis of the PI-cgt open reading frame encodes 719 amino acid residues with a
calculated molecular mass of 78.2 kDa (Figure 3.7). The amino acid sequence from 1 to
23 was deduced to be a signal peptide involved in secretion of the protein (SignalP v3.0)
(Bendtsen et al., 2004). After removal of the signal motif, the mature CGTase would
contain 696 amino acid residues with a predicted molecular mass of 75.66 kDa.

To assay the cloned PI-cgt for CGTase activity, we expressed it in E. coli strain
DHSa. The source bacterium P. illinoisensis isolate C36 (PI-C36), native E. coli DHSa,
and E. coli transformed with the pBS-PI-cgt construct (E. coli-PIcgt) were cultured in
adjacent rows of colonies on LB agar medium, BM+starch, and BM+starch and BCG. All
bacterial lines grew well on all media (Figure 3.8), though only PI-C36 and E. coli-PIcgt
colonies formed clear zones on the starch degradation I2-indicator plates (Figure 3.8b) or
the BCG CD-indicator plates (Figure 380). Clear zones indicated effective CGTase
secretion for starch conversion and CD biosynthesis by both the native and cloned cgt
gene products.

Development of Transgenic Tobacco Engineered for PI-CGTase Expression

Tobacco was genetically engineered with the subcloned P. illinoisensis C36 cgt
(PI-cgt) isolate gene. 15 independently transformed tobacco lines were regenerated from
pEl778-PIcgt (pE-PIcgt) transformed leaf discs and 6 lines regenerated from pCAMBIA-
PIcgt (pC-PIcgt) transformed leaf discs. To regenerated plants were cultivated under
greenhouse conditions for production of T1 generation seed for further analysis. T1 seeds
were screened on antibiotic selective media; pE-PIcgt transformed tobacco seeds (PE
lines) on kanamycin and pC-PIcgt transformed tobacco seeds (PC lines) on hygromycin.

For 10 of the 12 tested transgenic T1 tobacco lines, antibiotic resistant gerrninants

148

MFKWTKRIIL
IYQVFTDRFL
SDLGITALWI
DFQNLINTAH
YTNDTNKLFH
LWLDMGIDGI
ADNTNFANQS
VNDQVTFIDN
TGNGDPDNRA
DVYVYERKFG
NNITSTGGNV
DGRGFGTTKG
SGVNSNTYNN
GPLAIGPAFN
HTFTTPASGT

STTLSFSLLA
DGNPSNNPTG
SQPVENIYSL
AKGIKVIIDF
HNGGSDFSTL
RVDAVKHMPM
GMSLLDFRFN
HDMDRFKTSA
KMPSFSKTTT
NNVAVVAVNR
ANFTLAAGAT
TVYFGTTAVT
FTILSGNQVS
QVIYSYPTWY
ATVTVNWQP

GSALPLFPAA
GAYDASCSNL
INYSGVNNTA
APNHTSPAME
ENGIYKNLYD
GWQKNWMSSI
NEVRNVFRDN
VNNRRLEQAL
AFNVISKLAP
NLSTPTSISG
AVWQYTANTT
GSAITSWEDT
VRFVINNAST
YDVSVPAGTS

SVFADADTAV
KLYCGGDWQG
YHGYWARDFK
TDTSFAENGK
LADLNHNNST
YGYKPVFTFG
TSTMVALDSM
AFTLTSRGVP
LRKTNPAIAY
LTTSLPSGTY
TPTIGHVGPV
QIKVTIPAVA
TLGQNLYLTG
LEFKFFKKNG

SNKQNFSTDV
LINKINDNYF
KTNPAFGTMT
LYNNGTLLGG
IDTYFKDAIK
EWFLGSSASD
ITSTAADYAQ
AIYYGTEQYM
GTTQQRWINN
NDVLAGALSG
MGKAGNTVTI
AGNYAVKVAA
NVAELGNWST
STITWENGNN

Figure 3.7. Paenibacillus illinoisensis strain C36 CGTase amino acid sequence,

calculated molecular mass of 78.2 kDa derived from the nucleotide sequence. Putative

signal peptide cleavage site indicated by arrow.

149

     
   

Figure 3.8a

 

P. illinoisensis E. coli cgt-E. coli

Figure 3.8c

 

P. illinoisensis E. coli cgt-E. coli

Figure 3.8 (a) LB, (b) starch degradation assay, and (c) BCG CD-indicator assay of
Paenibacillus illinoisensis, E. coli strain DHSa, and transformed cgt-E. coli DHSa.

Images in this dissertation are presented in color.

150

segregated in a ~3:l ratio to sensitive seeds indicating single T-DNA copy numbers for
these lines (Table 3.1). By contrast, all plated seeds from PC4 and PEl tobacco lines
were tolerant to antibiotic selection, suggesting that these two lines may possess multiple
T-DNA integration events, since single transgene insertions display Mendelian
segregation in the T1 generation. Antibiotic resistant T1 progeny were analyzed further
for the presence of the PI-cgt gene.

T1 generation PI-cgt transformed tobacco plants displayed accelerated growth and
development relative to wild type plants. 45 days after planting, several transformed lines
had larger leaf area values (Figures 3.9a and 3.9b) and plant height (Figures 3.10 and
3.12a) than wild type plants. In addition, transformed plant lines began flowering an
average of 3-10 days faster than wild type plants (Figure 3.11). This was especially true
for the PE-PIcgt-tobacco lines, which appeared to be developmentally precocious relative
to PC-PIcgt or nontransgenic tobacco plants. By the time of flower development (~60-
70days) (Figure 3.12b), wild type plants had grown to the same height and possessed
similar leaf areas as transformed plants.

Using primer pairs specific to the PI-cgt gene sequence, a 1695bp fragment was
PCR amplified from genomic DNA of transformed plants, demonstrating integration of
T-DNA containing PI-cgt into the plant genome (Figure 3.13). 13 of 15 lines regenerated
pEl778-PIcgt (PE-PIcgt) lines and all 6 regenerated pCAMBIA-PI—cgt (PC-PIcgt) lines
tested positive for cgt transgene integration via genomic DNA PCR analysis.

Detection of cgt mRNA transcript by RT-PCR assay
For RT-PCR analysis we isolated RNA and produced cDNA from leaf tissue of

confirmed transgenic, T1 generation cgt-tobacco plants. Using DNA primers

151

Table 3.1. Selective marker segregation analyses of T1 generation cgt-tobacco seeds

 

R

S

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

To Parent Lines Hyg Hyg Segregation ratio
PCI 43 18 ~3:1
PC2 52 15 ~3:1
PC3 47 18 ~3:1
PC4 59 0 all positive
PCS 53 14 ~3:1
PC6 55 12 ~3:1

To Parent Lines KanR KanS Segregation ratio
PE] 63 0 all positive
PE2 39 14 ~3:1
PE3 54 17 ~3:1
PE4 44 15 ~3:1
PE5 50 14 ~3:1
PE6 55 15 ~3:1

 

PE = cgt-tobacco lines transformed with pE-PIcgt construct
PC = cgt-tobacco lines transformed with pC-PIcgt construct
Hng = transformed plant lines resistant to hygromycin
HygS = transformed plant lines susceptible to hygromycin
KarlR = transformed plant lines resistant to kanamycin

KanS = transformed plant lines susceptible to kanamycin

152

 

Figure 3.93.

—O
N

O

1___l___z;

00

h

Leaf area (cmz)
O\

N
1

Figure 3.9b.

 

;

 

Ho‘
93

 

 

 

 

j—lm

 

 

HC‘

 

 

Hm

 

 

HU"

 

 

Hm

 

 

 

Ho‘

 

 

 

l-—-Icr

 

 

 

 

 

1

1

PC 1 PC2 PC3 PC4 PCS PC6 WT PE] PE2 PE3 PE4 PE5 PE6

Hm

 

 

 

 

j—ic‘

 

 

l—W‘

a)

 

 

 

 

j-Icr

 

 

7—10‘

 

 

Her

1—t~

 

 

 

 

 

 

]-—|O"

 

 

fl-w‘

 

 

 

l

l

r

Plant lines

PCl PC2 PC3 PC4 PCS PC6 WT PEl PE2 PE3 PE4 PE5 PE6

Figure 3.9. Average leaf areas of the (a) 3rd leaf and (b) 4th leaf (counted from the base of

stem) of wild type and cgt-tobacco plant lines 45 days after planting. WT =

untransformed wild type, PC = independently transformed pC-Act2-PIcgt lines, PE =

independently transformed pE-(super promoter)-PIcgt lines. Different letters indicate

statistically different treatments (or = 0.05). Error bars = standard deviation.

153

127

10'1 l l

o.
a.
o.
0"
H”
J—lcr

Plant Height (cm)
O\
H
j—l
1!
O)
j—I m

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

0 I l l l I l l T l l I
PCl PC2 PC3 PC4 PCS PC6 WT PEI PE2 PE3 PE4 PE5 PE6

Figure 3.10. Average plant height of wild type and cgt plant lines 45 days after planting.
WT = untransformed wild type, PC = independently transformed pC-Act2-PIcgt lines, PE

= independently transformed pE-(super promoter)-PIcgt lines. Different letters indicate

statistically different treatments (or = 0.05). Error bars = standard deviation.

154

m

0"
|_.
H»

Days to flower
0\
N

max
GOO

 

 

 

 

 

 

 

 

l
l
l
l
l
l
l
l
564‘ _j
l
54‘ l' T— i" l l l l

 

 

PCl PC2 PC3 PC4 PCS PC6 WT PEI PE2 PE3 PE4 PE5 PE6
Plant lines
Figure 3.11. Average number of days from planting to flower of wild type and cgt plant
lines. WT = untransformed wild type, PC = independently transformed pC-Act2-Plcgt
lines, PE = independently transformed pE-(super promoter)-Plcgt lines. Different letters

indicate statistically different treatments (or = 0.05). Error bars = standard deviation.

155

Figure 3.12a.

 

Figure 3.12b.

 

Figure 3.12. Growth and development comparison of wild type and transgenic cgt-
tobacco lines. Shown are representative tobacco plants displaying differences in (a)
shoot height (~45 days after planting) and (b) flowering morphology (~65 days after

planting). Images in this dissertation are presented in color.

156

MWM WT PE9 PEIO PCS PE15 PE5 PE13 PE11PE12 PE14

 

Figure 3.13. Genomic PCR analysis of putative cgt-tobacco transforrnants.

Lane 1 = DNA marker (MWM); lane2 = wild type (WT); lanes 3-11 = kanarnycin-
resistant, putative transgenic cgt-tobacco lines. The cgt gene specific primers generate an
expected PCR amplification product of 1695bp in length. The 1636bp DNA marker band

is indicated by the arrow.

157

complementary to each end of the PI-cgt gene coding sequence, the expected 2160bp RT-
PCR product confirmed the presence of cgt mRNA in tested transgenic leaf tissues
(Figure 3.14). All cgt-tobacco lines that had tested positive for PIcgt transgene insertion
via genomic PCR analysis were confirmed to transcribe cgt mRNA. RT-PCR analysis of
mRNA not having undergone the cDNA reaction resulted in no amplification products
indicating there was no residual genomic DNA contamination in these extracts.

Plant CGTase Activity Assay

Crude leaf protein extracts were separated by non-denaturing PAGE, electro-
transferred to a starch-acrylamide gel, which was then stained with iodine solution (3.15).
The three tested cgt-tobacco lines produced distinct unstained, clear areas in the
zymograph gel, indicating starch degradation (Figure 3.15b). Faint, cleared bands were
also detected at other molecular weights in both transgenic and untransformed wild type
plant extracts, likely from additional amylolytic enzymes naturally occurring in tobacco
tissues.

Whole plants were also analyzed for starch degradation in rooted starch-agar
medium. Two independent lines of cgt-tobacco plants were observed to create unstained,
clear zones around roots growing in iodine-treated, semisolid medium (Figure 3.16). In
contrast, wild type tobacco planted starch-medium stained very dark upon addition of
iodine solution with no observed clear areas (Figures 3.16b and 3.16d). These results
showed that cgt-tobacco plants effectively secreted CGTase from roots into the
surrounding medium with subsequent starch degradation and apparent conversion to
cyclodextrin.

PAH Biodegradation

158

 

PEI PEll PCl PC12 PC2 WT

 

 

Figure 3.14. RT-PCR analysis of cgt tobacco leaf extracts. Tested putative lines were
tested via for RT—PCR and negative control reactions and samples for individual line ran
in adjacent lanes. For each plant line are shown RT-PCR products (+ ; Lanes 3, 5, 7, 9, 11
and 13) and in adjacent lefi lanes (- ; Lanes 2, 4, 6, 8, 10 and 12) are negative controls
(RNA extracts without 1St strand cDNA reaction). Lane 1 = DNA marker (MWM); Lane
12 and 13 = wild type (WT). RT-PCR expected size = 2160bp. The 2036bp marker band

is indicated by the arrow.

159

Figure 3.15a. Native PAGE Figure 3.15b. Zymogram of leaf extracts

\\'I' I’ll] l’ICZ l’(‘l

   

Figure 3.15. Native-PAGE and zymogram analysis of CGTase enzyme from cgt-tobacco
leaves: (a) PAGE of leaf protein extract stained with coomassie brilliant blue, (b)
Zymogram of native leaf proteins electro-transferred from PAGE to a polyacrylamide-
starch gel and stained with iodine solution (10mM KI, 14mM 12) to test for starch
degradation via CGTase activity. Lane 1 = wild type tobacco (WT); Lane 2 = PEI , Lane
3 = PE2, and Lane 4 = PCl independently transformed lines of cgt-tobacco. Images in

this dissertation are presented in color.

160

     

Figure 31.6212- Fioure 3.16c.

ll v

 

Figure 3.16b. Figure 3.16d.

   

Figure 3.16. Starch degradation analysis of wild type and cgt-tobacco plants. Petri plate
of WT and PCI plants grown on MS+starch medium (a) before and (b) afier staining for
medium starch content with iodine solution. GA7 tissue culture boxes of WT and PE6
plants grown on MS+starch medium (c) before and (d) afier staining for medium starch
content with iodine solution. Note: plant shoot tissues were excised prior to staining to
assist with visualization of media coloration. WT = non-transgenic wild type plants, PC]
= independently transformed pC-Act2-Plcgt line, PE6 = independently transformed pE-

(super promoter)-Plcgt line. Images in this dissertation are presented in color.

161

The starting tPAH concentration of the GC-FID analysis of the spiked sand-soil mix was
determined to be 275mg/kg soil DW with equivalent concentrations of the individual
PAH compounds phenanthrene, fluoranthene, and pyrene (~92mg/kg soil DW). After 30
days (T1) treatment, tPAH concentrations were decreased 95, 91 and 81% in unplanted,
wild type planted, and cgt-tobacco PE6 planted soils, respectively (Figure 3.17). After 45
days (T2), all treatments had removed most (~95-97%) of the added PAH content of the
soil. Soil tPAH concentrations were significantly different between each of the unplanted
and planted treatments at both sampled treatment times, though minimally so (~2-4ppm)
at the T2 time point. Rates of the reduction among the individual compounds appeared to
be inversely related to compound molecular weight (phen < flra < pyre), though not
significantly so due to substantial variation among treatments (Table 3.2). Comparison of
average individual PAH compounds demonstrates an apparent tendency for more rapid
removal of the lower molecular weight, 3-ringed phenanthrene relative to the 4-ringed
PAHS, fluoranthene and pyrene, particularly among the T1 planted soil samples.

Plant biomass was not significantly different between the two genotypes at either
sampling event. Wild type and cgt-tobacco whole plant DW biomass was 0.6g and 0.5 g
at T1 and 1.5g and 1.2g at T2, respectively.

DISCUSSION

We obtained a variety of cyclodextrin producing bacterial isolates from PAH
contaminated and uncontaminated, clean soils. Paenibacillus illinoisensis (formerly
classified as Bacillus circulans) is a gram-positive bacterium and was found to express a

CGTase observed to produce mostly BCD fiom soluble starch. The cgt gene from PI-C36

was cloned and expressed in E. coli strain DHSOL. The first 20-35 amino acid residues of

162

g _ g 30 days 1345 days

 

 

 

 

’53
h
3 ,,7‘A, ,4 4___._
9:1 2,, ,,_~v,z
'8 7,,,__-,,____ a,
m

_ a AI

-= -2 -,

UNP

 

Figure 3.17. tPAH concentration in spiked sand-soil mix after planted and unplanted
treatments for 30 days (T1) and 45 days (T2). TZ = Time zero, untreated soil; UNP =
unplanted; WT = non-transgenic wild type tobacco; PE6 = pE-(super promoter)-P1cgt
tobacco. Different letters indicate significant difference (or 5 0.05) between treatments
(lower case for T., upper case for T2); * indicates significant difference (or 5 0.05)

between T1 and T2 values (observed only for PE6 treatment). Error bars = SD.

163

Table 3.2. Phenanthrene (phen), fluoranthene (flra), pyrene (pyre), and summed PAH
(tPAH) concentrations in spiked sand-soil mix (100 ttg/g each at T0) after planted and
unplanted treatments for 30 days (T1) and 45 days (T2). PE6 = pE-(super promoter)-PIcgt
tobacco. Different letters indicate significant difference (or 5 0.05) between soil

treatments (within a column) for each individual PAH compound.

 

 

 

 

 

 

 

 

 

 

T1 phen flra Pyre tPAH
unplanted 3-7 5-0 5.5 14.2
a a a a
wild e 5.7 11.2 12.3 29.2
typ a ab ab ab
PE6 9.6 19.9 22.8 52.3
a b b b
T2 phen flra Pyre tPAH
unplanted 2-6 3-1 3 .2 8.9
a a a a
b a a ab
PE6 4.5 4.2 4.0 12.7
c a a b

 

 

 

 

 

164

 

the cgt product is an N-terrninal signal peptide directing translocation of CGTase to the
endoplasmic reticulum (ER) and subsequent secretion from the cell (Binder et al., 1986;
Fujiwara et al., 1992; Kim et al., 1998; Lawson et al., 1994). The cgt-transformed E. coli
were demonstrated to secrete CGTase into the medium, though with smaller clear zones
than PI-C36, possibly caused by partial trapping of the expressed PI-CGTase in the E.
coli periplasmic space due to an additional outer membrane in gram-negative bacteria
(Kim et al., 1998). CGTases from Bacillus sp. E1 (Yong et al., 1996), Brevibacillus
brevis CD162 (Kim et al., 1998), and Klebsiella pneumoniae (Binder et al., 1986) have
been expressed in E. coli and shown to be secreted into the growth medium. However,
secretion efficiencies likely vary among different CGTases, including the PI-C36 cgt
product.

The PI-cgt was subcloned into plant transformation vectors and expressed in
tobacco under control of the Arabidopsis ath promoter and the synthetic, Agrobacterium
multi-gene “super promoter”. CGTase was detected in crude protein preparations from
transgenic tobacco leaf tissue by an amylolytic zymograph assay. Iodine staining of
rooted culture medium demonstrated enhanced starch degradation by the intact roots of
two different transgenic lines, PC] and PE6. Starch medium clearing indicates that cgt-
transformed tobacco plants effectively secrete CGTase and the native signal peptide of
PI-C36 is functional in the plant secretory pathway.

Some transgenic plant lines displayed accelerated vegetative grth and
development relative to wild type plants. Nearly all PIcgt-tobacco had higher leaf area
and height than wild type plants during the vegetative grth stage (<60days).

Additionally, the PIcgt-tobacco reached flowering stage about 3 to 10 days faster than

165

wild type. Plant expression of the bacterial CGTase could possibly perturb plant starch
metabolism affecting plant growth and development. Perturbation of starch and sucrose
metabolism could consequently alter plant grth and development (Zeeman et al.,
2004). We did not quantify plant tissue starch, reducing sugars, BCD or chlorophyll
content in this project, though these metabolic products should be analyzed in future
studies to improve our understanding of changes in development of transformed plants.
Another possible cause of transgenic plants growth enhancement is positive feedback
from biostimulated soil microbes. Transformed plants secrete CGTase to produce CDs in
rhizosphere, potentially serving as a readily available carbon source, supporting microbial
multiplication or altering community diversity and/or population structure. CDs may
further facilitate plant development by stimulation or stabilization of microbially
produced plant growth promoting compounds.

The PAH phytoremediation trial using cgt-tobacco was inconclusive. Contrary to
expectations, PIcgt-tobacco plants displayed delayed PAH dissipation relative to
unplanted and wildtype planted treatments. At the 30-day treatment time point (T1),
spiked soil tPAH concentrations had been reduced >80% by all treatments, though tPAH
in cgt-planted soils was about 10% higher than for wild type and unplanted treatments.
By the 45-day sampling event (T2), all treatments displayed nearly total elimination of the
added PAH with only 3-4.5% remaining in the treated soils. Cgt-tobacco displayed the
least contaminant reduction relative to wildtype or unplanted soil treatments. Since
bacterial biodegradation is the primary mechanism in PAH phytostimulation, it appears
that tobacco root may have at least temporarily inhibited or altered bacterial PAH

metabolism relative to the unplanted soil treatment. Given the very high bioavailability of

166

the contaminant in the spiked soils, both plant genotypes — wild type and cgt-tobacco -
may have temporarily stabilized PAHS by sorption to root biomass. In addition, cgt-
tobacco could have further stabilized the target pollutants by sequestration in
cyclodextrin inclusion complexes.

The extremely labile nature of the PAH amendment in the experimental soil
medium is in sharp contrast to PAH contaminant behavior under field conditions. In most
contaminated field sites, PAHS are very stably sorbed to soil organic matter, resulting in
prolonged environmental persistence (Amellal et al., 2001; Cemiglia, 1993; Chiou et al.,
1998). For example, PAH spiked soils aged for 6 months were observed to dramatically
decrease PAH dissipation rates relative to freshly spiked soils (Binet et al., 2000). In our
study, spiked PAHs in the prepared soil medium were highly bioavailable and easily
dissipated, due most probably to both the very low soil organic matter (~0.6%) and lack
of “weathering” pretreatment. In addition, all treatments were augmented with a coke
oven soil bacterial extract inoculum of a relatively high density of PAH biodegrading
cells (~104 ZFU/ g soil DW) with minimal other C substrates available. The rapid PAH
dissipation observed in the unplanted treatments indicate that the reduction observed
among all treatments, including the planted soils, was principally via independent
bacterial biometabolism with some fraction possibly lost by compound volatilization.
Due to lack of uninoculated control treatments, we can not distinguish endogenous versus
bioaugrnented or biotic from abiotic influences on PAH reduction in this trial.
Experiments using weathered, contaminated field soils and weathering-pretreated, spiked

soils should be performed in combination with gnotobiotic and bioaugrnented treatments

167

 

to more clearly evaluate the effect of cgt-tobacco on PAH biodegradation in
contaminated soil media.
CONCLUSION

We isolated a cyclodextrin glycosyltransferase (CGTase) expressing soil
bacterium identified as Paenibacillus illinoisensis isolate C36 (PI-C36). The PI-C36
isolate was demonstrated to secrete CGTase enzyme into its grth medium and convert
starch preferentially to B-cyclodextrin (BCD). The PI-C36 cgt gene, encoding CGTase,
was cloned and expressed in transgenic E. coli and tobacco. Both cgt-E. coli and cgt-
tobacco lines secreted functional CGTase into growth medium with subsequent starch
conversion to BCD. Preliminary PAH phytoremediation studies using transgenic cgt-
tobacco were inconclusive for evidence of enhanced biodegradation of PAH-impacted
soils. Additional experiments are required to better evaluate the influence of transgenic
plant expression of bacterial CD biosynthetic proteins on bacterial community response

and PAH contaminant biodegradation rates.

168

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