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Master’s dissertation submitted in partial fulfilment of the requirements for the joint degree of

International Master of Science in Environmental Technology and Engineering

an Erasmus+: Erasmus Mundus Master Course jointly organized by Ghent University, Belgium University of Chemical Technology, Prague, Czech Republic UNESCO-IHE Institute for Water Education, Delft, the Netherlands

Academic year 2014 – 2015.

Induction of biphenyl degradation pathway genes by plant secondary metabolites

Host University:

University of Chemical Technology, Prague, Czech Republic

Binyam Woldehawariat

Promotor: Assoc. Prof. Ondřej Uhlík, MSc., Ph.D.

This thesis was elaborated at University of Chemical Technology, Prague, Czech Republic and defended at University of Chemical Technology, Prague, Czech Republic within the framework of the European Erasmus Mundus Programme “Erasmus Mundus International Master of Science in Environmental Technology and Engineering " (Course N° 2011-0172)

© 2015 Prague, Czech Republic, Binyam Woldehawariat, Ghent University, all rights reserved. ii

Deze pagina is niet beschikbaar omdat ze persoonsgegevens bevat. Universiteitsbibliotheek Gent, 2021.

This page is not available because it contains personal information. Ghent University, Library, 2021. DECLARATION

This thesis/dissertation was written at the Department of Biochemistry and Microbiology of the University of Chemical Technology, Prague, Czech Republic from February to August 2015.

I hereby declare that this thesis is my own work. Where other sources of information have been used, they have been acknowledged and referenced in the list of used literature and other sources.

I have been informed that the rights and obligations implied by Act No. 121/2000 Coll. on Copyright, Rights Related to Copyright and on the Amendment of Certain Laws (Copyright Act) apply to my work. In particular, I am aware of the fact that the University of Chemical Technology in Prague has the right to sign a license agreement for use of this work as school work under §60 paragraph 1 of the Copyright Act. I have also been informed that in the case that this work will be used by myself or that a license will be granted for its usage by another entity, the Institute of Chemical Technology in Prague is entitled to require from me a reasonable contribution to cover the costs incurred in the creation of the work, according to the circumstances up to the full amount.

I agree to the publication of my work in accordance with Act No. 111/1998 Coll. on Higher Education and the amendment of related laws (Higher Education Act).

In Prague on August 12, 2015

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Acknowledgements

I would like to thank Assoc. Prof. Ondřej Uhlík for his professional guidance, support and patience throughout this diploma thesis and beyond. I am indebted for his helpful suggestions during the course of my research. His advice and expertise were invaluable to accomplish my diploma thesis.

I would like to express my deepest appreciation to my tutor, Michal Strejček, MSc. for his time, patience and all the support he offered me in realizing my diploma thesis. He went beyond in helping and guiding me through the most important and difficult part of my research work. Special thanks to Lucie Musilová for all the support you gave me to accomplish this work. I am also grateful to Eglantina, Serena, Jáchym, Honza and the rest, for the support, advice and comforting thoughts in these past 6 months.

My heartfelt gratitude to the IMETE Management Board for making IMETE program possible. Thank you to Ineke Melis of UNESCO-IHE, Ing. Jana Bartáčková, Ph.D. of ICTP and Evelien Vandevelde of Ghent for handling all arrangements and support in Delft, Prague and Ghent. Special gratitude to European Commission for sponsoring my study with Erasmus Mundus scholarship.

Finally, I would like to thank and give special tribute to my wife, Sossena whose constant moral and emotional support has guided me to reach at this stage of my career. Thanks also to my mother and my friends for always being there for me and encouraged me to keep on with my dreams.

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Abstract

Many previous studies have reported that polychlorinated biphenyls (PCBs) are degraded more efficiently in vegetated soils when compared to non-vegetated soils. These findings led to the hypothesis that the enhancement in degradation could be attributed to the presence of plant secondary metabolites (PSMs) in vegetated soils, released by plant roots. Recent laboratory studies have also demonstrated that PSMs may serve as natural substrates and/or inducers for the biphenyl catabolic pathway.

Thus, the main objective of the present study was to investigate biphenyl catabolic pathway induction potential of selected PSMs in four bacterial strains: Achromobacter denitrificans AD400, Pseudomonas alcaliphila JAB1, Achromobacter xylosoxidans S3, and Pseudomonas putida S9. By identifying PSMs that could be used instead of biphenyl, we hope to contribute in the development of new approaches for bioremediation of PCB-contaminated soils. In the present study, all the strains were grown co-metabolically on sodium pyruvate plus one of the 14 different PSMs investigated. We confirmed the induction of the biphenyl catabolic pathway by the PSMs using quantitative real-time polymerase chain reaction (RT-qPCR) of the bphA gene. A mathematical model from literature was used to quantify the relative expression of bphA gene with respect to the control (sodium pyruvate).

The RT-qPCR results showed that for strains JAB1 and S3 all the PSMs were able to induce the bphA gene, whereas in strain AD400 all except p-cymene were able to induce bphA. However, none of the PSMs were able to induce bphA in strain S9. The investigation of the relative quantification of bphA gene showed that a number of the PSMs tested were able to induce the bphA significantly higher (p < 0.05) than the control. Additionally, in strain AD400, coumarin induced bphA gene significantly higher (p < 0.05) than biphenyl, whereas in strain JAB1 ferulic acid and p-cymene were able to induce bphA gene significantly higher (p < 0.05) than biphenyl itself. Naringin, α-pinene, p-hydroxybenzoic acid, vanillic acid, caffiec acid, , and carvone are other significant inducers identified in the present study.

In conclusion, the present study demonstrated that a number of PSMs have the potential to induce the biphenyl catabolic pathway at the same level or even higher than biphenyl itself. Thus, considering the fact that some PCBs can also be degraded by the same catabolic pathway, PSMs have the potential to replace biphenyl as a proper inducer for the bioremediation of PCB- contaminated sites.

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Table of Contents

Acknowledgements ...... v

Abstract ...... vi

List of Figures ...... ix

List of Tables ...... x

Abbreviations ...... xi

Chapter 1: Introduction ...... 1

Chapter 2: Literature review ...... 2

2.1 Polychlorinated biphenyls (PCBs) ...... 2

2.2 Environmental fate of PCBs ...... 3

2.3 Potential adverse effects of PCBs ...... 4

2.4 Remediation techniques of PCB-contaminated soils ...... 4

2.4.1 Physicochemical methods ...... 4

2.4.2 Bioremediation ...... 6

2.4.3. Plant secondary metabolites ...... 9

2.4.4. Rhizodegradation ...... 10

Chapter 3: Objectives ...... 12

Chapter 4: Material and Methods...... 13

4.1 Materials ...... 13

4.1.1 Chemicals ...... 13

4.1.2 Primers, enzymes and commercial kits ...... 13

4.1.3 Bacterial strains ...... 14

4.2 Methods...... 14

4.2.1 Cultivation media preparation ...... 14

4.2.2 Preparation of liquid growth medium ...... 14

4.2.3 Preparation of PSMs and biphenyl solutions ...... 14

4.2.4 Preparation of lysozyme stock solution ...... 14

4.2.5 Cultivation of bacterial strains on biphenyl ...... 15

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4.2.6 Assessment of bacterial strains growth on PSMs ...... 15

4.2.7 Assessment of PSMs ability induce the biphenyl catabolic pathway ...... 15

4.2.8 RNA isolation ...... 16

4.2.9 cDNA synthesis ...... 17

4.2.10 Quantification of 16S rRNA and bphA genes ...... 17

Chapter 5: Results ...... 19

5.1. Assessment of bacterial strains growth on PSMs ...... 19

5.2. Induction of bphA gene by PSMs ...... 21

Chapter 6: Discussion ...... 29

Chapter 7: Conclusion ...... 37

References ...... 39

Appendix A ...... 46

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List of Figures

Figure 1.1. Molecular structure of biphenyl ...... 2 Figure 1.2. Pathway for biphenyl biodegradation ...... 8 Figure 5.1. Growth curve of Achromobacter denitrificans AD400...... 20 Figure 5.2. Growth curve of Pseudomonas alcaliphila JAB1...... 20 Figure 5.3. Growth curve of Achromobacter xylosoxidans S3...... 21 Figure 5.4. Growth curve of Pseudomonas putida S9...... 21

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List of Tables

Table 4.1. qPCR reaction mixtures ...... 18 Table 5.1. Relative expression of bphA gene induced by biphenyl and PSMs by strain AD400...... 23 Table 5.2. Relative expression of bphA gene induced by biphenyl and PSMs by strain JAB1...... 24 Table 5.3. Relative expression of bphA gene induced by biphenyl and PSMs by strain S3...... 25 Table 5.4. Amplification of 16S rRNA gene in strain S9...... 28

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Abbreviations

16S rRNA 16S Ribosomal Ribonucleic acid cDNA Complimentary Deoxyribonucleic acid

CP Crossing Points DNA Deoxyribonucleic acid dNTPs Deoxyribonucleoside Triphosphates EDTA Ethylenediaminetetraacetic acid HCl Hydrochloric Acid M-MuLV Moloney Murine Leukemia Virus OD Optical Density PAHs Polycyclic Aromatic Hydrocarbons PCA Plate Count Agar PCB POPs Persistent Organic Pollutants PSMs Plant Secondary Metabolites RNA Ribonucleic Acid RT Reverse Transcriptase RT-qPCR Real time Quantitative Polymerase Chain Reaction TE Tris-EDTA

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Chapter 1: Introduction

The expansion of modern chemical industries have resulted in the release of huge amounts of various synthetic persistent organic pollutants (POPs), such as pesticides, polyaromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), etc. PCBs are known to resist, to a varying degree, photolytic, chemical, and biological degradation. PCBs are of environmental concern because of their toxicities, persistence, bioaccumulative and biomagnification properties and potential for long-range atmospheric transport to remote areas and regions where they have never been used.

Humans can be exposed to PCBs through diet, occupational accidents and the environment. Exposure to PCBs can be associated with a wide range of adverse health effects both to humans and animals, including endocrine disruption, reproductive and immune dysfunction, neurological disorders and cancer. This issue has raised concern among the international community, which led to the adoption of the Stockholm Convention in 2011 with the aim of banning or controlling the production of POPs. However, the persistent nature of POPs, including PCBs, means these recalcitrant chemicals remain to be a public health treat.

Several methods, including physicochemical and biological methods have been tested for the remediation of PCB contaminated sites. While the physicochemical methods exhibited high potential for efficient remediation of highly contaminated sites in relatively short period time, they have proved to be expensive, energy demanding, and damaging to the integrity of the soil. In order to address these problems, researchers have been investigating biological methods: anaerobic dechlorination and aerobic degradation of PCBs by co-metabolism with biphenyl. While laboratory studies show that aerobic degradation of lower chlorinated PCBs with biphenyl as a co-substrate appear to be very promising, unfortunately due to potentially harmful effects of biphenyl to the environment such techniques cannot be applied in situ. As a result, recently many researchers have shifted their focus on testing alternative cometabolites or inducers such as plant secondary metabolites for in situ PCB bioremediation. The findings of these studies show that the induction of PCB degrading genes with plant secondary metabolites is very promising. Thus, the objective of the present study is to investigate the potential of selected plant secondary metabolites to induce biphenyl catabolic pathway in selected PCB- degrading strains.

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Chapter 2: Literature review

2.1 Polychlorinated biphenyls (PCBs) PCBs are a class of 209 chlorinated molecules consisting of two rings joined by carbon-carbon bonds at C-1, 1’ position. Each benzene ring can have up to 5 substituents of chlorine at the ortho, meta or para positions as shown in Figure 1 (Wiegel and Wu, 2000; Xu et al., 2014), which are called congeners. The congener groups range from the three monochlorinated to a fully chlorinated decachlorobiphenyl (Vasilyeva, and Strijakova, 2007). Depending on the number and position of the chlorine atoms, the different congeners of PCBs exhibit a wide range of physical/chemical properties (e.g. water , vapour pressure, lipophilicity, viscosity, odour and colour) (Montone et al., 2001; Passatore et al., 2014). The effects of such differences in properties of PCBs can be observed in their mobility and distribution in the environment (Eckhardt et al., 2007). PCBs are also known for their high stability under most environmental conditions, although there are substantial differences among different isomers and congener groups (Singer et al., 2003; Van Aken et al., 2009; Passatore et al., 2014). Generally, the degradation potential of PCBs in the environment tend to decrease as their chlorine content increases, resulting in increased persistence for higher chlorinated biphenyls.

Figure 1.1. Molecular structure of biphenyl (Wiegel and Wu, 2000)

Since their first commercial production in the late 1920s, it is estimated that more than 2 million tonnes of PCBs have been produced worldwide (Breivik et al., 2004). PCBs had got their popularity owing to their enhanced chemical stability, low water solubility, low flammability, and excellent insulating properties, which led into their wide industrial and commercial applications as heat transfer fluids, flame retardants, lubricating and hydraulic oils, additives in paints, and rubber products, and in many other applications (Breivik et al., 2004;

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Borja et al. 2005; Passatore et al., 2014). Their unrestricted production and use up until the 1970s resulted in the widespread distribution of PCBs in the environment causing environmental toxicity. The persistent nature of PCBs in the environment combined with their toxicity and biomagnification in food chains led to the ban in their production by the United States Congress in 1979 and by the Stockholm Convention on Persistent Organic Pollutants in 2001 (Wiegel and Wu, 2000; Gomes et al., 2013; Tang, 2013; Xu et al., 2014).

Although the use of PCBs is currently restricted or controlled, PCBs are still widely distributed in the environment mainly due to environmental cycling process of previously released PCBs (Van Aken et al., 2009; Abella et al., 2015; Putschögl et al., 2015). In addition, PCBs can still be released into the environment from poorly maintained hazardous waste sites that contain PCBs, leaks or releases from electrical transformers containing PCBs and disposal of PCB- containing consumer products into municipal or other landfills not designed to handle hazardous waste (Breivik et al., 2007; Lu et al., 2015). PCBs may also be released into the environment by burning some wastes in municipal and industrial incinerators (Putschögl et al., 2015).

2.2 Environmental fate of PCBs The fate of PCBs in the environment is a function of a number of chemical, physical, and biological processes and/or properties, including water solubility, octanol:water partitioning coefficient, vapour pressure, degree of chlorination, adsorption to soils, atmospheric oxidation, hydrolysis and oxidation in water, photolysis, biodegradation, etc. (Van Aken et al., 2009). If released to soil, PCBs and especially highly chlorinated congeners adsorb to the soil tightly due to their high octanol:water partitioning coefficient (Kow) (Faroon et al., 2003; Passatore et al., 2014). If released to water/sediment, PCBs with greater amounts of chlorine, readily adsorb to organic matter (dissolved, suspended, and fatty tissues of aquatic biota) and are removed from the water column into the sediment (Abella et al, 2015; Lu et al., 2015). This phenomenon plays an important role in the immobilization of PCBs in aquatic environments. On the other hand, PCBs can readily leach through the soil in the presence of organic . Vapour loss of PCBs from soil and water surfaces appears to be another important mechanism with the rate of volatilization decreasing with increasing chlorination (Jones and Voogt, 1999; Breivik et al., 2007). Generally, the less-chlorinated the congeners are the more likely they solubilize in water, become volatile, and more likely they are to be biodegraded. On the other

3 hand, high-chlorinated PCBs are often more resistant to degradation and volatilization and sorb more strongly to particulate matter (Passatore et al., 2014).

2.3 Potential adverse effects of PCBs The persistent nature of PCBs coupled with their tendency to accumulate in lipids have played a major role in their wider distribution in environmental media and consequently have become a major concern for public and environmental health (DeCaprio et al., 2005; Passatore et al., 2014). Any potential impact of PCBs on living organisms depend on the degree and pattern of chlorine substitution and the rate of exposure resulting from the concentrations of PCBs in the environment. It is reported that PCBs with five or above chlorine atoms have the greatest potential for bioaccumulation and toxicity (DeCaprio et al., 2005; Van Aken et al., 2009). On the other hand, those with fewer chlorine atoms are more readily metabolized and excreted. It is also reported that PCBs with non-ortho-substitution tend to have the greatest potency for enzyme induction and possibly toxic effects in aquatic and likely other organisms (Montone et al., 2001).

Various studies show that PCBs can be bioaccumulated by aquatic and terrestrial organisms and thus enter the food web (Abella et al., 2015; Putschögl et al., 2015). Humans and wildlife that consume contaminated organisms can also accumulate PCBs in their tissues, and in the long run this could lead to adverse health effects in humans and wildlife (DeCaprio et al., 2005). Generally, PCBs have been demonstrated to cause a variety of adverse health effects in animals, including disruption of immune system, reproductive dysfunction, cancer, liver and nervous system damage, acne and other health effects (Passatore et al., 2014; Bergkvist et al., 2015; Li et al, 2015; Putschögl et al. 2015). According to the U.S. Environmental Protection Agency, PCBs are also reported to cause cancer in animals and are listed as potential carcinogens to human (Mayes et al. 1998; Chen et al., 2015). Thus, to protect public health and other living organisms from the potential adverse effect of PCBs, all congeners of PCBs must be rendered harmless or removed from polluted sites available to human exposure.

2.4 Remediation techniques of PCB-contaminated soils 2.4.1 Physicochemical methods Incineration and landfilling are the two most commonly used physicochemical methods for remediation of PCB contaminated soils and sediments (Gomes et al., 2013). Incineration can

4 be applied for PCB treatment in liquid as well as solid forms by subjecting them to temperatures typically exceeding 760°C in the presence of oxygen (Rahuman et al., 2000). Incinerators destroying PCB liquids must meet technical requirements of 2 seconds residence time at 1200°C and 3% of excess oxygen, or 1.5 seconds residence time at 1600°C and 2% of excess oxygen in the stack gases and should be able to achieve removal efficiency of 99.9999% (U.S. EPA, 1997). However, the public acceptance of hazardous waste incineration is very poor due to fear of exposure to toxic emissions. In addition, incineration of PCBs waste is very expensive and could cost up to $2,300 per ton for a fixed PCB incinerator which makes this method less attractive from public acceptance and economics perspective (U.S. EPA, 1997).

Landfilling is a method of containment of buried waste materials to prevent contact with the environment, thereby minimizing human and ecological risks associated with those wastes (U.S. EPA, 1997). However, landfill disposal of PCB contaminated soil and sediment does not provide waste reduction or destruction, and consequently PCBs can remain in the landfills for long periods of time (Gomes et al., 2013). As a result, there is a danger of PCB volatilization and migration through surrounding air channels to the ambient atmosphere (Rahuman et al., 2000). PCB contamination of groundwater could also happen due to failure of leachate collection systems.

Thermal desorption is a technology that can be applied ex situ or in situ to physically separate volatile and semi-volatile contaminants from soil and sediment at temperatures high enough to volatilize the organic contaminants (U.S. EPA, 1993; U.S. EPA, 1997; Gomes et al., 2013). Since the chamber temperature and residence times are not sufficient enough to destroy the pollutants, thermal desorption technology needs to be equipped with waste gas treatment system (U.S. EPA, 1997). extraction is another physicochemical process that can be used for the remediation of PCB-contaminated soils and sediments (U.S. EPA, 1993, Gomes et al., 2013). The technology uses chemical solvents under controlled pressure and temperature conditions to remove PCBs from soil and sediment, thereby reducing the level of pollution (Gomes et al., 2013). Solvent extraction is commonly used in combination with other technologies, such as solidification/stabilization, incineration, and soil washing. Chemical dehalogenation is another approach which involves chemical reagents and reduction processes to either mineralize PCBs or chemically alter the PCB congeners to a less toxic form that will satisfy standards for ultimate disposal or reuse of the contaminated media (Gomes et al., 2013, U.S. EPA, 2013). The technology can be achieved by either the replacement of the chlorine

5 molecules or the decomposition and partial volatilization of PCBs using catalysts like Zero- Valent Iron or Base Catalyzed Decomposition (Gomes et al., 2013). Waste solidification is another physicochemical technology that involves binding agents, such as Portland cement or asphalt, to the PCB waste in order to confine it in a solid matrix, thereby reducing its mobility and bioavailability (U.S. EPA, 1993; Gomes et al., 2013).

While physicochemical methods have the potential to efficiently remediate PCB-contaminated soils and sediments from small and heavily polluted areas, there is a lot of limitation in their feasibility for remediation of diffused pollutants. Generally, these methods are expensive due to the need to excavate the contaminated soil or sediment, transport of contaminated materials for ex situ treatment, damage to the environment as well as high energy and chemicals consumption (U.S. EPA, 1997; Kurzawova et al., 2012; Toussaint et al., 2012; Gomes et al., 2013; Passatore et al., 2014). Given the fact that PCB contamination is diffused widely, the financial feasibility in implementing physicochemical methods for PCB remediation does not seem to be sustainable. In addition, there is also a resistance from the general public in the implementation of some of the physicochemical methods. Thus, it is necessary to develop more environmental friendly, economically feasible and publicly acceptable alternatives to effectively deal with PCB contamination. In this regard, researchers have been studying alternative technologies for the remediation of PCB-contaminated media, mainly through laboratory studies with limited field application. One of the alternative technologies tested and the focus of this study is bioremediation.

2.4.2 Bioremediation Microbial bioremediation is the application of biological treatment techniques for the degradation of organic pollutants to simpler and less toxic or non-toxic substances, including carbon dioxide and water (Bedard et al. 2006; Gomes et al., 2013). Several recent studies have shown that there are two distinct but complimentary biological processes which are capable of biodegrading PCBs in the natural environment: aerobic oxidative processes and anaerobic reductive processes (Bedard et al. 2006; Vasilyeva, and Strijakova, 2007; Mackova et al., 2009; Kurzawova et al., 2012; Passatore et al., 2014). The aerobic process involves degradation of PCBs in the presence of oxygen, while anaerobic process biodegrade PCBs in the absence of oxygen. The aerobes attack PCBs oxidatively, breaking open the carbon ring and destroying the compounds, whereas anaerobes leave the biphenyl rings intact while removing the

6 chlorines. The findings of several previous studies have shown that while congeners with fewer chlorine atoms are more susceptible to complete aerobic biodegradation, those with higher chlorine atoms undergo detoxification in a reductive environment (Vasilyeva, and Strijakova, 2007; Gomes et al., 2013).

Anaerobic Degradation of PCB In the presence of chloroorganic pollutants including PCBs, some anaerobic are capable to switch to the process of dehalorespiration where PCBs can be used as terminal electron acceptors instead of the traditional electron acceptors in anaerobic environment (Rosenthal et al., 2004). In anaerobic environment such as sediments, under the effect of dehalogenase, chlorine atoms in PCBs are replaced by hydrogen atoms (Wiegel and Wu, 2000; Gomes et al., 2013). PCB dechlorinators are obligate anaerobes and are known to form microbial consortia with other bacteria, which provide them with carbon sources and maintain optimal level of hydrogen (Rosenthal et al., 2004; Vasilyeva, and Strijakova, 2007).

In general, microbial reductive dechlorination of PCBs preferentially removes meta and para chlorines from highly chlorinated congeners, resulting in increased amount of lower chlorinated, ortho substituted mono- through tetrachlorobiphenyls (Wiegel and Wu, 2000; Furukawa and Fujihara, 2008). This reductive dechlorination is an important natural process which decreases the toxicity of highly chlorinated PCBs and enhances the potential for their complete degradation under aerobic conditions (Park et al., 1999; Borja et al., 2005; Correa et al., 2010). However, various studies have shown that the metabolic activities of PCBs dechlorinators are affected by a number of environmental factors such as temperature, pH, available carbon source, partial pressure of H2, and the availability of electron donors and electron acceptors, thereby influencing the extent and rate of dechlorination (Wiegel & Wu, 2000). In addition, the number of PCB dechlorinators in natural environments is usually known to be low (about 102 cells/g), resulting in very low rates of dechlorination. Consequently, these facts present major obstacles in applying PCB-dechlorination as a remediation technology.

Aerobic Degradation of PCB Microbial degradation of PCBs under aerobic condition in the natural environment occurs by co-metabolism of PCBs with biphenyl (Borja et al., 2005; Pieper & Seeger, 2008; Mackova et al., 2010). Such aerobic degradation of PCBs have gained interest among many researchers because the complete mineralization of PCBs can only be realized under aerobic conditions

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(Furukawa and Fujihara, 2008). Since Ahmed and Focht (1973) reported the ability of Achromobacter to degrade few lower chlorinated PCBs by co-metabolism with biphenyl, many researchers have followed suit in isolating biphenyl-degrading organisms for the purpose of PCB degradation. These organisms belong to both Gram-negative and Gram-positive genera and comprise various Pseudomonas, Burkholderia, Achromobacter, Alcaligenes, Arthrobacter, Comamonas, Ralstonia, Sphingomonas, Acinetobacter, Rhodococcus and Bacillus isolates (Pieper, 2005; Vasilyeva, and Strijakova, 2007; Kurzawova et al., 2012). It is now a widely known fact that bacteria utilizing biphenyl as the sole carbon and energy source play a key role in PCB degradation. An additional widely known fact is that lower chlorinated PCBs are broken down by biphenyl catabolic pathway (or bph pathway) (Sylvestre, 2004).

Biphenyl upper pathway involves a four-step enzymatic process that turns biphenyl into benzoic acid and 2-hydroxy-penta-2,4-dienoic acid as shown in Figure 2 (Pieper 2005; Pieper & Seeger, 2008). There are seven genes involved in the bph pathway. The genes encoding for biphenyl 2,3-dioxygenase (bphA1A2A3A4) initiates hydroxylation of two adjacent biphenyl carbons to form an arene-cis-diol. Then the cis-2,3-dihydro-2,3-dihydroxybiphenyl dehydrogenase (bphB) converts dihydrodiol to 2,3-dihydroxybiphenyl. Afterwards, 2,3- dihydroxybiphenyl 1,2-dioxygenase (bphC) cleaves the ring and converts 2,3- dihydroxybiphenyl to 2-hydroxy-6-oxo-6-phenyhexa-2,4-dienoic acid (HOPDA). The last step of the upper pathway involves HOPDA hydrolase, bphD, which cleaves HOPDA into two compounds: benzoic acid and 2-hydroxypenta-2,4-dienoate (Borja et al. 2005; Pieper, 2005; Van Aken et al., 2009). The 2-hydroxypenta-2,4-dienoate is further converted to acetyl-CoA and pyruvate by 2-hydroxypenta-2,4-dienoate hydratase (BphE), 4-hydroxy-2-oxovalerate aldolase (BphF), and acetaldehyde dehydrogenase (BphG) (Seeger et al., 1997; Furukawa & Fujihara, 2008; Pieper & Seeger, 2008).

Figure 1.2. Pathway for biphenyl biodegradation (Pieper, 2005)

Several studies have shown that many bacteria have the ability to degrade PCB congeners with up to three chlorines, some of them degrade tetraCB, but only few bacteria (Burkholderia

8 xenovorans LB400, Pseudomonas pseudoalcaligenes, KF707, Rhodococcus globerulus P6, etc.) slowly degrade congeners with 5 or more chlorine atoms (Furukawa & Fujihara, 2008). The inability of most naturally occurring bacteria to grow on PCBs as their sole sourse of carbon and energy presents a major challenge in the implementation of bioremediation in the attenuation of PCB contamination (Rodrigues et al., 2000). Thus, when aerobically degraded, most intermediately chlorinated PCBs must be co-metabolized to be successfully removed. In this regard, biphenyl have been the cometabolite ubiquitously chosen in laboratory as well as field research for the co-metabolism of PCBs (Ahmed and Focht, 1973; Furukawa & Fujihara, 2008). However, biphenyl cannot be applied to a PCB contaminated site due to low water solubility, necessity of repeated application, and concerns about its toxicity (Singer, 2003; Leigh et al., 2006), which explain why many recent research activities have been focusing on experimenting with alternative cometabolites or inducers such as plant secondary metabolites for in situ PCB bioremediation.

2.4.3. Plant secondary metabolites Plant secondary metabolites (PSMs) are organic compounds produced by plants that are not directly involved in the normal reproduction, development, photosynthesis, or of the organism (Singer, 2003; Mazid et al., 2011). These chemicals are extremely diverse; many thousands have been identified in several major classes (Singer et al., 2003). Each plant family, genus, and species produces a characteristic mix of these chemicals, and as a result they can sometimes be used as taxonomic characters in classifying plants (O’Reilly-Wapstra et al, 2014).

PSMs can be classified on the basis of chemical structure (for example, having rings, containing a sugar), composition (containing nitrogen or not), their solubility in various solvents, or the pathway by which they are synthesized (e.g., phenylpropanoid, which produces tannins) (Singer, 2003; O’Reilly-Wapstra et al, 2014). A simple classification includes three main groups: terpenes (made from mevalonic acid, composed almost entirely of carbon and hydrogen), phenolics (made from simple sugars, containing benzene rings, hydroxyl group), and alkaloids (nitrogen-containing compounds, extremely diverse, may also contain sulphur) (Rausher, 2001; Singer 2003; Mazid et al., 2011).

Many PSMs are observed to be antibiotic, antifungal, antiviral and toxic or repellent to herbivores and help defend the plants producing them (Bourgaud et al., 2001; Singer et al.,

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2003; O’Reilly-Wapstra et al, 2014). It is believed that the continued modification to PSMs have helped to foster an extremely complex and intimate relationship between insects and mutualistic bacteria, which is postulated to have led in the development of PSMs detoxifying enzymes by the endosymbiont bacteria (Singer et al., 2003). In this regard, the termite–microbe symbiosis was one of such microbe-insect relationship widely studied. It is reported that the endosymbiont bacteria were observed transforming and assimilating carbon and nitrogen from lignin and humus from the termite gut (Singer et al., 2003). Given the structural similarities of lignin to recalcitrant organic pollutants, it is expected for this ecosystem to encourage the proliferation of broad-substrate enzymes which can also be applicable to xenobiotic degradation. In this regard, Maeda et al., (1995) isolated a Gram positive bacterium, Rhodococcus erythropolis TA421, from a dry-wood termite ecosystem, which was capable of degrading several highly persistent xenobiotics, including PCBs. Several researchers have also demonstrated the ability of PSMs such as flavonoids and terpenes to promote degradation of PCBs by microbial community in the rhizosphere (Donnelly et al. 1994; Hernandez et al. 1997; Leigh et al., 2006; Narasimhan et al. 2003). Singer et al. (2003) reported many PSMs are structurally analogous to xenobiotics, and hence have a potential to induce catabolic pathway in some bacteria.

2.4.4. Rhizodegradation Rhizodegradation is a phytoremediation technique whereby degradation of organic contaminants in the rhizosphere (area of soil surrounding the roots of the plants) occurs by means of microbial activity which is enhanced by the presence of plant roots (Van Aken et al., 2009). Plants and microorganisms often have symbiotic relationships making the rhizosphere an area of increased microbial density and very active microbial activity (Singer et al., 2003; Leigh et al., 2006; Uhlik et al., 2013). In recent years, a number of articles have addressed how plants could contribute towards the remediation of a wide range of xenobiotic and recalcitrant chemicals (Singer et al., 2003; Mackova et al., 2009; Van Aken, 2009; Kurzawova et al., 2012). These processes include: (1) modifying the physical and chemical properties of contaminated soils; (2) releasing root exudates, such as sugar, amino acids, and organic acids, that can be used as electron donors to support aerobic co-metabolism or anaerobic dehalogenation of chlorinated compounds; (3) improving aeration by releasing oxygen directly to the root zone, as well as increasing the porosity of the upper soil zones; (4) intercepting and retarding the movement of chemicals; (5) effecting co-metabolic microbial and plant enzymatic

10 transformations of recalcitrant chemicals; and (6) plants can release inducers such as phenolic, terpenes, flavonoids that can enhance the activity of microbial degradation of xenobiotic chemicals.

In this regard, Donnelly et al. (1994) proposed that certain plant-derived compounds, such as flavonoids, may serve as growth substrates for PCB-degrading bacteria and may induce bph genes. The findings of the researchers showed a high PCB turnover capacity of biphenyl- degrading organisms after growth with certain flavonoids, suggesting that plant roots could serve as a natural injection system capable of inducing PCB degradation in indigenous microorganisms over long periods. Hernandez et al. (1997) also demonstrated that soils enriched with peel, ivy leaves, pine needles or eucalyptus leaves can significantly increase biphenyl utilizers (108 g-1) as compared to unplanted soils (103 g-1), thereby enhancing the potential of PCB congeners degradation. Singer et al. (2003) studied the interactive effects of different treatments on the degradation of Aroclor 1242 in soil, including bioaugmentation with PCB-degrading bacteria, biostimulation with inducers and surfactants, and vegetation with Brassica nigra. The researchers observed a significantly higher PCB degradation in vegetated soil as compared to non-planted controls. Young-In et al. (1999), Singer et al. (2000), Master and Mohn (2001), Tandlich et al. (2001), Luo et al. (2015), Mackova et al. (2009), Dudasova et al. (2012), and Uhlik et al. (2013) also demonstrated that certain PSMs may play an important role in removal of PCBs. The findings of these studies strongly suggest that some plant secondary metabolites have the potential to stimulate microbial activity and induce biphenyl catabolic pathway, thereby playing a promising role in the biodegradation of PCBs.

As discussed above, there are several research findings that have demonstrated the positive effect of PSMs on microbial activity and biodegradation of PCBs. The findings of these studies show that the application of PSMs instead of biphenyl offers many advantages, including good bioavaiIability, compatibility (nontoxicity) in natural environments, and their ubiquity in the environment. Thus, use of PSMs as inducers is a very promising technology which can have the potential for providing the most cost effective and resource conservative approach as a long term solution for remediating sites contaminated with a variety of hazardous chemicals including PCBs. Thus, it is the objective of the present study to investigate the potential of selected PSMs to induce the biphenyl catabolic pathway genes in selected pure cultures.

11

Chapter 3: Objectives

 Investigation of bacterial abilities to utilize selected plant secondary metabolite as sole source of carbon and energy in enrichment cultures.

 Investigating the potential of plant secondary metabolites to induce the biphenyl catabolic pathway.

 Comparison of the biphenyl pathway induction potential of the secondary plant metabolites by quantification of 16S rRNA gene and bphA gene using quantitative polymerase chain reaction (qPCR).

12

Chapter 4: Material and Methods

4.1 Materials 4.1.1 Chemicals Plate count agar, agar bacteriological, and agar noble difco were obtained from Oxford, UK. The secondary plant metabolites: (S)-(-)-limonene (≥99%), 4-hydroxy coumarin (98%), and flavone were obtained from Sigma Aldrich, Switzerland. Naringenin (≥96%) and p- hydroxybenzoic acid (98%) were obtained from Sigma Aldrich, UK. Ferulic acid (>98%) was obtained from Sigma Aldrich, Japan. (R)-(-)- carvone (98%), p-cymene (99%), trans-cinnamic acid (99%), vanillic acid (97%), and caffeic acid (98%) were obtained from Sigma Aldrich, Germany. Naringin (96%), 4-bromobiphenyl (98%), coumarin (97%), and α-pinene (98%) were obtained from Sigma Aldrich, USA. RNase and DNase free water (molecular biology grade), biphenyl, Tris (≥99.9%), EDTA (≥99%), glass beads (acid washed), and sodium acetate were also obtained from Sigma Aldrich, USA. Undenatured ethanol (molecular biology grade) (≥99.5%) was obtained from Emplura, Germany. Methanol (100%) was obtained from PENTA, CZ. Ethanol (100%) was obtained from Sigma-Aldrich, USA. Chemicals for liquid medium, including KH2PO4 (99%), K2HPO4 (99%), MnSO4 (99%), and MgSO4 (99%) were obtained from PENTA, CZ. NH4Cl (98%) was obtained from Chemapol, CZ. FeSO4.7H2O

(99%), and CaCl2 (97%) were obtained from Lach:ner, CZ. Sodium pyruvate (≥99%) was obtained from Sigma-Aldrich, Japan. Yeast extract was obtained from Sigma Chemical, France.

4.1.2 Primers, enzymes and commercial kits Lysozyme was obtained from SERVA Electrophoresis GmbH, Germany. Murine RNase inhibitor, dNTPs, DNase buffer, DNase I, M-MuLV Reverse Transcriptase (RT), M-MuLV RT buffer, ROX reference dye, and KAPA SYBR FAST Master Mix were obtained from Kapa Biosystems, USA. Primers for 16s rRNA: 786f (5’-GAT TAF ATA CCC TGG TAG-3’); 939r (5’-CTT GTG CGG GCC CCC GTC AAT TC-3’), primers for bphA: 512f (5’-TGR TBT TYG CVA AYT GGG A-3), and 674r (5’-GGT ACA TGT CRC TGC AGA AYT GC-3’) were obtained from Generi Biotech, CZ. Random hexamers was obtained from Thermo Scientific, CZ. KAPA ReadyMix PCR (1U/µl) was obtained from Kapa Biosystems, USA. RNeasy Mini Kit was obtained from Qiagen, Germany.

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4.1.3 Bacterial strains Bacterial strains isolated from PCB-contaminated sediment by enrichment on synthetic mineral medium with biphenyl as a sole carbon source (Koubeket al., 2012), were used for the present study. These bacterial strains are Achromobacter denitrificans AD400, Pseudomonas alcaliphila JAB1, Achromobacter xylosoxidans S3, and Pseudomonas putida S9 (Koubeket al., 2012 (Kurzawova et al., 2012). These bacterial strains were previously identified as sp, however, using MALDI-TOF, we were able to identify them up to the level of species.

4.2 Methods 4.2.1 Cultivation media preparation Plate count agar (PCA) medium was prepared according to the manufacturer’s instructions and by adding a few grams of agar bacteriological (2.5 g/l) for better solidification of the medium. The PCA medium was autoclaved before being poured into Petri dishes.

4.2.2 Preparation of liquid growth medium The liquid growth medium was prepared according to (Bedard et al. 1986). The medium was a phosphate buffered mineral salts medium supplemented with 0.005% yeast extract. This growth medium was prepared in distilled water and contained the following mixtures per litre:

56.77 g K2HPO4, 21.94 g KH2PO4, 27.61 g/l NH4Cl, and 50 mg of yeast extract and 5 ml of concentrated basal salts solution. The concentrated basal salts solution contained 39 g/l MgSO4,

10 g/l MnSO4, 2 g/l FeSO4.7H2O, and 0.6 g/l CaCl2. This concentrated salts solution was prepared in distilled water and filter sterilized using a 0.2-µm syringe filter while the phosphate buffered mineral medium with yeast extract was sterilized by autoclaving. Upon cooling of the autoclaved mixture, 5 ml of the concentrated salt solution was added into the growth medium.

4.2.3 Preparation of PSMs and biphenyl solutions The SPMs and biphenyl solutions were prepared by dissolving them in 100% methanol and filter-sterilizing through a 0.2 µm filter (Gelman).

4.2.4 Preparation of lysozyme stock solution Lysozyme stock solution (50 mg/ml) was prepared by dissolving lysozyme powder in nuclease free water. Afterwards, a TE buffer with lysozyme concentration of 400 µg/ml was prepared for the purpose of cell lysis. The TE buffer (10 mM Tris-HCl, 1 mM EDTA pH 8.0) was also

14 prepared by dissolving Tris and EDTA in nuclease free water. The pH of the buffer was adjusted to 8.0 by adding HCl and was autoclaved afterwards.

4.2.5 Cultivation of bacterial strains on biphenyl Bacterial strains were maintained as glycerol stocks kept in the freezer at -80°C and revived by inoculating them into 10 ml mineral medium with biphenyl (300 ppm) as their sole carbon source. The cultures were incubated at 28°C on a rotary shaker at 120 RPM for 72 hours. Enriched cultures were passaged 5 times with (1% of culture from previous enrichment and 99% of mineral medium and biphenyl (300 ppm)). Afterwards, the cultures were cultivated using mixture of biphenyl (300 ppm) and each of the 14 PSMs or 4-bromobiphenyl as their sole carbon source in order for the bacteria to adapt to the new carbon sources. The bacterial strains were cultivated 3 times in this manner before they were provided with each PSMs or 4- bromobiphenyl as their sole carbon source. Throughout the cultivation procedure, the PSMs, 4-bromobiphenyl or biphenyl solutions were added to the sterile falcon tubes and the methanol was evaporated under a laminar flow box prior to medium addition.

4.2.6 Assessment of bacterial strains growth on PSMs 14 different PSMs were investigated for their ability to serve as growth substrates for the bacterial strains. The bacterial strains (with approximate OD value of 1 from previous culture) were inoculated in the mineral medium with PSMs as the sole carbon source. The kinetics of growth of each strains on each PSMs was tested on 8 different concentration: 2, 50, 100, 200, 300, 500, 1000, and 2000 ppm. The test cultures were grown by incubating them at 28°C on a rotary shaker at 120 RPM. Cultures were run in triplicate for each PSMs at each concentration. At a given time interval, the growth of the bacterial strains was monitored by taking 1 ml aliquots aseptically from each culture tube in a laminar flow box and analysed spectrophotometrically at 600 nm using Spectrophotometer, DU730, Life Science.

4.2.7 Assessment of PSMs ability induce the biphenyl catabolic pathway The same cultivation procedure as before was followed, with the exception of the carbon source in this step. For the purpose of investigation of bphA gene induction potential of the PSMs, the cultures were grown co-metabolically on sodium pyruvate (0.5% w/v) plus each of the 14 PSMs (50 ppm) and on sodium pyruvate (0.5% w/v) plus biphenyl (50 ppm) or 4- bromobiphenyl (50 ppm). The cultures were also grown on sodium pyruvate (0.5% w/v) as

15 their sole carbon source, as negative control and on biphenyl (300 ppm) as their sole carbon source, as a positive controls.

4.2.8 RNA isolation A 100 μl aliquot of each bacterium culture was used to inoculate 10 mL of mineral medium amended with sodium pyruvate (0.5% w/v) (as negative control), or 300 ppm biphenyl (as positive control) or 0.5% sodium pyruvate plus 50 ppm of each PSMs, 4-bromobiphenyl or biphenyl (as inducers for bphA genes). Cells were grown over night at 28°C and by shaking at 120 rpm in a 40 mL falcon tubes. The cells were then harvested via mid-to late log phase (OD approximately = 0.8) for RNA isolation. The cells were harvested by centrifuging at 5000 x g for 5 minutes at 4°C, and the supernatant was discarded by aspiration. Then the pellets were resuspended in 100 µl lysozyme-containing TE buffer (400 µg/ml) by vortexing, and afterwards the mixture was incubated at room temperature for 10 minutes. 0.1 g of acid washed glass beads was added in the mixture, and subsequently vortexed. After the cell lysis, total RNA isolation was completed using Qiagen RNeasy kit, and following the manufacturer’s protocol. Finally, the RNA samples were collected in Eppendorf tubes and stored at -80oC.

Following the RNA isolation, residual DNA digestion was performed to ensure the RNA samples are free of DNA contamination. The DNase I reaction was performed in Eppendorf tubes for a total volume of 25 µl, which consisted of 2.5 µl DNase buffer 10 x, 0.6 µl Murine RNase inhibitor (40,000 u/ml), 2 µl DNase I (2,000 u/ml), total RNA between 0.5 and 2.5 µg, and nuclease free water to a final volume of 25 µl. The overall mixture was incubated in Eppendorf Thermomixer Comfort at 37°C for 1 hour. After the incubation, 69µl of 100% ethanol, and 2.5µL of 3M NaAcOH was added to each tube, and incubated at -20°C for 2 hours. Following the incubation, the samples were centrifuged at 4°C and 20,000 x g for 20 minutes, and the supernatant was discarded by aspiration. Next, 100µl of 70% ethanol (prepared by diluting undenatured ethanol in nuclease free water) was added into each tubes and subsequently they were centrifuged at 4°C and 20,000 x g for 10 minutes. After removal of the supernatant, 100µl of 100% ethanol was added into each tubes and afterwards they were centrifuged at 4°C and 20,000 x g for 10 minutes. Then, the supernatant was removed and the RNA pellets were dried in a flow box. Finally, the pellets were eluted in 25 μl of nucleases free water and stored at -80°C. The quality and concentration of isolated RNA samples were measured using a Nano drop 2000 spectrophotometer.

16

4.2.9 cDNA synthesis Reverse-transcription was performed using the M-MuLV Reverse Transcriptase enzyme. The reaction mixture consisting of M-MuLV reverse transcriptase buffer 10 x, nucleotides for DNA synthesis (dNTPs), random hexamers, Murine RNase inhibitor, nuclease free water, and M- MuLV reverse transcriptase was prepared in sterile strip tubes for a total volume of 20 µl. During the first stage of cDNA synthesis, a mixture consisting of total RNA between 0.5 to 2 µg, 2 µl random hexamers (300 ng/µl), 4 µl dNTPs (10mM), and nuclease free water up to 16 µl was prepared in sterile strip tubes and subsequently the reaction mixture was incubated at 80oC for 5 minutes. Then the tubes were spun briefly and put on ice immediately. Next, on top of the 16 µl reaction mixture, 2 µL M-MuLV RT buffer 10 x, 1 µL M-MuLV RT (200,000 u/ml) and 1 µl Murine RNase inhibitor (40,000 ul/ml) was added for a final volume of 20 µl. During this procedure, the tubes were removed from ice only to add components, one at a time, and were placed back on ice. Then the complete reaction mixture was incubated at 42oC for one hour and this followed by inactivation of the enzyme at 90oC for 10 minutes. The incubations and enzyme inactivation were performed in TProfessional basic thermocycler from Biometra GmbH. Finally, the synthesized cDNA was stored in freezer at -20oC. The quality and concentration of the cDNA was determined using a Nano drop 2000 spectrophotometer.

4.2.10 Quantification of 16S rRNA and bphA genes To confirm that the plant secondary metabolites induced the biphenyl catabolic pathway of the bacterial strain in present study, a quantitative real-time polymerase chain reaction (RT-qPCR) analysis was conducted to quantify the expression of 16S rRNA and bphA in cells grown in the presence of sodium pyruvate, sodium pyruvate plus the plant secondary metabolites or biphenyl or biphenyl alone. Quantitative polymerase chain reactions (qPCRs) were performed using CFX96 Touch™ thermal cycler from BioRad. qPCR reactions were performed in volumes of 12 μl as shown in Table 4.

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Table 4.1. qPCR reaction mixtures

qPCR reaction mixtures Kappa sybr master mix (2 x) 6µl Primers f: (0.1 mM) 0.05 µl and r: (0.1 mM) 0.05 µl for both 16s rRNA and bphA ROX reference dye (50 x) 0.3 µl cDNA ≈ 20 ng Nuclease free water Up to 12 µl Ʃ 12 µl

The qPCR conditions were performed as follows: i. Initial denaturing: for 3 minutes at 95°C ii. Denaturing: for 3 seconds at 95°C iii. Annealing: at 55°C for 30 seconds for 16S rRNA or at 53°C for 30 seconds for bphA iv. Elongation: at 72°C for 20 seconds v. Final elongation at 72oC for 5 minutes.

The qPCR reaction was conducted for 30 cycles for all the procedures. At the end of the qPCR cycling, the melting curve was measured with a gradient of temperature from 65°C to 95°C with an increment of 0.5 °C for 5 seconds. Two controls without cDNA were also run to ensure no foreign DNA was present in the qPCR reaction mixture; moreover, controls were run with the isolated RNA to ensure that the RNA samples were not contaminated with DNA. All reactions were run in triplicates. Data were processed according to Pfaffl (2001) to calculate the relative expression of bphA gene by each PSMs with respect to a control (sodium pyruvate). One-way ANOVA analysis and Tukey’s test were conducted to assess the difference in the expression level of bphA by each PSMs as compared to the control and biphenyl.

18

Chapter 5: Results

5.1. Assessment of bacterial strains growth on PSMs The ability of each PSM to serve as a growth substrate for the four bacterial strains was assessed in phosphate buffered liquid mineral medium. After the bacterial strains were repeatedly grown in the growth medium amended with few crystals of biphenyl as the sole growth substrate, they were supplied with PSMs as their sole carbon sources. The growth of each bacterial strains on 14 different PSMs as well as on 4-bromobiphenyl was tested at 8 different concentration: 2, 50, 100, 200, 300, 500, 1000, and 2000 ppm of PSM. After monitoring the growth of the strains every day for 3 weeks, it was observed that none of the PSMs were able to support the growth of the strains.

Following their inability to grow on any of the PSMs, it was necessary for the strains to be provided with supplementary carbon source for their growth. Sodium pyruvate was chosen in this study as a growth substrate. Accordingly, the bacterial strains were provided with sodium pyruvate (0.5% w/v) to support their growth. After the strains were grown to exponential phase with sodium pyruvate (0.5% w/v), they were further amended with one of the 14 PSMs (50 ppm), 4-bromobiphenyl (50 ppm) or with biphenyl (50 ppm). Biphenyl (300 ppm) with no sodium pyruvate was used as an additional positive control. The growth of the bacterial strains was monitored spectrophotometrically by measuring OD at 600 nm, as shown in Figure 5.1 to 5.4 for selected strain and PSMs combination. The strains showed relatively fast growth on sodium pyruvate, sodium pyruvate plus PSMs and sodium pyruvate plus biphenyl and reached OD more than 1.0 within 30 hours. On the other hand, the growth of the strains on biphenyl (300 ppm) was slower and reached OD more than 1.0 after 54 hours. A one-way ANOVA was performed to analyse the growth data of the strains at 24 hours since all the strains were harvested within 24 hours for RNA isolation for all combinations of sodium pyruvate and PSMs, the only exception being the growth on biphenyl as sole carbon source. Generally, there was a significant difference (p<0.05) in the growth of each strain on sodium pyruvate and biphenyl. A significant growth difference (p<0.05) was also observed between the growth of each strain on sodium pyruvate and sodium pyruvate plus PSMs. Additionally, a significant difference (p < 0.05) in the growth of each strain grown co-metabolically on sodium pyruvate plus PSMs and on biphenyl was observed. However, generally, no significant differences (p <

19

0.05) in the growth of each strain were observed between different combinations of sodium pyruvate plus PSMs.

2 0.5% sodium pyruvate (SP)

1.6 0.5% SP + 50 ppm carvone

1.2 0.5% SP + 50 ppm p- hydroxybenzoic acid

0.8 0.5% SP + 50 ppm coumarin O.D. at at O.D. nm 600 0.5% SP + 50 ppm flavone 0.4

0.5% SP + 50 ppm biphenyl 0 0 24 48 72 96 120 144 300 ppm biphenyl Time (h)

Figure 5.1. Growth curve of Achromobacter denitrificans, AD400.

2 0.5% sodium pyruvate (SP)

1.6 0.5% SP + 50 ppm carvone

1.2 0.5% SP + 50 ppm p- hydroxybenzoic acid 0.8 0.5% SP + 50 ppm coumarin

O.D. at at O.D. nm 600 0.4 0.5% SP + 50 ppm flavone

0 0.5% SP + 50 ppm biphenyl 0 24 48 72 96 120 144 300 ppm biphenyl hours

Figure 5.2. Growth curve of Pseudomonas alcaliphila JAB1.

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2 0.5% sodium pyruvate (SP)

1.6 0.5% SP + 50 ppm carvone

1.2 0.5% SP + 50 ppm p- hydroxybenzoic acid 0.8 0.5% SP + 50 ppm coumarin

O.D. at at O.D. nm 600 0.4 0.5% SP + 50 ppm flavone

0.5% SP + 50 ppm biphenyl 0 0 24 48 72 96 120 144 300 ppm biphenyl hours

Figure 5.3. Growth curve of Achromobacter xylosoxidans S3.

2 0.5% sodium pyruvate (SP)

1.6 0.5% SP + 50 ppm carvone

1.2 0.5% SP + 50 ppm p- hydroxybenzoic acid

0.8 0.5% SP + 50 ppm coumarin

O.D. at at O.D. nm 600 0.5% SP + 50 ppm flavone 0.4

0.5% SP + 50 ppm biphenyl 0 0 24 48 72 96 120 144 300 ppm biphenyl hours

Figure 5.4. Growth curve of Pseudomonas putida S9.

5.2. Induction of bphA gene by PSMs Real time quantitative PCR (RT-qPCR) has become the standard technology to quantify nucleic materials such as cDNA for gene expression profiling. Two strategies are commonly employed to enumerate the results obtained by RT-qPCR: absolute quantification and the relative quantification. Absolute quantitation results in determination of the actual quantity of the target gene expressed in copy number or concentration. Relative quantification, on the other hand, expresses the target gene in ratio to the mean of control samples which is designated as the calibrator (Phongsisay et al., 2007). Relative quantification does not require standards with known concentrations or number of copies of target gene, which makes it more appropriate for

21 analysis of cDNA where the level of target gene depends on environmental factors (Pffafl, 2004).

RT-qPCR analyses was carried out to determine the level of bphA expression in each of the strains grown on 0.5% sodium pyruvate, 0.5% sodium pyruvate plus PSMs (50 ppm), 0.5% sodium pyruvate plus biphenyl (50ppm) or 4-bromobiphenyl (50 ppm), and solely biphenyl (300ppm). Data from RT-qPCR were processed using a mathematical model developed by Pfaffl (2001) to determine the relative quantification of a target gene (bphA) in comparison to a reference gene (16S rRNA) as shown in Equation (5.1).

E CPt arg et (controlsample) ratio  t arg et (5.1) CPref (controlsample) Eref  where: Etarget is the real-time qPCR efficiency of the target gene,

Eref is the real-time qPCR efficiency of the reference gene, ∆CP target is the crossing point deviation of control minus sample of target gene, and ∆CP ref is the crossing point deviation of control minus sample of reference gene. The crossing point is defined as the point at which the fluorescence for each sample crosses the threshold fluorescence RT-qPCR procedure.

Real-time qPCR efficiencies were calculated according to Pfaffl, 2001 as follows:

1/ slope E 10 (5.2) where slope is the slope of a standard curve constructed from the RT-qPCR data for both the target and reference genes.

The standard curve was constructed using the concentration of the starting material for each gene and the corresponding crossing points as shown in Figure 1 and 2 in Appendix A for strain AD400. The CP differences between the control and the samples for both bphA and 16S rRNA genes were calculated according to the derived CP values from the RT-qPCR data. Sodium pyruvate was taken as a control for the calculation of the relative quantification of bphA genes induced by PSMs, 4-brombiphenyl and biphenyl and the corresponding result is presented in Table 5.1 for strain AD400.

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Table 5.1. Relative expression of bphA gene induced by biphenyl and PSMs by strain AD400.

CP mean values Expression of bphA Inducers (50 ppm) 16S rRNA bphA gene relative to solely pyruvate-based growth biphenyl 9.61 ± 0.021 22.89 ± 0.033 0.42 ± 0.007 (R)-(+)-Limonene 11.78 ± 0.047 22.92 ± 0.082 2.32 ± 0.113 p-Hydroxybenzoic acid 12.13 ± 0.018 22.85 ± 0.147 3.17 ± 0.186 Naringenin 13.40 ± 0.012 22.84 ± 0.074 8.57 ± 0.261 Ferulic acid 13.12 ± 0.018 23.00 ± 0.010 6.40 ± 0.092 (R)-(-)-Carvone 10.87 ±0.062 21.46 ± 0.047 2.19 ± 0.113 4-Hydroxycoumarin 11.44 ± 0.021 24.95 ± 0.129 0.70 ± 0.038 Trans-cinnamic acid 12.78 ± 0.028 20.82 ± 0.018 13.25 ± 0.308 Naringin 12.93 ± 0.011 21.76 ± 0.057 9.73 ± 0.237 4-Bromobiphenyl 11.63 ± 0.033 23.10 ± 0.035 1.90 ± 0.059 Coumarin 12.64 ± 0.045 17.82 ± 0.015 46.79 ± 1.730 α-Pinene 11.26 ± 0.041 19.14 ± 0.012 8.65 ± 0.283 Vanillic acid 10.45 ± 0.045 20.04 ± 0.020 3.01 ± 0.111 Flavone 9.90 ± 0.040 23.37 ± 0.041 0.42 ± 0.015 Caffeic acid 10.81 ± 0.053 16.79 ± 0.022 17.72 ± 0.768 Biphenyl (300 ppm) 10.63 ± 0.104 16.55 ± 0.100 17.20 ± 1.560

The same procedure was followed for the other strains to calculate the efficiencies and the relative expression of bphA genes. The figures for the standard curves are presented from Figure 3 to 6 in Appendix A. The relative expression of bphA gene for strains JAB1 and S3 is presented below in Table 5.2 and 5.3.

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Table 5.2. Relative expression of bphA gene induced by biphenyl and PSMs by strain JAB1.

CP mean values Expression of bphA Inducers (50 ppm) 16S rRNA bphA gene relative to solely pyruvate-based growth biphenyl 15.74 ± 0.186 28.45 ± 0.452 1.49 ± 0.330 (R)-(+)-Limonene 10.59 ± 0.031 21.69 ± 0.177 2.55 ± 0.206 p-Hydroxybenzoic acid 14.42 ± 0.043 24.80 ± 0.088 4.65 ± 0.216 p-cymene 22.05 ± 0.411 24.45 ± 0.029 470.64 ± 40.675 Naringenin 13.11 ± 0.016 26.19 ± 0.094 1.04 ± 0.049 Ferulic acid 18.94 ± 0.164 24.76 ± 0.259 69.80 ± 8.43 (R)-(-)-Carvone 11.61 ± 0.033 22.81 ± 0.055 2.61 ± 0.079 Trans-cinnamic acid 11.72 ± 0.028 25.13 ± 0.041 0.81 ± 0.020 Naringin 13.12 ± 0.018 23.84 ± 0.076 3.56 ± 0.132 4-Bromobiphenyl 12.69 ± 0.032 23.14 ± 1.119 5.64 ± 0.430 Coumarin 10.14 ± 0.045 22.29 ± 0.107 1.46 ± 0.079 α-Pinene 10.87 ± 0.135 21.08 ± 0.088 4.12 ± 0.369 Vanillic acid 11.39 ± 0.007 22.67 ± 0.085 1.87 ± 0.035 Flavone 11.23 ± 0.002 22.34 ± 1.426 2.24 ± 0.073 Caffeic acid 11.47 ± 0.033 22.24 ± 0.048 3.18 ± 0.094 Biphenyl (300 ppm) 10.07 ± 0.036 19.95 ± 0.126 4.67 ± 0.121

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Table 5.3. Relative expression of bphA gene induced by biphenyl and PSMs by strain S3.

CP mean values Expression of bphA Inducers (50 ppm) 16S rRNA bphA gene relative to solely pyruvate-based growth biphenyl 10.17 ± 0.036 17.42 ± 0.034 2.39 ± 0.069 (R)-(+)-Limonene 8.68 ± 0.098 15.98 ± 0.039 1.96 ± 0.133 p-Hydroxybenzoic acid 15.06 ± 0.346 21.90 ± 0.244 5.36 ± 1.390 p-cymene 11.86 ± 0.041 19.91 ± 0.077 1.87 ± 0.087 Naringenin 12.65 ± 0.186 20.25 ± 0.255 2.65 ± 0.374 Ferulic acid 12.22 ± 0.186 21.16 ± 0.271 1.21 ± 0.215 (R)-(-)-Carvone 9.31 ± 0.007 16.69 ± 0.098 2.01 ± 0.094 4-Hydroxycoumarin 7.35 ± 0.751 16.60 ± 0.057 0.58 ± 0.056 Trans-cinnamic acid 10.63 ± 0.016 18.54 ± 0.093 1.75 ± 0.077 Naringin 12.05 ± 0.110 19.84 ± 0.038 2.20 ± 0.163 4-Bromobiphenyl 11.22 ± 0.045 18.85 ± 0.015 2.18 ± 0.067 Coumarin 9.37 ± 0.093 16.23 ± 0.228 2.71 ± 0.350 α-Pinene 9.35 ± 0.028 18.01 ± 0.061 1.00 ± 0.034 Vanillic acid 8.64 ± 0.511 16.19 ± 0.125 1.79 ± 0.60 Flavone 15.60 ± 0.214 26.00 ± 0.118 0.68 ± 0.205 Caffeic acid 11.22 ± 0.034 20.82 ± 0.091 0.74 ± 0.038

The findings of this study show that all the inducers were able to amplify bphA gene in AD400, JAB1, and S3. The level of bphA expression in each of the 3 strains induced by the different PSMs, 4-bromobiphenyl and biphenyl was statistically analysed using one-way ANOVA and Tukey’s test. The analysis of ANOVA for strain AD400 showed that there was a significant difference (p < 0.05) in bphA induction among the different inducers investigated in this study. As expected, biphenyl (300 ppm) as sole carbon source was also able to induce bphA significantly higher (p < 0.05) than the control. However, bphA induction when the strain grew co-metabolically on sodium pyruvate plus biphenyl (50 ppm) was not significantly different from the control, which can be attributed to the significance of biphenyl concentration in the induction of the biphenyl catabolic pathway (Pham et al., 2015). Tukey’s test for each inducers showed that all the PSMs, except 4-hydroxycoumarin and flavone were able to induce bphA significantly higher (p < 0.05) than the control. 4-Bromobiphenyl was also observed to induce bphA significantly higher (p < 0.05) than the control. Induction of bphA gene by coumarin was

25 significantly higher (p < 0.05) as compared to the other inducers, including biphenyl (300 ppm). On the other hand, induction of bphA by 4-hydroxycoumarin and flavone was not significantly different from control and therefore are not likely to be inducers of bphA in strain AD400. The finding also showed that except for coumarin and caffeic acid, induction of bphA gene by the other PSMs was below biphenyl (300 ppm).

The ANOVA analysis for strain JAB1 also showed that there was a significant difference (p < 0.01) in bphA expression by the different PSMs tested in this study. As expected, biphenyl (300 ppm) as sole carbon source was able to induce bphA significantly higher (p < 0.05) than the control. Additionally, induction of bphA when the strain grew co-metabolically on sodium pyruvate plus biphenyl (50 ppm) was significantly higher (p < 0.05) than the control. Analysis of Tukey’s test for each PSMs showed that all PSMs except naringenin and trans-cinnamic acid were able to induce bphA significantly higher (p < 0.05) than the control. Induction of bphA when the strain grew co-metabolically on sodium pyruvate plus 4-bromobiphenyl also appeared to be significantly higher (p < 0.05) than the control. Ferulic acid and p-cymene were able to induce bphA significantly higher than solely biphenyl. 4-bromobiphenyl was also able to induce bphA higher than solely biphenyl, although the difference was not significant. The finding also showed that p-hydroxybenzoic acid, naringin and α-pinene were comparable inducers as solely biphenyl. On the other hand, induction of bphA by naringenin and trans- cinnamic was not significantly different from control and thus they are less likely to be inducers of bphA in strain JAB1.

For strain S3, the ANOVA analysis showed that there was a significant difference (p < 0.01) in the level of bphA expression by the different PSMs. Induction of bphA when the strain grew co-metabolically on sodium pyruvate plus biphenyl was significantly higher (p < 0.05) than the control. On the other hand, analysis of Tukey’s test for the individual PSMs showed that 9 of the PSMs investigated in this study, including (R)-(+)-limonene, p-hydroxybenzoic acid, p- cymene, naringenin, (R)-(+)-carvone, trans-cinnamic acid, naringin, coumarin, and vanillic acid induced bphA significantly higher (p < 0.05) than the control. On the other hand, the induction of bphA gene by 4-hydroxycoumarin, ferulic acid, α-pinene, flavone, and caffeic acid was not significantly different from the control, and thus these PSMs are not likely to induce bphA in strain S3. The finding showed that induction of bphA by p-hydroxybenzoic acid was significantly higher (p < 0.05) than the induction by other inducers, including biphenyl (50 ppm). Additionally, (R)-(+)-limonene, (R)-(-)-carvone, p-cymene, naringenin, naringin,

26 coumarin, vanillic acid and 4-bromobiphenyl appeared to be comparable inducers as biphenyl (50 ppm) in strain S3.

A comparison of bphA gene induction by the inducers for each strain showed that p-cymene and ferulic acid induced the first two highest relative bphA expression in strain JAB1, while the third highest inducer of bphA was observed to be coumarin in AD400 as shown on Figure 5.5.

3

2.5

2 expression

1.5 bphA

1

0.5

log of relative log relative of 0

AD400 JAB1 S3

Figure 5.5. Comparison of relative bphA expression by different inducers for AD400, JAB1, and S3 strains.

(R)-(+)-limonene, p-hydroxybenzoic acid, (R)-(-)-carvone, naringin, coumarin and vanillic acid were observed to induce bphA significantly higher (p < 0.05) than the control in strain AD400, JAB1 and S3. On the other hand, a number of the PSMs investigated exhibited mixed effect on induction of bphA at least in one of the strains. For example, α-pinene and caffeic acid were observed to induce bphA gene significantly higher (p < 0.05) than the control in strains AD400 and JAB1, however both of them did not seem to induce bphA gene in S3. Naringenin was also observed to induce bphA gene in strain AD400 and S3 significantly higher (p < 0.05) than the control, although it did not seem to induce bphA in strain JAB1. On the

27 other hand, flavone was observed to induce bphA gene significantly higher (p < 0.5) than the control, although the metabolite did not seem to induce bphA in AD400 and S3.

For strain S9, bphA amplification was below detection level for all PSMs, except for biphenyl (300 ppm). A qPCR procedure was conducted to test for the integrity of each cDNA sample in terms of 16S rRNA gene presence and melting curve profile. It was observed from the investigation that the 16S rRNA gene was amplified for the strain grown co-metabolically on sodium pyruvate plus each of the PSMs as shown in Table 5.4. Analysis of the melting curve also showed that no nonspecific products were present in the samples, which could demonstrate the integrity of the cDNA samples. Thus, the findings of these investigations might suggest the PSMs investigated in this study are not able to induce the bphA gene in strain S9.

Table 5.4. Amplification of 16S rRNA gene in strain S9.

Inducers (50 ppm) CP value biphenyl 12.51 (R)-(+)-Limonene 11.76 p-Hydroxybenzoic acid 29.80 p-cymene 11.79 Naringenin 12.59 Ferulic acid 12.75 (R)-(-)-Carvone 11.12 4-Hydroxycoumarin 12.64 Trans-cinnamic acid 11.88 Naringin 13.02 4-Bromobiphenyl 12.26 Coumarin 11.59 α-Pinene 11.02 Vanillic acid 11.09 Flavone 11.12 Caffeic acid 11.84 Biphenyl (300 ppm) 11.45 0.5% Sodium pyruvate 13.14

28

Chapter 6: Discussion

Over the past decades, rhizoremediation has become increasingly popular as a means to treat soils contaminated with xenobiotics, including PCBs (Sylvestre et al. 2009; Van Aken and Bhalla, 2011). Several previous reports have provided evidence that plants can promote bacterial PCB-degradation in soil (Hernandez et al., 1997; Narasimhan et al. 2003; Leigh et al. 2006; Toussaint et al., 2012). Many researcher have also reported that PSMs may allow the growth of bacteria in the rhizosphere as well as triggering their biphenyl catabolic pathway, resulting in enhanced PCB-degradation (Donnelly et al. 1994; Hernandez et al., 1997). Thus, identification of PSMs that could induce the biphenyl catabolic pathway is a first step in developing a practical bioremediation strategy for in situ treatment of PCB-contaminated soils. In this regard, the present study investigated various PSMs that might support growth and/or induce bphA, a gene that encodes for the large subunit of biphenyl dioxygenase, for bacterial co-metabolism of PCBs.

In the present study, despite our repeated effort to cultivate the bacterial strains on PSMs using different liquid mineral media, it was observed that the strains were not able to utilize the PSMs as their sole carbon sources, which agrees with some previous findings for different strains (Gilbert and Crowley, 1997; Park et al., 1999; Tandlich et al., 2001; Toussaint et al., 2012; Pham et al., 2015). Mazid et al. (2011) and Uhlik et al., 2013 have shown that some PSMs exhibit antimicrobial nature, which could have an inhibitory effect on the growth of some bacterial strains. Donnelly et al. (1994) also reported the inability of some bacterial strains to utilize PSMs as their sole carbon source with no inhibitory effect on the microbes. On the other hand, although some PSMs may exhibit antimicrobial effects on various bacterial strains, Gilbert and Crowley (1997) hypothesized that these PSMs could still be able to induce the biphenyl catabolic pathway in some strains by means of detoxification mechanism as long as the concentration of the PSMs remain below the inhibitory level.

Gilbert and Crowley (1997) have reported the inhibitory effect of carvone on the growth of strain Arthrobacter sp. B1B. The investigation of the authors showed that carvone at a concentration of 500 ppm or above resulted in the lysis of the cell. At lower concentrations (100, 200, 300, or 450 ppm of carvone), the strain was not also able to grow, even though the cells remained viable. Park et al. (1999) reported the inability of Ralstonia eutropha H850,

29

Arthrobacter sp. B1B Corynebacterium sp. Tl04, and Burkholderia xenovorans LB400 to utilize carvone as their sole carbon source. Tandlich et al. (2001) also demonstrated that Pseudomonas stutzeri was not able to utilize carvone and limonene as sole carbon sources.

Although the present and several previous studies reported the inability of various bacterial strains to utilize PSMs as their sole carbon source, it is noteworthy to mention that microbial utilization of PSMs, however, did occur, and appeared to be strain and PSM specific as reported by previous studies (Donnelly et al., 1994; Hernandez et al., 1997; van der Werf et al., 1999; Jung et al., 2001). Donnelly et al. (1994) reported that a number of PSMs, including naringin, coumarin, myricetin, and catechin were able to support growth and induce PCB-degradation pathway in Ralstonia eutropha H850, Burkholderia xenovorans LB400, and Corynebacterium sp. MB1. However, the authors highlighted that not all PSMs were able to support growth in all the strains. Hernandez et al. (1997) reported that cymene and limonene were both substrates and inducers for the PCB-degradative pathway some Gram-positive environmental isolates (Rhodococcus sp., Corynebacterium, and Cellulomonas sp.).

Although all of the PSMs investigated in the present study were not able to support growth of the four strains, there are strong literature evidences to hypothesize that some of them could serve as an inducers of the biphenyl catabolic pathway in some of the strains, thereby promoting the degradation of PCBs (Singer et al., 2003; Park et al., 1999; Toussaint et al., 2012; Pham et al., 2015). These previous findings indicated that the PCB-degraders should be provided with supplementary carbon source, therefore, sodium pyruvate was chosen in the present study as growth substrate while the PSMs were added in the growth medium to investigate their potential for the induction of bphA gene.

Following the use of sodium pyruvate as a growth substrate, the bacterial strains appeared to grow faster as observed from growth curve monitoring. However, it was observed that their growth was affected by the type of PSMs used in the growth media as shown from Figures 5.1 to 5.4. Investigation of the bacterial growth curves showed that all four bacterial strains appeared to have lower growth rate in the presence of carvone, flavone and coumarin as compared to their growth in the presence of the other PSMs investigated. Luo and Hu (2013) reported the inhibitory effect of coumarin on Burkholderia cepacia, FL5B and Ralstonia eutropha, WL7B. On the other hand, coumarin was found to foster the growth of Crynecacteriam sp. MB1 and R. eutropha H850 (Donnelly et al., 1994). Luo and Hu (2013)

30 also reported the inhibitory effect of carvone on Burkholderia cepacia, FL5B, when the strain grew on glucose. The addition of biphenyl as co-substrate, on the other hand, was observed to further improve the growth of JAB1, S3, and S9 as compared to the control. However, addition of biphenyl did not seem to influence the growth of AD400. Luo and Hu (2013) also reported that the enhancement of WL7B growth by the addition of biphenyl when grown on glucose, although addition of biphenyl had only limited effect on the growth of FL5B. The authors also highlighted that cymene, carvone, limonene, and naringin did not have similar effects as biphenyl on the growth of the strain. On the other hand, in the present study, p-hydroxybenzoic acid was observed to have a mixed effect on the growth of the bacterial strains. AD400 and S9 were observed to have lower growth rate in the presence of p-hydroxybenzoic acid while the growth of JAB1 and S3 observed to be improved by its presence. Generally, the findings of the present and many other previous studies have shown that PSMs have different effects on the growth of different microbes at different concentrations (Donnelly et al., 1994; Park et al., 1999; Luo and Hu, 2013).

In order to investigate the induction of the biphenyl catabolic pathway by the PSMs, the strains were first harvested for total RNA isolation. Harvesting of the strains was performed after the strains were grown until mid- to late-log phase in order to enhance the quality and quantity of total RNA (Pham et al., 2015). However, in addition to the physiological state of the strains, the methods of cell lysis and RNA isolation are also known to greatly affect the quality and quantity of isolated total RNA (Pfaffl, 2004; Jahn et al., 2008). Generally, isolation of high quality RNA from bacteria is known to be very problematic (Jahn et al., 2008). It often involves lengthy procedures such as centrifugation and the use of toxic or expensive chemicals in order to inhibit the omnipresent RNases (Bernstein et al., 2002). Many commercially kits have also been manufactured with the aim of decreasing time and enhancing the quality of the isolated RNA (Phongsisay et al., 2007).

In the present study, freezing and thawing was investigated as a method of cell lysis, before the total RNA isolation procedure was followed using RNeasy Kit from Qiagen. The findings of this study showed that the quantity and quality of total RNA yield with this method were low, probably due to poor lysis of bacterial cells before extraction (Jahn et al., 2008). Following the use of enzymatic (lysozyme) cell lysis and homogenization with glass beads inserted before proceeding through RNeasy, Qiagen method, the RNA quantity and quality increased significantly compared to when freezing and thawing was used as a method of cell lysis.

31

The presence of residual DNA in the RNA sample is very problematic since it interferes with the amplification of the target gene (Copois et al., 2007). Thus, all the RNA samples were treated with RNase-free DNase, to get rid of residual DNA. After residual DNA degradation was performed, an RT-qPCR procedure was conducted for each RNA sample in order to ensure complete digestion of residual RNA. The result from RT-qPCR showed no detectable amplification, which suggested that the isolated RNA was free from DNA contamination. Ensuring complete digestion of contaminating DNA is crucial in order to avoid any bias during quantification of 16s rRNA and bphA genes (Stark et al., 2014). Moreover, complete digestion of contaminating DNA also ensures correct RNA concentration measurements.

High integrity RNA samples are very essential for many molecular techniques used in gene expression studies in order to accurately quantify the target gene (Copois et al., 2007). However, RNA samples are very susceptible to degradation due to cleavage with RNases as a result of improper handling or through storing the RNA in sub-optimal conditions (Bustin, 2002). Thus, in order to preserve its integrity, the RNA sample was reverse-transcribed to cDNA and stored at -20 oC or stored at -80 oC until the reverse-transcription was performed.

In the present study, total RNA and cDNA purities were assessed based upon two ratios, the absorbance at 260 nm to the absorbance at 280 nm (A260/A280) and the absorbance at 260 nm to the absorbance at 230 nm (A260/A230) (Wilfinge et al., 1997). Nucleic acids such as RNA and cDNA have their absorption maximum at 260 nm, whereas proteins have maximum absorption at 280 nm. Additionally, contaminants from reagent carry-over such as guanidinium thiocyanate, phenol, or other salts can be assessed at 230 nm (Vikhe Patil et al., 2015). In this study, A260/A280 values between 1.8 and 2.2 (considered pure) were achieved for most of the samples in the present study. A260/A230 values of 1.8-2.2 (considered pure) (Chomczynski, Sacchi, 1987) were also achieved for most samples.

The induction potential of each PSM for biphenyl catabolic pathway was investigated using RT-qPCR procedure using cDNA samples. The investigation showed that except for S9, the bphA gene was able to be detected and amplified for strains AD400, JAB1, and S3 grown co- metabolically on sodium pyruvate and 14 different PSMs or 4-bromobiphenyl. In order to quantify the expression level of the bphA gene, relative quantification method was selected. This method is based on the expression levels of a target gene (bphA) versus a reference gene (16S rRNA). To calculate the expression of a target gene in relation to a reference gene various

32 mathematical models are established. Calculations are based on the comparison of the crossing points (CP) and threshold values at a constant level of fluorescence (Kubista et al., 2006). In the present study, the relative quantification of bphA gene was calculated according to a model presented by Pfaffl (2001), in which the relative expression ratio of bphA gene was computed based on its RT-qPCR efficiencies (E), and the crossing point (CP) difference (∆CP) of one unknown sample versus one control (sodium pyruvate). Additionally, the qPCR efficiency was calculated from the slopes of the calibration curve according to the equation provided by Pfaffl (2001).

In RT-qPCR, it is possible that nonspecific annealing and primer elongation events could happen leading to the formation of non-specific products (Monis et al, 2005). During qPCR process such nonspecific products could compete with formation of specific qPCR product, resulting in reduced amplification efficiency and formation of a less specific RT-qPCR product (Varga and James, 2006). Thus, to distinguish nonspecific products from the specific amplicon, a melting curve analysis (Ririe et al., 1997) was performed for all the runs, and consequently qPCR products with nonspecific primers were excluded from the calculation.

The calculation of the relative expression of bphA gene showed that most of the PSMs investigated were able to induce the gene higher than the control. In strain AD400, 11 out of the 14 tested PSMs were observed to induce the bphA gene significantly higher (p < 0.05) than the control. The induction of bphA gene by coumarin in strain AD400 was significantly higher (p < 0.05) than the other inducers, including when the strain grew solely on biphenyl. Caffeic acid was also observed to induce bphA gene comparable to solely biphenyl (300 ppm). Naringenin, ferulic acid, trans-cinnamic acid, naringin, and α-pinene were also other significant inducers, although their induction level was below biphenyl (300 ppm). There are also several previous studies which reported on the bphA induction capability as well as contribution in the co-metabolism of PCBs by some of these PSMs in different bacterial strains. Gilbert and Crowley (1997) have demonstrated the ability of carvone (50 ppm) to induce co- metabolism of PCBs in Arthrobacter sp. strain B1B, although the concentration was observed to be too low to support growth. The authors highlighted that when induced by carvone at a concentration of 50 ppm, the strain was able to degrade more than 15% of Aroclor 1242 during a 15-h incubation period. In subsequent work, Gilbert and Crowley (1998) also demonstrated that repeated application of carvone-induced bacteria led to rapid degradation of PCB in comparison to repeated application of biphenyl-induced bacteria.

33

Dudasova et al. (2012) reported that biodegradation ability of P. stutzeri toward PCBs was enhanced by the addition of PSMs as compared to control. They highlighted that the strain was able to degrade 18.9% of PCBs (control), while the addition of biphenyl, carvone, and limonene resulted in degradation of 32.7%, 33.6%, and 32.9% of PCBs, respectively. Tandlich et al. (2001) also demonstrated the ability of carvone and limonene to induce biphenyl catabolic pathway and biodegradation of PCBs in Pseudomonas stutzeri. The authors reported that the use of glycerol and xylose as growth substrates and carvone and limonene as inducers not only extended the spectrum of degraded PCB congeners but also increased the effectiveness of their degradation. Donnelly et al. (1994) reported that naringin and coumarin fostered the growth and co-metabolism of PCBs in Ralstonia eutropha H850 and Corynebacterium sp. MB, respectively. The authors highlighted that while naringin was able to foster the greatest metabolic activity of strain H850 towards PCBs, strain MB1 grown on coumarin was able to metabolize higher number of different congeners as compared to the biphenyl controls. The findings of previous and our studies suggest that many of the PSMs investigated in the present study could stimulate PCBs degradation in AD400 in the same manner as demonstrated when biphenyl was added to PCB-contaminated soil (Focht, 1995).

The calculation of the relative expression of bphA gene in strain JAB1 also showed that 12 out of the 14 investigated PSMs were able to induce bphA gene significantly higher (p < 0.05) than the control. Both p-cymene and ferulic acid were able to induce bphA significantly higher (p < 0.05) than biphenyl in strain JAB1. Except for these two PSMs, the level of expressions for the other PSMs was below that of biphenyl (300 ppm). While most of the PSMs tested observed to be inducers for strains AD400 and JAB1, few of them observed to have contrasting effects. For example, while coumarin was the highest bphA inducer for strain AD400, its induction in strain JAB1 was in the same range with the control, and therefore is not likely to induce bphA gene in JAB1. Flavone, on the other hand, while it was observed to be significant inducer in strain JAB1, it did not seem to induce bphA gene in strain AD400. The contrasting effects of a given PSM on different bacterial strains was also observed in several previous studies (Donnelly et al., 1994; Singer et al., 2003; Luo et al., 2007; Luo and Hu, 2013; Uhlik et al., 2013). Uhlik et al. (2013) demonstrated the ability of naringin, caffeic acid and limonene to induce changes in both bacterial community structure and ability of PCB degradation. The authors highlighted that naringin exhibited the most promising potential in degradation of majority congeners present in Delor 103. On the other hand, the finding of the authors showed that caffeic acid caused a significant decrease of higher chlorinated PCB congeners, although

34 its addition resulted in the reduction of microbial diversity in the soil. Limonene was also reported to stimulate PCB degradation, although its effect was limited compared to naringin and caffeic acid. Pham et al. (2015) demonstrated the ability of flavone to induce the biphenyl catabolic pathway in Rhodococcus erythropolis U23A when the strain grew co-metabolically on sodium acetate plus flavone. However, their findings showed that the level of the biphenyl catabolic pathway expression varied significantly depending on the concentration of flavone. Pham et al. (2015) reported that at 0.1 mM, the induction caused by flavone (relative expression of 13.4 in comparison to control) was significantly higher than for biphenyl (relative expression of 1.3 in comparison to control). However, the authors reported that when the strain grew co- metabolically on sodium acetate plus flavone at a concentration above 0.1 mM, the level of expression decreased. The reduction in expression was attributed to the formation of a dead- end product, 4-oxo-2-chromenecarboxylic acid, which was formed following metabolism of flavone (Pham et al., 2015). In the present study about twice the concentration used by Pham et al. (2015) was used, which could explain the repression effect on induction of bphA gene observed in strains AD400 and S3. The findings of Pham et al (2015) and our results suggest that the level of biphenyl catabolic pathway expression may vary considerably depending on the nature and concentration of the PSMs as well as the type of strain.

As compared to AD400 and JAB1, most PSMs happened to induce bphA at lower level in strain S3. S3 is also the strain for which a number of PSMs did not seem to induce. This finding further solidify the notion that the induction effect of PSMs is strain specific.

For strain S9, unlike strains AD400, JAB1 and S3 the level of bphA amplification for all PSMs were below detection level. Although none of the PSMs were able to induce bphA gene in strain S9, it is possible for some of these PSMs to potentially induce other genes such as bphC in the upper biphenyl catabolic pathway as demonstrated by previous studies for other strains (Park et al., 1999). Park et al. (1999) studied induction of PCB degradative pathway in Ralstonia eutropha H850, Arthrobacter sp. B1B Corynebacterium sp. Tl04, and Burkholderia xenovorans LB400 by carvone. The authors demonstrated that in the presence of supplementary carbon source to support growth, carvone (50 ppm) was observed to induce PCB degradation in B1B strain as witnessed by the accumulation of 4,4’-DCBp meta ring cleavage product when the strain was provided with 4,4'-DCBp as cometabolite. However, the authors highlighted that the effects of carvone in the strain were related to the expression of bphC gene encoding 2,3-dihydroxybiphenyl-1,2-dioxygenase in the upper biphenyl

35 degradation pathway. Additionally, previous studies have shown that growth substrates can significantly influence the expression of the bacterial biphenyl catabolic pathway (Parnell et al., 2010). Pham et al. (2015) demonstrated that when U23A strain was co-metabolically grown on sodium acetate, sucrose, or mannitol plus flavone, the induction of the biphenyl catabolic pathway was significantly higher than when the strain grew on glucose or mannose plus flavone. Tandlich et al. (2001) also demonstrated the differential effect of carbon sources: biphenyl, glucose and xylose on PCB degradation. The authors highlighted that carvone and limonene were not able to enhance PCB degradation compared to control when P. stutzeri grew co-metabolically on glucose, which could suggest that the PSMs were not able to induce the biphenyl catabolic pathway in the presence of glucose. The implications of these findings are that each PCB-degrading bacterium may respond differently to different growth substrates, thereby influencing the induction of biphenyl catabolic pathways by PSMs. Thus, further studies are needed to understand the relationship between different growth substrates and expression of the biphenyl catabolic pathway in the strains investigated in the present study to improve our knowledge and setup the platform to deal with PCB removal from the environment in a sustainable way.

36

Chapter 7: Conclusion

The present study has examined the ability of selected PSMs to induce the biphenyl catabolic pathway in strains Achromobacter denitrificans AD400, Pseudomonas alcaliphila JAB1, Achromobacter xylosoxidans S3, and Pseudomonas putida S9. These bacterial strains were previously isolated from PCB contaminated soil using biphenyl as a growth substrate. In the present study, the strains were investigated for their ability to utilize 14 different PSMs as their sole carbon source, however, none of them were able to grow on these PSMs. Thus, sodium pyruvate was used as growth substrate while the PSMs were investigated for their potential to induce the biphenyl catabolic pathway. The strains were harvested for RNA isolation in the mid-to-late log phase in their growth stage. The investigation of freezing and thawing as cell lysis method followed by RNA isolation using RNeasy kit, Qiagen produced low quantity and quality of RNA. On the other hand, the use of TE buffer together with lysozyme and glass beads as cell lysis method followed by RNA isolation using RNeasy kit, Qiagen produced RNA of high quality and quantity.

The RT-qPCR result showed that all of the PSMs, except for p-cymene in strain AD400, were able to induce bphA gene in strains AD400, JAB1 and S3. However, none of the PSMs were able to induce bphA in strain S9. As previous studies have shown, the inability of all the PSMs to induce bphA in S9 could be affected by the growth substrate. Previous studies also showed that some of these PSMs could induce other genes such as bphC in the upper catabolic pathway. Thus, it is worthy to conduct further experimental studies with other carbon sources as well as with bphC primers.

A relative gene expression approach was followed to compare the induction of bphA gene by the PSMs in comparison to the control (sodium pyruvate). The study have identified a number of PSMs, including (R)-(+)- limonene, (R)-(-) carvone, naringin, coumarin, ferulic acid, p- hydroxybenzoic acid, α-pinene, vanillic acid, trans-cinnamic acid, and naringenin which were able to induce bphA significantly higher (p < 0.05) than the control at least in one of the strains. Ferulic acid and p-cymene were able to induce bphA significantly higher (p < 0.05) than biphenyl itself in strain JAB1. Coumarin was able to induce bphA signicicantly higher (p < 0.05) than biphenyl in strain AD400. Caffec acid was also able to induce bphA comparable to biphenyl in AD400. It was also observed that the biphenyl degrading bacterial strains may

37 respond differently to various PSMs with respect to their ability to induce the biphenyl catabolic pathway.

The overall results show that PSMs have high potential for the induction of the biphenyl catabolic pathway, which agrees with the findings of various previous studies. As demonstrated in a number of previous studies, lower chlorinated PCBs can be degraded by the biphenyl catabolic pathway enzymes. Thus, there is a very high potential for AD400, JAB1, and S3 to be induced by PSMs and co-metabolize PCBs. In this regard, further studies need to be conducted in order to investigate the PCB degrading potential of these strains while being induced by PSMs. Additionally, further investigation on the effect of different carbon sources on the potential of biphenyl catabolic pathway induction is recommended. It would also be very important to investigate the induction potential of the PSMs at different concentrations, as this is also another important factor that influences the expression level of the biphenyl catabolic pathway. These studies would help to setup an optimum combination between growth substrates and PSMs for efficient bioremediation of PCB contaminated soils.

38

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Appendix A

25

20 y = -2.8847x + 22.454 R² = 0.9907 15

10 Crossing Crossing points 5

0 0 1 2 3 4 5 6 Log concentration of cDNA

Figure 1. A standard curve constructed for 16S rRNA for strain AD400 from RT-qPCR data. The slope of the line is -1/log (efficiency) giving an efficiency of 2.20.

30

25

20

15

10 y = -5.0307x + 30.611 R² = 0.9928

5 Crossing Crossing points

0 0 0.5 1 1.5 2 2.5 3

Log concentration of cDNA Figure 2. A standard curve constructed for bphA for strain AD400 from RT-qPCR data. The slope of the line is -1/log (efficiency) giving an efficiency in this case of 1.58.

46

25

20 y = -3.4752x + 27.021 R² = 0.9996 15

10

Crossing Crossing points 5

0 0 1 2 3 4 5 6

Log concentration of cDNA

Figure 3. A standard curve constructed for 16S rRNA for S3 from RT-qPCR data. The slope of the line is -1/log (efficiency) giving an efficiency of 1.94.

30

25

20

15 y = -4.179x + 27.425 R² = 0.9965 10

Crossing Crossing points 5

0 0 0.5 1 1.5 2 2.5 3 Log concentration of cDNA

Figure 4. A standard curve constructed for bphA for strain JAB1from RT-qPCR data. The slope of the line is -1/log (efficiency) giving an efficiency in this case of 1.73.

47

25

20 y = -3.973x + 23.263 R² = 0.9963 15

10

5 Crossing pointsCrossing

0 0 1 2 3 4 5

Log concentration of cDNA

Figure 5. A standard curve constructed for 16S rRNA for JAB1 from RT-qPCR data. The slope of the line is -1/log (efficiency) giving an efficiency of 1.79.

35

30

25

20

15 y = -4.3693x + 32.655 R² = 0.9463

10 Crossing Crossing points 5

0 0 0.5 1 1.5 2 2.5 3

Log concentration of cDNA

Figure 6. A standard curve constructed for bphA for strain JAB1from RT-qPCR data. The slope of the line is -1/log (efficiency) giving an efficiency in this case of 1.69.

48