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This is to certify that this study was carried out by Frederick O. Oshomogho (MAT ..... centralized Benin auto-mechanic workshops cluster (popularly called Uwelu ...
COMPARATIVE STUDIES ON THE DEGRADABILITY OF ANTHRACENE AND PYRENE BY SYNTHETIC AND BIOSURFACTANTS

BY

FREDERICK OKHAKUMHE OSHOMOGHO PG/ENG0203369

CHEMICAL ENGINEERING DEPARTMENT FACULTY OF ENGINEERING UNIVERSITY OF BENIN

SEPTEMBER 2015

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ROLE OF SYNTHETIC AND BIO-SURFACTANTS IN THE DEGRADATION OF ANTHRACENE AND PYRENE

BY

FREDERICK O. OSHOMOGHO (MAT. NO: PG/ENG0203369)

A PROJECT REPORT SUBMITTED IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF ENGINEERING

IN

CHEMICAL ENGINEERING DEPARTMENT FACULTY OF ENGINEERING UNIVERSITY OF BENIN

SEPTEMBER, 2015 2|Page

CERTIFICATION This is to certify that this study was carried out by Frederick O. Oshomogho (MAT. NO: PG/ENG0203369) towards the attainment of a Master of Engineering Degree in Chemical Engineering, Faculty of Engineering, University of Benin, Edo State, Nigeria.

Prof C. N. Owabor Supervisor

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…………………………… PG Coordinator

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Prof E. O. Aluyor Head of Department

…………………………… External Supervisor

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DEDICATION This work is dedicated to my parents Mr. and Mrs. James Oshomogho (of blessed memories) for their relentless efforts and support in giving me the best of education. It is also dedicated to my lovely son, Lucas Oshioke Oshiomogho Okhakumhe who was born during the course of this work and also to my lovely wife, Mrs. Justina Odufa Oshomogho for being there when I needed her most.

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ACKNOWLEDGEMENT My sincere gratitude goes to God almighty for making a way when the road does not seem smooth, for restoring hope in despair and for the wisdom. I would like to gratefully and sincerely thank my supervisor Prof C. N. Owabor for her patience, guidance in spite of her very busy schedule to monitor my progress, provide information and sharing her knowledge throughout the course and also, the opportunity to develop my own individuality. I would also like to gratefully thank the Head of Department, Prof E. O. Aluyor and the Post Graduate Coordinator, Dr. Frederick Iyayi for their very special advice and support. My special thank also goes to Prof S. E. Ogbeide, Prof F. O. Aisien, Prof K. O. Obahiagbon, Prof C. O. Okiemen, Dr. Chris Akhabue, Dr. N. A. Amenaghanwon. My special thanks goes to the Emeritus Professor, T. O. K Audu for his fatherly advice and encouragements. I am indeed very grateful to the Chief Technologist, Mr. Joe Iyayi, Principal Technologist Mr. Banabas Lamidi Afegbua, and Mr. Victor Eboigbe, Mrs. Justina Okhomona, Mr. Moses Okhuoya, Mr. Wisdom Osayuki, Mr. William Owabor, Mr. Collins Agbonghile, Mr Toba Salokun for their precious knowledge and supportive spirit as they have always been helping and assisting me on my experimental tasks. Not to forget, I would like to thank all my colleagues and family who always encourage and support me and last but not least, I would like to thank everyone who has contributed directly or indirectly and gave me moral support and inspiration over this period.

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ABSTRACT Polycyclic aromatic hydrocarbons (PAHs) are ubiquitous persistent semi-volatile organic compounds. They are contaminants that are resistant to degradation and can remain in the environment for long periods due to their low water solubility which limits their availability for microbial uptake. This study therefore seeks to investigate the role of synthetic surfactants and biosurfactants (produced by microbes indigenous to the soil) during the clean-up of soils polluted with diesel containing recalcitrant PAHs. PAHs degrading bacteria capable of producing biosurfactants where isolated from diesel contaminated soil by serial dilution method. The concentration of residual recalcitrant PAHs in the soil were determined with the aid of gas chromatograph fitted with flame ionization detector (GC-FID). The degradation experiment was conducted in bioreactors containing soils spiked with strains of isolated biosurfactant producing microbes were separately added with and without synthetic surfactant. Results of the study showed that the concentrations of anthracene and pyrene were very high in the soil, accounting for a combined mass concentration of 14.11mg/kg out of the 30.103mg/kg which constitute the total PAHs present in the soil. The bioreactors containing biosurfactants producing microbes alone were observed to have very significant decrease in the concentration of the residual hydrocarbon content when compared to reactors with a combination of both synthetic surfactant and biosurfactant producing microbes. Detailed analysis showed that pyrene had a total of 0.812mg/g mineralized in the overall degradation process by all microbes compared to anthracene with 0.429mg/g of total mass degraded, because pyrene with four benzene rings was observed to be more soluble in solvents than anthracene with three rings. The degrading ability of the study organisms showed Basillus substilis which produces surfactin biosurfactant degraded 0.110mg/kg pyrene and

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0.056mg/g anthracene followed by Pseudomonas aeruginosa which produces rhamnolipid biosurfactant and degraded 0.094mg/g pyrene and 0.053mg/g anthracene to be most effective in the degradation of both anthracene and pyrene when compared with the results for Staphilococus aureus 0.048mg/g and Micrococus leutus 0.047mg/g. This effectiveness was further demonstrated when these microbes were used in combination with synthetic surfactant such as Tween 80.

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LIST OF TABLES Table 2.1

Physio-chemical properties of selected PAHs

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Table 2.2

Basic properties of some synthetic surfactants

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Table 2.3

Classification and microbial origin of biosurfactants.

41

Table 2.4

Environmental conditions affecting degradation

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Table 3.1

Treatment combination in bioreactors

68

Table 4.1

Physical-chemical characteristics of the soil used for experimentation

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Table 4.2

UV Spectrometer absorbance of anthracene

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Table 4.3

UV Spectrometer absorbance of pyrene

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Table 4.4

Mass concentration of anthracene in different reactor combinations

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Table 4.5

Mass concentration of pyrene in different reactor combinations

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Table 4.6

Physical-chemical characteristics of the soil used for experimentation

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Table A.1

Cultural and morphological characteristics

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Table A.2

Biochemical Characteristics

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Table B.1

Absorbance-concentration for standard solution of Pyrene and anthracene

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LIST OF FIGURES Figure 2.1 Figure 2.2 Figure 2.3

Structures of the polyaromatic hydrocarbons identified by the U. S. EPA as priority pollutants Schematic diagram of a Gas Chromatograph

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Figure 4.2

Schematic diagram of biodegradation by a microbe (Source: A Citizen 45 Guide to Bioremediation- EPA) Mass concentration of PAHs in diesel polluted in Chemical Engineering 72 Department, University of Benin. GC-FID chromatogram of PAHs in an aged diesel polluted soil. 73

Figure 4.3

Anthracene degradation curve

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Figure 4.4

Pyrene degradation curve

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Figure B

Pyrene and anthracene calibration curve

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Figure 4.1

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NOMENCLATURE CMC

Critical Micelle Concentration

PAHs

Poly Aromatic Hydrocarbon

GC FID

Gas Chromatograph flame ionization detector

UV-Vis

Ultra-Violet Visible

LMW

Low Molecular Weights

HMW

High Molecular Weights

PACs

Polycyclic Aromatic Compounds

KOW

Octanol-Water Coefficients

SPE

Solid Phase Extraction

SPME

Solid Phase Micro extraction

LC-FD

Liquid Chromatography Fluorescence Detection

LIF

Laser Induced Fluorescence

USEPA

United State Environmental protection Agency

PHWE

Pressurized Hot Water Extraction

HLB

Hydrophile Lipophile Balance

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TABLE OF CONTENTS Certification ……………………………………………………………………… Dedication ……………………………………………………………… Acknowledgement ……………………………………………………………… Abstract ……………………………………………………………… List of Tables ……………………………………………………………… List of Figures ……………………………………………………………… Nomenclature ………………………………………………………………

III IV V VI VIII IX X

Chapter 1 ……………………………………………………………………... Introduction …………………………………………………………….... 1.1 Backgroung to the study ……………………………………………… 1.2 Statement of the problem ……………………………………………… 1.3 Relevance of the study ……………………………………………… 1.4 Aim and objectives ………………………………………………………. 1.5 Scope of study …….....……………………………………………………… 1.6 Limitations to the study ……………………………………………….

1 1 1 2 4 5 6 6

Chapter 2 ……………………………………………………………… Literature Survey ……………………………………………………… 2.1 Polycyclic Aromatic Hydrocarbons ……………………………………… 2.1.1 Occurrence and sources of PAHs in the environment……………… 2.1.2 Reactivity of PAHs…………………………………………………. 2.1.3 PAH sequestration (Aging)…………………………………………. 2.1.4 Determination of PAHs……………………………………………… 2.1.5 Degradation of PAHs………………………………………………… 2.1.6 Analytical procedure for the measurement of PAHs……………….. 2.2 Surfactants ……………………………………………………………… 2.2.1 Classification of surfactants………………………………………… 2.2.2 Synthetic surfactants………………………………………………… 2.2.3 Biosurfactants………………………………………………………. 2.2.4 Physiological role of biosurfactants ……………………………… 2.2.5 Effects of surfactants on biodegradation of PAHs ……………….. 2.2.6 Solubilization by surfactant ………………………………………. 2.3 Bioremediation ……………………………………………………… 2.3.1 Principles of biodegradation ………………………………………. 2.3.2 Methods of biodegradation ………………………………………. 2.3.3 Biodegradation of organic compounds …………………………… 2.3.4 Condition for biodegradation ……………………………………… 2.3.5 Biosurfactant enhanced biodegradation ………………………… 2.3.6 Application of biosurfactants in biodegradation technology.….…… 2.3.7 Role of surfactant in biodegradation of PAHs …………………… 2.4 Bioreactors ……………………………………………………………… 2.5 Microbial growth kinetics ……………………………………………… 2.6 Kinetics of biodegradation ………………………………………………

7 7 7 9 11 13 16 17 27 31 33 33 35 43 43 44 44 45 46 50 50 54 55 56 56 56 58 11 | P a g e

Chapter 3 …………………………………………………………………….. Materials and method ……………………………………………………… 3.1 PAHs identification experiment ……………………………………… 3.1.1 Equipments and materials ……………………………………… 3.1.2 Procedure ………………………………………………………. 3.2 Bacteria strain isolation ……………………………………………… 3.2.1 Equipments and materials ……………………………………… 3.2.2 Procedure ……………………………………………………… 3.3 Biodegradation experiment ……………………………………………… 3.3.1 Chemicals, reagents and equipment ……………………………... 3.3.2 Soil sample collection and preparation ……………………… 3.3.3 Procedure ……………………………………………………… 3.4 Soil Analysis …………………………………………………………… 3.4.1 Determination of moisture content ……………………………... 3.4.2 Determination of soil pH and conductivity (EC) …………….... 3.4.3 Total organic carbon (TOC) and total organic matter (TOM) contents

62 62 62 62 63 64 64 64 65 65 66 66 68 68 68 68

Chapter 4 ………………………………………………………………. Results and Discussion ………………………………………………. 4.1 Result of PAHs contamination of diesel soil …………………………… 4.2 Biodegradation analysis ……………………………………………… 4.3 Isolation and identification of biosurfactant producing bacterial ……….. 4.3.1 Cultural and morphological characteristics ……………………… 4.3.2 Biochemical identification of hydrocarbon degrading bacteria … 4.4 Soil characteristics ……………………………………………………….. 4.4.1 Electrical conductivity (EC) ……………………………………… 4.4.2 Total organic carbon (TOC) and total organic matter (TOM) contents 4.4.3 Nitrate-nitrogen (NO3-N) ………………………………………. 4.4.4 Soil pH ……………………………………………………………

70 70 70 72 78 80 80 81 81 82 82 83

Chapter 5 ………………………………………………………………. Conclusion and Recommendation ……………………………………...... 5.1 Conclusion ………………………………………………………………. 5.2 Recommendation ………………………………………………………

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References Appendix A A. Appendix B A.

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Bacteria Identification

Calibration curve and application of Beer’s law

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CHAPTER

1

INTRODUCTION 1.1 BACKGROUND TO THE STUDY In a never-ending search for improvement in quality of life, man has made great strides. The scientific and technological vehicle which has carried man in his odyssey towards ultimate enjoyment of natures has been fueled by various petroleum derivatives. Pollution of the soil with petroleum derivatives is often observed in municipal soils around industrial plants, through pipe line vandarization, along major high ways tankers, and in areas where petroleum and natural gas are obtained (Adam et al., 2002). Processing and distribution of petroleum hydrocarbons as well as the use of petroleum products leads to contamination of soil. Changes in soil properties due to contamination with petroleum-derived substances can lead to water and oxygen deficits as well as to shortage of available forms of nitrogen and phosphorus (Wyszkowska and Kucharski, 2000). Contamination of the soil environment can also limit its protective function, upset metabolic activity, unfavourably affect its chemical characteristics, reduce fertility and negatively influence plant production. The introduction by man directly or indirectly of petroleum derivatives into the environment result in deleterious effects on human health and that of the organisms that are dependent on the soil (Aboribe, 2001). The increasing use of diesel oil in diesel engines of cars, industrial trucks and generators has led to an increased demand for diesel oil (Ogbo, 2009), thereby resulting in accidental spillage of diesel within generator house surroundings and along Nigerian high ways. Diesel oil which contains a highly complex mixture of petroleum fractions of aliphatic and polycyclic aromatic hydrocarbons (PAHs) also contains a very recalcitrant PAHs which do not easily volatilize or biodegraded.

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Diesel oil is one of the major products of crude oil and it constitutes a major source of pollution to the environment (Nwaogu et al., 2008). Diesel oil can enter into the environment through leakage from storage containers, refueling of vehicles, wrecks of oil tankers and warships carrying diesel oil and through improper disposal by mechanics when cleaning diesel tankers. Diesel spills on agricultural land generally reduce plant growth, reduces soil fertility and soil microflora population. However, oil contaminated soils are amendable to bioremediation because micro-organisms capable of degrading petroleum hydrocarbons are present (Jone and Edington, 2005) and the fraction of total microbial population in soil able to degrade varies from bacteria and fungi (Ekpo and Nya, 2012).

1.2 STATEMENT OF THE PROBLEM In Nigeria, as in many other countries, petroleum hydrocarbon contamination is widespread. Pollution arising from the diesel fuel is one of the environmental problems in Nigeria and is more widespread than crude oil pollution. The prevalent mode of indiscriminate spillage of these fuels in the environment calls for urgent attention. Contamination results from generator houses, filling stations, mishandling, deliberate disposal, spilling and leakage of petroleum products, such as mechanical workshops. A survey of Benin city, the capital city of Edo State, Nigeria indicates that apart from the centralized Benin auto-mechanic workshops cluster (popularly called Uwelu Spare Parts Market), there are diesel generator houses scattered all over the town from which diesel leakages, used engine oils, lubricating oils and other solvents containing petroleum hydrocarbons are indiscriminately dumped or spilled on every available space. The local utilization of diesel fuel in Benin City has increased in the recent time owing to poor power supply and consistent increasing rural-urban migration. This results in the upsurge in the number of

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vehicles and other machines that makes use of these fuels. These unguided practices have worsened the rate at which diesel fuel spread and contaminate the soils and water around the town. These diesel fuel and used oils form part of the most hazardous wastes commonly generated in diesel generator houses and auto-repair shops around cities in Nigeria (Ipeaiyeda and Dawodu, 2008). Waste oil and diesel are a mixture of different chemicals including petroleum hydrocarbons, chlorinated biphenyls, chlorodibenzofurans, additives, decomposition products and heavy metals that come from engine parts as they wear away (ATSDR, 1997). Ekundayo et al., (1989) have shown that noticeable changes in properties occur in soils polluted with petroleum hydrocarbons arising from used engine oils and diesel spills. Diesel fuel easily migrates into the environment and eventually seeps into water bodies (Olugboji and Ogunwole, 2008). PAHs belong to a group of over 100 hazardous substances of organic pollutants consisting of two or more fusedbenzene aromatic rings. Formation of PAH is due to incomplete combustion of organic matter through the condensation of ethylenic radicals in the gas phase to form the larger polycyclic compounds. PAHs are to a certain degree resistant to biodegradation and are sometimes included in a class of persistent organic pollutants (POPs) (Wild and Jones, 1995). They are hydrophobic compounds and their persistence in the environment is also linked to their low water solubility and electro-chemical stability. Of the over 900 agents compiled by the International Agency for Research on Cancer (IARC), 400 have been classified as carcinogenic (this includes PAHs) (IARC, 2009). The properties and environmental fate of PAHs are dependent on the number of rings and molecular weight. High molecular weights (HMW) PAHs are compounds with four or more fused benzene rings, whereas the low molecular weight (LMW) compounds consist of two to three fused benzene rings (Law et al., 2002).

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The effects caused by PAHs have been grouped according to their carcinogenicity, mutagenicity, teratogenicity, direct toxicity and/or combinations of all. There are several hundred PAHs which often exist as mixtures rather than as single compounds (Chun et al., 2002). The most common PAHs are anthracene, benzo(a)pyrene, chrysene, fluorene and pyrene. However benzo[a]pyrene (BaP) is commonly used as an indicator species for PAH contamination and most of the available data refer to this compound (Bull, 2008.)

1.3

AIM AND OBJECTIVES

This study is aimed at evaluating the effects of synthetic and biosurfactants in the degradation of anthracene and pyrene constituent of diesel fuel polluted soil. The specific objectives of the study are: a. To study the influence of synthetic surfactant (Tween 80) and biosurfactants on the degradation of most recalcitrant PAHs. b. To investigate microorganisms that can exist in a soil contaminated with diesel fuel. c. To investigate the most recalcitrant PAHs in a diesel polluted soil. d. To investigate the separate effects of synthetic and biosurfactants producing microorganism on PAHs.

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1.4

SCOPE OF STUDY

This study focused on (i)

Determination of most recalcitrant PAHs present in an aged diesel polluted soil in Chemical Engineering Department, University of Benin.

(ii)

Isolation, identification and characterization of biosurfactants producing microorganisms indigenous to diesel polluted soil in Chemical Engineering Department, University of Benin.

(iii)

Investigation of the biodegradation potential of each strain of biosurfactants producing microorganism.

(iv)

Investigation of the roles of synthetic and biosurfactants in the biodegradation of recalcitrant PAHs present in a diesel polluted soil.

1.5

RELEVANCE OF STUDY

Many bioremediation technologies have been developed to remove these hydrocarbons contaminants. Some biological treatments are cheaper than chemical and physical treatments and sometimes result in complete mineralization. This study on surfactants and their influence on the corresponding behavior in the degradation of PAHs provide necessary information for proper assessment of their transport in soil and possibility of contamination of ground water, and for better understanding of the mechanism altering the bioavailability of PAHs. These information and mechanism are important for bioremediation of PAHscontaminated soil. The study on the separate and combined effects of synthetic and biosurfactants with PAHs on soils would lead to a correct assessment of PAH's risk in real environment where many pollutants coexist.

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Influence of Tween-80 on the mineralization and metabolism of pyrene and anthracene in soil will provide necessary information for better understanding on how surfactants affects the behaviour and fate of PAHs in soil systems.

1.6

LIMITATIONS TO THE STUDY

Bioremediation involves the use of organisms most active in the degradation of bioavailable petroleum hydrocarbons. Factors such as solubility and specific activity of biosurfactants from bacteria strain has been studied with the following specific limitations. 1. Time: time has been invaluable in this study. Detailed analysis of the critical micelle concentration (cmc) of specific biosurfactants produced by each microbial strain would have been done to test the solubilization capacity. This process has been ignored in the study to meet up with time scheduled for the study. 2. Finance: the processes involved in biodegradation study are highly capital intensive. The insitu measurement of petroleum hydrocabons degradation is most accurately done by chromatographic device. The cost of gas chromatography analysis is very high thereby causing the degradation measurement to be done with a UV-Vis spectrometer whereby the concentration depletion was monitored instead. 3. Power: frequent electrical power outage greatly affected the set time of the periodical degradation measurement.

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CHAPTER

2

LITERATURE SURVEY 2.1

POLYCYCLIC AROMATIC HYDROCARBONS

Polycyclic aromatic hydrocarbons (PAHs), also known as poly-aromatic hydrocarbons or polynuclear aromatic hydrocarbons are potent environmental pollutants that consist of fused aromatic rings and do not contain heteroatoms or carry substituents (Ukiwe et al., 2013). They are found almost everywhere in the environment and are typically formed during incomplete burning of organic materials such as wood, coal, oil, gasoline etc. PAHs are also found in crude oil, coal tar and asphalt (Fetzer, 2000). The presence of PAHs in the environment is related to human activities within the environment. Large concentrations of PAHs are expected to be present in urban areas and in areas where bush burning for agricultural farming is commonly practiced as well as in petroleum exploration and refining operations (Fetzer, 2000). PAHs are of great concern to humans since as pollutant some have been identified as carcinogenic, mutagenic and teratogenic. In the environment, PAHs are primarily found in soil, sediment and oily substrates. Natural crude oil and coal deposits contain significant amount of PAHs arising from chemical conversion of natural products molecules such as steroids (Roy, 1995). PAHs are also formed by incomplete combustion of carbon containing fuels such as fat. Forest fires and agricultural bush burning account for the largest volume of PAHs from any natural source in the atmosphere. The actual amount of PAHs emitted from forest fires and agricultural bush burning vary with the type and nature of the fire as well as the intensity of the fire (Fagbote and Olanipekun, 2013). Once exposed to the atmosphere, most of the pollutant do not degrade quickly and may thus reside in the environment for extended periods of time (Fagbote and Olanipekun, 2013). During this period, winds may distribute the pollutant over a large area in a global manner (Fagbote and Olanipekun, 19 | P a g e

2013). PAHs’ presence in fossil fuels such as coal and crude oil deposits occur due to low temperature combustion of organic material over a significant period of time. PAHs are hydrophobic (they mix more easily with oil than water) compounds with aqueous solubility decreasing almost linearly with increase in molecular mass. The compounds may be classified as low molecular weights (LMW) or high molecular weights (HMW). LMW PAHs are those containing two or three benzene rings, while those with four or more benzene rings are called HMW PAHs (fig. 2.1). LMW PAHs are relatively water soluble, but those containing four or more rings are quite hydrophobic and insoluble (Cerniglia and Heitkamp, 1989). However, the higher the molecular weight of a PAH, the more likely it is to absorb to soil organic matter. This tendency to strongly adsorb on particulate matter renders the HMW PAHs less available and thus less susceptible to remediation (Cerniglia and Heitkamp, 1989). HMW PAHs have high resonance energies due to the dense clouds of pi-electrons surrounding the aromatic rings making them persistent in the environment and recalcitrant to degradation (Johnson et al., 2005). Another reason for the compounds to exhibit recalcitrant nature is their low aqueous solubility and high soil sorption (Parish et al., 2009). PAHs are affected by several physical, chemical and biological processes which alter their fate and transport in the subsurface environment. The nature of the subsurface environment determines if the individual compound may remain as a non-aqueous phase liquid, metabolized by microorganisms, taken up by plants, be volatilized into the interstitial void spaces, or be sorb onto soil organic matter (Ukiwe et al., 2013). However, desorption from the organic matter is a critical step in the degradability of PAHs.

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Figure 2.1: Structures of the polyaromatic hydrocarbons identified by the U. S. EPA as priority pollutants. Source: Gong et al., 2003.

2.1.1 OCCURRENCE AND SOURCES OF PAHs IN THE ENVIRONMENT The release of PAHs into the environment is widespread since these compounds are ubiquitous products of incomplete combustion. PAHs have been detected in a wide variety of environmental samples, including air, soil, sediments, water, oils, tars and foodstuffs (Ukiwe et al., 2013). Industrial activities, such as processing, combustion and disposal of fossil fuels, are usually associated with the presence of PAHs at highly contaminated sites. The following are the most common sources and occurrence.

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(a). Poly Aromatic Hydrocarbon in the AIR: The main route for PAH transport is through the atmosphere. Results from ambient air monitoring programs have shown that PAH concentrations are usually of the order of a few nanograms per cubic metre of air. Motor vehicles, including spark emission and diesel automobiles, trucks and buses, also contribute to atmospheric PAH pollution through exhaust condensate and particulates, tyre particles and lubricating oils and greases. (b). Poly Aromatic Hydrocarbons (PAHs) in sediments: As PAHs are characterised by low water solubilities and high octanol-water partitioning coefficients, their concentrations in water are extremely low. However, due to their hydrophobic nature, PAHs accumulate in fine grain sediments, partitioning to organic carbon-coated particles. PAH concentrations in sediments may accumulate due to a number of sources including atmospheric deposition, marine seeps of petroleum hydrocarbons, offshore production or petroleum transportation, sewage disposal or boating. (c). Poly Aromatic Hydrocarbons in soils: PAH concentrations in soils from industrialised countries have also revealed an increasing PAH burden since the mid 1800s, with a peak in the 1950/1960s. Anthropogenic combustion of fossil fuels and long range atmospheric transport of PAHs has contributed to the dispersal of PAHs throughout the environment. The concentration of PAHs in contaminated soils from industrial sites can vary depending on the industrial activity associated with the site. (d). Poly Aromatic Hydrocarbons in marine organisms: In addition to air, sediment and soil, PAHs may accumulate in marine organisms. Uptake of PAHs by marine organisms is dependent on the bioavailability of the PAHs (i.e. partitioning of the compound between sediment, water and food), as well as the physiology of the organism. The organism size, ingestion rate, growth rate, membrane permeability, ventilator rate, gut residence time and osmo-regulation are biological processes that influence the organism’s uptake of PAHs.

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(e). Poly Aromatic Hydrocarbons in plants: PAHs may also accumulate in vegetation that could indirectly cause human exposure through food consumption (Wagrowski and Hites, 1997). Atmospheric deposition of PAHs onto plants occurs, and in some cases, various plants have been found to take up these compounds. A number of factors can influence the accumulation of PAHs in plants, such as the physical properties of the PAH, the plant species and structure, as well as environmental conditions including atmospheric PAH concentrations, temperature and wind conditions.

2.1.2 REACTIVITY OF PAHs Polycyclic aromatic hydrocarbons are classified as chemically inert and are mainly transformed into other polycyclic aromatic compounds (PACs) by electrophilic substitution reactions rather than addition. Thus, large amounts of energy are required to transform an aromatic compound into a nonaromatic product. The electron distribution over the PAH molecule determines the positions of the molecule that are most reactive (Lundstedt, 2003). Solubility: Solubility of PAH compounds in water is dependent upon temperature, pH, ionic strength (concentration of soluble salts), and other organic chemicals (i.e. dissolved organic carbon) (Wick et al., 2011). Solubility is estimated by: a. Chemical structure b. octonol-water partition coefficients (a). Chemical structure: In general, as the number of benzene rings in a PAH compound increases, solubility decreases (Wick et al., 2011). There are, of course, exceptions to the rule. Symmetry, planarity, and the presence of substituents affect PAH solubility in organic solvents. Solubility has been found to increase in linearly-fused PAH as the number of rings increase because the bonds become weaker (olifinic) in character, but has not observed in angularlyfused PAH (Wick et al., 23 | P a g e

2011). Planar PAHs are less reactive (i.e. less soluble) and biologically less toxic (Dabestani and Ivanov, 1999). Alternant PAH compounds that are planar and symmetrical require a relatively high energy of solubilization because of their ability to fit closely in a lattice. Thus, they tend to be less soluble. As the compounds deviate from planarity or symmetry they tend to be more soluble in organic solvents. Methyl and polar substitution may also increase the solubility of PAHs in certain solvents (Wick et al., 2011). Most byproducts of PAH biological and chemical degradation tend to be more polar and have higher solubility in the environment than the parent compounds. (b). Octanol-water partition coefficients: There are substantial amounts of data on the relationship between aqueous solubility and octonol-water partition coefficients (Kow) for the partitioning of PAH between water and organic matter in soils (Mackay and Callcott, 1998). There is an inverse relationship between Kow and solubility which is determined with the following equation:

The octanol-water coefficient is often expressed as the log Kow. Naphthalene has a log Kow of 3.37 (table 2.1), while Indeno[1,2,3-cd]pyrene Kow is 6.50. In this case, naphthalene is more soluble than Indeno[1,2,3-cd]pyrene. This is also in agreement with the influence of chemical structure on solubility (Wick et al., 2011). Solubility and Kow of select PAH are provided in Table 2.1. Table 2.1: Physio-chemical properties of selected PAHs Compound Mole weight Boiling point g mole-1 C Naphthalene 128 218 Phenanthrene 178 339 Anthracene 178 340 Fluoranthene 202 375 Pyrene 202 393 Benz(a)anthracene 228 435 Benz(a)pyrene 252 496

Water solubility mg l-1 31.7 1.29 0.075 0.26 0.135 0.014 0.004

Log Kow 3.5 4.45 4.46 4.90 4.90 5.61 6.50

USEPA’s priority pollutant PAHs and selected properties (adapted from Lundstedt, 2003; Bojes and Pope, 2007).

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Because of the very low water solubility and high K ow values, they will tend to be sorbed to the organic matter in the soil in-stead of being solubilized in the infiltrating water and through this be transported downwards to the groundwater reservoirs. The sorption process is therefore counteractive to efficient biodegradation since it will decrease bioavailability, since the compounds due to sorption will be located in microporous areas of the soil inaccessible to the bacteria, and the biodegradation will thus be controlled by the slow desorptive and diffusive mass transfer into the biologically active areas (Zhang et al., 1998). It has been claimed that a slow sorption following the initial rapid and reversible sorption lead to a chemical fraction that is very resistant to desorption (Hatzinger and Alexander, 1995). This phenomenon is called aging, and the existence of such a desorption-resistant residues may increase the time as the compound stay in the soil dramatically. PAHs have also been shown to be partitioned or incorporated more or less reversibly into the humic substances of the soil after partial degradation and thereby be even more immobilised in the soil (Kästner et al., 1999).

2.1.3 PAH SEQUESTRATION (AGING) The most recalcitrant PAH fraction consists of residual PAHs in soil or sediments and seems to increase with aging or soil-PAH contact time (Wick et al., 2011). Even distribution of two- to sixringed PAHs in a soil sample may indicate a “young” PAH profile, since environmental processes such as natural degradation and evaporation are known to cause a reduction in the concentrations of LMW PAHs. A higher proportion of HWM PAHs generally indicates that there has been more aging in the contaminated soils or sediments. Polycyclic aromatic hydrocarbon aging can be summarized as a combination of the following two processes (Wick et al., 2011): i. Adsorption to soil organic matter and black carbon ii. Diffusion into three dimensional micropores of soil particles or residual charcoal in the soil matrix 25 | P a g e

Although sequestration of PAHs is primarily controlled by soil organic matter (Wick et al 2011), this is not the sole factor governing PAH sequestration. Intra-aggregate diffusion of PAHs protects them from microbial degradation or solvent extraction by local sorption on hydrophobic pore walls or by entrapment in voids due to constricted geometry. With time, PAHs will adsorb to soil components and diffuse to be trapped in soil micropores (Semple et al., 2003), which will limit bioavailability. Microbial activities also affect sequestration of PAHs in soils.

2.1.4 DETERMINATION OF PAHS Many analytical techniques have been applied in the determination of PAHs and their metabolites in contaminated systems. Ter-Laak, Agbo, Barendregt, and Hermens (2006) examined the use of solidphase extraction (SPE) in measuring freely dissolved concentrations of PAHs in soil pore water. Some researchers have suggested that solid-phase microextraction (SPME) with gas chromatographic/mass spectrometric (GC/MS) analysis can achieve detection limit as low as nanograms per litre of PAHs and alkyl PAHs in sediment pore water (Hawthorne et al 2005). In soil samples, high resolution gas chromatography-mass spectrometry (HRGC-MS) and liquid chromatography-fluorescence detection (LC-FD) have been applied successfully in measuring the final concentrations of PAHs (Berset et at., 1999). It was revealed that HRGC-MS has a high linear range compared to LC-FD. Hence, under practical considerations, HRGC-MS was regarded as a superior analytical technique than LC-FD (Berset, Ejem, Holzer, & Lischer, 1999). A recent study had noted that laser-induced fluorescence (LIF) spectroscopy is a mature technique for PAHs determination in terrestrial sediments (Grundl et al., 2003). In a related study, SPME coupled with LIF was applied successfully in in-situ sampling in an effort to develop a simple field-portable method to determine total dissolved PAHs concentrations in sediment pore water (Hawthorne et al., 2008). It was evident that PAHs undergo a variety of processes resulting from chemical, biological or photochemical reactions. Hence, environmental 26 | P a g e

conditions usually affect intermediate products of PAHs formed during degradation processes. A notable study by Dabroska, Kot-Wasik, and Namiesnik (2005) has revealed that different degradation pathways of PAHs require specialized analytical tools in their degradation studies. (Jiang et al., 2007) also used GC-MS in analyzing sediment samples in a study aimed at characterizing the distribution of PAHs in sediment samples.

2.1.5 DEGRADATION OF PAHS A number of physiochemical processes including adsorption, incineration, and absorption can be used to treat PAHs but the costs for chemicals and fuels as well as further treatment or disposal of secondary wastes are increasingly inhibiting adoption of these solutions. The processes leading to the eventual removal of PAHs pollutant from the environment has been extensively documented and involves the trio of physical, chemical and biological alternatives. The currently accepted disposal methods of incineration or burial in secure landfills (USEPA 2001; ITOPF 2006) can become prohibitively expensive when the amounts of contaminants are large. This often results in cleanup delays while the contaminated soil continues to pollute groundwater resources if on land, and death of aquatic life if on waterways, thus necessitating speedy removal of the contaminants. The biodegradation of pollutants is not a new concept as it has been intensively studied in controlled conditions (Chaillan et al., 2004) and in open field experiments (Chaıneau et al., 2003; Gogoi et al., 2003), but it has acquired a new significance as an increasingly effective and potentially inexpensive cleanup technology. (i). Biodegradation: This is the use of microorganisms to degrade or detoxify environmental pollutants (Bamforth and Singleton, 2005). Biodegradation is also a clean-up method that presents the possibility to eliminate organic contaminants with the aid of natural biological activity available in the substrate (Zeyaullah et al., 2009). The microorganisms used for biodegradation should be indigenous 27 | P a g e

to the contaminated area or site (Das and Chandran, 2011). It is expected that the bacteria be able to degrade the contaminant by multiplying in population and then decline when the contaminant is finally degraded (Mougin, 2002). The products of complete mineralization of the pollutant by biodegradation process include; CO2, H2O and cell biomass (Gratia et al., 2006). The optimization of biodegradation process involves many factors among which are the existence of a microbial consortia capable of degrading the pollutant, the bioavailability of the contaminant to microbial attack and certain environmental factors (soil type, temperature, soil pH, oxygen level of soil, electron acceptor agents, nutrient content of soil) contributing to microbial growth (Gratia et al., 2006). Many researchers are of the opinion that certain bacteria isolates are capable of degrading PAHs. Of particular note are the studies carried out by Punapayak, Prasongsuk, and Messner (2009) and Abd-Elsalam, Hafez, and Hussain (2009). Punapayak, Prasongsuk, and Messner (2009), investigated the use of laccase enzyme produced from an isolate of the white rot fungus Genoderma lucidum. It was revealed that G. lucidum degrade anthracene completely with or without the addition of a redox mediator. However, AbdElsalam, Hafez, and Hussain (2009), concluded that Escherichia coli, Alcoligenes sp., and Thiobacter subterraneus were efficient isolates for degrading anthracene and phenanthrene. A number of studies have reported extensively on a variety of other bacterial species that have been isolated and noted to posses the ability to utilize PAHs as energy source (Wu and Wang, 2009; Zhang et al., 2009; Ling et al., 2011; Zhao et al., 2008). Some microorganisms mainly from the genera Pseudomonas and Mycobacterium have been found capable of transforming and degrading PAHs under aerobic conditions (Mrozik et al., 2008). It is also evident that anthracene could be completely mineralized by Sphingomonas, Nocardia, Beijerinckia, Paracoccus, and Rhodococcus with dihydriol as the initial oxygenated intermediate (Teng et al., 2010). The ultimate goal of any remediation process should not be limited to removing contaminants from polluted substrates but should also include restoring the

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ability of the soil to function according to its potential (Epelde et al., 2009). In this regard, the bioavailability of PAHs becomes a very important factor to consider. Bioavailability of PAHs in biodegradation is determined by certain complex interactions between biotic and abiotic factors. Biotic factors affecting bioavailability include; the metabolic ability of microorganisms to degrade PAHs as well as the ways in which the bacteria encourage PAHs accessibility. One way by which microbes encourage PAHs accessibility is by the production of biosurfactants. Biosurfactants are produced to increase the solubility of insoluble substrates or solubilized (emulsify) insoluble substrates in order to enhance their direct contact between microorganisms and contaminants (Mulligan, 2005). Evidence have been accumulating to suggest that certain microorganisms namely; Bacillus subtilis, Pseudomonas aeruginosa and Microccocus leutus could produce bioremediation surfactants such as surfactin, rhamolipid and sophorolipid capable of improving bioremediation by solubilizing PAHs into the aqueous medium and enhance their bioavailability for degradation (Cottin and Merlin, 2007; Kuyukina et al., 2005). (ii). Chemical Degradation: When PAHs undergo chemical reactions, they are transformed into other polyaromatic hydrocarbons (they do not lose their aromatic character). Their aromaticity is conserved since considerable amounts of energy are required to change an aromatic compound into a nonaromatic compound (Dewar, 1952). The localization energy concept has proven very effective in determining the positions of reaction molecules within the PAH structure. These positions could be determined by considering electron distribution over the PAH molecule. The localization energy isolates a pi-electron at the centre of a PAH molecule from the remaining pi-electron system (Dewar 1952). The attacking species may be an electrophile or a radical. This attack leads to a degradation process of complete mineralization of the PAH molecule resulting in carbon dioxide (CO2), water (H2O), and other inorganic and organic compounds (Barbas et al., 1996).

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The occurrence of PAHs in anaerobic conditions depends on certain factors which include substrate interaction, pH and redox conditions (Chang et al., 2002). Stimulation of PAHs degradation rate under sulphate reducing conditions has been examined (Bach et al., 2005). Oxidation reactions are viewed as the most effective in this regard, although some authors have suggested that photochemical reactions may also aid oxidative reaction processes (Kochany and Maguire, 1994). It must be noted that most oxidation reactions in the environment are initiated by oxidants such as peroxides (H2O2), ozone (O3) and hydroxyl radicals generated by photochemical processes. The degradation pathways are such that the oxidation reactions involving hydroxyl radicals or O3 reacts with aromatic compounds such as PAHs at near diffusion-controlled rates by abstracting hydrogen atoms or by addition to double bonds (Haag and Yao, 1992). Ozone may attack double bonds directly or it can form reactive hydroxyl radicals (which attack double bonds) by decomposing water (Legrini et al., 1993). The reaction proceeds with complex pathways producing numerous intermediates. However, the final reaction products include a mixture of ketones, quinones, aldehydes, phenols and carboxylic acids for both oxidants (Reisen and Arey, 2002). Photochemical degradation of PAHs often involves the same oxidative species that are produced during the pure chemical oxidation of PAHs. Consequently, the reaction products include similar complex mixtures (Mallakin et al., 2000). Effects of carbon (C), nitrogen (N), and phosphorus (P) on PAHs degradation have also been investigated by several authors including (Ley et al., 2005; Quan et al., 2009). Further studies have revealed that addition of N and P to soil can help to evaluate the response of phenanthrene degradation (Johnson and Scow, 1999). Nitrate reduction has been the focus of so many researches (Rockne and Strand, 1998). Another study by Leduc, Samson and Al-Bashir (1992) had noted that acenaphthrene, fluorene and anthracene were capable of being degraded in denitrifying environment. Lovley, Woodward, and Chapelle (1994) have reported that ferric oxide reduction process has a limiting factor

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when applied in PAHs degradation. This limiting factor stems from the fact that ferric oxide is sparingly soluble in water. However, the study went further to note that chelating agents are added to subsurface environment to enhance solubility of ferric oxides and hence increase its bioavailability. Several studies have stressed the importance of surfactants to increase the solubility of PAHs by decreasing the interfacial surface tension between PAH and the soil/water interphase (Li and Chen, 2009). In one study, the efficiency of surfactants in enhancing desorption of PAHs from contaminated soils relative to water was evaluated (Zhou and Zhu, 2009). Hence, PAHs-degradation process involving surfactant utilization need to be optimized for each of the factors influencing degradation including, surfactant type and concentration, PAH specificity as well as presence of host bacteria (Jin et al., 2007). (iii). Phytodegradation: This is a process whereby plants are used in-situ as well as their associated microorganisms to extract, sequester or detoxify pollutants from contaminated systems (Epuri and Sorenson, 1997). Phytodegradation also referred to as phytotransformation is an important phytoremediation process which indicates the role of internal plant mechanisms and processes in removing contaminant from substrates (Newman and Reynolds, 2004). Research has indicated that various grasses and leguminous plants are potential candidates for phytodegradation of organics (Adam and Duncan, 1999). Some tropical plants have also been reported to show effective degradation tendency due to inherent properties such as deep fibrous root system and tolerance to high hydrocarbon and low nutrient availability (Dzantor et al., 2000). Recently published data have revealed that the tall fescue grass (Festuca arundinacea) and switch grass (Pannicum virgatum) are capable of degrading about 38% of pyrene in 190 days (Chen et al., 2003). Other studies involving pyrene degradation include that of Chouychai et al., (2009) and Cheema et al., (2009) who documented evidence showing plant enhanced phenanthrene and pyrene degradation in acidic soil as

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well as degradation of phenanthrene and pyrene in spiked soils by single and combined plants cultivation. These authors all agreed that the presence of vegetation such as corn (Zea mays), alfalfa (Medicago sativa) and rape seed (Brassica napus) significantly enhanced the adsorption of PAHs from contaminated soils. Moreover, F. arundinacea has been reported to show promising efficiency in promoting phytodegradation of PAHs in contaminated soils (Cheema et al., 2008). Meanwhile, it has been suggested that rice (Oryza sativa) is able to degrade PAHs successfully (Du, et al., 2011). The effectiveness of phytodegradation has also been observed when the process was applied as a secondary treatment process for decontamination of PAHs composted soils (Parish, et al., 2004). However, in contrast to the study of Parish, Banks, and Schwab (2004), some studies have documented phytodegradation as a primary remediation technology and a final polishing step for treatment of soil contaminated with PAHs (Pradham, et al., 1998). Reduction of PAHs concentration was evident after six months of treatment with P. virgatum and little bluestem grass (Schizachyrium scoparium). In a related study, Campbell, Paquin, Awaya, and Li (2002) noted that using industrial hemp (Cannabis sativa) in treating PAHs contaminated soil, led to reduced concentrations of benzo(a)pyrene and chrysene. Recent advances in phytodegradation studies have come to light in the application of rye grass (Lolium multiflorum) and bermudagrass (Cynodon dactylon) in degrading alkylated two ring naphthalenes (White, et al., 2006). In their contribution on distribution of PAHs in sub-cellular root tissue, Kang, Chen, Gao, and Zhang (2010) and Ward, Singh, and Van Hamme (2003) had revealed that using L. multiflorium, pyrene was most adsorped in the root of the plant than other PAHs. Previous studies from Simonich and Hites (1994), Bakker, Vorenhou, Sum, and Kollôffel (1999) and Gao and Zhu (2004) are of the opinion that the efficiency of plant uptake and metabolism of PAHs depended on the morphology of the plant system.

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The efficiency of water hyacinth (Eichhornia crassipes) in degradation of organics has been investigated by Xia (2008). The plant is also reported to accumulate high levels of five and above rings PAHs as opposed to two and three rings compounds (Moustafa and Shara, 2009). Wolverton (1975) showed that E. crassipes was able to adsorb organics and rapidly metabolize these to other components, while in a follow-up study, Wolverton & McDonald (1976) had demonstrated that the plant can adsorb and metabolize phenols and other trace organics. According to Nesterenko, Kirzhner, Zimmels, and Armon (2012), E. crassipes devoid of rhizospheric bacteria reduced about 45% of naphthalene in waste water in 7 days. Some recognized studies have shown that E. crassipes could improve the quality of oil-refinery waste water by decreasing petroleum hydrocarbons by about 18% (Tang and Lu, 1993). Nor (1994), shared the opinion that the plant has tremendous capability to adsorb phenolic compounds as well as heavy metals simultaneously. Recent findings have suggested that adsorption of hydrocarbons by E. crassipes is dominated by Van Der Waals forces (Lin and Zheng, 2003). Hence, the plant has capability to adsorb other compounds such as inorganic nutrients and pesticides (Reddy et al., 2006; Ebel et al., 2007). (iv). Combined Degradation: This is a novel degradation method whereby two or more degradation approaches are applied to remove PAHs from contaminated systems. It has the advantage over other degradation methods since it is regarded as efficient and cost effective, leaving behind no dead-end products capable of further contamination of the substrate after the clean-up exercise is over. Combined chemical pre-oxidation and bioremediation is a notable degradation approach investigated by Kulik, Goi, Trapido, and Tuhkanen (2005). Other studies in this regard include anaerobic digestion and ozonation (Bernal-Martínez et al., 2005), biodegradation and modified Fenton reagent (Nam et al., 2001), biological, chemical and electrochemical treatment (Zheng et al., 2007) as well as Fenton reagent versus ozonation (Goi and Trapido, 2004). A growing body of data have been reporting on the

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use of combined vegetation establishment with chemicals in degradation of PAHs (White et al., 2006; Pan et al., 2009; White et al., 2006) reported that a combination of L. arundinacum and L. multiflorum mixture with fertilizer successfully degraded two rings PAHs such as naphthalene. In a separate study, Pan et al. (2009) had observed that B. napus degraded PAHs in humic acid environment. Successes in combined degradation of PAHs have been documented using pressure assisted ozonation (Hong et al., 2008) and integrated treatment using soil washing, ozonation, and biological treatment of substrate (Haapea and Tuhkanen, 2006). Oxidation using ozone in the presence of sand was examined by Choi et al., (2001), while combined treatment using sequence extraction with surfactant-electrochemical degradation was successfully employed by Alcántara, Gómez, Pazos, and Sanromán (2008). Recent developments in the study of combined degradation method have come to light following reports published by Bernal-Martínez, Patureau, Delgenè, and Carrière (2009) and Bernal-Martínez, Carrière, Patureau, and Delgenè (2007) in the use of anaerobic digestion with recirculation of ozonated digested sludge and anaerobic digestion using ozone pre-treatment. Combined chemical and phytodegradation using inorganic nutrients such as NaNO3, Na2SO4 and Na3PO4 each with E. crassipes is an evolving innovative approach that is beginning to gain wide public approval. Ukiwe, Egereonu, Njoku, and Nwoko (2013) had documented evidence showing that a combination of each of the above mentioned inorganic nutrients together with E. crassipes is capable of degrading about 95% of PAHs in soil leachate, while the plant uses same nutrients as energy source leaving behind no by products. (v). Thermal approaches to the removal of PAHs: Thermal treatment of polycyclic aromatic hydrocarbons under controlled conditions is one of the more effective methods of degradation (Pakpahan et al., 2009). The method has applications at different reaction temperatures and under various conditions. Thermal degradation can occur at low, intermediate or high temperatures. It can also occur in the presence of additives with or without the presence of oxygen. Extraction with water

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or degradation in the presence of water is also a feasible method of PAH degradation (Lui et al., 2001). 1. Low temperature processes: Low-temperature thermal processes are generally classified as operating below 177 oC, though temperatures slightly above this are acceptable (Health care without harm, 2001). Uncontrolled low temperature thermal treatment of organic waste leads to the formation of PAHs with more than 4 rings while degradation occurs for only those with less than four rings (Pakpahan et al., 2009). According to a study conducted on the thermal stability of benzo[a]pyrene (BaP), benzo[a]anthracene (BaA) and dibenzo[a,h]anthracene (DBahA) at temperatures 100 - 200oC, the loss of each PAH, whether found as a solid or in solution depends on time. More of the original compounds were lost at 200 oC than at 100 oC (Chen and Chen, 2001). Low temperature thermal degradation of PAHs is sometimes enhanced in the presence of biochemical processes (Eriksson et al., 2003). 2. Hydrothermal processes (Hydrous pyrolysis): Pyrolysis is a physico-chemical action belonging to the umbrella group of thermal processes. It is the thermal decomposition of organic material with no or very limited oxygen at temperatures ranging between 300 and 600oC (Brown and Stevens, 2011). Under hydrothermal conditions, toxic and refractory organic compounds undergo oxidation and reduction among other reactions, however oxidation is the most widely used method for waste remediation purposes (Rice and Steven, 1998). Interest in the use of water as a ‘green’ solvent and reaction medium is increasing. This has resulted in a lot of research work on the reactions of organic molecules at hydrothermal subcritical (100-350oC, 5-20MPa) and supercritical water (374oC, 22.1MPa) conditions (Savage, 1999 in Kim et al., 2008). The use of pressurised water (Pressurised Hot Water Extraction,

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PHWE) in removing PAHs from sludge and sediments has proved effective with maximum efficiency at 300oC. (vi). Other Established Degradation Techniques: Solar ultraviolet radiation has been shown to degrade and alter the quantity of organic pollutants such as PAHs (Bertilsson and Widerfalk, 2002). Photochemical degradation of anthracene and phenanthrene was noted to be possible under the influence of humic acid (Bertilsson and Widerfalk, 2002). Direct photolysis of PAHs has also been observed under simulated solar radiation with pyrene degrading at a faster rate than phenanthrene and naphthalene (Jacobs et al., 2008). Similarly, ultra sound frequencies have been revealed to generate complete degradation of PAHs under operating conditions of 40 oC and applied current of 150 W (Psillakis et al.,2004). Wheat and Tumeo (1997) reported on the effectiveness of using sonochemical method to degrade phenanthrene and biphenyls in aerated aqueous solutions in the presence of Fe3+ ions. Further studies by Wang, Chen, and Yao (2003) and Manariotis, Karapanagioti, and Chrysikopoulo (2011) investigated sonochemical degradation of PAHs using high frequency ultrasound. Published data by Wang, Chen, and Yao (2003) and Manariotis, Karapanagioti, and Chrysikopoulo (2011) noted that the presence of chlorinated solvent during sonication resulted in the formation of solvent radicals which react with PAHs leading to the build-up of chloro-PAHs byproducts which are degraded forms of the PAHs compound. Gordon and Cain (2003) documented observations supporting the already held belief that titanium film annular photocatalytic reactors are able to degrade PAHs in dilute water streams. Nevertheless, in their separate contribution, Alshawabkeh and Sarahney (2005) examined the effect of current density on enhanced transformation of naphthalene and concluded that almost 88% of the compound was degraded after 8 h when the current density was increased from 13-18.2 mAL-1, making this degradation approach one of the ideal novel processes of PAHs decontamination.

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2.1.6 ANALYTICAL PROCEDURES FOR THE MEASUREMENT OF PAHs Measurements of PAHs in environmental matrices mostly require difficult analytical chemical procedures as a result of the complexity encountered in environmental samples. This complexity is readily seen when one considers the general categories of phases into which environmental samples may be categorized, namely; aqueous, air (gaseous or condensates/particulate matter), oil or organic liquid, solids or sludge, biological samples and even multiphase samples (Poster et al., 2006). Sampling, sample preparation, isolation and concentration from sample matrices are critical in the analytical determination of PAHs. Typical sample preparation procedures include extraction, concentration and clean-up (Speight, 2005). Extraction methods explored over the years include: Soxhlet, sonication, liquid-liquid extraction, purge and trap, headspace, shaking, vortex, solid-phase, supercritical fluid extraction, solid-phase micro extraction (SPME), stir-bar sorptive extraction (SBSE), miniaturized solid-phase extraction (SPE), liquid-phase micro extraction (LPME), membraneassisted solvent extraction techniques (MASE) etc., (Tang and Isacsson, 2008). Sample concentration methods employed include; nitrogen blowing, vacuum evaporation, micro-Snyder column technique and adsorbent or cryogenic trapping (Tang and Isacsson, 2008; USEPA, 1996). Some identification and quantification methods include the use of Gas Chromatography (GC), Gas Chromatography-Mass Spectrometry (GC-MS) and High Performance Liquid chromatography (HPLC).

(a). Sampling Sampling procedures differ depending on the matrix in which PAHs are found. The purpose of soil sampling is to determine the presence of contaminants. There are two basic techniques for soil sampling. Samples can either be collected with (i)

some form of core sampling through the drilling of boreholes, or

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(ii)

By excavations or trenches in which the samples are cut from the soil mass with hand-held corers (Pawliszyn, 2012).

Soil samples collected from a backhoe excavation, the ground surface, soil stockpiles or by means of a manual soil coring device such as the auger tool are usually collected in a thin-walled stainless steel or brass cylinder at least 3 inches long. All samples should be packed in a cooler with dry or blue ice in a manner that should prevent damage during transport to the analytical laboratory. Temperature during transport should be maintained at 4°C or below (Pawliszyn 2012). A thermometer should be placed in the cooler. Samples are kept at 4°C or below at the laboratory until they are analysed. Holding time should not exceed 14 days from the time of collection (Mason, 1992). (b). Extraction Soxhlet Extraction, Ultrasonic Agitation/Sonication and Mechanical Agitation which are conventional methods of extraction are globally practiced but modern extraction methods include: 1. Solid Phase Extraction-SPE (which is like a ‘miniature chromatography’) (Skoog et al., 2004), 2. Supercritical Fluid Extraction (SCF) (mostly the use of CO2 at critical temperature and pressure as a ‘green’ replacement to solvents) (Bell, 2009), 3. Microwave- Assisted Extraction (MAE) (which utilises microwave energy to heat the solvent) (Lopez-Avila et al., 1994) and, 4. Membrane Extraction (ME) (Rawa-Adkonis et al., 2006). (c). Solvent Extraction PAHs are known to be soluble in a wide range of organic solvents. Solvent extraction comes highly recommended for solid environmental samples particularly particulate matter from air and combustion effluent collected on filters (Lee et al., 1981). Solvents like acetone, benzene, cylcohexane, chloroform, methanol and other alcohols, acetic acid, benzene-methanol, petroleum ether,

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dichloromethane and tetrahydrofuran have been used. Acetone, benzene, cylcohexane have proved to be 100 % efficient in Soxhlet Extraction of benzo[a]pyrene from filters among others (Lee et al., 1981). The use of ultrasonic vibration at room temperature has also been explored for extraction of atmospheric dust (Mitra, 2003). Accelerated Solvent Extraction (ASE) or Pressurized Fluid Extraction (PFE) is a modern solvent extraction technique similar to Soxhlet extraction, except that the solvents are used near their supercritical region where they have high extraction properties and the solvent below its boiling point, enables a high penetration of the solvent in the sample. This method is unique because it allows high extraction efficiency with a low solvent volume (15-40 ml) and a short extraction time (15-20 min).The method is applicable to the extraction of water insoluble or slightly water soluble organic compounds (USEPA, 2007). Liquid -Liquid Extraction (LLE) which involves the partitioning of an analyte between an organic solvent and aqueous solution has been widely used for the extraction of aqueous PAHs. Improved versions of LLE include Microscale Solvent Extraction-MSE, Single Drop Micro Extraction-SDE, Gulden Large Sample Extraction-GLSE and Continuous Liquid-Liquid Extraction-CLLE (Pino et al., 2002)

(d). Concentration and clean-up Clean-up methods for PAHs analysis include solvent partitioning and chromatographic procedures (flame ionization detector chromatography and mass spectroscopy chromatography). Adsorbents in use

include

silica,

alumina

(classical),

florisil,

BioBeads

S-X3,

polydimethylsiloxane,

PDMSdivinylbenzene (PDMS-DVB), PDMS-Carbowax (Rawa-Adkonis et al., 2006). The following are examples of some selected concentration and clean-up schemes that are adaptable to various

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matrices. Though these schemes were not directly employed in this research, they give a broad perspective on the pre-analysis treatment of PAH extracts.

(e). Identification and quantification Column, paper, gel permeation and thin layer chromatography, flame ionization detector chromatography (FID-GC), High performance liquid chromatography (HPLC), LC-MS or HPLC-MS, Supercritical-Fluid Chromatography (SFC) LC-GC-MS, GCXGC-TOFMS, LC-To FMS and GCIRMS have successfully been used for identification and quantification of environmental PAHs.

(f). Gas Chromatography (GC) Gas Chromatography (GC) is an analytical technique used to separate compounds based primarily on their volatilities and thermal stability. Gas chromatography provides both qualitative and quantitative information for individual Compounds present in a sample. If all or some of a compound or molecules are in the gas or vapour phase at 400-450°C or below, and they do not decompose at these temperatures, the compound can probably be analysed by GC. Compounds move through a GC column as gases, either because the compounds are normally gases or they can be heated and vaporised into a gaseous state. The compounds partition between a stationary phase which can be either a solid (Gas Solid Chromatography GSC) or a liquid (Gas Liquid Chromatography GLC) and a mobile phase (gas). The differential partitioning into the stationary phase allows the compounds to be separated in time and space. In GLC usually referred to as GC (figure 2.2), the stationary phase is almost always a relatively nonvolatile liquid. This liquid is coated on either solid particles or on the inside walls of a capillary tube.

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Figure 2.2: Schematic diagram of a Gas Chromatograph In the 1950s and throughout the 1960s GC was used in the separation of smaller aromatic hydrocarbons like separation of xylene isomers and alkylated benzenes. Since PAHs cover a wide range of volatility, for example 218oC and 525oC are boiling points for naphthalene and coronene respectively, it is necessary to maintain high column temperatures in order to ensure elution of all the different sizes of molecules. In the 1960s, advances in columns for GC analysis of PAHs led to the introduction of thermostable silicone solid phases (Lee et. al., 1981). With time conventional packed columns became undesirable due to their inability to separate complex isomeric mixtures of PAHs. Glass capillary columns were introduced which offered greater resolution and inertness. Thin- film coating of stationary phase also offered reduced temperatures required to elute high molecular weight molecules of PAHs. Typical stationary supports for columns for GC analysis have been silicone and carborane polymers (Lee et. al., 1981). The traditional GC determination of PAHs has undergone a lot of innovations all aimed at better output. Modern approaches to GC analysis include Large-volume injection GC, fast GC, thermal desorption and others.

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(g). GC Detection systems GC detection systems include Flame Ionisation Detector (FID), Thermal Conductivity Detector (TCD) and the Electron Capture Detector (ECD). FID is the most commonly used GC detector since it responds to any molecule with a carbon-hydrogen bond is mass sensitive and destroys the sample afterwards (Lee et al., 1981; SRI Instruments, 2011). TCD is not as sensitive as other detectors but it is non-specific and non-destructive (Grob, 2004). ECD is selective to electronegative compounds; it is as sensitive as the FID but has a limited dynamic range and finds its greatest application in analysis of halogenated compounds (SRI Instruments, 2011).

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2.2 SURFACTANTS The term surfactant was coined by Antara products in 1950. Surfactants are amphiphilic (possess both hydrophilic and hydrophobic properties) molecules with a hydrophilic polar head (Gao et al., 2007). At or above a certain concentration level called critical micelle concentration (CMC) (Franzetti et al., 2010), the hydrophobic parts of the surfactants will tend to associate together to form a micelle (an aggregate of surfactant molecules dispersed in a liquid colloid) with a hydrophobic core (Wick et al., 2011). When micelles form in water, their tails form a core that can encapsulate an oil droplet, and their (ionic/polar) heads form an outer shell that maintains favorable contact with water. The number of molecules necessary to form a micelle generally varies between 50 and 100; this is defined as the aggregation number. The CMC of surfactants in aqueous solution can vary depending on several factors, such as molecule structure, temperature, presence of electrolytes and organic compounds in solution. At soil temperatures, the CMC typically varies between 0.1 and 1 mM. The size of the hydrophobic region of the surfactant is particularly important for the determination of the CMC: in fact the CMC decreases with increasing hydrocarbon chain length, i.e. increasing hydrophobicity. The addition of a CH2- group to the chain has been shown to decrease the CMC by a factor of 3, according to the Traube’s rule. However, anionic surfactants have higher CMCs than nonionic surfactants even when they share the same hydrophobic group. At concentrations above the CMC, additional quantities of surfactant in solution will promote the formation of more micelles. The formation of micelles leads to a significant increase in the apparent solubility of hydrophobic organic compounds, even above their water solubility limit, as these compounds can partition into the central core of a micelle. The effect of such a process is the enhancement of mobilization of organic compounds and of their dispersion in solution (Perfumo et al., 2010). This effect is also achieved by the lowering of the interfacial tension between immiscible phases. In fact, this contributes to the creation of additional

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surfaces, thus improving the contact between different phases (Christof et al 2002). When surfactants assemble in oil, the aggregate was referred to as a reverse micelle where the heads are in the core and the tails maintain favorable contact with oil. Surfactants are often classified into four primary groups (Mulligan et al., 2001), viz; anionic, cationic, non-ionic, and amphoteric or zwitterionic. i) ANIONIC: The hydrophilic head is negatively charged. This was the most widely used type because it’s good for cleaning applications from household tasks, dry cleaning, shampoos and other personal cleaning products, to specialty industrial uses such as dispersing agent in dyes (Christofi and Ivshina 2002). ii) CATIONIC: The hydrophilic head is positively charged. Cationic surfactants adhere to and modify solid surfaces. They are used as ingredients in fabric softeners, fuel and lubricant additives, disinfectants, and corrosion inhibiters (Christofi and Ivshina 2002). iii) NONIONIC: The hydrophilic head is polar but not fully charged. They are low foaming and good for dishwasher detergents, textile applications and in combination with anionic surfactants for cleaning applications (Christofi and Ivshina 2002). iv) AMPHOTERIC: The molecule has both potential positive and negative groups depending on the pH of the solution. They are mild and have a low irritancy factor. They are used in dermatological and cosmeceutical applications and in combination with other surfactants (Christofi and Ivshina 2002).

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2.2.1 CLASSIFICATION OF SURFACTANTS Surfactants, of both biological and chemical origin, are amphipathic molecules that accumulate at interfaces, decrease interfacial tensions, and form aggregate structures that allow hydrocarbon solubilization. Due to these properties, surfactants modify interfacial behavior and impact the way other molecules behave at interfaces and in solutions. In general, surfactants can be classified into two types based on their origin namely, chemically synthesised surfactants and biologically derived surfactants (Makkar et al., 2003). Another group of surfactant is from biosurfactant producing microorganisms of the genus Pseudomonas, Candida and Bacillus that is known to produce rhamnolipids, glycolipids and lipopeptide surfactants (Rahman and Gakpe, 2008). From the current perspective of moving towards green chemistry and the environment, biological surfactants or biosurfactants have exhibited in several studies similar emulsification properties and high surface activity characteristics compared to chemically synthesised surfactants (Banat, 1995).

2.2.2 SYNTHETIC SURFACTANTS Synthetic Surface active agents (surfactants) are chemically synthesized amphiphilic molecules with both hydrophilic and hydrophobic parts, which have a wide range of properties, including the lowering of surface and interfacial tension of liquids, and the ability to form micelles and microemulsions between two different phases. The most common hydrophobic parts of chemically synthesized surfactants are paraffins, olefins, alkylbenzenes, alkylphenols and alcohols (Franzetti et al., 2010); the hydrophilic part is usually a sulphate, sulphonate or a carboxylate group in anionic surfactants, a quaternary ammonium group in cationic surfactants and polyoxyethylene, sucrose or a polypeptide in nonionic surfactants. An important descriptor of chemico-physical properties of surfactants is related to the balance between their hydrophilic and hydrophobic moieties. Thus, surfactants can also be classified according to their 45 | P a g e

Hydrophile-Lipophile Balance (HLB) (Franzetti et al., 2010). The HLB value indicates whether a surfactant will produce a water-in-oil or oil-in-water emulsion: emulsifiers with a lower HLB value of 3-6 are lipophilic and promote water-in-oil emulsification, while emulsifiers with higher HLB values between 10 and 18 are more hydrophilic and promote oil-in-water emulsions. A classification based on HLB values has been used to evaluate the suitability of different surfactants for various applications. For example, it has been reported that the most successful surfactants in washing oil contaminated soils are those with a HLB value above 10 (Christof et al., 2002). The following in table 2.2 are some of the most commonly used synthentic surfactants.LAS (linear alkylbenzene sulfonate), TDTMA (tetradecyltrimethyl-ammonium bromide), Tween-80 (polyoxyethylene sorbitan monooleate), Brij30 (polyoxyethylene-4-lauryl

ether),

10LE

(polyoxyethylene

-10-lauryl

ether),

and

Brij35

(polyoxyethylene-23-lauryl ether). Tween 80: Polysorbate 80 is derived from polyethoxylated sorbitan and oleic acid. The hydrophilic groups in this compound are polyethers also known as polyoxyethylene groups, which are polymers of ethylene oxide. In the nomenclature of polysorbates, the numeric designation following polysorbate refers to the lipophilic group, in this case the oleic acid. The full chemical names for polysorbate 80 are: 

Polyoxyethylene (20) sorbitan monooleate



(x)-sorbitan mono-9-octadecenoate poly(oxy-1,2-ethanediyl)

The critical micelle concentration of polysorbate 80 in pure water is reported as 0.012 mM (Chou et al., 2005).

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Table 2.2: Basic properties of some synthetic surfactants Trade name/ abbreviation

Chemical name

LAS

linear alkylbenzene sulfonate (dodecylbenzene sulfonate) Polyoxyethylene sorbitanmonooleate tetradecyltrimethyla mmonium bromide polyoxyethylene 4 lauryl ether polyoxyethylene 10 lauryl ether polyoxyethylene 23 lauryl ether

Tween-80 TDTMA Brij30 10LE Brij35

Symbol

C12BSoNa

Average mol Formula C21H35SO3Na

Type

Average MW

Anionic

390

HLB

CMC (mgl-1)

C18S6E20

C64H125O26

Nonionic

1309

NC14Br

C17H38NBr

Cationic

336.4

C12E4

C20H42O5

Nonionic

362

9.7

10,20

C12E10

C32H66O11

Nonionic

626

14.1

48,63

C12E23

C58H118O24

Nonionic

1198

16.9

40,110

ThOD (gO2g-1)

433.5

15.0

13,40

2.01

100 2.48

2.02

C: represents alkyl chain (-CH2-), B: represents a benzene ring, So: represents a sulfonic group, S6: represents a sorbitan ring, N: represents a trimethylamino group, and E: represents an ethoxylate group (-C2H4O-). MW: molecular weight, HLB: hydrophile-lipophile balance. ThOD: Theoretical oxygen demand. Source: Ziqing O. U., 2000.

2.2.3 BIOSURFACTANTS Biosurfactants are microbially produced surface-active compounds which have amphiphilic molecules. They are a structurally diverse group of surface-active molecules synthesized by microorganisms. Rhamnolipids from Pseudomonas aeruginosa, surfactin from Bacillus subtilis, emulsan from Acinetobacter calcoaceticus and sophorolipids from Candida bombicola are some examples of microbial-derived surfactants (Franzetti et al., 2010). These amphiphilic molecules have both hydrophilic and hydrophobic regions causing them to aggregate at interfaces between fluids with different polarities such as water and hydrocarbons (Mazaheri and Tabatabaee, 2010). Biosurfactants are produced mainly by aerobically growing microorganisms in aqueous media from a carbon source feedstock, e.g. carbohydrates, hydrocarbons, oils and fats or mixtures. Particularly in bacteria, which are in a state of growth on a water-immiscible substrate, which was a source of food, for example 47 | P a g e

crude oil spillage treated with selected microorganisms. By evolution, the bacteria have adapted themselves to feeding on water-immiscible materials by manufacturing and using a surface active product that helps the bacteria which are in the aqueous phase to adsorb, emulsify, wet, and disperse or solubilize the water immiscible material. The emulsifiers are secreted into the culture medium during the growth of the microorganism and assist in the transport and translocation of the insoluble substrates across cell membranes. Some synthetic surfactants can be toxic to microorganisms, which may decrease the number of degraders in the soil (Mazaheri and Tabatabaee, 2010). The addition of biosurfactants (either surfactant producing microorganisms or natural compounds that act as surfactants) is one way to get around the problem of toxicity. Biosurfactant-producing microbes have been proposed as an alternative to chemical surfactants to enhance availability of hydrophobic compounds (Wick et al., 2011). Surfactants that are produced by microorganisms tend to have lower toxicities and are effective at wider temperature, pH, and electrical conductivity ranges (Bordas et al., 2005).

It has been reported that biosurfactant-producing indigenous bacteria achieved higher

hydrocarbon degradation rates than those achieved by nutrient addition alone (Straube et al., 2003) showed that the introduction of Pseudomonas aeruginosa and basillus substilis strains, a rhamnolipid and a lipoprotein producing bacteria, enhanced the biodegradation of PAHs in highly contaminated soil. Conte et al., (2005) found that humic acid can be used as an effective surfactant for PAH desorption. Bogan and Sullivan (2003) reported that the addition of fulvic acid to soils that had low humic acid/fulvic acid content greatly enhanced pyrene mineralization by Mycobacterium austoafricanum. They also reported slower progress in PAH sequestration in a soil with high fulvic acid content. Cyclodextrins: Cyclodextrins are cyclic oligosaccharides that are able to form complexes with hydrophobic molecules. They are the product of the action of cyclodextrin glycosytransferases on

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starch and are non-toxic and biodegradable in the environment. Cyclodextrins can be added to washing waters to solubilize PAHs and increase desorption from soil (Stokes et al., 2006). Vegetable oil: Vegetable oil has been proposed as an economic and environmentally friendly solvent to dissolve PAHs and has been shown to be as effective as organic solvents like acetone and dichloromethane (Gong et al., 2005). Glycolipids: The best studied microbial surfactants are glycolipids. Among these, the most known compounds are rhamnolipids, trehalolipids and sophorolipids, which are disaccharides combined with long-chain aliphatic acids or hydroxyaliphatic acids. Rhamnolipids are composed of one or two molecules of rhamnose linked to one or two molecules of β-hydroxydecanoic acid. The hydroxyl group in one of the acids has a glycosidic linkage with the reducing end of the rhamnose disaccharide, while the hydroxyl group of the second acid is involved in ester formation. As one carboxylic group is free, the rhamnolipids are anions above pH 4. Production of rhamnolipids was firstly reported in Pseudomonas aeruginosa and was then extensively studied also in other Pseudomonas species. Rhamnolipids can lower the surface tension of water to 25- 30 mN m-1 and the interfacial tension against n-hexadecane to 1 mN m-1; their CMC value range from 10 to 30 mgL-1. Rhamnolipid and rhamnolipid (L-rhamnosyl-L-rhamnosyl-β-hydroxydecanoyl-β-hydroxydecanoate and L-rhamnosyl-βhydroxydecanoyl-β-hydroxydecanoate, respectively) are the main glycolipids produced by P. aeruginosa. At present, they are the only microbial surfactants fully commercialized as a mixture for bioremediation purposes. Trehalolipids are a wide group of glycolipids, constituted by the disaccharide trehalose linked at C-6 and C-6’ to mycolic acids, which are long-chain α-branched-β-hydroxy fatty acids. Trehalolipids are produced by a number of different microorganisms, such as Mycobacterium, Nocardia and Corynebacterium. However, the most extensively studied compounds in this class are trehalose

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dimycolates produced by Rhodococcus erythropolis. Trehalolipids produced by different microorganisms differ in their structure, size and degree of saturation. The minimal values for interfacial tension of water against n-hexadecane achieved with different trehalolipids range between 1 and 17 mN m-1, while the surface tension is lowered to 25 and 40 mN m-1 by trehalose lipids produced by R. erythropolis and Arthrobacter sp. The CMC for trehalolipids is quite low, about 2mgL-1. Sophorolipids consist of two glucose molecules linked β-1, 2 (sophorose), with 6- and 6’-hydroxyl groups, generally acetylated, linked to a long-chain hydroxy fatty acid. The terminal carboxyl group can be in the lactonic form or hydrolyzed to give an anionic surfactant. They are produced mainly by yeasts, such as Torulopsis bombicola, T. petrophilum, T. apicola and Candida bogoriensis. Both lactonic and anionic sophorolipids were demonstrated to lower the interfacial tension of water against nhexadecane or vegetable oils to 1-5 mN m-1 over a wide range of pH, temperature and salt concentration (Franzetti et al., 2010). Lipopeptides: Most Bacillus species synthesize a number of cyclic lipopeptide antibiotics during the early stages of sporulation. For example, B. polymyxa produces polymixin, a decapeptide in which amino acids 3-10 form a ring structure, linked to a branched fatty acid, while B. brevis produces gramicidin S, a cyclic decapeptide consisting of a rigid ring with two positively charged ornitine sidechains on one side and the hydrophobic side-chains of the other residues on the other side. B. licheniformis produces a mixture of several lipopeptides acting synergistically; one of these surfactants can lower the interfacial tension between water and n-hexadecane to the very low value of 0.36 mN m1. The most relevant cyclic lipopeptide is surfactin produced by B. subtilis, because of its very high activity. Surfactin has a CMC of 25-50 mgL-1 and can lower the surface tension of water to 27 mNm-1, while the lowest interfacial tension against n-hexadecane is 1 mNm-1.

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Bioemulsifiers: A wide variety of microorganisms, including some Archaea, produce highmolecularweight polymers having the property to stabilize emulsions. Such polymeric compounds are generally exocellular polysaccharides, proteins, lipopolysaccharides or lipoproteins, in some cases combined in complex mixtures (Perfumo et al., 2010). The best studied bioemulsifiers are those synthesised by various species of Acinetobacter. Among these, the first studied compound was RAG-1 emulsan, produced by A. calcoaceticus RAG-1. It is a polyanionic amphiphilic heteropolysaccharide which contains a repeating trisaccharide, with long-chain fatty acids covalently linked through ester bonds. The hydrophobic groups are distributed across the molecule, forming a comb-type polymer. It is different from most of the other bioemulsifiers, since the latter are rather composed by mixtures of hydrophilic and hydrophobic polymers.

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Table 2.3: Classification and microbial origin of biosurfactants. Surfactant Class Trehalose lipids

Rhamnolipids Glycolipid (rhamnoselipid) Sophorose lipids

Microorganism Arthrobacter paraffineus spp. Corynebacterium spp. Mycobacterium spp. Rhodococus erythropolis Pseudomonas aeruginosa

Candida apicola Candida bombicola Candida lipolytica Glucose, fructose, saccharose lipids Arthrobacter spp. Corynebacterium spp. R. erythropolis Cellobiose lipids Ustilago maydis Polyol lipids Rhodotorula glutinous Rhodotorula graminus Diglycosyl diglycerides Lactobacillus fermentii Lipopolysaccharides Acinetobacter calcoaceticus (RAGI) Pseudomonas spp. Candida lipolytica Lipopeptides Arthrobacter sp. Lipoprotein (surfactin) Bacillus pumilis Bacillus subtilis Bacillus licheniformis Pseudomonas fluorescens Omithine, lysine peptide Thiobacillus thiooxidans Streptomyces sioyaensis Gluconobacter cerinus Phospholipids T. thiooxidans Corynebacter alkanolytic Sulfonylipids Capnocytophaga spp. Fatty acids (corynomycolic acids, Penicillium spiculisporum spiculisporic acids, etc.) Corynebacterium lepus Arthrobacter parafineus Talaramyces trachyspermus Adapted from Mulligan C. N., Yong R. N., Gibbs B. F. 2001.

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2.2.4 PHYSIOLOGICAL ROLES OF BIOSURFACTANTS The main physiological role of biosurfactants was to permit microorganisms to grow on Waterimmiscible substrates by reducing the surface tension at the phase boundary, therefore making the substrate more readily available for uptake and metabolism. However, the molecular mechanisms of the uptake process of these substrates (e.g: Alkanes) are still unclear. In addition to emulsification of the carbon source, they are also involved in the adhesion of microbial cells to the hydrocarbon. The cellular adsorption of the hydrocarbon-degrading microbes to water-immiscible substrates, and the excretion of surface-active compounds together allow growth on such carbon sources. Biosurfactants are classified according to their biochemical nature while the synthetic surfactants are usually classified according to the nature of their polar groups. The microorganism and the type of biosurfactant produced are given in Table 2.3. The rhamnose containing glycolipids produce by pseudomonas species are among the biosurfactants have been studied more extensively.

2.2.5 EFFECTS OF SURFACTANTS ON BIODEGRADATION OF PAHs Mass transfer of PAH compounds to the aqueous phase in the soil solution can be a major limiting factor in the bioremediation of PAHs (Wick et al., 2011).

Surfactants like cyclodextrins, and

vegetable oil may be used to enhance PAH solubility. Properly applied surfactants have been shown to improve desorption, apparent aqueous mobility and bioavailability of hydrophobic organic compounds such as PAHs (Mata-Sandoval et al., 2002). Although it is agreed that surfactants can enhance the solubility and dissolution of hydrocarbons from contaminated soil, contradictory results have been reported on the ability of surfactants to enhance the biodegradation of hydrocarbons. The focus is whether solubilization is conducive or binhibitory to the microbial uptake of hydrocarbons. The enhanced biodegradation in the micellar solution can be attributable to the increased solubility and bioavailability of substrate to bacteria, surfactant-enhanced substrate transport through the microbial 53 | P a g e

cell wall, increased interfacial area in the presence of surfactant, enhanced contact of bacteria with the hydrocarbon water interface, facilitated direct contact between cells and non-aqueous liquid phase, and decreased diffusion path length between the site of adsorption and site of bio-uptake by the microorganism due to enhanced adsorption of cells to hydrocarbon occupied soil particles in the presence of surfactant (Wick et al., 2011). Tiehm and Frizsche studied the biodegradation of both single and mixture of PAHs presolubilized by surfactant. Accelerated biodegradation rates were found for both single and mixed PAHs presolubilized compared with the rate of PAHs in crystal form. This indicated that solubilization increased the bioavailability of PAHs. The inhibitory effect was normally observed at surfactant concentrations approaching and exceeding the CMC. Potential mechanisms of inhibition include toxicity of surfactant to the microorganism, preferable microbial uptake of surfactants as substrate (Mukesh-kumar et al., 2012), and inhibition of the direct contact between cells and hydrocarbon by surfactant micelles. It was also observed that the effect of surfactant was also dependent on the specific bacteria involved (Bustamante et al., 2012), which means that the specific interactions between bacteria and surfactant also play an important role.

2.2.6 SOLUBILIZATION BY SURFACTANTS Surfactants solubilize hydrophobic contaminants by partitioning them into the hydrophobic core of the micelle. If the concentration of surfactant exceeds the critical micelle concentration, solubility of hydrophobic compounds can increase by an order of magnitude over normal aqueous solubility (Mulligan et al., 2001; Gao et al., 2007). The critical micelle concentration of a specific surfactant depends on temperature, ionic strength and surfactant chemistry. In mixed pollutant systems, the extent of solubilization will differ from those in single solutes. For example, Guha et al. (1999) found that naphthalene solubilized by surfactants increased the solubilization of other PAHs such as phenanthrene. Surfactants are thought to be able to solubilize 54 | P a g e

sorbed organic compounds in a soil-water system only after critical micelle concentration is attained (Gao et al., 2007). Several interactions affect the solubilization of hydrophobic compounds by surfactants. These include the micellular phase-organic interactions, surfactants monomer-organic interactions in the aqueous phase, and the interactions of surfactants and organic compounds with the solid phase (Franzetti et al, 2010). Soils with predominantly fine particles are generally found to reduce desorption efficiency of surfactants (Mulligan et al., 2001). Organic matter content and clay mineralogy also affect surfactant performance, but the effect differs depending on the type of surfactant used (Rodriquez-Cruz et al., 2005). Stimulatory effects, no effect, and inhibitory effects have been reported (Kim and Weber, 2003). Some of the negative effects of surfactants on biodegradation may be due to toxicity to microorganisms, prevention of bacterial access to contaminants through sequestration of micellar solubilized organics, or preferential biodegradation of the surfactant rather than the contaminant. For instance, phenanthrene solubilized by the surfactant Tween-80 was found to be unavailable for biodegradation by Sphingomonas paucinobilis because this microorganism preferred to use the hydrophobic portion of the surfactant as a carbon source rather than destabilizing the micelles. Because of the above interactions, the determination of a critical micelle concentration for a surfactant in a complex medium such as soil can be difficult. Desorption occurs at surfactant concentrations greater than the critical micelle concentration, but at lower concentrations, admicelles (surface aggregates of surfactants, also called hemimicelles) may form, sorb onto soil, and act as additional sorption sites that can enhance PAH sorption instead of reducing it (Wick et al., 2011). Anionic and nonionic surfactants are more commonly used in remediation because they are less likely to sorb onto soil surfaces (Mulligan et al., 2001). Nonionic surfactants are also advantageous because they have a low critical micelle concentration, high cold water solubility, and low microbial toxicity

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(Kim and Weber, 2003; Zhao et al., 2005). (Conte et al., 2005) have shown that PAHs can be effectively desorbed using nonionic surfactants like dodecylbenzene sulfonate.

2.3 BIOREMEDIATION Bioremediation is defined as the process whereby organic wastes are biologically degraded under controlled conditions to an innocuous state, or to levels below concentration limits established by regulatory authorities. It uses naturally occurring bacteria and fungi or plants to degrade or detoxify substances hazardous to human health and/or the environment (Figure 2.3). The microorganisms may be indigenous to a contaminated area or they may be isolated from elsewhere and brought to the contaminated site. Contaminant compounds are transformed by living organisms through reactions that take place as a part of their metabolic processes releasing water and harmless gases such as carbon dioxide.

Figure 2.3: Schematic diagram of biodegradation by a microbe (Source: A Citizen Guide to Bioremediation- EPA)

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2.3.1 PRINCIPLES OF BIODEGRADATION Biodegradation of a compound is often a result of the actions of multiple organisms. For bioremediation to be effective, microorganisms must enzymatically attack the pollutants and convert them to harmless products. As bioremediation can be effective only where environmental conditions permit microbial growth and activity, its application often involves the manipulation of environmental parameters to allow microbial growth and degradation to proceed at a faster rate. Like other technologies, biodegradation has its limitations. Some contaminants, such as chlorinated organic or high aromatic hydrocarbons, are resistant to microbial attack. They are degraded either slowly or not at all; hence it was not easy to predict the rates of clean up for a bioremediation exercise. The control and optimization of biodegradation processes is a complex system of many factors (Table 2.3) such as the existence of a microbial population capable of degrading the pollutants, the availability of contaminants to the microbial population and the environmental factors (type of soil, temperature, pH, the presence of oxygen or other electron acceptors, and nutrients). Further, it also depends on the concentration of the contaminants (high concentrations may be toxic to the microorganism), the presence of substances toxic to the microorganism and inhibitors to the metabolism of the contaminant. Oxygen level in ex-situ applications was easy to control than in-situ applications and was typically maintained by mechanical tilling, venting and sparging. Anaerobic conditions may be used to degrade highly chlorinated contaminants. This can be followed by aerobic treatment to complete biodegradation of the partially dechlorinated compounds as well as the other contaminants. Water serves as the transport medium through which nutrients and organic constituents pass into the microbial cell and metabolic waste products pass out of the cell (Figure 2.3). Moisture levels in the range of 20% to 80% generally allow suitable biodegradation in soils.

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Nutrients required for cell growth are nitrogen, phosphorous, potassium, sulfur, magnesium, calcium, manganese, iron, zinc and copper. If nutrients are not available in sufficient amounts, microbial activity will stop. Nitrogen and phosphorous are the nutrients most likely to be deficient in the contaminated environment and thus are usually added to the bioremediation system in a useable from (e.g., as ammonium for nitrogen and as phosphate. for phosphorous) Table 2.4: Environmental conditions affecting degradation Parameters

Conditions required for microbial activity

Optimum values for biodegradation

Moisture

25–28% of water holding capacity 30–90%

pH Oxygen content

5.5 – 8.8 Aerobic, minimum air-filled pore space of 10%

6.5–8.0 10–40%

Nutrient content

N and P for microbial growth

C:N:P = 100:10:1

Temperature (oC)

15–45

20–30

Contaminants

Not too toxic

5–10% of dry weight of soil

Heavy metals

Total content 2000 ppm

700 ppm

Type of the soil

Low clay or silt content

SOURCE: M.Vidali 2001; IUPAC; Vol 73, NO: 7. Pp1163-1172

2.3.2 METHODS OF BIODEGRADATION Biodegradation, which involves the use of microbes to detoxify and degrade environmental contaminants, has received increasing attention as an effective biotechnological approach to clean up a polluted environment. In general, the approaches to biodegradation are environmental modification, such as through nutrient application and aeration, and the addition of appropriate degraders by seeding. Three types of biodegradation are predominant in the industry today: natural attenuation, biostimulation, and bioaugmentation. 58 | P a g e

1. Natural attenuation: Natural attenuation was a collection of biological, chemical and physical processes that occur naturally resulting in the containment, transformation or destruction of undesirable chemicals in the environment. In contrast to biostimulation, monitored natural attenuation (MNA), if effective, provides significant benefits in terms of cost and effort. It can also be termed as intrinsic remediation, bioattenuation and intrinsic bioremediation. In this case, the contaminants are left on site and the naturally occurring processes are left to clean up the site. Processes include combination of both biotic and abiotic mechanisms like sorption, volatilization, dilution, and dispersion coupled with biodegradation (Vasilyeva and Strijakova 2007). In abiotic process, various factors such as the soil texture and type, physical and chemical properties of the pollutants determine the bioavailability of the organic chemicals in the soil. The inorganic contaminant-soil interactions, the existence of surface active fractions in the soil such as soil organic matter (SOM), amorphous non-crystalline materials and clays, can significantly enhance oil retention in soils because of large surface areas, high surface charges and surface characteristics. Volatilization may be an important attenuation mechanism for volatile organic contaminants like spilled petroleum products. The abiotic reactions and transformations are sensitive to at least two factors (a) the physicochemical properties of the pollutants itself, and (b) the physico-chemical properties of the soil (i.e. soil fractions comprising the soil). Similar to the inorganic contaminants, abiotic chemical reactions with organic compounds occur and include (a) hydrolysis, (b) formation of a double bond by removal of adjacent groups and (c) oxidation/reduction, or dehydrohalogenation reactions. Chemical mass transfer was responsible for partitioning of contaminants in the fate and transport of contaminants. Reduction –oxidation reactions can also play an important role in the fate of

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the contaminants. Assessment of retention or retardation processes was responsible for determining the partitioning and the attenuation of contaminants within the soil. In biotic processes microorganisms like bacteria, fungi, protozoa, and algae (Mulligan et al 2005) play a key role in degradation of contaminants. Understanding the types of chemicals that can be biodegraded or transformed, and the pathways of conversion was important as well as the toxicity and bioavailability of several chemicals since this will serve as the foundation knowledge required for determining the potential for natural attenuation. Substrates can become less bioavailable via interaction with negatively charged clay particles and organic material (Kumari 2012). Sorption and sequestration can be influenced by pH, organic matter content, and temperature and pollutant characteristics. 2. Biostimulation: If natural degradation does not occur or if the degradation was too slow, the environment has to be manipulated in such a way that biodegradation was stimulated and reaction rates are increased. Biostimulation involves the stimulation of indigenous microorganisms to degrade the contaminant. A biostimulation project requires that adjustments be made to the soil to enhance the microbial populations already present. These include adding a nitrogen source, a phosphorous source, and a myriad of trace minerals and making appropriate pH adjustments. Thus the concentration of these chemicals during biostimulation should be carefully monitored. In general for the biodegradation of the cell mass of microbes having the C: N ratio of about 5:1 to 10:1, C: N: P ratio of 100:10:1 was required (Kumari 2012). Addition of nutrients such as N-P-K or oleophilic fertilizers has therefore been observed to enhance biotransformation (Margesin and Schinner 2001). Normally, inorganic orthophosphates or polyphosphates provide P while inorganic nitrates or ammonium salts provide nitrogen. Whereas when organic nutrients are added, the organic group gets adhere to

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the organic contaminants at the organic-aqueous (Chen et al., 2000). Sometimes other ratios have been used because the carbon in the organic contaminants may mineralize into carbon dioxide rather than assimilate into biomass or the inadequate bioavailability of organic contaminants and nutrients (Kwok and Loh, 2003). For an on site treatment, the nutrients are spread over the site and worked into the soil. For an in situ treatment, the nutrients are added to the water upstream in the hydrostatic gradient. 3. Bioaugmentation: Bioaugmentation involves adding preselected organisms to the site to degrade the contaminant. It was the process in which there was controlled addition of specially formulated biocultures to assist those found naturally in the soil. It was done in conjunction with the development and monitoring of an ideal growth environment in which these selected bacteria can live and work. In most cases, the targeted organic contaminants either serve as the food source or are co-metabolized. Essential elements are added to the "food source" to provide the required nutrient levels, and water provides the media in which the bacteria function. Bioaugmentation was considered to be an effective approach in the case of contamination with high recalcitrant chemicals where natural attenuation and biostimulation was not applicable. However, much attention has been paid to the application of bioaugmentation considering the effects on humans and ecosystem on the addition of large amount of degradative bacteria to the contaminated sites. Moreover, the perishability after remediation and the long term affect on the indigenous microbial community of the injected bacteria must be clarified prior to the application. For many years, bioaugmentation has been practiced intentionally as it enhances a specific biological activity in the remediation of contaminated sites. Due to the possible advantages, bioaugmentation was generally accepted as an efficient technique for soil/sediment bioremediation. The incompetence of

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indigenous microbes in some cases were apparent to enhanced bioremediation rate after the addition of competent microorganisms in some bioaugmentation studies. It was suggested that the reinoculation of soil with indigenous microorganisms directly isolated from the same soil was often included in the term bioaugmentation.

The factors like adaptation, distribution and applicability of indigenous

microorganisms, compound characteristics and physico-chemical characteristics of the contaminated soil/sediments /sludges influences the various process in bioremediation. Microbial ecology was very important in evaluating both the potential success of bioaugmentation and its possible advantages over biostimulation. Microorganisms are affected by maintenance energy, the production of, and resistance to antibiotics and toxic metabolites, predation,etc. The advantages of employing mixed cultures as opposed to pure cultures in bioremediation have also been widely demonstrated attributing to the effects of synergistic interactions among members of the association.

2.3.3 BIODEGRADATION OF ORGANIC COMPOUNDS Biodegradation of organic compounds leads to mineralization where organic substrates are converted into inorganic products. Thus, in the mineralization of organic C, N, P, S or other elements, CO2 or inorganic forms of N, P, S or other elements are released by the organism. In this processes microorganisms destroy numerous organic molecules and the mineralization of synthetic chemicals to inorganic products. Many synthetic chemicals due to their biomagnification when discharged into the environment become directly toxic or hazardous. In such cases, bioremediation was beneficial as a result of mineralization in the total destruction of the parent compound to inorganic products.

2.3.4 CONDITIONS FOR BIODEGRADATION Several conditions must be satisfied for biodegradation to take place in an environment. These include the following:

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(a) An organism that has the necessary enzymes to bring about biodegradation must exist. The mere existence of an organism with the appropriate catabolic potential was necessary but not sufficient for biodegradation to occur. (b) Organism must be present in the chemical containing environment. Although some microorganisms are present in essentially every environment near the earth surface, particular environment may not contain an organism with the necessary enzymes, (c) The chemical must be accessible to the organism having the requisite enzymes. Many chemicals persist even in environments containing the biodegrading species simply because the organism does not have access to the compound that it would otherwise metabolize. Inaccessibility may result from the substrates being in a different microenvironment from the organism, in a solvent not miscible with water or sorbed to solid surfaces. (d) If the initial enzyme bringing about the degradation was extra cellular, the bonds acted upon by that enzymes must be exposed for the catalyst to function. This was not always the case because of sorption of many organic molecules. (e) Should the enzymes catalyzing the initial degradation be intracellular, that molecule must penetrate the surface of the cell to the internal sites where the enzymes acts. Alternatively, the products of an extra cellular reaction must penetrate the cell for the transformation to proceed further. (f) Because the population or biomass of bacteria or fungi acting on many synthetic compounds are initially small, conditions in the environment must be conducive to microorganisms. are frequently are the major and occasionally the sole means for degradation of particular compounds, the absence of a microorganism from particular environment, or its inability to function, often means that the compound disappears very slowly.

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2.3.5 BIOSURFACTANT ENHANCED BIODEGRADATION In 1997 the enhancing effect of biosurfactants on the biodegradation of hydrocarbons was demonstrated (Franzetti et al., 2010). Subsequently, several investigations were published describing the use of biosurfactants in different systems and environments (i.e. liquid, slurry and solid phases, soil, water) (Liu et al., 1995). Also considering the publication bias which probably led to an overpublication of successful applications, the main emerging feature of this large body of literature is the contrasting results reported. In fact, one key point in the application of biosurfactants to environmental remediation is their specificity, due to the fact that different microbial strains produce different molecules. In some cases biosurfactants have enhancing effects on the same producing strain or related organisms. For examples degradation of nhexadecane was stimulated by rhamnolipid in Pseudomonas aeruginosa, but not in Rhodococcus strains, and the same P. aeruginosa was stimulated only by its own rhamnolipid (Li and Chen, 2009). Amphiphiles are able to alter the physico-chemical conditions at the interfaces affecting the distribution of the chemicals among the phases. A hydrocarbon contaminated soil contains at least six phases: bacteria, soil particles, water, air, unsoluble liquid and solid hydrocarbon. The hydrocarbons can be partitioned among different states: solubilised in water phase, ab/adsorbed to soil particle, sorbed to cell surface, as free/unsoluble phase. The addition of biosurfactants alters the hydrophobicity of the surfaces, solubilises organic matter and hydrocarbons within the micelles, thus dramatically complicating an already complicate system. For these reasons at the current stage of knowledge, the accurate modelling of the effect of biosurfactants addition in bioremediation treatment is not possible and the feasibility has to be evaluated experimentally. However, the understanding of the natural roles of biosurfactants and the interaction with biosurfactants and the environment is crucial for our ability to forecast the effects of the addition of

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amphiphilic molecules, either of biological or chemical origin, on the biodegradation of pollutants. In the following paragraphs the current knowledge about these interactions is reviewed.

2.3.6 APPLICATIONS OF BIOSURFACTANTS IN BIOREMEDIATION TECHNOLOGY Hydrocarbons are hydrophobic compounds with low water solubility, thus microorganisms have developed several mechanisms to increase the bioavailability of these compounds in order to use them as carbon and energy source. Therefore one of the major factors limiting the degradation of hydrocarbons such as n alkanes is their low availability to the microbial cells. Microorganisms employ several strategies to enhance availability of those hydrophobic pollutants, such as biofilm formation and biosurfactant production (Christofi and Ivshina, 2002). In this sense, growth of microorganisms on oil hydrocarbons has often been related to their capacity of producing polymers with surfactant activity named biosurfactant. These biopolymers can either be low molecular weight polymers such as glycolipids (Guerra-Santos et al., 1986) and Lipopeptide or high molecular weight polymers such as emulsan (Zoumis, 2001), alas an (Kumari, 2012) or biodispersan (Rosenberg et al., 1999). Biodegradation of hydrocarbons in soil can be efficiently enhanced by addition or by in situ production of biosurfactants. It was generally observed (Kosaric, 2001) that the degradation time and particularly the adaptation time, for microbes are clearly shortened. It was clearly showen that rhamnolipids stimulate different processes involved in degradation of organic substrates (Norman et al., 2002). Bioemulsifiers have been often reported as enhancers of hydrocarbon biodegradation in liquid media, soil slurries and water and soil microcosms. In soil microcosms, treatment of waste crude oil with Halomonas bioemulsifiers produced a selective enhancing of indigenous hydrocarbon degrading bacteria suggesting the utility of bioemulsifiers as biostimulating agent of a bioremediation process.

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The application of biosurfactant on the remediation of metal, phenanthrene and polychlorinated biphenyl contaminated soil has recently been showed. For instance, a study on to determine the potential positive effect of novel biosurfactants on the enhancement of Aroclor 1248 metabolization in both in vitro and in situ experiments. Biosurfactants appeared to act very specifically, i.e. depending on strain and concentration used. It was due to the three-way interaction between biosurfactant, substrate (Zhang et al., 2000). Taking into account the specificity of surface- and biological activities of various biosurfactants, it was suggested that they might promote the mineralization of sorbed PCBs in polluted soils, when the optimized biosurfactant-degrader combination was used. A biosurfactantproducing Alcaligenes sp. strain MM-1, Arthrobacter sp. Strains EK1 and S-II from the North Sea was isolated while screening for oil-degrading marine microorganisms (190). It was demonstrated that the use of a mixture of Hydrocarbon-degrading microbes for bioaugmentation of soil contaminated with slop oil from a petrochemical industry resulted in the bioreclamation of soil, and showed an enhancement in bioremediation of polyaromatic hydrocarbons. In addition to these, it was also shown biosurfactant production by polycyclic aromatic hydrocarbon-metabolizing Pseudomonas aeruginosa 19SJ. Biosurfactants are also of great value to bioremediation of sites heavily contaminated with toxic heavy metals such as uranium, cadmium, lead and etc. These observations suggest the application of natural microbial consortia and their products was a feasible approach to resolve certain environmental related problems caused by mankind.

2.3.7 ROLE OF SURFACTANT IN BIODEGRADATION OF PAHS This work gives a review of surfactant effects on the biodegradation of solubilization of PAHs, a family of common and toxic pollutants that have raised significant environmental concerns. Surfactants have been proven to be important vehicles for the recovery of these compounds from contaminated soil or aquifers due to the solubilization process. Both positive and negative effects have 66 | P a g e

been reported on surfactants on microbial utilization of PAHs. The positive effects are generally attributable to the increased solubility/dissolution these compounds by surfactants which enhances their bioavailability. The negative effects are contributed by a variety of factors, which include toxicity of surfactants to microorganism, preferential degradation of surfactants and limited bioavailability of substrate solubilized in surfactant micelles. Nonionic surfactants are normally less toxic to microorganisms than ionic surfactants due to the weaker interactions between the neutral surfactant molecules and charged cell membrane. For a bioremediation application, solubilization efficiency is a prior criterion for the selection of a surfactant. However, its biodegradability and toxicity to the microorganism have to be considered to ensure an efficient remediation and the environmentally friendly application of the surfactant. Other important parameters of surfactant to be considered include its soil adsorption and cloud point. Surfactants with moderate biodegradability to the microorganisms should be considered. Firstly, sufficient solubilization capacity of such a type of surfactant can be maintained during a bioremediation process. Secondly, the reduction of effective surfactant concentration can increase the bioavailability of the substrate by releasing them into the aqueous phase. Thirdly, surfactants with a certain degree of biodegradation are more environmentally benign. Selection of surfactants that are nontoxic or with minimal toxicity to microorganisms is also essential to achieve a successful bioremediation. Biosurfactants are good alternative to synthetic commercial surfactants in term of low cytotoxicity. However, their application is limited due to their small scale production. To make the surfactant-mediated bioremediation a cost effective technique, efforts should be taken on the development of synthetic surfactants that biologically compatible with cells.

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2.4 BIOREACTORS Treatment of contaminated sediments in a slurry phase is often done in a bioreactor. Advantages of the slurry phase include: complete mixing of nutrients into sediment, increased contact between microorganism and contaminant, control (Eweis et al., 1998). Bioreactors are commercially available for this type of treatment but are expensive. Sediment is often put into the bioreactors in small batches; however, continuous flow operations are possible. With treatment, the slurry is mixed with nutrients and microbial cultures and aerated. The sediment then settles and the water is transferred to a treatment plant while the sediment is returned to a contained area (Eweis et al., 1998). For large quantities of sediment, this is a difficult option.

2.5 MICROBIAL GROWTH KINETICS The relation between the specific growth rate (µ) of a population of microorganisms and the substrate concentration (S) is a valuable tool in biotechnology. This relationship is represented by a set of empirically derived rate laws referred to as theoretical models. These models are nothing but mathematical expressions generated to describe the behaviour of a given system. The classical models, which have been applied to microbial population growth, include the Verhulst and Gompertz function (Okpokwasili and Nweke, 2005). The Gompertz function was originally formulated for actuarial science for fitting human mortality data but it has also been applied deterministically to organ growth. The Gompertz function is based on an exponential relationship between specific growth rate and population density. Equation 2.1 represents one of its parameterization. N(t) = Cexp{exp[-B(t - M)]} - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - 2.1 Where t = time,

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N(t) = population density at time t, C = upper asymptotic value, that is; the maximum population density, M = time at which the absolute growth rate is maximal, and B = relative growth rate at M. time. Gibson et al. (1987) modified the Gompertz function to a function which could be applied to the description of cell density versus time in bacterial growth curves in terms of exponential growth rates and lag phase duration (equation 2.2) Log N(t) = A + Dexp{ - B(t-M)]}

- - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - 2.2

Where N(t) = population density at time t, A = value of the lower asymptote (Log N(-t)), D = difference in value of the upper and lower asymptote [Log N(t) – log N(-t)], M = time at which the exponential growth rate is maximal. The idea of microbial growth kinetics has been dominated by an empirical model (equation 2.3) originally proposed by Monod (1942). The Monod model introduced the concept of a growth limiting substrate.

µ = µ𝑚𝑎𝑥

𝑆 𝐾𝑠 + 𝑆

− − − − − − − − − − − − − − − − − − − − − − − − 2.3

Where µ= specific growth rate, µmax = maximum specific growth rate, S = substrate concentration, Ks = substrate saturation constant (i.e. substrate concentration at half µmax).

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In Monod’s model, the growth rate is related to the concentration of a single growth-limiting substrate through the parameters µmax and Ks. In addition to this, Monod also related the yield coefficient (Yx/s) (equation 2.4) to the specific rate of biomass growth Y and the specific rate of substrate utilization (q) (equation 2.5).

𝑌𝑥 𝑑𝑥 = 𝑠 𝑑𝑠

µ=

− − − − − − − − − − − − − − − − − − − − − − − − − − 2.4

𝑌𝑥/𝑠 𝑑𝑠

𝑋

.

𝑑𝑡



𝑌𝑋

𝑠

𝑞

− − − − − − − − − − − − − − − − − − − − − − 2.5

2.6 KINETICS OF BIODEGRADATION The basic hypothesis of biodegradation kinetics is that substrates are consumed via catalyzed reactions carried out only by the organisms with the requisite enzymes. Therefore, rates of substrate degradation are generally proportional to the catalyst concentration (concentration of organisms able to degrade the substrate) and dependent on substrate concentration characteristic of saturation kinetics (e.g. Michaelis-Menten and Monod kinetics). Saturation kinetics suggests that at low substrate concentrations (relative to the half-saturation constant), rates are approximately proportional to substrate concentration (first order in substrate concentration), while at high substrate concentrations, rates are independent of substrate concentration (zeroorder in substrate concentration) (Azeez et al., 2012). In the case of substrates that contribute to the growth of the organisms, rates of substrate degradation are linked to rates of growth (i.e. the concentration of the biomass increases with substrate depletion). The mathematical analysis of such growth-linked systems is more complex than those situations where growth can be ignored. There are a number of situations where it may not be possible to quantify the concentration of substrate-degrading organisms in a heterogeneous microbial community. 70 | P a g e

However, the rate of substrate depletion can be measured. There are also situations in which the organism concentration remains essentially constant even as the substrate is degraded (i.e. no growth situation). Given these various features of biodegradation kinetics, different models including firstorder, zero-order, logistic, Monod (with and without growth) and logarithmic models can be used to describe biodegradation. Biodegradation kinetics is used to predict concentrations of chemical substances remaining at a given time during ex situ and in situ bioremediation processes. In most cases, information is based on loss of parent molecule targeted in the process (Azeez et al., 2012). The key interest is frequently the decrease in toxicity concentration. Nevertheless, toxicity measurements require bioassays, which are always very difficult and tedious. Therefore, efficacy of biodegradation is based on chemical measurements, e.g. disappearance of parent molecule, appearance of mineralization products or disappearance of other compounds used stoichiometrically during biodegradation of a compound, for instance, electron acceptors. There are several scenarios by which a compound can be transformed biologically. This includes when the compounds serve as: (1) Carbon and energy source (2) Electron acceptor (3) Source of other cell components. Other scenarios are the transformation of a compound by non-growing cells (the compound does not support growth) and the transformation of a compound by cometabolism, that is; transformation of a compound by cells growing on other substrate. The simplest case is where the compound serves as source of carbon and energy for the growth of a single bacterial species. The compound is assumed to be water-soluble, non-toxic and other substrates or growth factors are limiting. In the case of single-substrate limited process, the Monod equation (equations 2.6 and 2.7) is often used to describe microbial growth and biodegradation processes.

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µ=

.𝑞 =

µ𝑚𝑎𝑥 𝑆

- - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - - 2.6

𝐾𝑆 + 𝑆

𝑞𝑚𝑎𝑥 𝑆

− − − − − − − − − − − − − − − − − − − − − − − − − − − 2.7

𝐾𝑆 + 𝑆

Where µ = specific growth rate (1/X.dX/dt), q = specific substrate utilization/removal rate (1/X .dS/dt), Y = true growth yield [mass of biomass (X) synthesized per unit of substrate (S) utilized or removed, S = aqueous phase concentration of the compound, Ks = affinity constant or half saturation constant for the compound (meaning the concentration of compound when q is maximum). The hyperbolic equation proposed by Monod was modified by Lawrence and McCarty (1970) to describe the effects of substrate concentration (S) on the rate at which a given microbial concentration (X) removes the target substrate (-dS/dt) (equation 2.8). Alternatively, Monod equation can be written in terms of microbial growth by incorporating the net yield coefficient (Y) (equation 2.9).

. .

𝑑𝑆 𝑑𝑡 𝑑𝑋 𝑑𝑡

=−

𝑞𝑚𝑎𝑥 𝑆𝑋 𝐾𝑆 + 𝑆

= −𝑌

𝑑𝑆 𝑑𝑡

=

− − − − − − − − − − − − − − − − − − − − − − − − − 2.8 𝑌𝑞𝑚𝑎𝑥 𝑆𝑋 𝐾𝑆 + 𝑆

− − − − − − − − − − − − − − − − − − − − − − − 2.9

The Monod equation has frequently been simplified to an equation, which is either zero or first order in substrate concentration and the kinetics, has been widely used to describe biodegradation of organic contaminations in aquifer systems (Okpokwasili and Nweke, 2005). The versatility of Monod’s equation is attributed to its ability to describe biodegradation rates that follow zero- to first-order

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kinetics with respect to the concentration of the target substrate. Moreso, Monod’s model describes the dependence of biodegradation rate on the concentration of biomass.

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CHAPTER

3

MATERIALS AND METHODS This chapter is sub-divided into the following sections: 1. PAHs Identification Experiment 2. Bacteria Strain Isolation Experiment 3. Biodegradation Experiment 4. Soil Analysis

3.1

PAHs IDENTIFICATION EXPERIMENT

3.1.1 EQUIPMENT/MATERIALS Apparatus: the major equipments used in this study are: 1. Sonicator: this is an ultrasonic sound extractor of solvent. Produced by Thermo Fisher Scientific England. 2. Rotary evaporator: this is Rotavapor R-100, used in the evaporation of volatile liquids. Produced by Buchi Group, Switzerland. 3. Compressed air /clean-up cartages/glass wool (for cleaning up bed). Channeled through pipes and was used to concentrate extract to the required volume for injection into a Gas Chromatograph. 4. GC-FID: Agilent HP 6890 Gas Chromatography - Flame Ionization Detector Series. Used for the detection PAHs in the diesel contaminated soils. Produced by Hewlett-Packard, California United State.

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Reagents: All reagents used in this study were of highest purity and analytical grades. 1. Sodium sulfate anhydrous: Produced by Sigma-Aldrich used in the removal of residual water molecules in sample. 2. Cilical gel: Produced by Sigma-Aldrich used as an absorbent. 3. Dichloromethane: produced by JHG Chemicals, Guangzou, China. Used as an extraction solvent.

3.1.2 PROCEDURE a. Collection of Samples Sample was collected from diesel polluted soil around the Department of Chemical Engineering generator house, University of Benin. The contaminated soil sample was wrapped with aluminum foil paper, kept in dry ice inside a lagged bottle and immediately taken to “Earth Quest Laboratory”, Warri Delta State for analysis. b. PAHs Extraction 10g of the aged diesel polluted soil sample was weighed into a glass jar and immediately mixed with 20ml of dichloromethane (used as an extraction solvent) and placed into a Sonicator and allowed to sonicate for 1 hr. The extract was then immediately withdrawn with a pipet and injected into a treatment bottle containing 5g of sodium sulphate anhydrous (used for removal of residual water), 5g silica gel (used as an absorbent of residual water) and glass wool where the contaminant was cleaned with 20ml of dichloromethane. c. Methods of Analysis The withdrawn extract was concentrated to 1ml and then injected into Agilent 6890N Gas Chromatography - Flame Ionization Detector, where the PAHs were identified and quantified. Extracts

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peaks were the displayed on the screen. Detection time was studied to correspond to various PAHs present according to their volatilization capacities (Fagbote and Olanipekun 2013).

3.2

BACTERIA STRAIN ISOLATION

3.2.1 EQUIPMENTS/MATERIALS The materials and major equipments used for this study are: Auger tool, foil paper, Petri dish, distilled water, biochemical reagents (kovax reagents), gram staining reagents (crystal violet, safranine, lugor iodine and ethanol), Autoclave, microscope, laminar flow hood.

3.2.2 PROCEDURE i. Soil Sample Collection Soil samples were collected from diesel polluted soil around the Department of Chemical Engineering generator house, University of Benin from three different depths (top: 0cm, subsurface: 25cm and deep aquifer: 50cm and the homogenized) with the help of an Auger tool. Samples were transferred to sterile plastic bags and maintained in aseptic conditions. ii. Isolation of bacterial cultures Isolation of soil hydrocarbon degrading bacteria was performed by serial dilution method. 1gram of each sample was weighed into a Mcarthney (sample bottle) bottle and dissolved with 10ml distilled water.

One gram of soil sample was then serially diluted in sterilized distilled water to get a

concentration range from 10-1 to 10-3. Nutrient agar solution was prepared and autoclaved for 1 hour at 120oC. A volume of 0.1 ml of each dilution was transferred aseptically to starch agar plates (Petri dish). The sample was spread uniformly. The plates were incubated at 37°C for 24 hr. The bacterial isolates were further sub cultured to obtain pure culture. Pure isolates on nutrient agar slants were maintained at 4ºC in a refrigerator to avoid overgrowth. 76 | P a g e

iii. Identification of biosurfactant producing bacteria a. Cultural and morphological characterization The bacterial isolates were gram stained and observed under a high power magnifying lens in light microscope to obtain the colony morphology i.e. colour, shape, size, nature of colony and pigmentation. b. Biochemical characterization The bacterial isolates were characterized biochemically by Simmons citrate test, urease test, catalase test, oxidase test, glucose test, indole test, spore, motility, and MR.

3.3

BIODEGRADATION EXPERIMENT

3.3.1 CHEMICALS, REAGENTS AND EQUIPMENT All chemical and reagents used in this study were purchased from Pyrex Chemicals, Benin City, Nigeria. They were all used in the pure form as purchased at Pyrex Chemicals Ltd, Benin City, Edo State. 1. Pyrene and Anthracene: Pure analytical grade anthracene and pyrene produced by Merck Chemicals, USA. The choice of these two chemicals is based on the desire to degrade the most recalcitrant PAHs in a diesel polluted soil. 2. Ethanol: analytical grade ethanol with 99.5% purity produced by JHD Chemicals, Guangdong China. 3. Tween 80: This is polyoxyethylene sorbitan mono-oleate. A biodegradable non-ionic synthetic surfactant produced by Sigma-Aldrich, USA. It has an average molecular weight of 1310g and a critical micelle concentration (CMC) value of 15mg/l. 4. Rhamnolipid biosurfactant produced by Pseudomonas aeruginosa, Surfactin produced by Bacillus subtilis (Mazaheri-Assadi and Tabatabaee, 2010) 77 | P a g e

5. The analytical equipment used in this study was the UV- Vis Spectrometer. This is Hack DR/2010 Particle Data Logging Spectophotometer. Used mainly for the determination of concentrations of PAHs in bioreactors during degradation.

3.3.2 SOIL SAMPLE COLLECTION AND PREPARATION Surface and subsurface uncontaminated soils were obtained from Chemical Engineering Department. The soils were sun dried for 12 hours to reduce moisture content. Lumps of soils were then crushed and sieved through a 2.0 mm mesh, thereby roots and other unwanted materials were removed. The sieved soil was then autoclaved at 130oC for 1hr to kill all indigenous microbes. The pretreated soil was then analyzed to determine the organic matter, organic carbon, porosity, nitrogen, phosphorus contents, water holding capacity and pH by the method according by Mathieu and Pieltian (2003).

3.3.3 PROCEDURE 200g of the pretreated soil sample was separately placed into 20 plastic plates (bioreactors) labeled A1, B1, C1, D1, E1, F1, G1, H1, I1, J1 which was added 2g of anthracene each and A2, B2, C2, D2, E2, F2, G2, H2, I2, J2 which was added 2g of pyrene each. The plates (bioreactors) were added 3ml of microorganisms’ nutrient broth and 5ml of Tween 80 as in Table 3.1 and then covered with perforated lid to enable aerations. The experiment lasted 6 weeks at room temperature.

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Table 3.1: Treatment combination in bioreactors Contents Sample Plates

PAHs

Microorganism

Synthentic Surfactant

A1

Anthracene

Basillus substilis

Tween 80

B1 C1 D1 E1 F1 G1 H1 I1 J1(control)

Anthracene Anthracene Anthracene Anthracene Anthracene Anthracene Anthracene Anthracene Anthracene

Basillus substilis Pseudomonas aeruginosa Pseudomonas aeruginosa Staphilococus aurius Staphilococus aurius Micrococus letus Micrococus letus None None

None Tween 80 None Tween 80 None Tween 80 None Tween 80 None

A2 B2 C2 D2 E2 F2 G2 H2 I2 (control) J2

Pyrene Pyrene Pyrene Pyrene Pyrene Pyrene Pyrene Pyrene Pyrene Pyrene

Basillus substilis Basillus substilis Pseudomonas aeruginosa Pseudomonas aeruginosa Staphilococus aurius Staphilococus aurius Micrococus letus Micrococus letus None None

Tween 80 None Tween 80 None Tween 80 None Tween 80 None None Tween 80

a. PAHs extraction and estimation of THC Ten grams (10 g) of each soil sample in the bioreactors was weighed out every seven days interval and transferred into a sample bottle. Into this was added 10 ml of ethanol. The ethanol/soil mixture was shaken vigorously for two minutes with an orbital shaker and filtered into another sample bottle. The ethanol-oil extract was thereafter placed in cuvette wells and its absorbance was determined using Hack DR/2010 Particle Data Logging Spectophotometer.

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3.4

SOIL ANALYSIS

3.4.1 Determination of moisture content A constant weight of watch glass was obtained and thereafter, 20 g of sample was weighed into the watch glass, and transferred into the oven for 1 h at 110oC. The samples were cooled inside desiccators for 30 min before a constant weight of the sample and watch glass after heating and cooling was recorded. Moisture content was estimated as: % Moisture Content = [W1 - (W3-W2)] X100 W1 Where W1 = weight of sample; W2 = Constant weight of watch glass; and W3 = Weight of sample + watch glass after heating and cooling.

3.4.2 Determination of soil-pH and electrical conductivity (EC) Five grams (5.0 g) of each soil sample (in a sample cell) was added 50 ml of distilled water. The lump of the soil was stirred to form homogenous slurry, then pH meter (Jenway 3015 model) and EC meter (Jenway 4010 model) probes were immersed respectively into the sample and allowed to stabilize at 25oC and pH of sample was recorded.

3.4.3 Total organic carbon (TOC) and total organic matter (TOM) contents TOC and TOM were determined using a Muffle Furnace. 5.0g of soil sample was weighed into crucible and placed into a furnace for 1.0 hr at 400oC to calcine thereby burning off all the organic matter constituents. The weight of the calcined soil was determined and the difference before and after was then used to calculate the total organic matter content.

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Total organic carbon was similarly determined by carbonizing 5.0g of the soil in a furnace 3.0 hr at 700oC. The difference in weight was then used to estimate the total organic carbon.

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CHAPTER

4

RESULTS AND DISCUSSION 4.1

RESULT OF PAHs CONTAMINATION OF DIESEL POLLUTED SOIL

Figure 4.1 and Table 4.1 shows the total mass concentration of PAHs in the diesel polluted soil sample as 30.103mg/kg. This quantity of PAHs in the soil represents intermediate level of contamination on the polluted soil. The background concentration reveals that anthracene 7.127mg/kg and pyrene 6.983mg/kg which has least water solubility had the highest concentrations in the soil. These concentrations show that they were therefore most recalcitrants PAHs in the soil. Naphthalene and Benzo(g,h,i) perylene

with the highest water solubility were present in a highly negligible

concentrations. Figure 4.1 clearly shows the mass concentrations of PAHs in the diesel fuel contaminated soil. The result also showed that anthracene and pyrene constituted the largest group of compounds with high concentrations in the soil sample. However the total PAH concentrations were very high when compared with the maximum background limit of 15mg/kg in polluted soils set by Dutch Environment Ministry and USEPA (2002) respectively, (Polish Environment Ministry, 2002). Because of their persistent nature, anthracene and pyrene are listed individually among the USEPA’s seven priority chemicals.

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contaminated sample (mg/kg) 8

PAHs Conc. mg/kg

6 contaminated sample (mg/kg)

4 2 0

Figure 4.1: mass concentration of PAHs in diesel polluted soil in Chemical Engineering Department, University of Benin. FID1 A, (TT010515\SIG10029.D)

700

pyrene

anthracene

pA

PAH profile of contaminated sample 600

fluoranthene

500

naphthalene

100

5

10

15

20

25

Indeno(1,2,3-cd)pyrene

200

chryseDne-pybrenze(aI)Ianthracene

acenaphthylDen-epyrenaeceInaphthene fluorene

300

benbzeon(kz)ofl(ubo)rflaunotrhaenntheene benzo(a)pyrene D-pyreneIV

phenanthrene

400

30

35

40

min

Figure 4.2: GC-FID chromatogram of PAHs in an aged diesel polluted soil. Figure 4.2 shows the chromatogram of the GC-FID result. The peak time for each component was observed to be different from others. This was related to the flash points of the component as 83 | P a g e

naphthalene which was the most volatile of all components was observed to produce the first peak just before 10 minutes after sample was injected into the GC column. Pyrene produced a peak about 20 minutes and Indeno(1,2,3) perylene produced the last peak after sample had been retained in the column for 33 minutes.

4.2

BIODEGRADATION ANALYSIS

The estimated mass concentrations of anthracene and pyrene were respectively plotted against time to produce the curve of Figure 4.3 and Figure 4.4. It was observed from Table 4.4 and Table 4.5 that the initial concentration at the start of the degradation study was varied in all the reactors. This was attributed to the spiking of the soil 7 days before the first reading was taken at day 0. This was done to enable efficient dissolution and sequestrations of the PAHs into the soil matrices because both anthracene and pyrene were poorly soluble in solvents. During this period, the various strains of the microbes would have started the mineralization of the bioavailable hydrocarbons in the reactors because PAHs are utilized only in the dissolved state.

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0.58 0.57 A1 Conc.

Mass Concentration g/kg

0.56

Control

0.55

B1 Conc. C1 Conc. D1 Conc.

0.54

E1 Conc. F1 Conc.

0.53

G1 Conc.

0.52

H1 Conc. I1 Conc.

0.51 B1

J1 Conc.

0.5 0

5

10

15

20

25

30

35

40

45

TIME (DAYS)

Figure 4.3: Anthracene degradation curve.

It was observed from Figure 4.3 that the degradations process was consistently very fast till about day 30 in B1, D1, F1 and H1. This curves which were produced in reactors containing only strain of microbe without synthetic surfactant (Tween 80) was a clear indication of by biosurfactant efficiencies in relation to the availability of hydrocarbons for degradation. The curves were seen to start turning towards the horizontal direction from day 34 indicating that substrate (hydrocarbon) were no longer bioavailable after continuous mineralization by the microbe. Curves A1, C1, G1 and E1 were obtained from reactors which contained both biosurfactant producing microbes and synthetic surfactant (Tween 80) and shows that the degradation of anthracene became slow from day 26. This is related to the activity of the synthetic surfactant. The reduced bioavailability of hydrocarbon molecule may have been caused by synthetic surfactant ability to form micelles which would as well engulf hydrocarbons molecules within these micelles thereby making the engulfed molecules not to be bioavailable for

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degradation by the microbes. This effect can also be the cause of the reduction in the overall amount of the hydrocarbon degraded in all reactor containing both synthetic and biosurfactants when compared with reactor containing biosurfactant producing microbes alone. It was observed also that the highest total concentration depletion occurred in the degradation of anthracene in reactor B1 which contain Basillus substilis without synthetic surfactant (Tween 80) when compared with the degradations in other reactors. Reactor D1 which contain Pseudomonas aeruginosa was observed to be the second highest in overall degradation efficiency when compared to reactors containing Staphilococus aurius and Micrococus letus. Reactors I1/I2 and J1/J2 were used as control reactors. I1/12 contain PAH contaminated with synthetic surfactant alone while J1/J2 contained PAH without both synthetic and biosurfactant producing microbes. Though, there was a very small depletion in their concentrations. This reduction may have resulted from natural attenuation caused by residual microbes which could result from contaminations from air and water.

Mass Concentration g/kg

0.59 0.57

A2 Conc. B2 Conc.

0.55

C2 Conc. 0.53

D2 Conc. E2 Conc.

0.51

F2 Conc. G2 Conc.

0.49

H2 Conc. 0.47

I2 Conc. J2 Conc.

0.45 0

10

20

30

40

50

TIME (DAYS)

Figure 4.4: Pyrene degradation curve 86 | P a g e

Figure 4.4 shows the degradation of pyrene by all four microbes. Reactors A2, C2, E2 and G2 which contained pyrene with synthetic surfactant and biosurfactant producing microbes showed a slower degradation process when compared with reactors B2, D2, F2 and H2 which contained pyrene with biosurfactant producing microbes alone. As in the degradation of anthracene (Figure 4.3), reactor B2 which contained Basilus substilis without tween 80 was found to produce the highest degradation when compared with other reactors. Detailed analysis of the degradation results showed that the quantity of anthracene degraded in the bioreactors are 0.0518mg/g in A1, 0.0557mg/g in B1, 0.0467mg/g in C1, 0.0527mg/g in D1, 0.0476mg/g in E1, 0.0476mg/g in F1, 0.0467mg/g in G1, 0.0487mg/g in H1 and in the control reactors as 0.0202mg/g and 0.0111mg/g in I1 and J1 respectively. The two control reactors shows the degradability potentials of all the microbes and also, the role of synthetic surfactant alone on sorption of pyrene into the soil matrices thereby making it non-bioavailable. The quantity of pyrene degraded also revealed in the bioreactors to be 0.1028mg/g in A2, 0.1099mg/g in B2, 0.0887mg/g in C2, 0.0939mg/g in D2, 0.0781mg/g in E2, 0.0851mg/g in F2, 0.0922mg/g in G2, 0.0975mg/g in H2 and in the control reactors as 0.0354mg/g and 0.0284mg/g in I1 and J2 respectively. The overall degradation of pyrene with 0.812mg/g was higher than anthracene with 0.429mg/g degradation in all bioreactors. This was reasoned to have been caused by the better solubility potential of pyrene. These results suggest that in both cases, PAH degradation was related to the type and nature of PAH present in the contamination and the type of biosurfactant producing organisms. By increasing the solubility of the hydrophobic substrates it facilitated their transport to the microbial cells and enhanced their metabolism.

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Basillus substilis and Pseudomonas aeruginosa which respectively produced surfactin and rhamnolipid biosurfactants therefore are the most active biosurfactants producers for PAHs degradation operations.

4.3

ISOLATION AND IDENTIFICATION OF BIOSURFACTANT PRODUCING BACTERIA

Selection of bacteria was based on ability to grow in the presence of petroleum diesel as sole source of carbon. Five colonies of the bacteria which showed higher growth rates were selected for further characterization and identification. Selected isolates were characterized by colony morphology and biochemical characteristics according to Cappuccino and Sherman (2010). Basillus substillis, Micrococus leutus, Pseudomonas aeruginosa and Staphilococus aurius strain were isolated, identified and selected because they could grow in the presence of petroleum hydrocarbon such as anthracene and pyrene as sources of carbon and energy. Bacillus substilis was observed to occur in all three depths of the soil because of their spore forming ability; the spores have the ability to survive harsh conditions in soil environment. Bacillus subtilis produces surfactin and lipopeptides surfactants, while Pseudomonas aeruginosa produces glycolipids. P. aeruginosa also showed a rapid growth in the culture media but was only observed to be present in the top soil (0.0 cm depth) because it is an aerobic organism. The high incident rate of Bacillus sp. and Pseudomonas sp. in diesel contaminated soil when compared to other organisms isolated may be a direct correlation to their ability to produce surfactants to degrade hydrocarbons. Peudomonas aeruginosa is a gram negative, rod shaped, motile bacteria and its biochemical characters are shown in appendix A Table A1. Staphylococcus Aureus was equally found only in the top soil while E. coli, and Micrococcus Letus were found in the top and sub-surface of the diesel contaminated soil.

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4.3.1 CULTURAL AND MORPHOLOGICAL CHARACTERISTICS The colonies were culturally and morphologically identified with the aid of a high power magnifying lens in light microscope after gram staining and the shape, margin, elevation, size, colour, cell arrangement, cell type and colony surface and gram reactions resulted in the suspicion of E. coli, Micrococcus Letus, Staphylococcus Aureus, Basillus substilis and Pseudomonas aeruginosa was made. Results of these experiments are as shown in Appendix A Table A1.

4.3.2 Biochemical identification of hydrocarbon degrading bacteria The morphologically identified bacterial strains were confirmed with biochemical tests by performing Catalase test, coagulase test, oxidase test, urease test, glucose test, citrate test, and indole test and the experiments further confirmed the presence of B. substilis, M. letus, E. coli, and P. aeruginosa in the diesel contaminated soil samples. Result of this experiments are as shown in appensix A Table A2

4.4

SOIL CHARACTERISTICS

Table 4.1 shows the physico-chemical parameters of the soil used for the biodegradation experiment. Nitrogen, phosphorus, organic matter and organic carbon were observed to be affected by the degradation process. Particle size analysis was carried out using the hydrometer method.

Table 4.6: Physical-chemical characteristics of the soil used for experimentation Ph Electrical conductivity (EC) (dS/m) Soil texture (Particle Size Distribution) Nitrogen concentration (%) Total organic matter (TOM) (mg/kg) Total organic carbon (TOC) (mg/kg) Porosity (%) Water holding capacity (%)

7.5 2.29 – 2.04 7% Clay, 10.8% Silt and 75% Sand 0.10 – 0.08 1.28 – 1.2 4.4 – 4.1 35 1.3

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Temperature readings ranged from 29oC to 30.8oC for control soil, 29oC to 31.4oC for pyrene bioreactors, 30.9oC to 31.8oC for anthracene bioreactors.

4.4.1 Electrical conductivity (EC): Electrical conductivity (EC) is a measure of ionic concentration in the soils and is therefore related to dissolve solutes. EC was significantly lower in the PAHs contaminated soils than in the control soils at the end of the degradation process (Table 4.1). It is not likely that the degraded PAHs were directly responsible for the observed changes in EC since organic compounds like crude oil cannot conduct electrical current very well. However, it is possible that the anoxic biodegradation mechanism through direct dehydrogenation allowed the anaerobic metabolism of hydrocarbons in the presence of an electron acceptor such as nitrate ion, which may be partially responsible for the observed differences in EC.

4.4.2 Total organic carbon (TOC) and total organic matter (TOM) contents: Total organic carbon and total organic matter contents were slightly lower than the 4.4 mg/kg obtained for the control soils. Organic matter content should normally increase following the addition of such levels of carbonaceous substances but results obtained herein show that there is rather a reduction in organic carbon and organic matter contents of the polluted soils (Table 4.1). The most plausible reason perhaps might be that the surfactant-oil mixture impaired the metabolic processes that would have facilitated the addition of organic carbon from the petroleum hydrocarbons by reducing the carbon-mineralizing capacity of the microorganisms. Thus, two decomposition processes are of significance to the present discussion: the decomposition of the soil organic matter and the decomposition of the added petroleum hydrocarbons. Both decomposition processes are however the prerogative of heterotrophic organisms. It is most likely that while these organisms might have been stimulated by the presence of the spilledoil on site, their proliferation did not adequately cope with the business of breaking down the excess 90 | P a g e

carbonaceous substrate, perhaps due to various factors that might include the environmental conditions of weathering and climatic predispositions as well as the physico-chemical properties earlier discussed.

4.4.3 Nitrate-nitrogen (NO3-N) content: The reduction in the concentration of NO-N in the contaminated PAHs polluted soils suggests that the process of nitrification might have reduced PAHs contamination. hydrocarbon- utilizing microbes such as Azobacter spp normally become more abundant while nitrifying bacteria such as Nitrosomonas spp become reduced in number. This probably explains the relatively lower values of NO3-N obtained after degradation.

4.4.4 Soil-pH: The pH of the contaminated soils in the bioreactors was significantly lower than the control soils at the end of the degradation process. The bioreactors containing pahs and tween 80 had a much reduced pH than that without tween 80. Pahs and Tween 80 must have discouraged the leaching of basic salts which are responsible for raising pH in the control. The binding of the pahs with soil particulate matter in the soil probably posed a major resistance to the removal of such basic ions.

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CHAPTER

5

CONCLUSION AND RECOMMENDATION 5.1

CONCLUSION

It is evident from this study that the surrounding soils of Chemical Engineering Department generator house is contaminated with PAHs arising from the generator fumes, diesel spill and engine oil spill. Therefore, the following conclusions can be drawn from the overall study: 1. Anthracene and Pyrene were the most recalcitrant polycyclic aromatic hydrocarbons in aged diesel fuel polluted soil. 2. Bioavailability depends on the solubilization potential of PAHs and not molecularity because, Pyrene with four benzene rings was observed to be more soluble in solvents than anthracene with three rings and therefore was more mineralized in all bioreactors compared to Anthracene. 3. Bacillus substilis and Pseudomonas aeruginosa are potent biosurfactant producing microorganisms capable of solubilizing very lowly soluble PAHs such as anthracene and pyrene thereby increasing their bioavailability for mineralization. 4. This study further confirms that Pseudomonas and Bacillus are potential biosurfactants producers and useful tools for bioremediation processes. 5. Combination of biosurfactants producing microbes with synthetic surfactant in degradation process is less active compared to biosurfactant operation alone.

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5.2

RECOMMENDATION

Whenever surfactants are added to soil, some of the surfactant will sorb to soil constituents and therefore will be removed from the aqueous phase. This removal will in turn influence the amount of surfactant required to reach the critical micelle concentration (cmc) in the aqueous phase, which must be reached because solubilization properties of surfactants become significant only at surfactant concentrations above the cmc. In view of these, the following recommendations are drawn: 1. Hydrophobic surfactants (lower HLB value) should be used in degradation as they sorb to a greater extent than hydrophilic surfactants. However, since lower HLB surfactants also will solubilize more of the hydrophobic pollutants per unit of surfactant added, there will be a trade-off in selecting a surfactant with an optimum HLB value for a given system. 2. The higher the final interfacial tension, the lower the probability that mobilization will occur. It is important to consider surfactant sorption in estimating the minimum amount of surfactant required to achieve mobilization in a given soil system, it also is important not to overdose the surfactant.

3. As shown in this project, it is possible to reduce biodegradation rates of PAH in the presence of surfactants by sequestering most of the PAH in micelles. We have not yet developed a complete quantitative model of the effects of surfactant concentration on overall biodegradation rates of hydrophobic compounds.

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Shahid GH (2007). Potential hazards of gasoline additives in altering soil environment in favour of harmful microorganisms. Int. J. Environ. Res. Public Health, 1(1): 1-4. USEPA. (2007): Pressurized Fluid Extraction (PFE). Method 3545A. Ukiwe N. Luke, Ubaezue U. Egereonu, Njoku C. Pascal, Nwoko I. A. Christopher and Allinor I. Jude (2013): Polycyclic Aromatic Hydrocarbons Degradation Techniques: A Review. International Journal of Chemistry; Vol. 5, No. 4; 2013 ISSN 1916-9698 Vasilyeva G. K. and Strijakova E.R. (2007): Bioremediation of Soils and Sediments Contaminated by Polychlorinated Biphenyls. Microbiology; 76; 6; pp 639–653. Wick F. Abbey, Haus W. Nicholas, Sukkariyah F. Beshr, Haering C. Kathryn and Lee W. Daniels (2011): Remediation of PAH-Contaminated Soils and Sediments: A Literature Review. Virginia Tech, Environmental Soil Science, Wetland Restoration and Mined Land Reclamation. Pp 57 – 59. Wild S.R. and Jones, K.C. (1995): Polynuclear aromatic hydrocarbons in the United Kingdom environment: a preliminary source inventory and budget. Environ. Pollut., 88(1):91-108. Wyszkowska J. and Kucharski J. (2000): Biochemical Properties of soil contaminated by petrol. Polish J. Environ. Stud., 9(6): 476-485. Wyszkowski M. and Ziolkowska A. (2008): Effect of Petrol and Diesel oil on content of organic carbon and mineral components in soil. Am-Eur. J. Sust. Agric., 2(1): 54-60. Zhang Y., Wu R. S. S., Hong H. S., Poon K. F. and Lam M. H. W. (2000): Field study on the desorption rates of polynuclear aromatic hydrocarbons from contaminated marine sediment Environ Toxicol Chem; 19(10):5; pp 24-31. Ziqing O. U. (2000): Separate and Combined Environmental Behaviour of Surfactants and Polycyclic Aromatic Hydrocarbons (PAHs) ZoumisT., Schmidt A., Grigorova L. and Calmano W. (2001): Contaminants in sediments: remobilisation and demobilisation. Sci Total Environ; 266; pp 195– 202.

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APPENDIX

A

A. BACTERIA IDENTIFICATION

TABLE A.1: CULTURAL AND MORPHOLOGICAL CHARACTERISTICS Isolates

Shape

Margin

Elevation

Size

Colour

Colony surface

Gram reaction

Cell type

A

Round

Rough

Flat

1mm

Green

Moist

-ve

Rod

Cell arrang ement Singly

B

Round

Smooth

Raised

1mm

Moist

+ve

Cocci

Cluster

C D

Round Round

Smooth Rough

Raised Raised

0.5mm 0.5mm

Golden yellow Pink Yellow

Moist Dry

-ve +ve

Rod Cocci

Chain Cluster

E

Round

Rough

Raised

1mm

Cream

Dry

+ve

Rod

Chain

F

Round

Smooth

Raised

1mm

White

Dry

+ve

Cocci

Cluster

Suspected bacteria Pseudomonas aeruginosa Staphylococcu s aureus E. coli Micrococcus Letus Bacillus subtilis Staphylococcu s Epididymis

TABLE A.2: Biochemical Characteristics Isolates

Catalase

Coagulase

Oxidase

Urease

Glucose

Citrate

Motility

MR

Spore

Indole

Suspected Bacteria Pseudomonas aeruginosa Staphylococcus Aureus E. coli Micrococcus Letus

A

-ve

-ve

+ve

-ve

+ve

+ve

+ve

-ve

-ve

-ve

B

+ve

+ve

-ve

-ve

-ve

+ve

-ve

-ve

-ve

-ve

C D

+ve +ve

-ve -ve

-ve +ve

-ve -ve

+ve -ve

+ve +ve

+ve -ve

+ve -ve

-ve -ve

+ve -ve

E

+ve

-ve

+ve

+ve

-ve

+ve

-ve

-ve

+ve

-ve

Bacillus Subtilis

F

+ve

-ve

+ve

+ve

-ve

+ve

-ve

-ve

-ve

-ve

Staphylococcus Epididymis

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APPENDIX

B

A. CALIBRATION CURVE AND APPLICATION OF BEER'S LAW Three standard solutions were prepared separately with pyrene in ethanol and anthracene in ethanol (Table B1). The absorbance of the standard solutions was determined by UV-Vis spectrometer. The calibration curve was used to relate the absorbance of the standard solutions to their known concentrations. Table B.1: absorbance-concentration for standard solution of Pyrene and anthracene PAH

Concentration 0.1mol/l 0.5mol/l 1.0mol/l 0.1mol/l 0.5mol/l 1.0mol/l

Pyrene

Anthracene

Absorbance 0.382 0.761 1.510 0.638 0.237 2.610

Pyrene Absorbance/Concentration Curve

Anthracene Absorbance/Concentration Curve 3 Absorbance

1.5 Abs= 0.986Cconc - 0.477 R² = 0.951

1 0.5

0 0

2

4

Concentration mol/dm3

Absorbance

Absorbance

2.5 2

Mass of solute in solvent 0.10g in 5ml ethanol 0.51g in 5ml ethanol 1.01g in 5ml ethanol 0.09g in 5ml ethanol 0.45g in 5ml ethanol 0.89g in 5ml ethanol

1.6 1.4 1.2 1 0.8 0.6 0.4 0.2 0

Absorban… Abs = 0.564Cconc - 0.243 R² = 0.965

0

2

4

Concentration mol/dm3

Figure B: pyrene and anthracene calibration curve

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APPENDIX

C

ANTHRACENE AND PYRENE ABSORBANCES AND CONCENTRATION Table 4.2 and Table 4.3 show the absorbance of anthracene and pyrene respectively as determined by uv-spectrometer. The concentrations were estimated from the absorbance using the linear equations of the calibration curves for anthracene and pyrene with the equations Abs= 0.986Cconc - 0.477 and Abs = 0.564Cconc - 0.243 respectively (appendix A). Table 4.4 and Table 4.5 show the result of the estimated mass concentrations of anthracene and pyrene respectively, from the absorbance in accordance with Bear Lamberts law during the 42 days of degradations.

Table 4.2: UV Spectrometer absorbance of anthracene SAMPLES DAYS 0 7 14 21 28 35 42

A1 Abs. 0.077 0.066 0.053 0.045 0.033 0.028 0.026

B1 Abs. 0.075 0.064 0.052 0.041 0.03 0.022 0.020

C1 Abs. 0.076 0.065 0.053 0.047 0.040 0.032 0.030

D1 Abs. 0.080 0.064 0.052 0.046 0.036 0.030 0.028

E1 Abs. 0.082 0.068 0.061 0.050 0.044 0.037 0.035

F1 Abs 0.081 0.067 0.060 0.051 0.045 0.036 0.034

G1 Abs. 0.079 0.068 0.062 0.050 0.043 0.034 0.033

F2 Abs 0.079 0.064 0.057 0.044 0.035 0.032 0.031

G2 Abs 0.079 0.062 0.054 0.047 0.034 0.030 0.027

H1 Abs. 0.078 0.066 0.061 0.050 0.041 0.031 0.03

I1 Abs. 0.082 0.077 0.073 0.069 0.065 0.063 0.062

J1 Abs. (Control) 0.085 0.084 0.082 0.079 0.078 0.075 0.074

Table 4.3: UV Spectrometer absorbance of pyrene SAMPLES DAYS 0 7 14 21 28 35 42

A2 Abs 0.080 0.059 0.047 0.032 0.027 0.025 0.022

B2 Abs 0.081 0.056 0.046 0.034 0.025 0.022 0.019

C2 Abs 0.077 0.067 0.057 0.042 0.037 0.029 0.027

D2 Abs 0.078 0.062 0.055 0.048 0.041 0.029 0.025

E2 Abs 0.077 0.069 0.057 0.047 0.036 0.034 0.033

H2 Abs 0.078 0.060 0.052 0.046 0.032 0.024 0.023

I2 Abs 0.083 0.081 0.078 0.073 0.069 0.065 0.063

J2 Abs (Control) 0.085 0.082 0.079 0.077 0.075 0.074 0.069

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Table 4.4: Mass concentration of anthracene in different reactor combinations SAMPLES TIME(DAYS) 0 7 14 21 28 35 42

A1 Conc. 0.5619 0.5509 0.5375 0.5294 0.5172 0.5121 0.5101

B1 Conc. 0.5598 0.5487 0.5365 0.5254 0.5142 0.5061 0.5041

C1 Conc. 0.5609 0.5497 0.5375 0.5314 0.5243 0.5162 0.5142

D1 Conc. 0.5649 0.5487 0.5365 0.5304 0.5203 0.5142 0.5122

E1 Conc. 0.5669 0.5527 0.5456 0.5345 0.5284 0.5213 0.5193

F1 Conc. 0.5659 0.5517 0.5446 0.5355 0.5294 0.5203 0.5183

G1 Conc. 0.5639 0.5527 0.5467 0.5345 0.5274 0.5183 0.5172

H1 Conc. 0.5629 0.5507 0.5456 0.5345 0.5254 0.5152 0.5142

I1 Conc. 0.5669 0.5619 0.5578 0.5538 0.5497 0.5477 0.5467

J1 Conc. 0.5699 0.5689 0.5669 0.5639 0.5628 0.5598 0.5588

Table 4.5: Mass concentration of pyrene in different reactor combinations

SAMPLES TIME(DAYS) 0 7 14 21 28 35 42

A2 Conc. 0.5727 0.5355 0.5142 0.4876 0.4787 0.4752 0.4699

B2 Conc. 0.5745 0.5301 0.5124 0.4911 0.4752 0.4699 0.4646

C2 Conc. 0.5674 0.5496 0.5319 0.5053 0.4964 0.4823 0.4787

D2 Conc. 0.5691 0.5408 0.5284 0.5159 0.5035 0.4823 0.4752

E2 Conc. 0.5674 0.5532 0.5319 0.5142 0.4947 0.4911 0.4893

F2 Conc. 0.5709 0.5443 0.5319 0.5089 0.4929 0.4876 0.4858

G2 Conc. 0.5709 0.5408 0.5266 0.5142 0.4911 0.4840 0.4787

H2 Conc. 0.5691 0.5372 0.5230 0.5124 0.4876 0.4734 0.4716

I2 Conc. 0.5780 0.5745 0.5691 0.5603 0.5532 0.5461 0.5426

J2 Conc. 0.5816 0.5762 0.5709 0.5674 0.5638 0.5621 0.5532

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APPENDIX

D

Table 4.1: PAHs Concentration in Diesel Contaminated Soil PARAMATER Naphthalene Acenaphthylene Acenaphthene Florene Phenathrene Anthracene Fluoranthene Pyrene Benzo(a)anthracene Crysene Benzo(b)fluoranthrene Benzo(a)pyrene Benzo(k)fluoranthrene Indeno(1,2,3) perylene Benzo(g,h,i) perylene TOTAL PAHs (mg/kg)

METHOD USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B USEPA 8015B

contaminated sample (mg/kg) 0.085 0.624 1.125 2.754 1.924 7.127 2.392 6.983 2.337 2.660 0.835 0.210 0.548 0.498 0.000 30.103

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