Microbial Community Changes in Methyl Tert-Butyl Ether (MTBE)- Contaminated Soils

University of Tennessee, Knoxville TRACE: Tennessee Research and Creative Exchange

Masters Theses Graduate School

8-2004

Microbial Community Changes in Methyl tert-Butyl Ether (MTBE)- Contaminated Soils

Bo Yang University of Tennessee - Knoxville

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Recommended Citation Yang, Bo, "Microbial Community Changes in Methyl tert-Butyl Ether (MTBE)-Contaminated Soils. " Master's Thesis, University of Tennessee, 2004. https://trace.tennessee.edu/utk_gradthes/2223

This Thesis is brought to you for free and open access by the Graduate School at TRACE: Tennessee Research and Creative Exchange. It has been accepted for inclusion in Masters Theses by an authorized administrator of TRACE: Tennessee Research and Creative Exchange. For more information, please contact [email protected]. To the Graduate Council:

I am submitting herewith a thesis written by Bo Yang entitled "Microbial Community Changes in Methyl tert-Butyl Ether (MTBE)-Contaminated Soils." I have examined the final electronic copy of this thesis for form and content and recommend that it be accepted in partial fulfillment of the requirements for the degree of Master of Science, with a major in Environmental Engineering.

Kung-Hui Chu, Major Professor

We have read this thesis and recommend its acceptance:

Chris D. Cox, Paul D. Frymier

Accepted for the Council: Carolyn R. Hodges

Vice Provost and Dean of the Graduate School

(Original signatures are on file with official studentecor r ds.) To the Graduate Council:

I am submitting herewith a thesis written by Bo Yang entitled “Microbial Community Changes in Methyl tert-Butyl Ether (MTBE)-Contaminated Soils.” I have examined the final electronic copy of this thesis for form and content and recommend that it be accepted in partial fulfillment of the requirements for the degree of Master of Science, with a major in Environmental Engineering.

Kung-Hui Chu _ Major Professor

We have read this thesis and recommend its acceptance:

Chris D. Cox _

Paul D. Frymier _

Accepted for the Council:

Anne Mayhew _ Vice Chancellor and Dean of Graduate Studies

(Original signatures are on file with official student records.)

Microbial Community Changes in Methyl tert-Butyl Ether (MTBE)-

Contaminated Soils

A Thesis

Presented for the

Master of Science

Degree

The University of Tennessee, Knoxville

Bo Yang

August 2004

ACKNOWLEDGEMENTS

I would like to thank the following individuals:

z Dr. Kung-Hui Chu for her advice and guidance.

z Dr. Paul Frymier and Dr. Chris Cox for their helpful suggestions.

z Dr. Chang-Ping Yu for his helpful suggestions and technical assistance.

ii Abstract

This study addresses two questions regarding the microbial community structure shift in MTBE-contaminated soils: Are known MTBE-degradaers/utilizers present in MTBE- contaminated soils? How is the microbial community affected by the addition of oxygen and the addition of cometabolic substrate?

A new quantitative fingerprinting method, Real-time-t-RFLP, was applied to study the effects of aeration and cosubstrate (toluene) on the microbial community structure of

MTBE-contaminated soils. Real-time-t-RFLP, developed from two novel molecular methods (Real-time polymerase chain reaction (PCR) and terminal restriction fragment length polymorphism (t-RFLP)), is capable of quantifying each ribotype in a microbial community. Results of this study showed that microbial community shifted dramatically after the addition of MTBE, oxygen, and toluene. Results indicated that there might be some known MTBE-degraders/utilizers present in the MTBE-contaminated soils. And cloning of the PCR products is needed to confirm their positive identifications in the future study.

iii Contents

Page

1 Introduction 1

2 Background 4

2.1 Properties of MTBE 4

2.2 MTBE environmental occurrences 6

2.1.1 MTBE in surface water, groundwater, and drinking water 6

2.2.2 Effects of MTBE on human health 9

2.2.3 Actions against MTBE 10

2.3 Biodegradation of MTBE 11

2.3.1 Aerobic degradation pathways of MTBE 12

2.3.2 Microorganisms using MTBE as a sole carbon and energy source 16

2.3.2.1 Mixed cultures using MTBE as a sole carbon and energy

source 16

2.3.2.2 Pure cultures using MTBE as a sole carbon and energy source 18

2.3.3 Microorganisms using MTBE as a cometabolic substrate 21

2.4 In Situ bioremediation of MTBE 22

3 Materials and Methods 24

iv 3.1 Materials 24

3.1.1 Chemicals 24

3.1.2 Soil microcosms 24

3.2 Experimental approaches 25

3.2.1 DNA extraction from soil samples 27

3.2.2 Real-time-t-RFLP assay 29

Step 1 Real-time PCR amplification of 16S rDNA 29

Step 2 Gel purification of the PCR products 31

Step 3 Restriction digestion of the PCR products 32

Step 4 Terminal restriction fragment length polymorphism (t-RFLP)

analysis 33

Step 5 Data analysis 34

4 Results and Discussion 36

4.1 Microbial community of MTBE-contaminated soils at different depths 36

4.2 Changes of microbial community in MTBE-contaminated soils due to the

aerobic MTBE enrichment and a cosubstrate 44

5 Conclusions and Future Studies 62

5.1 Conclusions 62

5.2 Future studies 63

List of References 65

Appendix 72

Vita 75

vi List of Tables

Table 2.1 Basic physical and chemical properties of MTBE and BTEXs. 5

Table 2.2 Aerobic MTBE degradation characteristics of pure cultures. 13

Table 4.1 Characteristics of soils at four depths. 36

Table 4.2 Calculated 16S rDNA copies with terminal restriction fragment (T-RF) bases of original soil samples. 38

Table 4.3 Characteristics of known MTBE-utilizers. 42

Table 4.4 Calculated 16S rDNA copies with terminal restriction fragment (T-RF) bases of soil samples (SET I). 51

Table 4.5 Calculated 16S rDNA copies with terminal restriction fragment (T-RF) bases of soil samples (SET II). 53

Table 4.6 Characteristics of MTBE-degraders (cometabolic degradation). 57

vii List of figures

Figure 2.1 Proposed pathways for aerobic MTBE biodegradation. 15

Figure 3.1 Flow chart showing steps taken in this research. 26

Figure 3.2 Flow chart showing steps taken in Real-time-t-RFLP assay. 30

Figure 4.1a Microbial community profiles of original soils based on t-RFLP analysis. 39

Figure 4.1b Microbial communities of original soils analyzed by Real-time-t-RFLP. 40

Figure 4.2 Microbial community profiles of original soil based on TAP analysis. 43

Figure 4.3 MTBE concentrations in soil microcosms. 45

Figure 4.4a Microbial community profile (SET I) based on t-RFLP analysis. 47

Figure 4.4b Microbial community (SET I) analyzed by Real-time-t-RFLP. 48

Figure 4.5a Microbial community profile (SET II) based on t-RFLP analysis. 49

Figure 4.5b Microbial community (SET II) analyzed by Real-time-t-RFLP. 50

Figure 4.6 Effects of aerobic MTBE enrichment on microbial communities (SET I) 54

viii Figure 4.7 Effects of aerobic MTBE enrichment on microbial communities (SET II) 55

Figure 4.8 Effects of toluene on microbial communities (SET I) 58

Figure 4.9 Effects of toluene on microbial communities (SET II) 60

Figure A.1 Linear regression of standard curve. 74

Chapter 1

Introduction

Methyl tert-butyl ether (MTBE) has been used in US gasoline to replace lead and other toxic substances as an octane enhancer since the late of 1970’s. In the 1990’s, the Clean Air Act (CAA) Amendments (U.S. Environmental Protection Agency. 1999) required that oxygenates be used in US gasoline to reduce air pollution. Since then,

MTBE has been more widely used in the gasoline as oxygenates to reduce vehicle emissions, such as air carbon monoxide and volatile organic compounds. Due to its low cost and favorable transfer and blending characteristics compared to other gasoline oxygenate additives such as ethanol, MTBE has become the most commonly used oxygenate, and is currently added at concentration of 15% (V/V) to 30% (V/V) in US gasoline (Squillace et al., 1997).

The wide usage of MTBE increases opportunities for it to leak into the environment through gasoline underground storage tank leakages, recreational watercraft operations, and accidental spills. It has been reported that approximately

250,000 of the 385,000 confirmed leaking underground storage tank releases involve

MTBE (Kane et al., 2001). Since MTBE is extremely water soluble (50,000 mg/L

@ 25 ℃) and has a low tendency to adsorb to soils, it moves rapidly in groundwater

and potentially contaminates drinking water resources. MTBE is classified as a

possible human carcinogen by the US Environmental Protection Agency (EPA), 2000.

The drinking water advisory document reported that acceptable concentration of

MTBE in drinking water is 20-40 μg/L (US EPA 2002).

MTBE (C5H12O) contains two organic functional groups, an ether link and the

branched moiety. The ether bond (C─O─C) has been reported as an obstacle for

compounds’ biodegradability (White et al., 1996). MTBE used to be considered

relatively recalcitrant. However, biodegradation of MTBE has been reported in laboratory and field studies. Several bacterial strains including Rubrivivax strain

PM-1 (Hanson et al., 1999), Hydrogenophaga flava ENV 735 (Hatzinger et al., 2001),

Mycobacterium austroafricanum IFP 2012 (Francois et al., 2002), and others (Mo et

al., 1997) have been identified as using MTBE as a sole carbon and energy source.

Other organisms, such as Mycobacterium vaccae JOB5 (Smith et al., 2003),

Pseudomonas mendocina KR1 (Whited et al., 1991), and propane-grown strain ENV

425 (Steffan, et al., 1997) have also been reported to degrade MTBE as a cometabolic

substrate. However, little is known about how MTBE-degrading/utilizing

communities change as biodegradation of MTBE occurs. As MTBE is commonly

co-contaminated with benzene, toluene, ethyl-benzene, and xylenes (BTEXs), which

have a potential to serve as co-metabolic substrate for MTBE degradation, a better

2 understanding of the microbial community change in MTBE-contaminated sites will be beneficial to engineers in selecting an effective remedial strategy.

This study attempts to answer two questions: Are there known

MTBE-degraders/utilizers present in MTBE-contaminated soils? How is the microbial community affected by aerobic MTBE enrichment and the additions of a cometabolic substrate? The hypothesis of this research is that the community shifts of cultivated and uncultivated microorganisms in MTBE-contaminated soils, can be characterized qualitatively and quantitatively by molecular methods, and microbial community changes in response to additions of oxygen and cometabolic substrates.

Chapter 2 describes the physical/chemical properties of MTBE, environmental occurrences of MTBE, biodegradation of MTBE, characteristics of known

MTBE-degraders, and in situ bioremediation for MTBE. Chapter 3 describes the materials, experimental setup, and analytical methods; particularly, Real-time-t-RFLP assay was detailed. Real-time-t-RFLP, a quantitative fingerprinting method, is capable of quantifying each ribotype in a microbial community (Yu et al., 2004).

Results and discussion are summarized in chapter 4. Conclusions from this study and future study, such as molecular quantitative active MTBE-degraders in the

MTBE-contaminated soils, are described in chapter 5.

Chapter 2

Background

2.1 Properties of MTBE

MTBE is a synthetic chemical without natural sources. It is a clear liquid with low viscosity that is flammable and has an offensive odor. Chemically, MTBE is an ether based molecule containing 18% by weight oxygen with physical characteristics akin to other common gasoline constituents such as BTEXs. A comparison of physical and chemical properties of MTBE and BTEXs is presented in Table 2.1.

MTBE has a high solubility and a low Henry’s constant. Its water solubility is about 50,000 mg/L @ 25℃ and its dimensionless Henry’s constant ranges from 2.16

X 10-2 to 1.23 X 10-1 (Squillace et al., 1997). Therefore, MTBE is not a volatile

chemical. When considering the partitioning of MTBE between the water phase and

soils or subsurface solids, it prefers to be in the water phase as well.

Table 2.1 Basic physical and chemical properties of MTBE and BTEXs. Physical and MTBE Benzene Toluene Ethyl-benzene o-Xylene chemical properties Chemical C H O C H C H C H C H Formula 5 12 6 6 7 8 8 10 8 10

Chemical structure

Molecular weight 88.15 78.11 92.14 106.17 106.17 (g/mol) Vapor density @ 1 3.80 3.36 3.97 4.57 4.57 atm; 10 ℃ Specific gravity @ 0.744a 0.88a 0.8669a 0.867a 0.8802a 25 ℃ Water solubilitya 50,000 1730 534.8 161 175 (mg/L) Vapor pressura 245-276 95.19 28.4 9.53 6.6 (mm Hg) @ 25℃ Henry’s Law 0.022a 0.23b 0.272b 0.336b 0.212b constant (-) a Log KOC 1.049 1.50-2.16 1.56-2.25 1.98-3.04 1.68-1.83 a b b a b Log KOW 1.20 2.36 2.73 3.24 3.10 a. Office of Science and Technology Policy (OSTP). June 1997. Executive Office of the president. National Science and Technology Council. Committee on Environmental and Natural Resources. Interagency Assessment of Oxygenated Fuels. Washington, DC. b. Crittenden, Dave, Dave Hand, et al. Environmental Technologies Design Option Tools (ETDOT) for the Clean Process Advisory Systems (CPAs) Adsorption, Aeration and Physical Properties software. National Center for Clean Industrial and Treatment Technologies. 1997

2.2 MTBE environmental occurrences

2.2.1 MTBE in surface water, groundwater, and drinking water

In the 1970’s, MTBE was first added into gasoline as an alkyl lead additive to aid performance through enhancing the octane rating. In areas that did not meet the ambient air quality standard for carbon monoxide during the winter of 1992, the use of oxygenated gasoline was required. In 1995, nine metropolitan ozone nonattainment areas were required to use reformulated gasoline (RFG), which contains oxygenates, all year. The CAA Amendments also permitted other ozone nonattainment areas to take part in the RFG program. The most widely used oxygenate in RFG is MTBE.

MTBE is added at concentrations from15% to 30% by volume to oxyfuels and

RFG. The reasons for MTBE use being widespread are that it is cheap, it is relatively easy to produce with compounds readily available at most refineries, it blends readily with gasoline, and it can be shipped through existing pipelines

(Squillace et al., 1997). MTBE is produced from the chemical reaction of methanol and isobutylene.

While beneficial in reducing carbon monoxide in air, MTBE has been detected frequently as an environmental contaminant. It has been found in urban air (Pankow et al., 1997), surface water (Reuter et al., 1998), and shallow groundwater (Squillace et al., 1999).

MTBE has been commonly found to migrate ahead of the gasoline components

such as BTEXs at a gasoline spill site. The reasons are that MTBE has a high

mobility and it is less subject to biodegradation during transport (Fiorenza et al.,

2003). In the air, MTBE tends to enter atmospheric water through precipitation. In

areas where the atmospheric concentration of MTBE is locally high, as near parking

garages and gasoline stations, MTBE could occur in measurable quantities in

groundwater and stormwater in the immediate vicinity (Squillace et al., 1997). It has

been estimated that MTBE may have been released from up to 250,000 leaking

underground storage tanks in the United States (Johnson et al., 2000).

MTBE can be found in surface waters due either to a direct release, fallout from precipitation, discharge from contaminated groundwater, stormwater runoff, or emissions from motorized watercraft. Watercraft emissions are the primary source of MTBE in lakes and reservoirs. Volatilization at the air-water boundary is the major mechanism of MTBE loss from lakes and reservoirs (Fiorenza et al., 2003). It has been observed that the loss of MTBE follows an exponential decay, with more than 80% volatilized 2 to 3 weeks after discharges end. However, if MTBE is transported into deeper water after turnover, the rate of loss became much lower.

In a study of volatile organic compounds (VOCs) in ambient groundwater in the

U.S. conducted by the United States Geological Survey (USGS) National Water

Quality Assessment Program (NAWQA), MTBE was detected at a frequency of

16.9% in wells in urban areas and 3.4% in wells in rural areas. Among 60 VOCs,

only trichloromethane (chloroform) had a higher frequency than that of MTBE. The

common use of MTBE as a gasoline additive is the likely cause of its common

presence in water sources.

Nationwide, the city of Santa Monica, California is considered the place most

influenced by MTBE. In Santa Monica, about 50% of the drinking water supply

depends on the groundwater from two well fields, the Charnock and the Arcadia

(Stephanie et al., 2003). In August 1995, at the Charnock well field, MTBE was

detected. By April 1996, the concentrations of MTBE at Charnock and Arcadia had

increased to 610μg/L and 86μg/L respectively, concentrations that are dramatically higher than the standard suggested by EPA. In response, all of the city’s wells in the two fields were shut down immediately to avoid further consumption, and the city began to purchase replacement water. The investigation of this incident showed that

25 underground storage tanks were the direct reason for the contamination of the

Charnock well field. Meanwhile, the Arcadia well field was contaminated by a single service station (U.S. EPA, 2000). This incident was the first water contamination that attracted public attention to MTBE.

2.2.2 Effects of MTBE on human health

There are taste and odor problems associated with the presence of MTBE in drinking water. It can be detected at 2.5 ppb for odor and 2.0 ppb for taste (U.S.

EPA, 2000).

Most of the research on the effect of MTBE on human health has been focused on the effects of inhalation. MTBE is classified as a possible human carcinogen by the

EPA as a result of inhalation cancer tests, but no quantitative estimate of its cancer potency has been determined by EPA because of the limitations of the available data

(U.S. EPA, 2000). In 1997 the EPA announced that there was not enough data on

MTBE health effects and exposure to establish a national primary drinking water regulation. The drinking water advisory document provided by the EPA in 1997 concluded there was a low probability that MTBE between 20 and 40 μg/L would cause adverse health effects (U.S. EPA, 2000). The MTBE contamination problems are more serious in California than any other place in US. In 1998, the California

EPA proposed 14 μg/L as the concentration at which the increased risk of developing cancer from oral exposure is one in a million (Keller et al., 1998).

2.2.3 Actions against MTBE

In response to growing concerns regarding MTBE in water, the EPA appointed an independent Blue Ribbon Panel in 1998. The objective of this panel was to investigate the air quality benefits and water quality concerns associated with oxygenates in gasoline and to provide independent advice and recommendations on ways to maintain air quality while protecting water quality. The Blue Ribbon Panel issued a report with their recommendations (U.S. EPA, 1999) in September, 1999.

They recommended the removal of the current congressional Clean Air Act requirement for 2% oxygen in RFG to improve the nation’s water protection programs, and to reduce the use of MTBE substantially nationwide to maintain current air quality benefits and accelerate research on MTBE and its substitutes (U.S. EPA,

1999).

Some states, especially those that have experienced damage to their drinking water supplies by MTBE, have acted against MTBE use in gasoline. In March 1999

California and Maine began to remove MTBE from gasoline. The governor of

California announced an executive order to remove MTBE by December 31, 2002

(Fiorenza et al., 2003). In July 1999 a New Hampshire law was enacted to reduce the use of MTBE. The law also required that the state request a waiver from the

EPA allowing them to meet RFG requirements until 2002. Meanwhile, other states such as Arizona, Kansas, Missouri, New York, and South Dakota also have suggested

10 bans, phase-outs, or limiting the use of MTBE (U.S. EPA, 2000).

One of alternatives to the use of MTBE as an oxygenate in reformulated gasoline

is to use ethanol. Compared to the regulated aromatic compounds of BTEXs,

ethanol would be degraded faster. However, the ethanol-blended gasoline could

cause additional groundwater problems. Ruiz-Aguilar et al. (2002) reported that a

preferential use of ethanol will rapidly drive a source zone anaerobic. Powers et al.

(2001) showed that an extraordinary high concentration at an ethanol spill site would

cause order-of-magnitude increases in BTEX concentrations in groundwater. Since

the water in pipelines will cause the ethanol to separate from the gasoline, ethanol

reformulated gasoline cannot be transported by pipelines. Storage areas are needed

for ethanol blending and therefore the expense will be an increase in price between

2.4 to 2.9 cents per gallon of gasoline (U.S. EPA, 2002).

2.3 Biodegradation of MTBE

In the past 20 years, researchers have changed opinions on the biodegradability

of MTBE. MTBE was originally considered recalcitrant in the biosludges,

sediments, and groundwater. The ether bond in MTBE was viewed as an

impediment to the chemical compound’s biodegradation. Modrzakowski and

Finnerty (1989) observed that the terminal carbons of dioctyl ether were utilized for

growth by Acinetobacter sp., but the ether bond was not broken. Dimethyl ether

11 inhibited methane oxidation in soil (Ormeland and Culbertson, 1992). Salanitro et al.

(1994) also speculated that MTBE might inhibit metabolism or electron transport, or uncouple adenosine triphosphate (ATP) formation. However, other ethers, such as diphenyl ether (Liaw et al., 1989), cyclic ether (Bernhardt and Diekman, 1991), and dialkyl ether (Modrzakowski and Finnerty, 1989) have been observed to be biodegraded.

2.3.1 Aerobic degradation pathways of MTBE

MTBE degradation has been reported to occur aerobically and anaerobically.

Numerous laboratory studies have reported aerobic biodegradation of MTBE. Under aerobic conditions, MTBE is either used as a sole carbon and energy source by mixed consortia (Salanitro et al., 1994) and pure cultures (Hanson et al., 1999; Deeb et al.,

2001; Hatzinger et al., 2001; Francois et al., 2002), or cometabolically degraded

(Steffan et al., 1997; Hardison et al., 1997; Liu et al., 2001). Characteristics of known MTBE degraders are presented in Table 2.2. There are reports showing the degradation of MTBE under anaerobic conditions (Bradley et al., 2001; Finneran et al., 2001; Somsamk et al.; 2001, Kolhatkar et al., 2002). Aerobic biodegradation of

MTBE is discussed more in the following section.

Table 2.2 Aerobic MTBE degradation characteristics of pure cultures (adapted from Fiorenza et al. 2003) Strain Metabolism MTBE Cell Yield Reference Type Degradation Rate Rubrivivax Heterotrophic 0.07, 1.17, 3.56 0.18 mg Hanson et al., gelatinosus PM1 g/mL/h by cells/mg 1999 2X106 cells/mL MTBE for MTBE=5, 50, 500 mg/L Deeb et al., 50 mg/g cells/h; 2001 19 nmol/min/mg protein Hydrogenophaga Heterotrophic 86 nmol/min/mg 0.4 mg/mg Hatzinger et flava F735 protein MTBE al., 2001 Mycobacterium Heterotrophic 0.6 nmol/h/g dry 0.44 g/g Francois et austroafricanum wt; MTBE al., 2002 IFP 2012 20 nmol/min/mg protein ENV 425 Cometabolic 4.6nmol/min/mg Steffan et al., with propane cell protein 1997 Graphium sp. Cometabolic 1.9 nmol/h/mg Hardison et with n-butane dry wt without al., 1997 butane incubation; 0.6 nmol/h/mg dry wt with butane Arthrobacter Cometabolic 6.78 Liu et al., (ATCC 27778) with butane nmol/min/mg 2001 protein

A mechanism for MTBE degradation by a propane-oxidizing organism, bacterial strain ENV425, was proposed by Steffan et al. (1997). In this pathway (Pathway I in

Figure 2.1), the hemiacetal dismutes into tert-butyl-alcohol (TBA) and formaldehyde.

Subsequently the TBA is oxidized to 2-methly-2-hydroxy-1 propanol (MHP) and then oxidized to 2-hydroxy isobutyric acid (HIBA), 2-propanol, and ultimately pyruvic acid. Here HIBA degradation is a rate-limiting step, because the inhibitors of monooxygenase P450 will decrease MTBE and TBA degradation. Another study, however, suggests that MTBE is transformed to tert-butyl formate (TBF) initially

(Pathway II in Figure 2.1) and subsequently undergoes abiotic or biotic hydrolysis to

TBA (Hunkerler et al., 2001, Hardison et al., 1997, and Hyman et al., 1998).

So far, most researchers and scientists agree that the two mechanisms described above, oxygenative cleavage by monooxygenase and oxidation by cytochrome P450 enzymes, provide most applicable pathways of MTBE degradation. Both pathways have been confirmed in research. For instance, Francois et al. (2002) noted the evidence of monooxygenase activity in Mycobacterium austroafricanum IFP 2012, and cytochrome P450 activity in Graphium sp. has been confirmed in the research of

Hardison et al. (1997).

CH3 CH3 Pathway II CH3 O H3C C OCH3 H3C C OCH3 H3CCOCH

CH3 CH3 CH3 tert-butoxy-methanol MTBE TBF Pathway I

CH3

TBA H3CCOH + Formaldehyde

CH3

MHP H3CCOH

CH2 OH

2-Propanol

O CH3 O H3CCCH3 H3CCC HIBA OH OH

Acetone ? ?

Hydroxyacetone OH CH2 O OH CH2 O Methacrylic Acid H3CCC H3CCC OH OH Pyrovicacid OH OH 2,3-Dihydroxy-2 Methyl Propionic Acid

CO2 + Biomass

Figure 2.1 Proposed pathways for aerobic MTBE biodegradation. Adapted from Steffan et al., 1997, Hunkerler et al., 2001, Hardison et al., 1997, and Hyman et al., 1998.

2.3.2 Microorganisms using MTBE as a sole carbon and energy source

In the past decade, mixed cultures (Salanitro et al., 1994; Cowan et al., 1996;

Eweis et al., 1998; Bradley et al., 2001a) and pure cultures (Deeb et al., 2001;

Hatzinger et al., 2001; Francois et al., 2002) have been reported to metabolize MTBE

under aerobic laboratory conditions.

2.3.2.1 Mixed cultures using MTBE as a sole carbon and energy source

In 1994, Salanitro et al. enriched a mixed culture, BC-1, from the biotreater

sludge of a chemical plant. BC-1 is the first reported mixed culture involved in

aerobic biodegradation of MTBE in the laboratory. BC-1 has been maintained in the continuous culture with 120 to 240 mg/L MTBE as the sole carbon source in a simple

+ 3- 2＋ 2+ feed containing NH4 , PO4 , Mg , and Ca nutrients. The MTBE degradation rate

remained steady at 34 mg/g cells/h when there was a long cell retention time (80 to 90

days). However, at a cell retention time less than 50 days, the MTBE degradation

rate decreased, indicating that MTBE degraders are slow-growing microorganisms.

Salanitro et al. (1996) have confirmed that BC-1 is capable of degrading MTBE

+ widely. First, a continuously oxygen-sparged culture was able to oxidize NH4 to

- NO3 in the culture medium and obtain CO2 from the metabolism of the alkyl ether.

14 Second, BC-1 could degrade radiolabled methoxy carbon, CH3O-MTBE to 40%

14 CO2 and 40% C-labeled cells. Third, the result of substrate depletion experiments

16 showed that the intermediate TBA was produced during biodegradation and then metabolized by isopropanol and acetone. The study showed a first-order MTBE degradation rate of 0.31/day.

MTBE degradation is affected by both temperature and dissolved oxygen (DO).

In an experiment constructed by Cowan et al. (1996), MTBE and other oxygenates were observed to be biodegraded by an MTBE degrader, which was enriched from a petroleum refinery’s wastewater treatment plant. The maximum specific growth rate of MTBE decreased from 0.036 h-1 to 0.015 h-1 when temperature decreased from 30

℃ to 20 ℃. The half-saturation constants of the five target compounds ranged from

6 to 13 mg/L, with growth yields of these compounds of only 0.33 to 0.43 mg biomass COD formed/mg substrate COD consumed, suggesting the slow growth on the alkyl ethers. In addition to temperature, DO also affected the MTBE degradation.

The study of Park et al. (1997) showed that MTBE biodegradation decreased at less than 0.9 mg/L DO.

Eweis et al. (1998) also proved that MTBE and tert amyl methyl ether (TAME) were biodegraded by a mixed microbial culture, but ethyl tert butyl ether (ETBE) and diisopropyl ether (DIPE) could not be degraded by this culture. This indicates that the initial enzymatic attack is at the oxygen-methyl bond, which is common on MTBE and TAME, but not on ETBE and DIPE. Eweis et al. also assumed that some MTBE degraders use only the methoxy carbon, or they are prevented by a downstream

metabolite of the tertiary part of the ether, which causes a low cell yield.

2.3.2.2 Pure cultures using MTBE as a sole carbon and energy source

The major pure cultures used in the research of MTBE aerobic biodegradation

are Rubrivivax gelatinosus strain PM1 (Hanson et al., 1999), Hydrogenophaga flava

ENV 735 (Hatzinger et al., 2001), and Mycobacterium austroafricanum IFP 2012

(Francois et al., 2002). Other pure cultures involved in this research also include

Rhodococcus sp. nov. (Salanitro et al., 2001), Rhodococcus ruber (Hernandez-Perez et

al., 2001), Pseudomonas aeroginosa (Garnier et al., 1999), Mycobacterium vaccae

and Xanthobacter sp. (Hyman et al., 1998), Graphium sp. (Hadison et al., 1997), ENV

425 and ENV 421 (Steffan et al., 1997).

A bacterial Rubrivivax gelatinosus strain PM1 (Hanson et al., 1999) was isolated

from a mixed microbial community in a compost biofilter. Hanson et al. have

reported that strain PM1 could degrade MTBE and use it as the sole carbon and

energy source. Morphological analysis showed that PM1 is a gram-negative,

uniflagellated rod that produces an extracellular matrix. The result of 16S rDNA

analysis indicates that PM1 belongs to the β1 subgroup of Proteobacteria.

This study showed that initial linear rates of degradation increased with increasing MTBE concentration. At initial concentrations of 5, 50, and 500 mg/L

MTBE, the MTBE degradation rates by 2X106 cells/mL were 0.07, 1.17, and 3.56

µ g/mL/h, respectively. In the mineralization study, PM1 converted 46% of

14 14 14 14 C-MTBE to CO2 and 19% of C-MTBE to C-labeled cells in 120 hours when incubated with 20 µ g of uniformly labeled 14C-MTBE/mL.

Protein analysis was used to verify the growth of PM1 on MTBE. The

concentration of protein increased with decreasing MTBE concentration, which

indicates that PM1 used MTBE as its energy source for biomass growth. The mass

of protein produced was 2.7, 4.5, and 57 mg/mL at initial MTBE concentrations of 0,

5, and 50 µg /mL MTBE, respectively. This result gives a yield of 0.18 mg

cells/mg MTBE, which is consistent at different MTBE concentrations. The yield

is lower than cell yields on other substrates such as aromatics, sugars, and aliphatics.

White et al. (1996) suggested that the higher energy required to cleave the ether

bond could be the reason for the low cell yields. Salanitro et al. also explained that

MTBE may act as an uncoupler to ATP synthesis or that intermediates may be toxic

to cells.

In the study of MTBE bioaugmentation potential in groundwater samples, strain

PM1 was inoculated into sediment core material collected from a contaminated groundwater plume at Port Hueneme, California. Results showed that the degradation rate of MTBE increased with each MTBE spike. PM1 degraded an initial addition of 20 µg MTBE/mL in 40 hours, and then it took 20.5 and 12.5 hours to degrade the second and third addition of the same amount of MTBE. Hanson et al.

19 also found that strain PM1 remained active on MTBE for up to 83 hours after it was inoculated into soil samples. These results suggest that PM1 has potential for bioaugmentation of MTBE-contaminated groundwater sites.

A pure methylotrophic bacterium, Mycobacterium austroafricanum IFP 2012, which was isolated from an activated sludge sample, has been identified as an efficient MTBE degrader (Francois et al., 2002). Morphological analysis showed that Mycobacterium austroafricanum IFP 2012 is a gram-positive, nonsporulated rod and formed yellow colonies on plates. The result of 16S rDNA analysis indicates that Mycobacterium austroafricanum IFP 2012 had a 100% identity with

Mycobacterium austroafricanum IFP 2173 (Solano et al., 2000).

Studies of degradation capability showed that Mycobacterium austroafricanum

IFP 2012 could degrade MTBE, TAME, TBA, TAA, methanol, and formate, as sole carbon and energy source. The growth of Mycobacterium austroafricanum IFP 2012 on MTBE endured two phases. In the first phase, MTBE was degraded while TBA accumulated, with a constant cell concentration. In the second phase, cell concentration increased with TBA degradation. Two intermediates, TBF and HIB), also were detected during growth on MTBE.

2.2.3 Microorganisms using MTBE as a cometabolic substrate

Cometabolic degradation refers to a reaction in which microbes transform a substrate but this substrate cannot serve as a carbon and energy source for microbial growth. In cometabolic degradation, a primary substrate is required to support the growth of the microorganisms. A few microorganisms have been reported to degrade MTBE cometabolically. These organisms include fungi (Hardison et al.,

1997), propane-utilizing bacteria (Smith et al., 2003 and Steffan et al., 1997), and pentane-utilizing bacteria (Garnier et al., 1999).

Hardison et al. (1999) showed that a filamentous fungus, Graphium sp. strain

ATCC 58400, could degrade MTBE using propane or n-butane as a cosubstrate. The

MTBE biodegradation rate was 0.92 mg MTBE/g cells/h when growing on n-butane.

The metabolic intermediate TBF was initially detected in the study and further hydrolyzed both biotically and abiotically to TBA. Results indicated that

Cytochrome P-450 was responsible for the cleavage of ether bonds in compounds.

Steffan et al. (1997) reported a propane-oxidizing bacterium, ENV425, which was isolated from uncontaminated turf soil, converted 60% of the labeled [14C]

14 MTBE into CO2 within 30 hours. Results showed that MTBE was initially oxidized to TBA. TBA was further oxidized to 2-methyl-2-hydroxy-1-propanol and then 2-hydroxy isobutyric acid. Results also indicated that P-450 enzyme involved in the oxidation of MTBE and TBA.

Smith et al. (2003) reported cometabolism of MTBE by Pseudomonas

mendocina KR-1 grown on C5 to C8 n-alkanes. The previous studies reported that

Pseudomonas mendocina KR-1, which grows well on toluene, was capable of degrading another groundwater pollutant, trichloroethylene (TCE) (McClay et al.,

1995). It was also reported that Pseudomonas mendocina KR-1 cannot degrade

MTBE (Steffan et al., 1997). However, the current study showed that Pseudomonas mendocina KR-1 oxidized up to 94% of the added MTBE to TBA when cells were

grown on the C5 to C8 n-alkanes in the presence of MTBE (Smith et al., 2003).

MTBE was oxidized directly to TBA without generating TBF.

2.4 In Situ bioremediation of MTBE

In Situ MTBE biodegradation of MTBE may potentially occur through natural attenuation and engineered approaches. Three engineered approaches are commonly practiced: addition of oxygen to promote more rapid MTBE biodegradation, enrichment of degrading microorganisms at the MTBE-contaminated sites, and provision of an alkane and oxygen to enhance cometabolic biodegradation of MTBE.

Because quick aerobic degradation of MTBE has been detected in lab studies

(Salanitro et al., 1994), adding oxygen to promote rapid biodegradation has been commonly chosen for field studies. Jacanmardian et al. (1997) reported that MTBE concentrations decreased to nondetectable after the injection of oxygen (9ppm DO) at

22 a contaminated site with 5 to 10 mg/L MTBE, but there was no MTBE concentration change in the low oxygenated area (DO<0.5 ppm).

Other studies have showed the heterogeneous distribution of MTBE degraders after addition of oxygen (Mormile et al., 1994). To overcome the heterogeneous distribution of MTBE degrading microorganisms, the method of adding microorganisms capable of degrading MTBE in the field is employed. This method is also called bioaugmentation, which means to enrich the degrading microbial population in the MTBE-contaminated sites. One drawback of bioaugmentation is that most of the MTBE-degraders exist in the subsurface and they need time to adapt before performing degradation.

Another strategy is to add an alkane, such as propane, and oxygen, to enhance cometabolic biodegradation of MTBE (Fiorenza et al., 2003). Adding oxygen and alkanes will result in a high cost. In the study of Steffan et al. (2001), propane oxidizer ENV 425 was added at a concentration of 1 X 1011 cells/mL to three sparge wells. After 5 months, MTBE decreased from 320mg/L before the addition of ENV

425 to 190 mg/L and 25 mg/L at two upgradient wells, respectively, and in the downgradient well, MTBE decreased from 88 mg/L to 1.7 mg/L.

Chapter 3

Materials and Methods

3.1 Materials

3.1.1 Chemicals

MTBE (99%) and toluene (99%) were obtained from ACROS ORGANICS (New

Jersey, USA). TBA (>99%) was obtained from AlfaAesar (Ward Hill, MA).

3.1.2 Soil microcosms

Soils samples at depths of 35 feet to 52 feet were collected from an

MTBE-contaminated site. Ten grams of homogenized soil samples were used to inoculate 20 ml of nitrate mineral salts medium (NSM) (Chu et al., 1999) in 160 ml serum bottles (Alltech) and then capped with 20 mm aluminum seals and Teflon/Gray

butyl liner (Alltech). The compositions of NSM are listed as follow: NaNO3, 0.995

g; Na2HPO4, 0.866 g; K2SO4, 0.1708 g; MgSO4.7H2O, 0.037 g; CaSO4.2H2O,

0.0121 g; FeSO4.7H2O, 0.0222 g; KI, 0.0002 g; ZnSO4.7H2O, 0.0006 g; MnSO4,

0.0003 g; H3BO3, 0.0001 g; CoSO4, 0.0011 g; per liter of de-ionized water. MTBE

24 was added in each bottle to make the initial liquid MTBE concentration around 60 mg/L. Killed controls were prepared similarly; the controls were autoclaved twice, before the addition of MTBE. All experiments were conducted in duplicate.

Serum bottles were incubated at 30 oC in a rotary shaker at 150rpm. Every week, 50

µL gas phase samples were injected into a Hewlett Packard 5890 Series II Gas

Chromatograph equipped with a 30-M DB1 capillary column (Supelco) and a flame ionization detector. The temperatures of the injector and detector were maintained at

150 oC and 300 oC, respectively. The column temperature was set initially at 50 oC for 2 minutes, then ramped up to 100 oC at the rate of 30 oC per minute, and held constant for 10 minutes.

3.2 Experimental approaches

Figure 3.1 shows a flow chart describing the steps taken in this research. Part I was conducted to check whether known MTBE-degradaers/utilizers exist in

MTBE-contaminated soil samples. The experiment starts with the DNA extraction from original soil samples collected from an MTBE-contaminated site and is then followed by the Real-time-t-RFLP assay. Part II was conducted to examine whether the microbial community is affected by aerobic MTBE enrichment and the additions of a cometabolic substrate. The experiment starts with MTBE enrichment, aeration, and toluene additions to original soil microcosms. This was followed by DNA

Soil microcosms Part I Part II

MTBE enrichment, aeration, and toluene

DNA extraction

MTBE degradation detected in soil samples

DNA extraction Total genomic DNA

Total genomic DNA

Real-time-t-RFLP assay Real-time-t-RFLP assay

Microbial community of Changes of the microbial soils at different depths community

Figure 3.1 Flow chart showing steps taken in this research.

26 extraction to get the total genomic DNA. This is then followed by the

Real-time-t-RFLP assay.

3.2.1 DNA extraction from soil samples

Genomic DNA was extracted from soil samples using the Q-Biogene FastDNA

SPIN Kit (for soil) (Bio 101, Vista CA, USA). Two µL of soil sample were transferred into a 2.0 mL Eppis vial and spun down at 10,000 rpm at room temperature for 2 min using a microcentrifuge (Eppendorf5 415D, WESTBURG, NY).

The pellets were resuspended in 978 µL of sodium phosphate buffer and were transferred into a lysing tube, to which was added 122 µL of MT Buffer.

The reagents were homogenized by securing the lysing tube in FastPrep FP120

(BIO 101 SAVANT) and processing it for 30 seconds at speed 5.5. After cooling the sample on ice for 2 min, homogenization was repeated once. The lysing tube was cooled on ice for 5 min and then was centrifuged at 14,000 rpm at 4℃ for 15 min by using Micromax RF (Thermo IEC, NEEDHAM HEIGHTS, MA). Next, the supernatant was transferred to a clean 2.0 ml Eppis tube, 250uL protein precipitation solution (PPS) reagent was added, and the solution was mixed by shaking the tube by hand 10 times.

The tube was centrifuged at 14,000 rpm at room temperature for 5 min to pellet precipitate by using Micromax RF (Thermo IEC). The supernatant was transferred

27 to a clean 2 mL Eppis tube and 1mL binding matrix suspension was added. The tube was shaken gently by hand for 10 min to allow the binding of DNA to the matrix, the tube was placed in a rack for 3 min to allow settling of the silica matrix. The silica matrix was centrifuged for a short time, and then 500 µL of supernatant was discarded carefully to avoid discarding the settled binding matrix. The binding matrix was resuspended in the remaining supernatant and was transferred into a spin filter, which was centrifuged at 14,000 rpm at room temperature for 2 min by using a 5415D microcentrifuge (Eppendorf).

Following the above procedures, the catch tube was emptied, the matrix was dried for 2 min, and 500 µL salt/ethanol wash solution (SEWS-M) was added to the spin filter. The sample was centrifuged at 14,000 rpm at room temperature for 2 min by using the 5415D microcentrifuge (Eppendorf), then the catch tube was emptied, and matrix was again dried for 2 min. After washing with 500 µL 80% EtOH, the matrix was centrifuged at 14,000 rpm at room temperature for 2 min, then the filter was removed and placed in a new Catch Tube. After drying the filter for 5 min at room temperature, the DNA was eluted from the binding matrix by resuspending the binding matrix in 100 µL water, followed by a rest for 10 min at 37 oC in an incubator

(IC-62, American Scientific Products). The sample was then centrifuged at 14,000 rpm at room temperature for 2 min to transfer DNA to the catch tube, and the process

28 was then repeated. After this, the spin filter was discarded and the catch tube was stored at -20 oC for further research.

3.2.2 Real-time-t-RFLP assay

Real-time-t-RFLP assay (Yu et al., 2004) is shown in Figure 3.2. The assay was carried out as described in steps 1-5.

Step 1 Real-time PCR amplification of 16S rDNA

The 16S rDNA gene contained in the extracted DNA samples was quantified on a

DNA Engine Option continuous fluorescence detection system, using the primers

16S1055f (5’-(HEX)-ATGGCTGTCGTCAGCT-3’) and 16S1392r

(ACGGGCGGTGTGTAC-3’) and Taqman probe 16STaq1115f

(5’-(6-FAM)-CAACGAGCGAACCC-(TAMRA)-3’). Each real time PCR reaction

(25µL) consisted of 2.5 µL of PCR Buffer (Fisher Bioreagents, Pittsburg, PA), 0.5 µL of Bovine serum albunin (BSA) (10 mg/L); 0.2 µL of deoxynucleotides triphospates

(dNTP) (25 mM), 1.5 µL of MgCl2 (50 mM), 0.75 µL of forward primer (HEX)-1055f

(20 µM), 0.75 µL of reverse primer 1392r (20 µM), 0.625 µL of Taqman probe 16S

Taq1115f (10 µM), 0.25 µL of Taq Polymerase (Fisher Bioreagents, Pittsburgh, PA) (5

U/µL), 5 µL of DNA template, and HPLC water to increase the volume to 25 µL.

For pure cultures, 5.0 µL of 2.0 ng/µL and 5.0 ng/µL DNA solutions were used as

Real-time-t-RFLP Assay

Extract total genomic Real-time PCR t-RFLP DNA from soil samples

Amplify a target gene Digest labeled PCR by real-time PCR with products with a restriction a fluorescence-labeled enzyme primer and probe

Analyze labeled terminal restriction fragments (T-RFs)

Quantify initial DNA amount Determine relative used in PCR amplification abundance of each ribotype and determine DNA copies in in a microbial community (% sample (DNA copies/mL) of each ribotype)

1.E+08 3.E+06 L 7.E+05 1.E+06 5.E+04

1.E+04 1.E+03

1.E+02 DNA copies /m DNA copies 80 120 200 450 Terminal restriction fragment (T-RF) (base)

Figure 3.2 Flow chart showing steps taken in Real-time-t-RFLP assay (Adapted from Yu et al., 2004).

DNA templates. In order to minimize PCR inhibition for the soil samples due to co-extracted PCR inhibitors such as humic acid, the extracted DNA solutions were diluted a ten, a hundred, and a thousand fold to be used as DNA templates. The real-time PCR was implemented on a Continuous Fluorescence Detector (MJ

Research, DNA Engine, OPTICONTM). The thermal protocol included 10 min of initial denaturation at 95 ℃, followed by 30 cycles of 95 ℃ for 30 sec, 54 ℃ for 1 min, a plate read step, 72℃ for 2 min, and finally a hold temperature at 4 ℃.

Step 2 Gel purification of the PCR products

The amplified PCR products were separated along with a 1 Kb Plus DNA Ladder

(Invitrogen, Carlsbad CA) on a 1.5% agarose gel (containing 1% EtBr) using Tris acetate EDTA (TAE) as the electrophoresis buffer. Each sample included 20 µL of

DNA and 5 µL of dye (6X). The marker included 10 µL of 1 Kb Plus DNA Ladder,

15 µL of HPLC water, and 5 µL of dye (6X). According to the number of samples, the gel purification was carried out using a Mini-Horizontal Unit FB-SB-710 or

Midi-Horizontal System FB-SB-1316 (Electrophoresis System, Fisher Scientific).

The agarose gel was illuminated under UV350 and DNA bands of the expected size (~350 bp) were excised using a fine scalpel. Gel slices were transferred into a

MicroSpinTM column (Amersham Biosciences, Piscataway, NJ) and gently broken into small pieces by using a sterile disposable 20-200 µL pipette tip. The column

31 was then placed in a sterile 1.5 mL microcentrifuge tube and centrifuged at 3,000 rpm for 4 min at room temperature in order to transfer the TAE buffer containing DNA to the microcentrifuge tube. Centrifugation was repeated once and the spin column was discarded. To increase the volume of the DNA sample to 100 µL, HPLC water was added followed by an additional 10 µL of 3.0 M sodium acetate. Two volumes of ice cold ethanol (220 µL) equal to twice the amount of the NDA solution were then added and the tube was stored on ice for 3 to 4 hours. After this, the tube was centrifuged at 12,000 rpm at 4℃ for 15 min by using a Micromax RF (Thermo IEC) to pellet the precipitated DNA. After gently discarding the supernatant, the DNA pellets were washed twice with fresh 80% ethanol and dried at 37 ℃ for 30 min in the incubator.

Step 3 Restriction digestion of the PCR products

The dried DNA was resuspended with 40 µL of HPLC water in a clean 1.5 mL tube to which we added 0.5 µL of the restriction enzymes Msp 1 (5’C▼CGG3’)

(Fisher Bioreagents, Pittsburgh, PA), 5 µL of buffer (10X), and 5 µL of BSA (10X).

The tube was then placed in 37 ℃ water bath overnight. After this, the digestion reaction was stopped by incubating the tube in 70 ℃ water bath for 20-25 min. The tube was taken out from the water bath and spun for a short time to collect condensation drops from the top of the tube, and then 50 µL of HPLC water were

32 added to increase the volume to 100 µL. The digested DNA in the reaction mixtures was ethanol precipitated and desalted by ethanol two times as described above. Ten

µL of 3.0 M sodium acetate were added in the tube. Two volumes of ice cold ethanol (220 µL) were then added and the tube was stored on ice for 30 min. After this, the tube was centrifuged at 12,000 rpm at 4℃ for 15 min by using a Micromax

RF (Thermo IEC) to pellet the precipitated DNA. The supernatant was discarded gently and the remaining DNA pellets were then washed twice with fresh 80% ethanol and dried at room temperature for 30min. The desalted DNA pellet was resuspended in 100 µL of HPLC water, and then quantified by using a Hoefer DyNa Quant 200

Fluorometer.

Step 4 Terminal restriction fragment length polymorphism (t-RFLP) analysis

To determine the precise length of terminal restriction fragments from the amplified 16S rDNA products, t-RFLP analysis was carried out by using an ABI prism 310 Genetic Analyzer (Applied Biosystems Instruments (ABI), Foster City, CA) in a GeneScan mode. Each reaction contained 5 µL of desalted digested PCR

product, 10 µL of deionized formamide CH3NO, and 0.5 µL of Genescan ROX 500 size standards. Reaction mixtures were gently added in a clean Genescan tube to prevent any air bubbles from forming in the Genescan tube, since Genescan tubes cannot be centrifuged. Samples were denatured at 95 ℃ for 10 min followed by

33 rapid chilling on ice for 5 min. After this, the sample was analyzed by the ABI prism 310 Genetic Analyzer. Final DNA concentrations in the range of 0.05 ng/µL to 0.7 ng/µL were introduced electrokinetically with an injection time of 10 sec at 15

KV into the capillary containing polymer (GS STR POP4). Electrophoresis was carried out for 25 min at 60℃, at 15 KV. The matrix file selected was GS(D)6-FAM

HEX NED (primer) ROX, and the Local Southern Method was used to prepare a standard curve. After electrophoresis, the size of the terminal restriction fragments was determined by comparison with the internal standards by using GeneScan software.

Step 5 Data analysis

The threshold of real-time PCR was determined as 10 times the standard deviation of the background fluorescence. The threshold cycle (CT) of each PCR reaction was automatically determined by the computer software when the fluorescence exceeded the calculated threshold. During each PCR run, the CT values obtained from the plasmid standards were used for standard curves, and gene copies of samples were calculated with the standard curve.

The initial load of DNA in the community sample, as determined by real-time

PCR, was taken as 100%, and the initial copy number of each population within the community was calculated. Areas of all the major peaks detected in the t-RFLP

34 electropherograms were added together and calculated as 100%. Peaks below 30 bp found in all the samples, whose intensities were independent of input DNA, were excluded from the analysis. Peak area calculated below 1% was regarded as background noise, and was also excluded from the analysis. Relative amount of

DNA present in each population, within a community sample, was determined by calculating the percentage contribution of their peak area to the total peak area.

Chapter 4

Results and Discussion

4.1 Microbial community of MTBE-contaminated soils at different depths

Original soils at the depth of 35 to 52 ft were collected at an

MTBE-contaminated site. In this study, we applied Real-time-t-RFLP assay to study the microbial community of original soils at different depths. Soil samples obtained at four different depths were conducted to extract genomic DNA. Table 4.1 shows the characteristics of soils collected at four different depths.

Genomic DNA was conducted for Real-time PCR amplification and gel purification. The combined results of Real-time PCR and t-RFLP in the later figure, the 16S rDNA copies with terminal restriction fragment (T-RF) bases are shown

(calculations are presented in APPENDIX A) in the later table. The microbial

Table 4.1 Characteristics of soils at four depths.

Field MTBE Depth concentration Characteristics of soils (ft) (mg/L) 35-37 0.1-1 White and grey color, dirt. 40-42 0.1-1 Grey and brown, dirt and sand. 45-47 0.1-1 Dark brown, sand and little dirt. 50-52 0.1-1 Light brown, sand.

36 community structures of four soil samples are shown in Figure 4.1b.

Figure 4.1a, Figure 4.1b, and Table 4.2 indicate that both 16S rDNA copies and dominant ribotypes of original soils at different depths are different. In the original soil at the depth of 35-37 ft, the dominant terminal restriction fragment sizes are 83 bp,

100 bp, 103 bp, 105 bp, 245 bp, and 328 bp with the 16S rDNA copies of 1.99×108

/mL, 1.30×108 /mL, 4.03×108 /mL, 5.00×108 /mL, 2.50×108 /mL, and 4.62×109 /mL, respectively. In the soils at the depth of 40-42 ft, the dominant terminal restriction fragment sizes shifted to 83 bp, 98 bp, 102 bp, 103 bp, 105 bp, 242 bp, 245 bp, and

328 bp with the 16S rDNA copies of 7.67×107 /mL, 2.38×107 /mL, 4.36×107 /mL,

2.13×107 /mL, 6.86×107 /mL, 6.80×107 /mL, 8.98×107 /mL, and 3.54×107 /mL, respectively. In the soils at the depth of 45-47 ft, the dominant terminal restriction fragment sizes shifted to 83 bp, 98 bp, 102 bp, 105 bp, 111 bp, 189 bp, 207 bp, 242 bp,

245 bp, 328 bp, and 330 bp with the 16S rDNA copies of 7.23×106 /mL, 2.38×107

/mL, 4.27×107 /mL, 6.00×107 /mL, 3.55×106 /mL, 5.68×106 /mL, 6.13×106 /mL,

1.08×107 /mL, 9.88×107 /mL, 4.16×107 /mL, and 1.27×107 /mL, respectively. In the soils at the depth of 50-52 ft, the dominant terminal restriction fragment sizes shifted to 83 bp, 98 bp, 102 bp, 105 bp, 189 bp, 242 bp, 245 bp, and 328 bp with the 16S rDNA copies of 2.80×107 /mL, 2.06×107 /mL, 7.23×107 /mL, 1.11×108 /mL, 2.96×107

/mL, 3.83×107 /mL, 1.94×107 /mL, and 1.25×107 /mL, respectively.

Table 4.2 Calculated 16S rDNA copies with terminal restriction fragment (T-RF) bases of original soil samples. T-RF 16S rDNA (copies/mL) (base) 35-37 ft 40-42 ft 45-47 ft 50-52 ft 83 1.99E+08 7.67E+06 7.23E+06 2.80E+07 98 0.00E+00 2.38E+07 2.38E+07 2.06E+07 100 1.30E+08 0.00E+00 0.00E+00 0.00E+00 102 0.00E+00 4.36E+07 4.27E+07 7.23E+07 103 4.03E+08 2.13E+07 0.00E+00 0.00E+00 105 5.00E+08 6.86E+07 6.00E+07 1.11E+08 111 0.00E+00 0.00E+00 3.55E+06 0.00E+00 189 0.00E+00 0.00E+00 5.68E+06 2.96E+07 207 0.00E+00 0.00E+00 6.13E+06 0.00E+00 242 0.00E+00 6.80E+06 1.08E+07 3.83E+07 245 2.50E+08 8.98E+07 9.88E+07 1.94E+08 328 4.62E+08 3.54E+07 4.16E+07 1.25E+08 330 0.00E+00 0.00E+00 1.27E+07 0.00E+00

Terminal Restriction Fragment (T-RF) bases

35-37 ft

40-42 ft

45-47 ft Fluorescence

50-52 ft

Figure 4.1a Microbial community profiles of original soils based on t-RFLP analysis.

1.E+10

1.E+09 35-37 ft

1.E+08

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330 1.E+10

1.E+09 40-42 ft ) 1.E+08 mL 1.E+07 es/ i 1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330 (cop 1.E+10 1.E+09 45-47 ft

1.E+08 rDNA 1.E+07 6S 1 1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

1.E+10

1.E+09 50-52ft

1.E+08 1.E+07 1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

Terminal Restriction Fragment (TR-F) bases

Figure 4.1b Microbial communities of original soils analyzed by Real-time-t-RFLP.

As mentioned in the background, a few pure strains are known to use MTBE as a sole carbon and energy source, while others are capable of degrading MTBE as a cometabolic substrate. In this study, analysis was conducted to examine whether there are ribotypes which are similar to that of known MTBE-degraders. Based on the t-RFLP Analysis Program (TAP) (Marsh et al., 2000), the predicted fragment sizes of three known MTBE-degraders (Table 4.3), Rubrivivax strain PM-1,

Hydrogenophaga flava ENV 735, and Mycobacterium austroafricanum IFP 2012 are

100 bp, 108 bp, and 88 bp. The expected fragment sizes are assumed 3 bases shorter than the predicted T-RFs according to our previous studies (Yu et al. 2004).

Therefore their expected fragment sizes are 97 bp, 105 bp, and 85 bp, respectively.

Analysis indicated that the same fragment sizes as that of Hydrogenophaga flava

ENV 735 have been detected in microbial community profiles (Figure 4.2) at four depths. In the soils at the depths of 40-42 ft, 45-47 ft, and 50-52 ft, 97 bp is close to the ribotype of Rubrivivax strain PM-1 (98 bp). This result indicates that there might be known MTBE-utilizers Hydrogenophaga flava ENV 735 and Rubrivivax strain

PM-1 existing in MTBE-contaminated soils.

Table 4.3 Characteristics of known MTBE-utilizers.

Expected Accession Degradation Strain Fragment References Number Mechanism Size*(base) Rubrivivax strain Growth Hanson et al. AF176594 97 PM-1 Substrate 1999 Hydrogenophaga Growth Hatzinger et N/A 105a flava ENV 735 Substrate al. 2001 Mycobacterium Growth Francois et al. austroafricanum AF487529 85b Substrate 2002 IFP 2012 * Predicted length based on TAP analysis. MspI is used as restriction enzyme. Expected fragment sizes are assumed 3 bases shorter than the predicted T-RFs, according to our previous studies (Yu et al. 2004). a. Based on Hydrogenophaga flava (AF190800). b. Based on Mycobacterium austroafricanum (AB021420).

Terminal Restriction Fragment (T-RF) bases

105 bp (T-RF base is the same as that of ENV735) 35-37 ft

40-42 ft 105 bp (T-RF is the same as that of ENV735) 97bp (T-RF is close to that of PM1)

105 bp (T-RF is the same as that of ENV735) 45-47 ft Fluorescence Fluorescence 97 bp (T-RF is close to that of PM1)

105 bp (T-RF is the same as that of ENV735) 50-52 ft 97bp (T-RF is close to that of PM1)

Figure 4.2 Microbial community profiles of original soils based on TAP analysis.

4.2 Changes of microbial community in MTBE-contaminated soils due to the aerobic MTBE enrichment and a cosubstrate

A few pure strains are known to use MTBE as a sole carbon and energy source

(Hanson et al., 1999; Deeb et al., 2001; Hatzinger et al., 2001; Francois et al., 2002); while some microorganisms are capable of degrading MTBE aerobically as a cometabolic substrate (Smith et al., 2003; Steffan et al., 1997; Hardison et al., 1997;

Liu et al., 2001).

In this study, experiments were conducted to examine the effects of aerobic

MTBE enrichment and co-substrate (toluene) on the microbial community structure of

MTBE-contaminated soils. Experiments using soils at the depth of 35-37 (SET I)

and the depth of 40-42 ft (SET II) were incubated in the presence of O2 and MTBE

(around 70 ppm) in the first 92 days. On the 72nd day, all samples were aerated to

replenish O2. Aeration was carried out replenishing air into the sample bottle for 1 hr. On the 93rd day, 10 mL liquid samples were taken out from bottles and MTBE was spiked again with a concentration of 40 ppm. On the 129th day, toluene (around 5 ppm) was added in samples as a cosubstrate and other three additions in SET I and two additions in SET were followed in 8 days. Four liquid samples were collected at different time intervals (Figure 4.3) for Real-time-t-RFLP assay: A (original soil), B

(after aeration), C (before toluene addition), and D (after 4 additions of toluene in

SET I and after 3 additions of toluene in SET II).

MTBE in Soil Microcosms (Depth:35-37 ft) SET I Field MTBE Concentration: 0.1~1ppm 120

100 Effects of aerobic MTBE enrichment Effects of MC

Toluene M1 80 M2

TC 60 A* T1 * 40 B T2 Concentration (mg/L) C* D*

20 Aeration*

0 0 20 40 60 80 100 120 140 160 Time (day)

MTBE in Soil Microcosms (Depth: 40-42 ft) SET II Field MTBE Concentration: 0.1~1ppm

80 MC 70 Effects of aerobic MTBE enrichment Effects of M1 60 Toluene

50 * M2 B C* 40 A* TC 30 * 20 Aeration* D

Concentration (mg/L) T1

T2 0 0 102030405060708090100110120130140150 Time (day)

Figure 4.3 MTBE concentrations in soil microcosms. * A= 0.5 g of original soil samples were analyzed (start point); B =10 mL of liquid was withdrawn for molecular analysis. MTBE was added on day 93 (after aeration); C= 10 mL of liquid was withdrawn for molecular analysis. MTBE and toluene were again added on day 129; D= 10 mL of liquid was withdrawn for molecular analysis (end point). MC=MTBE concentration in control. M1=MTBE concentration in sample 1. M2=MTBE concentration in sample 2. TC=toluene concentration in

45 control. T1=toluene concentration in sample 1. T2=toluene concentration in sample 2.

The t-RFLP electropherograms of SET I and SET II are shown in Figure 4.4a and

Figure 4.5a, respectively.

Based on the Real-time-t-RFLP analysis (Figure 4.4b), the dominant TR-F sizes

(SET I) in the original soil (Point A) are 83 bp, 100 bp, 103 bp, 105 bp, 245 bp, and

328 bp. The dominant ribotypes shifted to 100 bp, 103 bp, 108 bp, 115 bp, 325 bp, and 328 bp after the first MTBE enrichment (Point B). The microbial community continued to change to 8 dominant ribotypes after aeration and the second MTBE addition (Point C). Additions of toluene did not result in significant decrease of

MTBE (Figure 4.3) but resulted in dominant ribotypes change, from 83 bp, 103 bp,

105 bp, 111 bp, 113 bp, 245 bp, 328 bp, and 330 bp (Point C) to 83 bp, 98, 103 bp,

105 bp, 111 bp, 113 bp, 245 bp, and 328 bp after three additions of toluene (Point D).

16S rDNA copies (SET I) with T-RF bases are presented in Table 4.4.

Based on Real-time-t-RFLP analysis (Figure 4.5b), the microbial community

(SET II) shifted from 8 dominant ribotypes in the original soil (Point A) to 12 dominant ribotypes after the first MTBE enrichment (Point B). The microbial community continued to change to 14 dominant ribotypes after aeration and the second MTBE addition (Point C). Additions of toluene did not result in significant decrease of MTBE but resulted in microbial community change, from 14 dominant

Terminal Restriction Fragment (T-RF) bases

A (original soil)

B (after aeration)

C (before toluene addition) Fluorescence

D (after 4 additions of toluene)

Figure 4.4a Microbial community profile (SET I) based on t-RFLP analysis.

1.E+10

1.E+09 A (original soil)

1.E+08

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

1.E+10

1.E+09 B (after aeration)

)

1.E+08

1.E+07

1.E+06 ies/mL

p 1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

co 1.E+10 (

1.E+09 C (before toluene addition)

1.E+08

1.E+07

1.E+06 16S rDNA 16S rDNA

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

1.E+10

1.E+09 D (after 4 additions of toluene)

1.E+08

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

Terminal Restriction Fragment (TR-F)

Figure 4.4b Microbial community (SET I) analyzed by Real-time-t-RFLP.

Terminal Restriction Fragment (T-RF) bases

A (Original soil)

B (after aeration)

C (before toluene addition) Fluorescence

D (after three additions of toluene)

Figure 4.5a Microbial community profile (SET II) based on t-RFLP analysis.

1.E+10

1.E+09 A (original soil)

1.E+08

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

1.E+10

1.E+09 B (after aeration)

1.E+08

)

1.E+07

1.E+06

1.E+05

ies/mL 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330 p 1.E+10

co 1.E+09 ( C (before toluene addition)

1.E+08

1.E+07

1.E+06

1.E+05

16S rDNA 16S rDNA 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

1.E+10

1.E+09 D (after 3 additions of toluene)

1.E+08

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

Terminal Restriction Fragment (TR-F) bases

Figure 4.5b Microbial community (SET II) analyzed by Real-time-t-RFLP.

Table 4.4 Calculated 16S rDNA copies with terminal restriction fragment (T-RF) bases of soil samples (SET I). T-RF 16S rDNA (copies/mL) (base) A B C D 83 1.99E+08 0.00E+00 1.20E+08 3.09E+06 98 0.00E+00 0.00E+00 0.00E+00 2.31E+06 100 1.30E+08 1.35E+08 0.00E+00 0.00E+00 103 4.03E+08 7.19E+08 5.60E+08 2.86E+07 105 5.00E+08 3.01E+08 2.37E+08 1.58E+07 111 0.00E+00 0.00E+00 3.13E+07 1.79E+06 113 0.00E+00 0.00E+00 3.58E+07 2.38E+06 115 0.00E+00 5.55E+07 0.00E+00 0.00E+00 245 2.50E+08 0.00E+00 2.86E+08 2.60E+07 325 0.00E+00 3.36E+08 0.00E+00 0.00E+00 328 4.62E+08 2.08E+08 3.58E+08 3.57E+07 330 0.00E+00 0.00E+00 9.81E+07 0.00E+00 A=original soil; B=after aeration; C=before toluene addition; D=after 4 additions of toluene.

51 ribotypes to 10 dominant ribotypes after three additions of toluene (Point D). 16S rDNA copies (SET II) with terminal restriction fragment (T-RF) bases are presented in

Table 4.5. Slower and even no degradation of MTBE was observed after the second addition of MTBE. The reasons for the loss of degradation ability were unclear.

The results of microbial community profiles of original soil samples indicated that there might be known MTBE-utilizers Hydrogenophaga flava ENV 735 present in MTBE-contaminated soils. According to the microbial community profile of original soils, in the soil at the depth of 35-37ft, there could be MTBE-utilizer

ENV735 present at the fragment size of 105 bp. Theoretically, if bacteria indeed exit in the soil and its degradation ability indeed work, the peak at 105 bp should increase after MTBE enrichment. However, t-RFLP results showed that the peak at 105 bp did not increase instead of a light decrease after aerobic MTBE enrichment (Figure

4.6).

According to the microbial community profile of original soils, in the soil at the depth of 35-37ft, there could be MTBE-utilizer ENV735 present at the fragment size of 105 bp. However, Real-time-t-RFLP results showed that in the sample at the depth of 40-42 ft, there were no increases at the 98 and 105bp after MTBE enrichment

(Figure 4.7).

In these two sets, the possibilities of the results could be: there might be no

MTBE-utilizers PM1 and ENV 735 present, or their degradation abilities were

Table 4.5 Calculated 16S rDNA copies with terminal restriction fragment (T-RF) bases of soil samples (SET II). T-RF 16S rDNA (copies/mL) (base) A B C D 80 0.00E+00 0.00E+00 6.43E+06 0.00E+00 83 7.67E+06 3.36E+06 1.28E+07 0.00E+00 85 0.00E+00 4.99E+06 0.00E+00 3.60E+07 91 0.00E+00 1.20E+07 2.42E+07 0.00E+00 92 0.00E+00 0.00E+00 0.00E+00 1.17E+07 93 0.00E+00 0.00E+00 1.29E+07 0.00E+00 98 2.38E+07 1.41E+07 7.62E+06 6.52E+07 100 0.00E+00 1.31E+07 0.00E+00 1.45E+07 102 4.36E+07 0.00E+00 2.02E+07 0.00E+00 103 2.13E+07 4.39E+07 1.88E+07 4.62E+07 105 6.86E+07 4.77E+07 3.35E+07 5.27E+07 107 0.00E+00 0.00E+00 0.00E+00 0.00E+00 108 0.00E+00 3.75E+06 6.99E+06 2.81E+07 111 0.00E+00 6.79E+06 0.00E+00 2.17E+07 113 0.00E+00 0.00E+00 5.01E+06 0.00E+00 115 0.00E+00 0.00E+00 0.00E+00 8.61E+06 189 0.00E+00 4.50E+06 4.44E+06 0.00E+00 207 0.00E+00 0.00E+00 2.82E+06 0.00E+00 242 6.80E+06 0.00E+00 0.00E+00 0.00E+00 245 8.98E+07 0.00E+00 0.00E+00 0.00E+00 309 0.00E+00 0.00E+00 8.63E+06 0.00E+00 325 0.00E+00 8.37E+06 0.00E+00 0.00E+00 328 3.54E+07 1.29E+07 0.00E+00 0.00E+00 330 0.00E+00 0.00E+00 1.11E+07 1.11E+08 A=original soil; B=after aeration; C=before toluene addition; D=after 3 additions of toluene.

1.E+10 A (original soil) 105bp (Hydrogenophaga flava ENV735) 1.E+09

1.E+08

1.E+07

)

1.E+06

ies/mL 1.E+05 p 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330 co

( 1.E+10 B (after aerobic MTBE enrichment) 1.E+09

1.E+08 16S rDNA 16S rDNA

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

Terminal Restriction Fragment (TR-F) bases

Figure 4.6 Effects of aerobic MTBE enrichment on microbial communities (SET I).

1.E+10 98bp (close to Rubrivivax PM1) A (original soil) 1.E+09 105bp (Hydrogenophaga flava ENV735)

1.E+08

1.E+07

)

1.E+06 ies/mL

p 1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330 co

( 1.E+10 B (after aerobic MTBE enrichment) 1.E+09

1.E+08 16S rDNA 16S rDNA

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

Terminal Restriction Fragment (TR-F) bases

Figure 4.7 Effects of aerobic MTBE enrichment on microbial communities (SET II).

55 inhibited by other unclear factors. Another possibility could be that since we observed a 15% concentration decrease of MTBE which was subtracted that of the control in the sample, there are might be some MTBE-utilizers which are uncultivated or have not been isolated.

Analysis also was conducted to examine the existence MTBE-degraders

(cometabolic degradation) in soils. Table 4.6 presents characteristics of known

MTBE-degraders which involved in MTBE cometabolic degradation. Bacteria have been reported that can cometabolic with propane, toluene, and butane.

In this study, we chose toluene as the cosubstrate because it is easy to operate compare to propane and pentane. A known MTBE-degrader is cometabolic with toluene called Pseudomonas mendocina KR1. TAP analysis shows that the expected fragment size of KR1 is 83 bp.

If MTBE-degrader KR1 which can use toluene as a cosubstrate presents in the sample, the peak at 83bp might increase after toluene additions. However, real-time-t-RFLP results showed the peak decreased (Figure 4.8). Therefore, there are probably not KR1 present in the sample at the depth of 35-37 ft or its biodegradation ability was inhibited by some unclear factors.

Toluene additions also caused both 16S rDNA copies and dominant ribotypes changed significantly. For example, the dominant ribotype at 330bp disappeared and the dominant ribotype at 98bp came out after toluene additions. 16S r DNA copies

Table 4.6 Characteristics of MTBE-degraders (cometabolic degradation).

Expected Degradation Strain Fragment References Mechanism Size*(base) Mycobacterium Cometabolic with Smith et al. 85 vaccae JOB5 propane 2003 Pseudomonas Cometabolic with Whited et al. 83a mendocina KR1 toluene 1991 Cometabolic with Steffan et al., ENV 425 N/A propane 1997 Cometabolic with Hyman et al., Xanthobacter sp. 103b propane 1998 Cometabolic with Hardison et al., Graphium sp. N/A propane 1997 Arthrobacter Cometabolic with 85 Liu et al., 2001 (ATCC 27778) butane Pseudomonas Cometabolic with Garnier et. al, N/A aeruginosa pentane 1999 * Predicted length based on TAP analysis. MspI is used as restriction enzyme. Expected fragment sizes are assumed 3 bases shorter than the predicted T-RFs, according to our previous studies (Yu et al. 2004). a. Based on Pseudomonas mendocina (AJ308310). b. Based on Xanthobacter sp. (AY436816).

1.E+10 C (before the addition of toluene) 1.E+09 83bp (Pseudomonas mendocina KR1)

1.E+08

) 1.E+07

1.E+06 ies/mL p 1.E+05

co 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330 (

1.E+10 D (after 4 additions of toluene) 1.E+09 16S rDNA 16S rDNA 1.E+08

1.E+07

1.E+06

1.E+05 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

Terminal Restriction Fragment (TR-F)

Figure 4.8 Effects of toluene on microbial communities (SET I).

58 of most ribotypes decreased after toluene additions.

In SET II, based on the real-time-t-RFLP results, we can see that the peak at

83bp disappeared after additions of MTBE (Figure 4.9). That means probably there are not KR1 present in the sample at the depth of 40-42 ft.

Additions of toluene did not result in significant decrease of MTBE but resulted in microbial community change, from 14 dominant ribotypes to 10 dominant ribotypes after three additions of toluene.

Slower and even no degradation of MTBE was observed after the second addition of MTBE. The reasons for the loss of degradation ability were unclear.

Analysis indicated that the same fragment sizes as that of Mycobacterium vaccae

JOB5, Pseudomonas mendocina KR1, Xanthobacter sp., and Arthrobacter (ATCC

27778) have been detected in microbial community profiles (Figure 4.4b and Figure

4.5b). The expected fragment size of Xanthobacter sp. was detected in all the samples. The expected fragment sizes of Mycobacterium vaccae and Arthrobacter were detected in SET II but not in SET I. The expected fragment size of

Pseudomonas mendocina was detected in both sets but not in all the samples. Results indicate that there might be known MTBE-degraders (cometabolic degradation)

Mycobacterium vaccae JOB5, Xanthobacter sp., and Arthrobacter (ATCC 27778), which can cometabolic with propane, present in MTBE-contaminated soils. Cloning of the PCR products is needed to confirm their positive identifications in the future

1.E+10 C (before the addition of toluene) 1.E+09 83bp (Pseudomonas mendocina KR1)

1.E+08

) 1.E+07

1.E+06 ies/mL p

1.E+05 co

( 80 83 85 91 92 93 98 100 102 103 104 105 107 108 111 113 115 162 189 207 242 245 309 325 328 330

1.E+10 D (after 3 additions of toluene) 1.E+09 16S rDNA 16S rDNA

1.E+08

1.E+07

1.E+06

1.E+05 80 83 85Terminal 91 92 93 98 Restriction 100 102 103 104 105 Fra 107 108gment 111 113 ( 115TR-F 162 189) bases 207 242 245 309 325 328 330 Terminal Restriction Fragment (TR-F) bases

Figure 4.9 Effects of toluene on microbial communities (SET II).

60 study.

Chapter 5

Conclusions and Future Studies

5.1 Conclusions

The objective of this research was to study the microbial community at a

MTBE-contaminated site. A newly developed molecular method (Real-time-t-RFLP) was applied to study the effects of aerobic MTBE enrichment and a co-substrate

(toluene) on the microbial community structure of MTBE-contaminated soils.

Real-time-t-RFLP, combining Real-time PCR and t-RFLP, is capable of quantifying each ribotype in a microbial community. Results of this study showed that microbial community (at the depths of 35-37 ft and 40-42 ft) shifted dramatically according to additions of MTBE, aeration, and cosubstrate (toluene). Although expected ribotypes of known MTBE-utilizers (metabolic degradation) (Hanson, et al., 1999;

Deeb et al., 2001; and Hatzinger et al., 2001) and MTBE-degraders (cometabolic degradation) (Smith et al., 2003; Whited et al., 1991; Hyman et al., 1997; and Liu et al., 2001) were detected in microbial community profiles, cloning of the PCR

62 products is needed to confirm their positive identification in the future study.

The commonly used engineered approaches for MTBE bioremediation include direct metabolism, cometabolism, bioaugmentation. A better understanding of microorganisms at an MTBE-contaminated site is beneficial to engineers in selecting an effective bioremediation strategy. This study has provided some preliminary information of this MTBE-contaminated site. Both expected known MTBE-ulitizers

(metabolism) and MTBE-degraders (cometabolism) were detected in the microbial community profile. However, there was no significant MTBE degradation after aerobic MTBE enrichment. It is likely that an aerobic cometabolism, bioaugmentation, or some combination, could be applied a feasible approach for

MTBE bioremediation at this contaminated site.

5.2 Future studies

Cloning of the PCR products is needed to confirm known

MTBE-utilizers/degraders’ positive identification in the future study. Experiments needed to be conducted to test the existence of MTBE-degraders cometabolic with propane.

It has been reported that more than 99% of organisms in the environment are not cultivated by routine techniques (Amann et al., 1995 and Pace, 1997). To confirm uncultivated clones in soil samples which may directly or indirectly related to MTBE

63 degradation, stable-isotope probing (SIP) (Manefiled et al., 2002) is a novel culture-independent method capable of linking metabolically active microorganisms to the microbial community function. However, the utilization of SIP may be limited due to the lack of quantitative information of metabolically active microorganisms within the community. A quantitative assay for linking microbial community function and structure (Q-FAST), which combines two modern molecular techniques, stable-isotope probing (SIP) and Real-time-t-RFLP, has been developed in our lab (Yu et al., 2004). The Q-FAST will be applied to confirm whether uncultivated clones are involved in specific MTBE degradation in the future study.

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Appendix

Appendix A

Calculations of the Real-time-t-RFLP assay

Total DNA copies were calculated by the following equation: 200µL × C(ng / µL) Total 16S rDNA Copies = ×16SrDNAcopies 5µL × C(ng / µL) ×V (mL)

where: 200µL was the total volume of genomic DNA sample;

5µL was the volume of genomic DNA used for each real-time PCR

reaction;

C (ng/µL) was the genomic DNA concentration obtained from DNA

extraction in the soil samples;

V (mL) was the sample size used for Genomic DNA extraction. For the

original soil sample, V=0.5 mL. For other samples, V = 10 mL;

16S rDNA copies were the DNA copies measured in each real-time PCR

reaction. The copies were determined from real-time PCR standard

curves (Figure B.1);

Total 16S rDNA Copies were the calculated DNA copies using

Real-time-t-RFLP assay.

Figure A.1 Linear regression of standard curve.

VITA

Bo Yang was born on April 17, 1973 in Kunming, China. She attended No. 3

High School and graduated in June 1991. She then pursued a Bachelor of Science degree in Civil Engineering at Wuhan University, HuBei, China. After completing her degree at Wuhan University in July 1995, she joined Yunnan Provincial Electrical

Power Design Institute in Kunming, Yunan, China and worked there as an engineer.

By July 2002, she attended University of Tennessee and sought a Master of Science degree in Environmental Engineering.