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Microbial characterization and biogeochemical cycling of iron in a creosote contaminated aquifer along… Ross, KellyAnn 2007

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M I C R O B I A L C H A R A C T E R I Z A T I O N A N D B I O G E O C H E M I C A L C Y C L I N G O F I R O N I N A C R E O S O T E C O N T A M I N A T E D A Q U I F E R A L O N G A G R O U N D W A T E R F L O W P A T H b y K E L L Y A N N R O S S B . S c , The U n i v e r s i t y o f Water loo , 2002 A T H E S I S S U B M I T T E D I N P A R T I A L F U L F I L M E N T O F T H E R E Q U I R E M E N T S F O R T H E D E G R E E O F M A S T E R O F S C I E N C E i n T H E F A C U L T Y O F G R A D U A T E S T U D I E S ( M i c r o b i o l o g y ) T H E U N I V E R S I T Y O F B R I T I S H C O L U M B I A September 2007 © K e l l y A n n Ross , 2007 11 A C K N O W L E D G E M E N T S M a n y thanks go to people w h o have assisted me w i t h the research i n this thesis. I w o u l d l i ke to thank the members o f the M o h n L a b , H a l l a m L a b and Grays ton L a b for their advice and support. In particular, m y gratitude is directed to G o r d Stewart for his infinite technical support; D o m i n i c F r i g o n for his perception, vast knowledge , and wi l l ingness to guide me through t ry ing circumstances; and my, supervisor B i l l M o h n for his patience. A d d i t i o n a l l y I w o u l d l i ke to thank two people, w h o I hope to meet someday, for their assistance: G o r d o n Webster f rom C a r d i f f Un ive r s i t y , a pioneer i n the m i c r o b i a l characterizat ion o f sediment from unconvent ional environments; and E l i zabe th Jones f rom the U S G e o l o g i c a l Survey, who empathizes w i t h me that i ron is one o f Ear th ' s most f ick le elements. M o s t important ly, I w o u l d l ike to thank m y collaborators i n the Ear th and O c e a n Sciences Department:. Roge r B e c k i e for his mentorship; M a r i o B i a n c h i n for his tenacity i n the f ie ld , w h i c h I brought back to the lab; and T i l m a n R o s c h i n s k i for his knowledge and friendship. A s w e l l , I w o u l d l ike to thank G l e n n B u d d e n and the crew o f the Ocean Venture , w h o made long hours o f sampl ing enjoyable. T h i s thesis is dedicated to m y mother and sister; two scientists w h o preceded me. A B S T R A C T Na tu ra l attenuation o f a contaminant p lume i n an aquifer is an attractive clean up opt ion that el iminates the expense o f engineering processes. A creosote-contaminated aquifer located beneath the Fraser R i v e r at B r a i d Street i n N e w Westminster B C was characterized to deduce the presence o f natural attenuation. A geochemica l and m i c r o b i a l characterization o f the contaminated aquifer was conducted to infer b iogeochemica l processes. Geochemica l analysis o f the pore water and the corresponding sediment-bacterial commun i ty characterization, v i a the sma l l subunit 16S r R N A , were implemented to elucidate the microorganisms and b iogeochemica l processes i n v o l v e d i n the degradation o f creosote compounds i n the p lume. F r o m these results it appears that the b iogeochemica l processes at the B r a i d Street site were dominated by Fe oxida t ion , w h i c h increased along the contaminant f lowpath that goes f rom anaerobic to aerobic i n nature. In i t ia l ly , it was observed that the geochemica l and bacterial analysis support the idea o f convect ive f l ow at the groundwater - r iver water interface, and spatial va r iab i l i ty i n hyporheic flowpaths. Secondly , the divers i ty o f the bacterial communi ty increases i n the hyporheic zone relative to the anaerobic zone o f the p lume, where this is attributed to more act ivi ty that i n turn, is a consequence o f steeper gradients i n redox solutes found i n the hyporheic zone. T h i r d l y , net F e 2 + ox ida t ion is occur r ing over F e 3 + reduct ion a long the contaminant f lowpath, where this process was la rge ly attributed to the presence o f P-proteobacteria, whose apparent richness increased a long the contaminant f lowpath i n the hyporheic zone. A d d i t i o n a l l y , the Burkholderia-related 2+ phylo type dominant a long the f lowpath is l i k e l y responsible for anaerobic Fe oxidat ion. The presence o f F e 2 + ox ida t ion at the B r a i d Street, site is attributed to the phys icochemica l processes associated w i t h a large t ida l ly inf luenced r iver such as the Fraser R i v e r ; where convect ive f low and spat ial ly variable hyporheic f lowpaths increase the extent o f the hyporheic zone i n the aquifer. I V T A B L E OF CONTENTS LIST OF T A B L E S vi LIST OF FIGURES vii Chapter 1. INTRODUCTION 1 The Study of Microbial Ecology 1 Natural Attenuation of Poly Aromatic Hydrocarbons 2 The Hyporheic Zone 9 Hydrogeology and Geomorphology in Hyporheic Zones. . . . . 1 0 Geochemistry and Redox zonation in Hyporheic Zones 11 Iron cycling in Groundwater and Hyporheic Zones 20 The Microbiology behind Iron Cycling 22 Microbial Community Studies in Hyporheic Zones 25 Thesis overview 26 2. M A T E R I A L S A N D METHODS 28 Site Description and Sample Handling 28 Geochemical Analysis 29 Microbial Biomass Analysis 31 D N A Extraction and PCR Amplification 32 Bacterial Community Analysis 33 Enrichment Cultures 35 Clone Library Construction and Analysis 38 V 3. R E S U L T S A N D D I S C U S S I O N . . .38 Geochemis t ry . . . . .38 M i c r o b i a l B iomass . . . . . . . . 51 Bac te r ia l C o m m u n i t y C o m p o s i t i o n and Structure. . . . . . 56 Denatur ing Gradient G e l Electrophoresis . 5 7 Enr ichment Cultures . . . 64 C l o n e Librar ies . . . . . - 65 Phylogenet ic Tree , 6 8 Bac te r ia l C o m m u n i t y A n a l y s i s Summary . 76 Shannon D i v e r s i t y Index 78 4. T H E S I S S U M M A R Y A N D F U T U R E D I R E C T I O N 80 Thesis S u m m a r y . 80 Future Di rec t ion . 82 R E F E R E N C E S 84 A P P E N D L X 132 P H O S H O L I P I D F A T T Y A C I D M A S S S P E C T R A . . . 1 3 2 VI L I S T O F T A B L E S Table 1. Table 1. Sediment P A H s from H Z - 2 0 0 4 . 1 0 4 2. Table 2. Sediment P A H s from H Z - 2 0 0 4 - . . 1 1 0 3. Tab le 3. Stratigraphy descr ipt ion o f samples used for bacterial c o m m u n i t y compos i t ion . . ., . • . . . 1 1 1 4. Tab le 4. Closest 16S r R N A sequence match to exc ised bands us ing the B L A S T N search tool 112 5. Table 5. A compar i son o f phylogenet ic groups f rom H Z clone l ibrary classif ied i n R D P II database to sequences i n the N C B I database us ing the B L A S T N search too l " 114 6. Tab le 6. A compar i son o f phylogenet ic groups f rom B H Z clone l ibrary classif ied i n R D P II database to sequences i n the N C B I database us ing the B L A S T N search tool 119 7. Table 7. A compar i son o f phylogenet ic groups f rom S Z clone l ibrary classif ied i n R D P II database to sequences i n the N C B I database us ing the B L A S T N search tool 121 8. Table 8. C o m p a r i s o n o f Shannon index (H) and associated Shannon equitabi l i ty ( E H ) calculated from the relat ive abundance values o f phylotypes i n clone l ibraries and corresponding D G G E profiles 123 vii L I S T O F F I G U R E S F igure 1. . F igure 1. Geographic locat ion o f the B r a i d Street site. . . . 1 2 4 2. F igure 2. Seasonal groundwater gradients and diurnal-r iver water ingression . . . . 1 2 5 3. F i g u r e 3 . L o c a t i o n o f sampl ing i n the B r a i d Street site 126 4. F igure 4, D i s s o l v e d i ron , F e 2 + , total i ron , indane, benzothiophene and naphthalene and locat ion o f sediment samples corresponding to profiles . 127 5. F igure 5. D G G E analysis o f bacterial 16S r R N A gene us ing pr imer 357F-907R. . . . . . . . . . . . . . 1 2 8 6. F igure 6. Burkholderia band intensity versus indane concentration.. . 129 7. F igure 7. D i s t r ibu t ion and relative abundance o f c lone l ibraries. . . . 130 8. F igure 8. Phylogen t ic tree o f clones that dominated at least 2 % o f the c lone libraries 131 1 Chapter 1 INTRODUCTION The Study of Microbial Ecology Prokaryotic species make up the vast majority of biomass on Earth and may exceed all other life forms on the planet in numbers and biomass (144). There are predicted to be 10 3 0 individuals globally (116). Their phylogenetic and physiological diversity is greater than that of plants and animals and their interactions with other organisms are even more complex. Microbes are essential to geochemical cycling, which makes understanding their community structure and function critical to understanding ecosystem sustainability (116). Characterization of a microbial community can yield information on the transformation of solutes taking place within the community's ecosystem. In particular, microorganisms can affect the chemistry of a contaminated environment by transforming the contaminant directly, or the environment, which will in turn affect the fate of the contaminant (62). The fundamental understanding of microbial ecology has proven to be challenging. In order to make quantitative predictions about a microbial community's role in an ecosystem we need to have quantitative information on the links between microbial community structure and activity. Microbiologists have traditionally used a reductionist approach, most notably enriching a single population or a community in a laboratory to understand physiological mechanisms in a natural environment. Due to the complex nature of microorganisms and their interactions with other organisms in the environment, observing microorganisms in the laboratory often cannot reveal microbial activity in nature. 2 More recently genomic approaches have been developed to assess the physiology of microorganisms obtained from natural environments. However, we are still a long way from relating the nature of bacterial genomes to the ecology of bacterial communities (116). Natural Attenuation of Poly Aromatic Hydrocarbons Characterizing a microbial community in a contaminated environment can help determine biodegradation processes. Within a subsurface environment, the collective processes that act to degrade contaminants and retard their movement is referred to as 'natural attenuation' of the contaminant (130). Biodegradation and the onset of natural attenuation are important concepts for subsurface environments where the contamination is inaccessible to ex situ remediation processes. A n objective of my thesis was to better understand the process by which microorganisms degrade polycyclic aromatic hydrocarbons (PAHs) in subsurface environments. PAHs in creosote PAHs are a common contaminant of aquifers and marine sediments (11). The Fraser River located in British Columbia is an example of a P A H contaminated environment where PAHs are ubiquitous throughout the sediments in the river (11, 58). The greatest contributor to P A H contamination in the Fraser River has been high-temperature wood treatment processes, which generate creosote (17). Creosote consists of a complex mixture of PAHs (85%), monoaromatic hydrocarbons (10%), and nitrogen, sulfur, and oxygen containing heterocyclics (5%) (55). In the literature, the biodegradation of homocyclic and heterocyclic aromatic compounds are categorized along with PAHs and the term P A H will be used in this manner throughout this thesis. PAHs can have harmful effects on animals and may inhibit the ability of organisms to fight disease. PAHs are known to be endocrine disruptors, mutagenic, reprotoxic, 3 carcinogenic and teratogenic. The biocidal properties of PAHs are what account for the harmful effects of creosote on the environment (96). Biodegradation of PAHs Many different bacteria are capable of using PAHs as nutrients through respiration processes (130). Microorganisms obtain energy for growth and cell maintenance by transferring electrons from an electron donor to an electron acceptor, which is referred to as the terminal electron acceptor process (TEAP). In a system, i f oxygen depletion by microbial metabolism exceeds oxygen input, the system will become anaerobic. At this point some facultative anaerobes switch from oxygen to NO3" as the next thermodynamically favorable electron acceptor. As NO3" concentrations decline, obligate anaerobes begin to use other electron acceptors. In P A H contaminated subsurface environments, an excess of available organic carbon often exists and microbial metabolism is limited by availability of electron acceptors. In subsurface sediments, it is generally observed that available electron acceptors are consumed sequentially in the order: dissolved oxygen>N03"(denitrification)>Fe3+>S042" >C02(methanogenesis) (84). Microbial respiration can convert PAHs to less harmful compounds that are sometimes mineralized by microorganisms to C O 2 . Bacteria capable of degrading low molecular weight PAHs are usually Gram-negative and include the genera Pseudomonas in addition to other Proteobacteria. Many of the high molecular weight PAH-degrading bacteria described in the literature are Actinomycetes; Gram-positive bacteria with a high GC content (130). A complicating factor in studying PAH-degrading microbial communities in geographically diverse sites is that the intrinsic communities of microorganisms differ in relation to the geochemistry and physical dynamics of the site (84). •4 Due to the interest in application of bacteria for bioremediation, the conditions under which they degrade PAHs, as well as the metabolic pathways involved, have been investigated in many studies. Biodegradation of PAHs has been observed under both aerobic and anaerobic conditions. There are several disadvantages to the promotion of anaerobic P A H bioremediation. For example, aerobic biodegradation kinetic rates are reported to be two orders of magnitude greater relative to anaerobic biodegradation (11). Furthermore, not all environments have an autochthonous population of anaerobes that are able to degrade the suite of PAHs found within the contaminated environment. This phenomenon has been observed in creosote-contaminated sediment, where limited biodegradation of PAHs was seen under denitrifying, SO4 2"-reducing and methanogenic conditions (55). In this study, three creosote P A H compounds: naphthalene, benzothiophene and indane, were monitored at the Braid Street site along the north arm of the Fraser River located in New Westminster, BC (Fig. 1). Of these three compounds, only naphthalene is included in the USEPA-regulated PAHs. In general, cleanup criteria at contaminated sites are based on a sum of sixteen USEPA-regulated PAHs, the sum of seven carcinogenic PAHs or both (130). The contamination at the Braid Street site is located in a plume within an aquifer below the Fraser River, where naphthalene degradation was previously determined to be active (6, 16). The TEAPs within the plume were determined to be predominately Fe 3 + reduction and methanogenesis (19). However, oxygen and NO3" could be potential TEAPs where the aquifer discharges, at the point where anaerobic groundwater mixes with aerobic river water (19). Naphthalene biodegradation The biodegradation of naphthalene has been extensively studied and can be thought of as a model process to help us understand the metabolic pathways used by 5 microorganisms in degrading PAHs. The more thermodynamically favorable aerobic process, whereby molecular oxygen is incorporated into the aromatic ring prior to the dehydrogenation and subsequent P A H ring cleavage, has been the focus of most studies. During aerobic P A H metabolism, oxygen is integral to the mechanism of mono- and dioxygenase enzymes in the initial oxidation of the aromatic ring (11). Diverse bacteria are able to oxidize naphthalene using dioxygenases, including members of the genera, Pseudomonas and Rhodococcus. A few bacteria are also known to oxidize PAHs by metabolic pathways using cytochrome P450 monoxygenases, such as members of Mycobacterium and Burkholderia (11, 71). Rockne and colleagues (122) reported the ability of marine methanotrophs to degrade PAHs via the action of the methane monoxygenase. However, the latter are considered minor mechanisms of degradation compared with the activity of the dioxygenases (11). It is widely recognized that microorganisms are also capable of degrading PAHs in the absence of molecular oxygen. Studies have proposed a mechanism for the anaerobic degradation of naphthalene (90, 151), where the first step is the carboxylation of the aromatic ring, which may activate the ring prior to hydrolysis. Degradation then proceeds by reduction via a series of hydrogenation reactions. There may be other mechanisms for anaerobic naphthalene degradation; however these have not been elucidated. For example, it is proposed that the initial step in anaerobic naphthalene degradation under S042"-reducing conditions occurs via an initial hydroxylation reaction (90). Overall, delineation of the pathway(s) of anaerobic P A H degradation is still at its infancy. Naphthalene degradation under SO4 2 " and N03_-reducing conditions is well documented yielding degradation rates two orders of magnitude lower than those reported under aerobic conditions. Some examples of studies that determined anaerobic degradation rates of naphthalene in microcosms using different TEAPs are given here: Naphthalene degradation in different studies under denitrifying conditions in microcosms containing aquifer material yielded rates of 0.055 uM/day (24) and 0.42 uM/day (39), and in river sediment at 384 uJVI/day (80). In microcosms where SO42" was a TEAP, degradation in river sediment yielded rates of 336 u,M/day (80), and in aquifer material at 0.23 uM/day (39). Furthermore, naphthalene degradation via Fe reduction in microcosms containing marine sediments yielded a rate of 0.05 jxM/day (13, 34). Anaerobic enrichment cultures have demonstrated that complete naphthalene degradation by methanogenesis is also possible where long-term exposure to PAHs exists (29). It is the variance observed in the rate of naphthalene degradation in different sediments that is why it is important to study the biogeochemical processes involved in P A H degradation. Studying P A H biodegradation will enable us to understand the factors that influence degradation rates and therefore help us to predict rates in geographically diverse ecosystems. : Benzothiophene and indane biodegradation Contrary to naphthalene, little information is available concerning the biodegradation of the homocyclic and heterocyclic PAHs in creosote. The aerobic desulphurization of benzothiophene, including the microorganisms and the enzymes involved, has been well documented due to its presence in diesel fuel, where a decrease in the sulfur component will ideally lead to a decrease in pollution (28, 77). The results of anaerobic degradation of benzothiophene seem less promising: Annweiler et al (2001) determined that incubating benzothiophene with a SO42" reducer did not yield growth with benzothiophene as the only carbon source after 100 days (5). As well, Dyeborg et al (1997) found complete persistence of benzothiophene in denitrifying and SCV'-reducing microcosms (40). However, co-metabolic growth of benzothiophene in anaerobic 7 microcosms with naphthalene as an auxiliary substrate yielded degradation products (5, 92). In creosote-contaminated sites, benzothiophehe has been reported to inhibit naphthalene degradation (5, 92). Mundt et al (2003) investigated the transformation of typical tar oil compounds including benzothiophene and indane under NO3" and SO4 2 "-reducing conditions: indane and benzothiophene were transformed in assays with polluted soil samples following a 426-day lag phase, whereas they were persistent in assays on highly polluted soil (98). Bacteria are able to transform indane under aerobic respiration (37), but there has been very little investigation as to its anaerobic degradation. Anaerobic degradation of indane has only been studied in subordination to research on the degradation of other PAHs (5, 98, 99, 148). Richnow et al (2003) studied the microbial in situ degradation of PAHs in a contaminated aquifer through monitoring carbon isotope fractionation. Here they revealed that indane concentration decreased downstream of the groundwater flowpath, suggesting microbial degradation, but its carbon isotope ratios remained constant (121). It is indicated from these results of benzothiophene and indane degradation studies that for the decision whether natural attenuation can be implemented, the fate of indane and benzothiophene has to be considered. Factors affecting natural attenuation of PAHs One of the main factors in determining the degradation rates of PAHs is their bioavailability, which is a function of physicochemical and microbiological factors that control the ability of a compound to be degraded. Bioavailability is one of the most important factors affecting natural attenuation (97, 130). The persistence of a pollutant is often linked to its bioavailability in a given environment (56). 8 Hydrophobic organic contaminants, including PAHs, can have a low bioavailability (129). The larger the molecular weight of the P A H , the lower its solubility, which in turn reduces the accessibility of the P A H for metabolism by the microbial cell (11). PAHs can undergo rapid sorption to mineral surfaces (e.g. clays) and organic matter (e.g. humic and fulvic acids) in the subsurface. The sorption of uncharged organic compounds is dominated by hydrophobic interactions, and therefore the solubility of these compounds can be used to estimate their affinity for sediment (76). In addition to solubility, microbial respiration processes can significantly affect the fate and transport of a contaminant (130). Microorganisms can affect the redox conditions of the subsurface environment through changes to geology, available electron acceptors and nutrients. These changes will in turn affect the fate of a contaminant by -altering sorption, co-precipitation and biodegradation processes (98). Indane, benzothiophene and naphthalene (in order of decreasing concentration) are the dominant PAHs in the contaminant plume at the Braid Street site. Additional studies have determined indane and benzothiophene as prominent PAHs at a site of creosote-derived contamination where degradation of other less persistent components such as naphthalene has been identified (121, 148). For naphthalene, rates of desorption are typically lower than rates of sorption, but considerably less is known about the fate and transport processes of heterocyclic components of creosote. The solubilities of benzothiophene and indane are 130 ppm and 108 ppm respectively (111). The solubilities of these PAHs are approximately three times greater than that of naphthalene and they are therefore expected to sorb less strongly than naphthalene. It is observed that different compounds of creosote travel at different speeds in groundwater, with more tightly sorbed components moving more slowly and dominating at later times of plume evolution (19). 9 It is important to ask, whether the apparent attenuation of a contaminant in the subsurface is consistent with our understanding on the fate and transport of that contaminant with the redox conditions present in the subsurface (32). Only then can we conclude the incidence of natural attenuation. The Hyporheic Zone Creosote contamination at the Braid Street site is introduced to the Fraser River via the hyporheic zone (HZ). This is the zone where the anaerobic groundwater from the aquifer mixes with the aerobic surface water from the river. A redox gradient is produced in the HZ from this mixing, which controls chemically and microbially mediated solute transformations occurring on the sediment surfaces (22, 95). Generally, the upwelling groundwater supplies microorganisms with nutrients while the downwelling river water provides dissolved oxygen and organic matter to microorganisms in the HZ. These factors are what make the HZ an excellent environment for rapid bacterial respiration, which could promote P A H biodegradation (22, 115). A number of properties influence the ecological processes occurring in the HZ including: water movement, permeability of sediment, substrate particle size, resident biota, land use and physiochemical features of the overlying river and underlying aquifer (22). For example, the rate at which groundwater and dissolved solutes enters the river water is controlled by river stage elevation and duration, water table gradient, and riverbed hydraulic properties (95). These factors are what make the biogeochemical properties of large river HZs, such as the Braid Street site HZ, complicated and difficult to study and therefore are not understood in detail. Many studies have employed methods to characterize biogeochemical processes taking place within the HZ. Some of these methods involved the collection of samples of HZ sediment or pore water, and others involved simulations in the form of laboratory 10 experiments or computer models (131). To date, however, there have been relatively few studies looking directly at microbial ecology and microbial processes in hyporheic sediments, as microbial ecologists have tended to focus their attentions on riverbed, lake, marsh/wetland and marine sediments (134). There is still much to be discovered through linking the relationships between physical and microbiological processes in HZs. Hydrogeology and Geomorphology in Hyporheic Zones Spatially variable hyporheic flowpaths The factors that determine which biogeochemical processes will take place in the HZ depend on the proportion of groundwater and surface water present and whether upwelling or downwelling is occurring (131). Depending on the relative head differences between the groundwater and river water, flow in the HZ may be upwards into the river, downwards from the river into the aquifer, or horizontal. In most HZs Darcy's law can be assumed to apply and the flow of water across the interface is a function of the hydraulic conductivity of the sediments and the hydraulic gradient across the HZ (131). The HZ matrix is porous medium, which normally includes locally significant heterogeneities (e.g. Conant, 2004). Water can only move freely through sediments that have sufficient porosity and therefore the hydraulic conductivity of sediments in the aquifer can play a role in the development of spatially variable hyporheic flowpaths (35). The production of CO2 and organic acids can lead to increased mineral solubility, which can lead to development of porosity and permeability (131). Conversely, microbial production of carbonate, Fe 2 + , and sulfide can result in the precipitation of calcite or pyrite that reduce porosity and permeability (32). One of the biggest impacts of pollutants on the microbiology of the HZ is the blocking of interstitial spaces by precipitates. These blockages can create local hydraulic heterogeneities and anaerobic conditions in areas that were previously aerobic, thus inhibiting microbial activity (60). 11 The hydraulic gradient and direction of flowpaths across the HZ can be controlled by riverbed form and channel slope irregularities. At the local scale, grain size variation, bedding, jointing and other sedimentary structures are important, whereas at a regional scale, changes in geological terrain and the effects of pressure, in regards to depth of the river, significantly affect HZ mixing (54, 131). For example, the deposition of gravel and sand in point bar and riffle structures result in a more permeable riverbed than in areas where clays and silts are located (109). This concept can be extrapolated to the circumstance at Braid Street where Bianchin hypothesized that the contaminant discharge is likely occurring along the face of a moderately sloped submerged bar (17). Furthermore, a Br- tracer analysis conducted by Bianchin (2001) concluded that tidal forcing on the plume is spatially variable (16). Bianchin hypothesized that this pressure variation at Braid Street in combination with the submerged bar is probably yielding convective hyporheic flowpaths and that high river stage elevation is not expected to enhance advective/horizontal flow (17). Sediment sorption capacity In addition to spatially variable hyporheic flowpaths and varying hydraulic conductivity, sediment sorptive capacity will also vary in the aquifer. Younger et al. (1993) studied attenuation of pollutants in a HZ through integration of fieldwork and a three-dimensional model. They discovered that parameters affecting the HZ, such as riverbed thickness, roughness and hydraulic conductivity, had the most influence on the model output. The authors discuss the effects of the hydrogeology of riverbed sediment, where a fine grain size provides a low hydraulic conductivity and high clay and organic matter content makes the sediments highly sorptive. The authors note that adsorption is likely to be the primary attenuation process operating in the riverbed sediments (146). The mineral content of aquifer sediment has been the subject of considerable 12 investigation to determine intrinsic potential for attenuation of pollutants, principally by sorption processes. Clay content is thought to have a significant affect on sorption processes; however, Richardson (2003) identified organic matter composition as being a priority in the attenuation of organic contaminants (120). It should be noted that although hydrodynamic dispersion and sorption processes to mineral and organic content both act to reduce the concentration of a pollutant, they do not reduce the total mass. A decrease in the pollutant mass is accomplished through degradation processes and therefore understanding the biogeochemical processes occurring in the HZ are of vital importance. Fraser sands hydraulic conductivity and gradients The Fraser sands are fluvial deltaic deposits with discontinuous interbedded finer and coarse-grained lenses. The sands are about 20-30 m thick and act as an aquifer locally confined by a silt layer that is approximately 5-10 m thick. Golder's (1997) determined that the sand hydraulic conductivity was SxlO^m/s and the silt was 5xl0" 7 m/s. Furthermore, Golder's determined that the flow in the aquifer is horizontal except beneath the river, where groundwater discharges upwards into the river (1). Bianchin (2001) determined through a Br- tracer study low-flow average groundwater velocities that ranged from 32 to 69 m/year (16). Bieber (2003) determined through groundwater modeling a net hydraulic gradient of 6x10"* m/s towards the river that reverses with tidal fluctuations and seasonally when river stage rises during freshet (19). Extent of the HZ Figure 2 depicts the general groundwater flow and the extent of the HZ seen at the Braid Street site determined from previous studies on the site (16, 19). Particle tracking performed by Bieber (2003) revealed particle movement that is directed downwards and 13 away from the river during periods of high river stage elevation. At freshet, the particles migrate up to 4 m away from the discharge point in the offshore region of the plume, which could represent the extent of the HZ at this site (19). Through measurements of resistivity and temperature profiling of the site Bianchin (2006) determined groundwater and river water mixing in the HZ zone up to 1.5 m below the riverbed, where groundwater principally discharges into the river approximately 100 m offshore in a 10 m wide band (18). Similar methods have been successfully applied previously in HZs including a study in a tidal creek in Australia: Acworth and Dasey (2003) used a combination of borehole electrical tomography and cross-creek electrical imaging to delineate the zone in which saline waters from the creek mixed with freshwater in the underlying aquifer. Here, they revealed the pattern of mixing was not homogeneous (2). As per the case of temperature analysis, this is based on the relatively constant temperatures of aquifers. Aquifer temperatures recorded around England and Wales indicated an average of 11.2°C with fluctuations of only 3°C (2). The surface water temperatures fluctuate daily, nightly and seasonally, and the difference in temperature between the groundwater and river results in a temperature gradient. Conant (2004), combined temperature profiling with hydraulic testing and geochemical measurements to delineate groundwater fluxes across a HZ. Here Conant discovered different flow behaviors that were dependent on hydraulic conductivity and the structure of the riverbed (35). Source contamination At the Braid Street site, the source contamination is a dense non-aqueous phase liquid (DNAPL) creosote plume, which extends into the Fraser sands aquifer to approximately 22 m in depth in a localized area. The deep D N A P L source is approximately 120 m from the shore of the Fraser River and is located up gradient of the 14 site (19). The contaminant-plume in the Fraser sands is estimated to be 200 m wide (17). As part of a proposed site management plan, a pump and treat program has been in operation since July 1996 to contain the D N A P L source. The capture zone of this well is at a maximum during low gradient periods during freshet and at high river stage, and at minimum during periods of high aquifer gradients during the rainy season (149). Overall, the pumping well decreases the hydraulic gradient, thereby increasing the residence time of solutes and the potential for biodegradation (19). Geochemistry in Hyporheic Zones Geochemistry along the hyporheic flowpath A number of processes can occur as water enters the HZ sediments into the aquifer and mixes with groundwater. In one study, the most significant changes were found to occur within the first few meters of infiltration and were related to microbial degradation of organic matter changing redox conditions. Further changes along the flowpath were attributed to solubility adjustments partly controlled by alkalinity and to mixing with a deeper aquifer of different chemical composition (54, 131). In reduced zones, the activity of metal-reducing bacteria can provide a mechanism for metal mobilization. The microbial reduction of Fe 3 + and manganese oxides can result in the release of a wide variety of metals (e.g. arsenic) previously sorbed to or co-precipitated with the oxides. Generally metals are more mobile under reducing conditions and less mobile under oxidizing conditions (54). Alkalinity, which buffers pH, can also control adsorption - desorption reactions. In general, retardation processes including sorption and precipitation slow the rate at which a pollutant moves through an aquifer (131). 15 Redox zones; aquifer vs. HZ The redox environments present in a contaminated aquifer depend on the geochemistry of the pollution source, the natural geochemistry of the aquifer and the physics of their interaction, including temporal and spatial variability (7). Redox zones may be dominated by one redox process, but other redox processes may also take place in the same zone (32). Respiration by microorganisms and oxidative degradation of organic compounds will lead to the creation of a reduced zone, within which chemical species are consumed, such as dissolved oxygen, NO3" and Fe 3 + , and others are produced, such as Fe 2 + and C 0 2 (30, 84). The river water and groundwater mixing in the HZ differ in oxygen content, pH, temperature, and nutrient content. The ingress of river water into the aquifer can replenish lost nutrients such as NO3" and reoxidize reduced species, such as Fe 2 + (95). In general, as long as microbial activity consumes more oxygen than is replenished from recharge of the aquifer or river water ingress, the environment becomes reduced. As the aquifer becomes more reduced, dissolution of iron oxides and methanogenesis can become prominent TEAPs (53). Braid Street site redox zones in 1996. 1999 and 2002 In 1996 at the Braid Street site, concentrations of naphthalene recorded along the plume axis were found to greatly decrease in the offshore zone of the plume. In 1996 and 1999 Fe , C H , and C 0 2 levels were shown to increase with the corresponding decrease in naphthalene along the flowpath from the source zone to offshore (6, 16). Reactive transport modeling (6) suggested that significant mass loss must occur in the offshore region of the plume in order to explain the reduced contaminant concentrations found in this zone. Significant mass loss was confirmed by a 14C-naphthalene tracer study performed in 1999(16). 16 At the Braid Street site in 1996 and 1999, the increase in concentrations of CH4 and dissolved iron along the plume axis with a decrease in contaminant concentration suggested pollutant degradation via the TEAs methanogenesis and Fe 3 + reduction (6, 16). In 1999 the concentrations of dissolved contaminants were below the detection limit at a point further along the flowpath where discharge was expected to occur. Bianchin (2001) suggested that this decrease appears to be due to a separate process than that which is responsible for mass loss in the rest of the offshore zone. Bianchin hypothesized that degradative processes are spatially variable and are likely moving from anaerobic to aerobic in nature towards the aquifer - river interface (16). In 2002, Bieber also observed high dissolved iron and CH4 concentrations along the plume flowpath, and she also attributed this to Fe3+-reducing and methanogenic degradation processes (19). Furthermore, Bieber conducted geochemical modeling and revealed that degradation of naphthalene might be taking place by a combination of Fe 3 + reduction and methanogenesis or by Fe 3 + reduction, alone. Bieber agreed that a separate degradative process was taking effect at the discharge point compared to upgradient in the plume (19). However, in contrast to previous studies by Banchin and Anthony, Bieber did not see an increase in iron concentrations along the plume flowpath (19). In addition to iron, C H 4 and CO2, other redox sensitive solutes such as NO3/NO2", SO4/H2S and M n 2 + were also analyzed at the Braid Street site. Bianchin (2001) and Bieber (2003) found these solutes in relatively low concentrations and concluded that they are therefore likely not involved in electron transfer processes (16, 19). High N 0 3 " values were sampled within the plume during profiling by Bianchin (2001) in 1999 but were attributed to contamination of samples with nitric acid during collection (16). Bieber (2003) also observed higher NO3" concentrations than is normally measured in waters with high dissolved iron and C H 4 . Bieber attributed these high values to sampling during freshet, when aquifer gradients directed towards the river are at a minimum and . . - • 17 the influence of the pumping well is at a maximum. Furthermore, Bieber suggested that as a result of freshet, plume water is mixed with higher NO3" water downgradient of the plume (19). Overall, the analyses conducted from 1996 to 2002 concluded that Fe3 +reduction and methanogenesis are the dominant TEAPs at the Braid Street site and are attributed to the apparent receding naphthalene plume. Furthermore, below the river water -groundwater interface at the discharge point, mass loss is likely attributed to a combination of physical and microbiological processes not associated with these processes. In this thesis, it is hypothesized that the mass loss at the discharge point is attributed to enhanced microbial processes in the HZ, where this enhancement will induce secondary oxidation reactions such as Fe 2 + oxidation. Modeling of redox conditions in plumes (65) suggests that secondary oxidation reactions (oxidation of e.g., CH4, sulfides, Fe 2 + , and N H / ) may be very important in contaminant plumes but has been dealt with only rudimentarily in studies on HZ biogeochemical processes, including studies at the Braid Street site. Fate of naphthalene at the Braid Street site Reactive transport modeling led Anthony (1998) to predict the ti/2 of complete anaerobic degradation of naphthalene in the aquifer to be 1 year (6). Additionally a 1 4 C -naphthalene tracer study led Bianchin (2001) to estimate a tj/2 of 0.72 to 2.9 years (16). Lab microcosms were also set up to determine the degradation capabilities of the aquifer material from the anaerobic zone at the Braid Street site. Anthony (1998) saw complete aerobic degradation of naphthalene in microcosms in one week. However, he did not see any naphthalene degradation following 240 days under S0 4 2 " and N 0 3 " reducing conditions (6). Lesser (2000) set up microcosms from Braid Street in triplicate to measure naphthalene degradation under S0 4 2 " and N03-reducing environments 18 amended with PAHs. Here he saw 20 to 30% degradation over unamended controls after 288 days. From this Lesser determined a ti/2 of naphthalene to be 2.31 years, in agreement with Bianchin's field based estimates (82). No microcosms were set up under Fe3+"reducing or methanogenic conditions. As well, the products of Fe 3 + reduction and methanogenesis were not monitored in the field tracer experiment. Furthermore, Bieber (2003) looked for anaerobic degradation products of naphthalene in the pore water at Braid Street, but found inconclusive results (19). Therefore the degradative pathway of naphthalene attenuation has not been ascertained at the Braid Street site. Braid Street site contaminant plume The offshore profiling conducted in 1999 did not demonstrate naphthalene concentrations reduced from 1996. Bianchin (2001) concluded that the rate of naphthalene desorption processes could be similar to the rate of biodegradation and dispersion resulting in apparent stability of the plume (16). Contrary to Bianchin's conclusions, Bieber (2003) demonstrated that naphthalene concentrations measured show that the plume is not at steady state (19). This might indicate that the naphthalene plume is receding back towards the source contamination. Bieber determined in 2003 that indane and benzothiophene were the significant components of the off shore plume, and not naphthalene. Solid phase redox processes Aquifer sediments contain large pools of redox-sensitive species including minerals, precipitates and ions associated with exchange sites on particle surfaces. Important oxidized solid species include Fe 3 + oxides and SO4 2 " present on exchange sites. Important reduced species include organic matter, sulfides, and Fe 2 + carbonates. Reduced species such as Fe 2 + , M n 2 + and N H 4 + may also be present on exchange sites 19 (32). Organic carbon may have significant capacities, both as an electron acceptor or donor i f the pools are reactive, but this reactivity is normally considered to be low (54). Roschinski (2007), performed sediment extractions to quantitate the reactive fraction of the minerals that are present on the surface of sediment grains by measuring the relative amounts of Fe 3 + and Fe 2 + on the sediments from Braid Street site. Furthermore, Roschinski used a scanning electron microscope (SEM) and x-ray diffraction (XRD) for qualitative analyses of the reactive fractions of the sediments (125). Roschinski's sediment extractions showed no distinct increase of Fe 3 + with depth,. however he did see variability in Fe 3 + concentration when he conducted duplicate, triplicate and quadruplicate extractions. This variability corresponded to an increase in contaminant, dissolved iron and total iron concentrations seen in Figure 4. Roschinski suggested that the lack of noticeable iron accumulation in the HZ might indicate that iron precipitates on shallow sediments are subsequently scoured from the riverbed during freshet (125). Scouring at the Braid Street site was a phenomenon originally proposed by Bianchin in 2001 (16). Through S E M analysis, Roschinski determined the presence of a single iron-sulfide crystal on the sediment from depth to 1.2 m below the riverbed (125). The bacterial reduction of S 0 4 to sulfide leads to pyrite precipitation in the presence of Fe 2 + concentrations that are found at the Braid Street site (32), therefore the presence of pyrite may indicate that S 0 4 reduction is or was a dominant TEAP. Roschinski also observed amorphous iron mineral coating on several samples. These were assumed to be iron oxides or iron carbonates, which could be indicative of Fe 2 + oxidation or Fe 2 + precipitating out as a consequence of saturation (125). 20 Iron Cycling in Groundwater and Hyporheic Zones Precipitation and adsorption processes are dependent on redox conditions (32). Downgradient from the plume when additional oxygen can be supplied by the river, the water environment becomes oxidized and dissolved iron and manganese is removed from solution by a combination of adsorption to precipitates and oxidation reactions partly catalyzed by bacteria (54). Oxidation of Fe 2 + forms hydrous oxides and hydroxides that settle as amorphous gels in the sediment pores or form coating on sediment particles, which eventually crystallize into water-free oxides such as hematite (61, 145). Iron cycling at the Braid Street site Geochemical modeling of the Braid Street site in 2003 suggested that the lack of increase in dissolved iron along the flowpath downgradient from the plume might be due to surface complexation or mineral precipitation (19). The oxidation of dissolved iron leading to mineral precipitation is considered to be a secondary redox reaction. The premise that oxidation of Fe 2 + might be occurring near the discharge point and that Fe 3 + reduction is a predominant microbial process leading to the degradation of the contaminant plume at the Braid Street site, leads to the conclusion that biogeochemical cycling of iron at the Braid Street site is prevalent along the contaminant flowpath. In groundwater systems, iron occurs in one of two oxidation states: reduced soluble Fe 2 + or oxidized insoluble Fe 3 + (54). The fate of Fe 3 + in groundwater Fe3+-hydroxide (Fe(OH)3) is the direct result of Fe 2 +oxidation and precipitation. With time, amorphous Fe3+-hydroxide is mineralized. Amorphous hydrous Fe 3 + oxides are amphoteric ion, meaning that as pH conditions change it has the capacity to offer hydrogen ions or hydroxyl ions for cation or anion exchange respectively. Fe 3 +can form 21 soluble complexes with many inorganic and organic ligands, including compounds that are the byproducts of the P A H biodegradation (32). The activation energy to transform Fe 3 + to Fe 2 + is relatively small. Conversely reductive dissolution of Fe 3 + oxyhydroxide minerals requires much more energy in order to change the coordination environment of iron. Structural Fe 3 + must form chemical bonds with adsorbed reductants, organic compounds, reduced metal complexes or with extracellular iron reductase. Multiple metal-oxygen bonds in the mineral crystal must then be broken to release structural Fe 2 + to solution. Reduction of Fe 3 + oxyhydroxide minerals is a much slower reaction than the analogous reaction between aqua-ions. It is these slow kinetics, due to high activation energies for structural rearrangement, that are important reasons that many redox processes in contaminant plumes require microbial catalysis to proceed at measurable rates. In general, as the crystallinity of the Fe 3 + mineral increases the microbiological availability decreases (85). The fate of Fe 2 + in groundwater Complexed ions typically react more slowly than uncomplexed ions. This impacts the behavior of Fe 2 + . Fe 2 + normally can be oxidized to Fe 3 + in minutes, but complexed Fe may take months to complete the same reaction. The time required for uncomplexed Fe 2 + to undergo oxidation to the Fe 3 + is dependent on many factors, the dominant being: pH; temperature; dissolved oxygen concentration, and the presence of other soluble ions. The lower the pH and temperature the longer the time required for completion of the oxidation reaction (32). Longer times for abiotic reactions to occur can yield a competitive environment to the advantage of microorganisms (30). Microorganisms can cycle iron Iron readily undergoes reduction or oxidation depending upon surrounding conditions. There are many microbiological metabolic pathways that utilize the redox 22 couple Fe 2 + and Fe 3 + (140). Organic material can be biodegraded with Fe 3 + as the terminal electron acceptor, resulting in the production of reduced soluble Fe 2 + . Anaerobic conditions are required as is the presence of Fe 3 + in a suitable form (85). Furthermore, microorganisms can oxidize Fe 2 + to Fe 3 + under aerobic or anaerobic conditions by methods including: the extraction of CO2 from Fe2+-bicarbonate leaving insoluble Fe3+-hydroxide, the utilization of Fe2+-organic acid complexes as a carbon source leaving precipitated Fe3+hydroxide, and the direct oxidation of dissolved Fe 2 + by lithotrophs (30). Iron cycling in HZs The presence and form of iron in groundwater are an indication of complex reactions in the subsurface (32). However, to date the literature does not address how iron behaves in HZs, what processes control its mobility, and the effects of those processes on river water and groundwater concentrations of iron. The Microbiology behind Iron Cycling Microorganisms conserve energy from both oxidation and reduction reactions involving the cycling of iron between Fe 3 + and Fe 2 + . Fe 2 + is used as an electron donor to provide reducing equivalents for respiration and the assimilation of carbon into biomass by lithotrophic Fe 2 +-oxidizing microorganisms under both aerobic and anaerobic conditions. Whereas Fe 3 + can be used as a respiratory T E A under anaerobic conditions by lithotrophic and heterotrophic Fe3+-reducing microorganisms (140). Archaeal and bacterial domains are capable of oxidation and reduction of iron. However, to date the majority of phylotypes associated with iron-cycling have been determined in the bacterial domain (140), and the research contained in this thesis focuses primarily on the bacterial microbial characterization at the Braid Street site. . . • 23 Fe 3 + reduction by FRB Under anaerobic conditions, Fe3 +-oxide minerals are reduced by Fe3+-reducing bacteria (FRB). These bacteria have been identified mainly in river and aquifer sediments, and have been noted in marine sediments to a lesser degree. During this v-respiration process FRB can use both organic and inorganic electron donors. The microbially mediated reduction of Fe3 +-oxide minerals can generate both aqueous and solid-phase Fe2+-minerals such as siderite (30). Fe2"1" oxidation by FOB Microbially mediated Fe 2 + oxidation is carried out by Fe 2 +-oxidizing bacteria (FOB). FOB are ubiquitous and have been identified in many different environments. The aerobic bacterial oxidation of Fe 2 +has been known for more than 100 years, and corresponds mainly to acidic environments where Fe 2 + is stable under a higher partial pressure of oxygen. However, FOB have been identified to compete with abiotic oxidation of iron in microaerophillic/neutrophilic environments (41). FOB are capable of competing with abiotic oxidation kinetics between oxygen and Fe 2 + and have been shown to contribute to iron cycling in microaerophillic environments, where this metabolic process has been coupled to growth (141). Oxidation of Fe 2 + by FOB can lead to the precipitation of Fe2+-oxides such as goethite and hematite, which can contribute to the immobilization of heavy metals and metalloids through co-precipitation or physical envelopment (140). Anaerobic FOB Anaerobic Fe2+-oxidation by FOB was only identified in the early 1990s. FOB can couple Fe 2 + oxidation to the reduction of N 0 3 " . The oxidation of Fe 2 + , including Fe (s), coupled to NO3" reduction is energetically favorable at neutral pH and should yield enough energy to support carbon fixation and microbial growth (41). In 24 environments where NO3" reduction and Fe3+-oxidation merge, this metabolism not only influences the iron cycle, but can also influence the nitrogen cycle. Several phylogenetically diverse mesophiles have been described as being capable of NO3" -dependent Fe 2 + oxidation. However, in most cases growth was either not associated with this metabolism or was not demonstrated in the absence of an additional electron donor or organic carbon as an energy source at circumneutral pH (140). Anaerobic microbial Fe oxidation closes the gap in the iron redox cycle. Together with microbial Fe 2 + reduction, these metabolisms are now known to be characteristic of phylogenetic diverse organisms (140). Microaerophillic FOB Since the deposition of iron-oxides typically takes place within the redox interfacial environments characterized by low dissolved oxygen concentration, the possibility exists that bacterial Fe 2 + oxidation and Fe 3 + reduction are tightly coupled in such environments and could even be performed by the same species in environments influenced by periodic fluctuations in the inputs of organic carbon and oxidants. For example, Geobacter sp. are capable of both dissimilatory Fe 3 + reduction and oxidation of Fe 2 + with reduction of N 0 3 " to N H 4 + (123, 141). As well, FOB appear to have the potential to induce coupling of aerobic Fe 2 + oxidation and reduction at aerobic -anaerobic interfaces by alternating bacteria. The juxtaposition of FOB and FRB and rapid microscale iron redox cycling within the Fe 2 + - oxygen reaction zone in co-cultures by Roden et al 2004, was hypothesized to maintain the majority of iron in the reduced state at the sand - water interface despite the detectable presence of oxygen (123). The result of FOB and FRB in the HZ: Adjacent aerobic and anaerobic zones such as that found in the HZ may promote the exchange of redox-sensitive metabolites like iron and NO3" between physiologically 25 diverse microorganisms. This will in turn produce an overall stimulatory effect on gross microbial metabolism and community diversity in the HZ (67, 95). It is therefore hypothesized that where steep gradients of Fe 2 + and Fe 3 + are seen in the Braid Street pore water, there should be a concomitant increase in microbial diversity adhered to the corresponding sediments. Microbial Community Studies in Hyporheic Zones It is now widely accepted that microorganisms primarily control redox chemistry in most environments. Microbial analysis will therefore confirm whether the potential for various redox processes is present in the plume at the Braid Street site and whether this potential will increase with diversity in the HZ portion of the plume. To my knowledge, only one previous study has related the microbial community diversity from a phylogenetic perspective to biogeochemical processes occurring in the HZ. This work conducted by Feris et al (2003), using a range of methods including DGGE, focused on determining the influence of heavy metal contamination on microbial community structure rather than investigating the role of such microbial groups in attenuating or mobilizing metal contaminants in the HZ (43). Feris established that heavy metal load was positively correlated with certain phylogenetic microbial groups, and inversely correlated with other phylogenetic groups (44). He also concluded that hydrogeological heterogeneity gives rise to seasonal variations in hyporheic habitats that are evident as a dynamic microbial community with marked seasonal variation. In brief, correlations to heavy metals were observed only in autumn and early winter, when organic matter deposition to streams is greatest. Feris et al. (2004) concluded that the abundance of organisms susceptible to heavy metals was at its greatest during the period of maximum bacterial growth (45). 26 The majority of HZ studies have focused on the physiological diversity of microorganisms in the HZ. Moser et al. (2003) employed a 'freeze core' technique to collect intact cobble-bed samples from the HZ of the Columbia River in Washington State in order to obtain information on the geochemistry and microbiology of HZ sediments. Here they discovered relatively large populations of SO4 2 - , NO3"-, and Fe 3 + -reducing bacteria; indicative of aerobic and anaerobic zones promoting the exchange of redox sensitive metabolites between a diverse consortia of microorganisms (95). Another study used a column experiment to delineate the biogeochemical activities in the HZ. Mermillod-Blondin et al. (2005) used slow filtration columns, composed of sand and gravel, to quantify small-scale biogeochemical processes and associated microbial activity within the HZ (91). Columns are shown to accurately recreate the aerobic - anaerobic gradient often observed in the HZ (20). They also reproduced the high heterogeneity of the HZ, with anaerobic portions in sediments enabling denitrification and fermentation processes (91). Overall, slow filtration columns were shown to be an appropriate tool to quantify in situ rates of biogeochemical processes and to determine the relationship between microbial activity and the physicochemical environment in hyporheic sediments. Thesis Overview Many questions still need to be addressed concerning the fate of a contaminant plume in an aquifer where river water ingression is occurring. The most prominent question that is addressed in this thesis is what redox changes are occurring along a contaminant flowpath, which affects the fate of PAHs, with the presence of river water ingression. Microorganisms primarily control changes in redox conditions in an aquifer; therefore this question is addressed via the analysis of geochemical conditions in conjunction with studying the microbial community of the aquifer at the Braid Street site. 27 It was hypothesized that at the Braid Street site the sediment-bacterial community structure will correlate with iron-speciation and creosote-compounds in the groundwater and that most of the diversity observed in the community structure will be found at the groundwater - river water interface. Here I have characterized the bacterial microbial community at the Braid Street site along the flowpath of the contaminant plume and at a point outside of the contaminant plume. I compared the bacterial community to the geochemistry of corresponding pore water, and applied previous knowledge of the site in regard to geomorphology, hydrogeology and geochemistry, to draw conclusions and new hypotheses on what biogeochemical processes are occurring at the Braid Street site. The experimental approach conducted included geochemical analyses of pore water, analyses of sediment using small subunit 16S rDNA, and a total microbial biomass analysis using phospholipids. 28 Chapter 2 . MATERIALS A N D METHODS Site Description and Sample Handling The Braid Street site was described thoroughly in the introduction of this thesis and procedures for sediment coring and pore water profiling have been described elsewhere (19, 125) and therefore are only described briefly here. The study site was located in New Westminster, British Columbia, Canada along the north arm of the Fraser River (Fig. 1). Five cores of two-inch diameter and 1.5-m in length were collected from four zones in the aquifer termed the surface zone (SZ), hyporheic zone (HZ), bottom of hyporheic zone (BHZ) and the anaerobic zone (AZ). The location of each zone including their depth and distance from the shoreline is depicted in Figure 3. Cores for sediment characterization in the HZ, BHZ and A Z were collected from a sampling vessel during freshet in 2004 and 2005 using a freezing-shoe drive point piston corer developed at the University of British Columbia by the Earth and Ocean Science department based on previous designs (100, 132). Furthermore, commercial divers were hired to collect cores in the SZ using a push-corer during freshet in 2006. Each core was cut vertically into 40-cm sections onboard the sampling vessel, which were capped under a stream of N2 and placed on ice until they were brought back to the lab. In an anaerobic glove box, the 40-cm sections were cut further into 10-cm sections vertically and sediment was sampled from each section using a sterile technique to collect the sediment and to not touch the walls of the core casing. The top depth of each 10-cm section was used to describe individual sediment samples and depths are described in meters below the riverbed (mbrb). Each sediment sample was homogenized 29 and split into two Falcon tubes for geochemical analysis and microbial characterization. Additionally, a sample from each 10-cm section of a zone was homogenized for enrichment culture inoculums. The Falcon tubes were placed into a -80°C freezer until directly preceding sediment collection for microbial analysis, and enrichment cultures were inoculated immediately proceeding sediment collection. Pore water profiles were collected using a Waterloo Drive Point Profiler (114). A minimum of three total volumes were purged from the system before sample collection for cations, anions and dissolved gases. Purging continued until parameters monitored in the flow-through cell (pH, conductivity and temperature) had stabilized. Bieber (2003) describes these analyses in more detail (19). Geochemical Analysis Cations, anions, total iron and Fe Analysis of cations was performed using an inductively coupled plasma mass spectrometer (ICP-MS) at A L S environmental in Vancouver, BC. Additionally, A L S measured anions in the pore water using an ion chromatograph (IC). Dissolved gases were analyzed at U B C using gas chromatography (GC) using methods and calculations described in detail by Bieber, 2003 (19). Analysis of total iron, Fe 2 + and dissolved oxygen (DO) were performed on the sampling vessel using a H A C H DR/2400 spectrometer that measures the absorption in the visible light spectrum, where chemical methods and instructions for analyses are included with the manual of the spectrometer. Prior to analysis of iron, samples were diluted by a factor of 4 to reduce the concentrations to within the dynamic range of the spectrometer. 30 The analysis of the PAHs indane, benzothiophene and naphthalene in sediment and pore water at the Braid Street site was developed by the Microbiology department at U B C and are described here in detail. Pore water PAHs Water samples were collected in 42-ml screw-cap vials. Initially, 7.0 ml of water was removed from each vial, and the vials were weighed. An internal standard consisting of 1.0 ppm each of indane, naphthalene, benzothiophene, 3-fluorotoluene, and 2-fluorobiphenyl dissolved in 2.0 ml of dichloromethane was added to each vial. The vials were reweighed to confirm the mass of solvent added. The vials were shaken at 250 rpm for 2 h and allowed to settle for 60 min. The dichloromethane layer (in the form of globules coated with amorphous Fe 3 +OOH) was removed from the bottom of the vial by pipetting, and placed in an 8-ml tube, with a PTFE-faced rubber liner screw cap to minimize adsorption of PAHs to the cap. Excess water was removed by pipetting, and the tubes were centrifuged (2000 x g) for 4 min. The remaining water layer was removed. The dichloromethane samples were placed in 2-ml autosampler vials and analyzed on an Agilent 6890 gas chromatograph with a 5973N mass selective detector (GC-MS), using a 30-m HP-5ms column, 250 um diameter, 0.25 urn film thickness. The carrier gas was helium, at 1.0 ml/min. The inlet temperature was 260°C, and the oven temperature was started at 40°C for 1.0 min, and then was ramped at 15°C/min to 115°C, then 5°C/min to 125°C, then 20°C/min to 220°C, then 30°C/min to 280°C. The transfer line temperature was 280°C. The MS was run in EI mode, scanning 50 - 400 m/z. Analytical standards were run to create standard curves. Furthermore, extraction standards, using 35 ml of water spiked with standards in methanol, were extracted by the 31 same method to determine extraction efficiency. A l l calculations used 2-fluorobiphenyl as an internal standard. Sediment PAHs The analysis of sediment PAHs was developed by the Microbiology department at U B C and is described here in detail. Initially 2.0 g of wet sediment were weighed into an 8-ml screw cap tube. The sediment PAHs were then extracted with 1.0 ml of acetone (each extraction requires a quick vortex and 15 min of soaking). The mixture was centrifuged and the supernatant was saved. This procedure was repeated twice and the supernatant was pooled. The sediments were then extracted an additional time with acetone, where they were left to soak overnight in the acetone. The following day the last acetone extraction was pooled with the previous and the solvent pool was weighed. Additional extractions were performed four times with 1.0 ml of hexane, where each extraction was left to soak overnight before centrifugation and pooling. The hexane solvent was weighed and added to the acetone solvent. At this point, the internal standard 2 fluorobiphenyl was added to the pooled solvent. To assure that the maximum amount of PAHs were extracted from the sediment, an additional extraction with 2.0 ml of hexane was performed, left stationary for 3 days. The supernatant here is kept separate from the pooled supernatant. The sediments are dried in the fume hood for several days, and then weighed. The volume of solvent was calculated by measuring the acetone and hexane weights. The P A H concentrations were determined by a GC-MS using the same oven temperatures and calibrations as for the water samples, and were calculated relative to internal standard for both the pooled solvent and the final hexane extract. The values of sediment PAHs, where the final extract had <2% of the total concentration, were reported. 32 Microbial Biomass Analysis Microbial biomass in the sediment samples was estimated by two methods chloroform fumigation extractions (CFE) and phospholipids fatty acid analysis (PLFA) and is briefly described here. Chloroform fumigation extraction CFEs were performed using the method described by Basiliko et al, 2005 (12). Phospholipid fatty acid analysis PLFAs were extracted from 8 g of sediment using the original method published by Frostegard et al, 1991 (51). Prior to F A esterification 200 ul of the P L F A 19:0 was added to the extract, following the silica column collection, at a concentration of 25 ppb. The F A methyl esters were dissolved in 20 ul of hexane and analyzed on an Agilent 6890 gas chromatograph with a 5973N mass selective detector (GC-MS), using a 30 m HP-5ms column, 250 um diameter, 0.25 um film thickness. The concentration of the cyl7:0 P L F A was calculated by integrating its peak in relation to the 19:0 internal standard. D N A Extraction and PCR Amplification D N A extraction Extraction of D N A for DGGE analysis was performed using the Fast+ method described by Webster et al, 2003 with slight modifications (142). Recently, this analysis has been implemented on subsurface sediment collected from the deep-sea margin of the ocean (52, 105, 110, 126, 143). The rich organic carbon observed in these sediments can be compared to the subsurface sediments at the Braid Street site. Due to the apparent importance of implementing this method to extract D N A from subsurface sediments, this method is described here in detail. The FAST+ method is the same as FAST method found in the Biol01" manufacturer's protocol, but with the following modifications: addition of 200 u,g of poly-adenylic acid (poly A) to the lysis mixture (64), extended spin and matrix binding times and elution of the crude D N A extract in 100 ul of sterile distilled water (SDW). D N A extractions were carried out on five 1-g replicate samples with the exception of enrichment cultures where no replicates were used. Replicate crude D N A fractions were pooled, purified and concentrated in a YM-30 Microcon centrifugal device (Millipore, Bedford, M A , USA) by washing three times with 200 (j.1 SDW, making sure not to dry the column, and eluting in 30 LII SDW. PCR amplification The PCR reaction mixtures consisted of the following: 33.95 u.1 of autoclaved, filter-sterilized SDW, 0.1 ul of forward and reverse primers targeting the 16S rRNA gene (each 100 uM), 1-3 ul of M g C l 2 (50 mM), 10 u.1 of 5x buffer (Stratagene; La Jolla, CA), 1.25 ul of deoxynucleotides (10 mM each), 0.6 u.1 of the proprietary Pfu-based Herculase®II enzyme, and 1 u.1 of unquantified D N A extract in SDW from sediment samples or excised bands. The amplification cycling was based on the recommendations given by Feris et al, 2004. A touchdown procedure was used to amplify D N A from the sediment samples or bands excided from D G G E gels using the following cycles: 95°C for 1 min followed by 94°C for 45s, 65°C for 45 s and 72°C for 1 min, repeated for 9 cycles. This was followed by 25 cycles of 94°C for 45 s, 55°C for 45 s and 72°C for 1 min, the latter increased 3 s per cycle, resulting in an extension time of 12 min in the final (twenty-fifth) cycle. f Bacterial Community Analysis 34 Denaturing gradient gel electrophoresis Denaturing gradient gel electrophoresis (DGGE) analysis of partial 16S rRNA gene sequences was performed on PCR amplified samples using primers targeting the V3-V5 region (357F-GC, 5'-C G C C C G C C G C G C C C C G C G C C C G T C C C G C C G C C C C C G C C C G C C T A C G G G A G G C A G C A G - 3 ' , and 907R-cy5, cy5-5'-CCGTCAATTCMTTTGAGTTT-3') . This is an extension of the region recommended by Muyzer et al, (1993), so that bands cut from the gel could be successfully queried in the NCBI database (101). Additionally, the fluorophore cy5 was linked to the 5' end of the reverse primer so that a low quantity of D N A (approximately 100 ng) could be visualized on the D G G E gel (103). Directly following amplification, 10-15 ul of the PCR mixture generated from each sample was separated by D G G E with the Bio-Rad D-GENE system (Bio-Rad Laboratories, Hercules, CA). The gels were run and analyzed using the method described by Neufeld and Mohn, 2005, with the addition of a pre-run at 85 V for 20 min, as recommended by Feris, 2003 (43, 103). D G G E band sequencing Bands were cut from the DGGE gel using a sterile razor blade. Each excised band was placed in a microcentrifuge tube containing 20 ul of SDW, vortexed and left at 4°C overnight. The following day, an additional 20 ul of SDW was added to the tube, which was vortexed. This solution was used as template for the above PCR. The PCR amplicons were sequenced using primers 357F and 907R and Sanger sequencing at the McGi l l University and Genome Quebec Innovation Centre (Montreal, Quebec, Canada) 35 D G G E fingerprint analysis Cluster analysis was performed on the fingerprints using Pearson correlation coefficients with an unweighted pair group method and arithmetic mean (UPGMA) algorithm. In addition to clustering, the integration of densitometric curves from bands that were manually picked in the gel was used to estimate the relative abundance of phylotypes. Enrichment Cultures Fe3+-reducing enrichment cultures were initiated from the Braid Street site sediments. The enrichments were amended with 25, 12.5, and 12.5 mg, respectively, of naphthalene, benzothiophene plus indane as carbon sources and electron donors. The PAHs were mixed and added to serum bottles with a sterile syringe in the presence of 10 g of oven-dried solid adsorber resin Amberlite X A D 7 (Fluka, Buchs, Switzerland), which was used to reduce the toxicity of the PAHs (94). Fe 3 + was added as an electron acceptor in the form of an amorphous Fe 3 + OOH or as 100 m M of Fe 3 + chelated with nitrilotriacetic acid (NTA) trisodium salt at a ratio of 1:9 with the sediment inoculum (86). The enrichments were made up in 125-ml serum bottles filled with 100 ml of bicarbonate-buffered mineral medium (pH 7.3) and a trace metal solution (87). The serum bottles were capped with Teflon stoppers due to PAHs readably adsorbing to rubber stoppers (5, 90). The medium was bubbled out with N 2 / C 0 2 (80:20) at 1 atm total gas and was reduced with 1 mM FeCh prior to autoclaving. Each culture was then inoculated with 20 g of sediment from their respective zones: the HZ/BHZ (HZ-Fe (III)); the A Z (AZ-Fe (III) + NTA) and a core further out in the HZ that was not analyzed in this thesis aside from the enrichment cultures (HZ-2-Fe (III)). Following a 426-day lag phase it appeared that the amorphous Fe 3 + in the cultures had been reduced, where a lack of reddish-brown color was observed in the cultures. At this point, sediment was collected for bacterial community analysis. Clone Library Construction and Analysis 36 Clone library construction Clone libraries were constructed using the 16S rRNA gene fragments amplified by the above PCR procedure using primers B27F ( 5 - A G A G T T T G A T C C T G G C T C A G -3') and U1492R (5 ' -GGTTACCTTAGTTACGACTT-3 ' ) obtained from the sediments in the aquifer. Amplicons were purified using the MinElute kit (Qiagen, Valencia, CA) according to the manufacturer's instructions with an extension of incubation times. For the A Z sample it was necessary to purify the 16S amplicon using the QIAquick Gel extraction kit (Qiagen, Valencia, CA) according to the manufacturer's instructions. Herculase®II does not add the deoxynucleotide adenosine to the ends of the purified amplicons, which is required for ligation of amplicons into the pCR®2.1-TOPO® plasmid used in the TOPO®TA® Cloning kits. Therefore prior to cloning 5 ul of the PCR amplicon, it was incubated at 70°C for 30 min with 2.45 ul of autoclaved and filter sterilized SDW, 0.3 ul of M g C l 2 (50 mM), 0.25 ul of deoxynucleotides (10 m M each), 1 ul of lOx buffer (Invitrogen, Carlsbed, CA), and 1 ul of taq D N A polymerase (5 units/ul). Cloning proceeded using the TOPO®TA® Cloning kit according to the manufacturer's instructions. Colonies containing transformed plasmids were randomly picked and transferred to 96-well plates containing 180 ul of LB K A N 50 and 10% glycerol for restriction-fragment length polymorphism analysis. Unique phylotypes were picked from the clone libraries and sequenced for phylogenetic analysis according to methods described by Stilwell, 2007(133). Phylogenetic analysis The 16S rRNA genes of the most abundant phylotypes (> 2% of library) were i used to construct a phylogenetic tree. The tree was constructed using sequences aligned 37 with the RDP aligner. To begin, a distance matrix was generated using the Jukes-Cantor corrected distance model (72), where only alignment model positions were used. The tree was then created using the Weighbour method, which gives less weight to significantly longer distance, with alphabet size 4 and length size 1000 and bootstrapping with 100 replicates (27). Diversity index The Shannon diversity index (H') and its related equitability index (E) were calculated, which take into consideration both the total number of observed phylotypes (S) and the relative abundance of each phylotype (p,) (49). 38 Chapter 3 RESULTS A N D DISCUSSION Geochemistry Pore water contaminants Bieber (2003) determined that indane, benzothiophene and naphthalene were the dominant contaminants of the Braid Street plume in 2002 and 2003 (19). The contaminant concentrations determined by Bieber were compared to concentrations measured in 2005 and 2006 to get an idea of the development of the Braid Street plume. The concentrations determined in 2005 and 2006 can be found in Table 1. Graphs illustrate the contaminant peaks in Figure 4, and the locations where pore water profiles were collected are illustrated in Figure 3. In 2002 the naphthalene concentrations at the Braid Street site peaked approximately 40 m offshore and 15 mbrb at 2,177 ppb, and in 2005 at approximately the same location significantly decreased to 42 ppb. In the plume where discharge is expected to occur at 90 m (HZ/BHZ) and 110 m (SZ) offshore there was no naphthalene detected in 2005 and 2006 respectively, whereas Bieber detected 5 ppb at about 80 m offshore in 2002. Bieber determined that indane and benzothiophene dominated the plume near the discharge point in June 2002. In 2003 indane concentrations were measured as high as 873 ppb in the core of the plume and at 348 ppb close to the discharge point. Bieber suggests that a significant amount of indane might discharge into the river near at the discharge point. When the aquifer was sampled in 2005, indane peaked at 640 ppb compared to the previously recorded 873 ppb in the A Z . Relative differing levels in 39 indane concentrations compared to naphthalene concentrations between 2003 and 2005 suggests that the attenuation of indane is a lot slower than that of naphthalene, with possibly no decrease in terms of total mass. Generally, indane peaks were found in deeper regions than those of naphthalene peaks (Fig. 4), and therefore these two pollutants disperse differently in the plume. Whereas benzothiophene peaks at the same depth as naphthalene (Fig. 4), and was measured to be 82 ppb at 90 m offshore in 2005 and 140 ppb at 75 m offshore in 2003. When comparing contaminant concentrations measured in 2003 to concentrations measured in the A Z in 2005, we see 1.9% of naphthalene, 58.5% of benzothiophene, and 73.3% of indane remaining in the plume in 2005. This is consistent with previous studies of creosote degradation based on natural attenuation in anaerobic aquifers, where indane is reported to be recalcitrant, benzothiophene less so and naphthalene degradable. The offshore component in the HZ/BHZ in 2005 is not comparable to Bieber's measurements in 2003 due to different sampling points. However, i f peaks are compared, we saw 0% of naphthalene, 11.5% of benzothiophene, and 22.7% of indane remaining, which still signifies indane as being more recalcitrant than benzothiophene and naphthalene. Samples were taken from the HZ portion of the plume, where discharge is expected to occur, in 2005 and 2004. In comparing the profile from the 2004-HZ (Table 1) to the 2005-HZ (Table 1, Fig. 4), we see approximately the same concentrations in the 2004 and 2005 profiles, where 79 ppb was measured for indane and 10 ppb for benzothiophene, and the concentrations in 2004 were 72 and 6 ppb respectively (Table 1). The slightly higher concentrations could be due to the profile being taken slightly further offshore in 2005, where groundwater would appear to be moving up because peaks are at higher depths in 2005 than in 2004 (Table 1). It should be noted that in the 2005 pore water profiles we see peaks in the mass spectrum of 5-indanol, mesitol and o-toluidine along with other amines that are 40 frequently greater than indane peaks. In particular 5-indanol was a dominant peak in many of the profiles, where this compound might originate from the hydroxylation of indane. It has been suggested that anaerobic degradation of PAHs under reducing conditions is thought to occur via an initial hydroxylation reaction (90). This could indicate that the decrease in indane concentration observed in the HZ relative to the A Z is due to microbial metabolism, and not to mass loss via discharge into the river or complete biodegradation. Contaminant sorption Contaminant peaks in a pore water profile can be compared to contaminant peaks in the sediment to get an idea of how much contaminant is being adsorbed to the sediment (Table 1 and Table 2). A pore water profile and sediment core were taken from the HZ in 2004, where indane, benzothiophene and naphthalene were measured both in the pore water and on the sediment. Indane peaked at 1.83 mbrb in the sediment (Table 2) and was aligned to the peak of indane in the pore water at 2.4 mbrb (Table 1). Realignment of sediment cores with pore water profiles revealed 2.77% of indane in the pore water where the concentration on the sediment was 2524 ppb. This concentration of indane sorbed to the sediment was comparable to the concentration Bieber measured in the A Z pore water in 2003. 0.25 m below where the indane concentration peaks on the sediment, 7.13% of indane was measured in the pore water. Furthermore, 0.6 m above the indane concentration peaks on the sediment, 4.28% of indane was found in the pore water. When comparing the concentration of benzothiophene on the sediment relative to the pore water, 1.5% of benzothiophene was found in the pore water. Naphthalene was not comparable because no naphthalene was detected in the groundwater at this point in 41 the aquifer. However, it should be noted that 290 ppb of naphthalene was measured on the sediment. These data reveal that there is still naphthalene and benzothiophene contamination existing on the sediment where it cannot be measured in the pore water; therefore it is possible that biodegradation processes are still degrading the contaminants in the aquifer at these points, but degradation may be limited by bioavailability. Where neither naphthalene nor benzothiophene are observed in the pore-water, more benzothiophene is sorbed onto the sediment than naphthalene (Table 2). Although these comparisons will not yield quantitative information on the sorption - desorption processes in the aquifer at the Braid Street site, they are useful to get an approximate idea of how strongly the contaminants sorb to the aquifer matrix. Iron: dissolved iron, Fe 2 + and total iron It is generally observed that Fe 2 + travels at the same velocity as groundwater along a contaminant plume flowpath, where relatively high concentrations of iron should indicate the preferential flowpath of groundwater through a point in the aquifer. Furthermore, dissolved iron concentrations are indicative of reductive dissolution of Fe 3 + by microorganisms oxidizing available carbon sources (30, 31). PAHs are an example of readily available carbon sources in a contaminated aquifer, therefore high dissolved iron concentrations are often observed with high concentrations of PAHs. However, cycling of iron by bacteria, ion exchange and precipitation could maintain elevated concentrations of iron in the groundwater after the redox processes no longer are active (e.g., Albrechtsen et al., 1995) (4). Therefore the presence of dissolved iron in the sampling point indicates that reduction is ongoing, or has happened, at the point or upgradient of the point (32). 42 Dissolved iron concentrations were sampled at Braid Street to delineate the flowpath of groundwater in the contaminant plume. As well, concentrations of Fe 2 + and total iron (which is used to calculate the concentration of Fe 3 +) were measured at the same points as dissolved iron to delineate redox gradients in the contaminant plume. An environment with a higher concentration of Fe 3 + is expected to be more oxidized than one with a lower concentration. Furthermore, dissolved iron concentrations were compared to concentrations previously measured at the site to indicate any differential activity in the contaminant plume. For example, the lower dissolved iron concentrations detected in the aquifer in 2005 compared to 2003 might indicate less reductive dissolution of iron in the A Z , and could be attributed to a receding plume. Iron concentrations sampled along the plume axis by Bieber varied between 21 and 112 ppm in June 2002, and in February 2003 between 23 and 95 ppm (19). The greatest iron concentrations were observed in the profiling point containing the most naphthalene and in two shallow sampling intervals where no contamination was detected. In the 2006 pore water profile of the SZ, dissolved iron concentrations were between 62 and 90 ppm, Fe 2 + between 55 and 91 ppm and total iron between 62 and 90 ppm (Table 1). The relationship of dissolved iron to total iron here is nearly 1:1 throughout, indicating little Fe 3 + production. Furthermore, the SZ is where the highest concentration of iron was observed is a located in a region outside the contaminant plume - no contamination was detected in the pore water here (Table 1). The high iron concentrations in shallow intervals are consistent with data collected in 1999 and 2002. Bieber suggested that these high iron concentrations could be attributed to degradation of organic matter by Fe 3 + reduction (19). In the HZ/BHZ dissolved iron concentrations between 8 and 56 ppm, Fe 2 + between 29 and 54 ppm and total iron between 32 and 171 ppm were observed in 2005 (Table 1). Here the peaks in dissolved iron, Fe 2 + and total iron correlate with peaks in 43 indane (Fig. .4). In the BHZ profile below 2.5 mbrb, Fe 2 + and total iron concentrations are nearly equivalent, and the concentration of indane decreases to 3 ppb (Table 1). At 1.3 and 1.9 mbrb in the HZ, the ratio of total iron to Fe 2 + is 3.33 and 2.08, and the ratio of Fe 3 + (calculated from subtracting the concentration of Fe 2 + from total iron) to indane is 1.48 and 0.96 respectively. Whereas at 4 and 4.6 mbrb in the BHZ the ratio for total iron to Fe 2 + is 1.10 and 1.08, and the ratio of total iron to indane is 8.8 and 16.25 respectively. This suggests that there is a different relationship between iron and indane in the HZ than the BHZ, where, this relationship increases by a factor of 2 with a increase in depth at these points. In the A Z , dissolved iron concentrations were detected between 17 and 72 ppm (Table 1), where these numbers are lower than those seen by Bieber in the same area. 2"i_ We see Fe concentrations between 25 and 71 ppm and total iron between 37 and 78 ppm (Table 1), where peaks in dissolved iron correlate to peaks in indane concentration (Fig. 4). In the A Z at 7.9 mbrb the ratio of total iron to Fe 2 + is 2.2, and the ratio between Fe and indane is 1.16, which is a similar relationship to that observed in the HZ. Conversely the ratio of total iron to Fe 2 + is 1.15,1.09 and 1.11 at depths 10.6,11.58 and 12.5 mbrb respectively, where the relationship of total iron to indane is 0.48, 0.18 and 0.09 respectively. The total iron to indane ratio is decreasing by a factor of 2 with respect to these depths in the A Z , which is inversely proportional to the relationship seen in the B H Z profile. The relationship of total iron to indane seen here could be attributed to increasing saturation of dissolved iron from reductive dissolution and concomitant precipitation of iron with depth. Under the circumstances that sorption - desorption of indane is at equilibrium in the aquifer, and that indane travels at the same relative speed to dissolved iron along the contaminant flowpath in the aquifer, an inference can be made concerning the degradation of indane from the result of measured iron concentrations. In particular, it 44 can be hypothesized that indane is being rapidly attenuated at the points in the BHZ where the concentration of Fe 2 + is much greater than indane. Furthermore, indane would be attenuated by a different process in the HZ leading to a concomitant increase in Fe 3 + . As well indane could be degraded in the A Z but rates would be slower than that found in the HZ with exception of the community at 7.93 mbrb, which is located on the fringe of the plume in the A Z , and represents conditions found in the HZ. It should be noted that in the mass spectra of the pore water profiles along the contaminant flowpath indanol is a dominant peak (data not shown). Therefore it appears more likely that indane attenuation is attributed to transformation to indanol rather than mineralization. pH, Alkalinity, Temperature, Conductivity. Metals and Salts Fe reduction and methanogenesis are thought to be the predominant TEAPs in the aquifer at the Braid Street site. Fe 3 + reduction coupled to naphthalene degradation is acid consuming, and therefore will have an effect on pH. Furthermore, methanogenesis is CO2 consuming and will decrease the buffering capacity that high alkalinity provides, and in concert with Fe 3 + reduction, may act to raise the pH further. In General, Fe 3 + reduction is acid consuming (equation 1), whereas Fe 2 + oxidation is acid producing (equations 2 and 3) (32): Therefore measuring pH could give an indication of the reactions occurring in the aquifer. The pH values measured ranged from 6.6-6.8 in the SZ, from 6.1-7.2 in the HZ/BHZ and from 6.3-7.2 in the A Z (Table 1). The pH values negatively corresponded with iron concentration, except for in the SZ where no relationship could be determined. Fe(OH) 3-X + e- -> Fe 2 + + 30H" + X 4Fe 2 + + 0 2 + 4 H + -> 4Fe 3 + + 2 H 2 0 4Fe 3 + + 12H 20 -> 4Fe(OH)3 + 12H + (1) (2) (3) 45 Oxidation of 1 mole of Fe 2 + could require 1 hour at pH 7, while this reaction at the same temperature and partial pressure of oxygen could take 100 hrs at pH 6 (30). Therefore, these small changes in pH observed in the aquifer should be considered. Where lower pH values correlate to higher dissolved iron values, this will mean that FOB will be able to compete with abiotic factors for the thermodynamically favorable oxidation of iron (141). Along with iron oxidation and reduction reactions, pH can be controlled by the ingress of water into the aquifer (15), where the river water pH is above 7 and the groundwater pH from inland is about 6.2 (6). River water ingression might explain why the pH values were consistently above 7 in the BHZ profile below 3.5 mbrb (Table 1). Total alkalinity (CaCOa) was not measured in 2005. However in the 2004-HZ profile, alkalinity positively corresponds with iron and negatively corresponds with pH. In the 2004-BHZ profile, alkalinity negatively corresponds with iron and positively corresponds with pH. Whereas further offshore in the SZ, alkalinity positively corresponds with iron and pH (Table 1). Conductivity values ranged from 466 to 3470 uS/s in the SZ, 382 to 475 LIS/S in the HZ, 519 to 1281 in the BHZ and 309 to 681 uS/s in the A Z (Table 1). Conductivity in the B H Z increased dramatically with depth beginning at 4 mbrb, which intersects the depth that iron reaches its minimum concentration (Table 1). High conductivity values corresponded to high dissolved Ca and Cl" concentrations suggesting that values are from groundwater mixing with saltwater from a deep region of the aquifer. This saltwater is determined to be present in the deepest regions of the aquifer, and it has been suggested that saltwater is coming from regional groundwater discharge beneath the plume that is most likely associated with the underlying lower bedrock and sandy silt (19). In an alternative theory, Bianchin suggested the possibility of it being a pool of saltwater that has been preserved from the ocean originally occupying this location of the river (18). 46 The measured temperature in the profiles appears to increase with increased iron concentrations in the HZ/BHZ and is likely attributed to the time it took for pore water to be collected (Table 1). Samples were collected in the summer and therefore the increase in temperature was linked to an increase in time for sample collection. Measuring temperature in aquifers could indicate the presence of groundwater mixing with surface water (131). A n in depth analysis via temperature profiling has been carried out by Bianchin (18). Br" was detected in areas only where dissolved CI" was above 30 ppm, and the highest Br" concentration was found in the SZ at 3 ppm. This suggests that Br originates from weathering processes in the regional groundwater (Table 1). Redox, hydrodynamic and sorption processes along a flowpath will induce changes in the concentration of other metals in addition to dissolved iron. Metals were compared in the HZ/BHZ to the A Z (Table 1). Higher concentrations of dissolved Zn and As were observed in the HZ/BHZ compared to the A Z and corresponded with dissolved iron concentrations in the HZ, which might be attributed to the slightly lower pH observed here (Table 1). Conversely, higher concentrations of dissolved Ba, Si, Na and Mg were observed in the A Z . Dissolved Ba, Si, and Na corresponded with iron until below 9.25 mbrb, which intersects the point where a lower value in total iron is measured. Dissolved Ca was observed in similar concentrations in the HZ/BHZ and A Z , and also corresponded with dissolved iron until below 9.25 mbrb. Furthermore, dissolved Ca increased dramatically and negatively corresponded to dissolved Na below 12 mbrb in the A Z . The higher dissolved Ca concentration observed here is likely attributed to ingress of groundwater from the regional groundwater zone. Saturation indices are normally calculated with the measurement of alkalinity, pH and temperature, to determine what minerals are likely present in an aquifer at equilibrium. However, 4 7 alkalinity was not measured at the time of this sampling date, therefore little information can be deduced from these data. Collectively, these data suggest that the four zones (SZ, HZ, BHZ and AZ) are made up of groundwater originating from three different sources in the aquifer. For example, the pore water in the SZ primarily originates from the regional groundwater; that in the HZ originates from contaminant groundwater plus river water; that in the BHZ originates from a convergence of river water plus contaminant groundwater near the top, and river water ingress plus regional groundwater near the bottom; and lastly, that in the A Z , originates mostly from contaminant groundwater mixed with regional groundwater near the bottom. Nutrients Nutrients such as NO3" and SO42" can be measured in the aquifer pore water. However, their presence does not indicate their use in TEAPs. Furthermore, when NO3" or SO4 " is not detected where their presence is expected, this could indicate denitrifying or SO4 "-reducing conditions in these zones (32). In freshwater sediments NO3" is generally expected to dominate over SO42" concentrations. NO3" was measured in one sample in the SZ at 1.83 mbrb and in two samples in the HZ at 0.99 and 1.91 mbrb, where the concentrations were 0.03, 2.5 and 0.4 ppm respectively (Table 1). The concentrations in the HZ in 2005 correlate to the same area where NO3" was detected in 2004, ascertaining that these values are not from contamination of nitric acid. In 2002, Bieber measured concentrations up to 7 ppm in low naphthalene zones and as high as 2.7 ppm in high naphthalene zones, which were equivalent to the concentrations measured by Bianchin in 1999. Bieber suggested that the concentrations in the high naphthalene zone were likely due to contamination by the metal preservative 48 nitric acid (19). However, these concentrations could be attributed to the reversal of the groundwater gradient during freshet carrying NO3" into these zones from its production in more aerobic zones, or carried in with the ingression of river water. Furthermore, the production of NO3" by nitrogen-cycling microorganisms in high naphthalene zones under microaerophillic conditions could be occurring. In low NO3" HZs, the mineralization of organic nitrogen may be the primary source of N 0 3 " for denitrifying bacteria (38), where ammonification and nitrification become closely linked in sub-surface flow lines (70). Microaerophillic microbes including bacteria and archaea can oxidize N H 4 + to produce NO3", NO2" or N2O whereas anaerobic microbes can oxidize N H 4 + and reduce NO2" to produce N2 (47). Furthermore, Clement et al 2005, proposed that microbes can oxidize N H 4 + using Fe 3 + iron as a TEAP to produce N 0 2 " (33). N0 2 " , a product of NO3" reduction and N H 4 + oxidation, was not observed in any of the samples, but is rarely found to accumulate in the environment (32). N H 4 + oxidation is thought to be the rate-limiting step in most environments where nitrogen cycling is dominant (47). Furthermore, Weber et al (2006) observed N H 4 + to be the stoichiometric end product of NO3" reduction coupled to Fe 2 + oxidation in wetland-sediment inoculated enrichment cultures (141). Therefore it is recommended for further studies conducted at the Braid Street site to measure N H 4 + to elucidate nitrogen redox cycling. S0 4 2 " was measured at one depth, at 1.52 mbrb in the SZ (Table 1). However, Fe -S (pynte) was seen bound to minerals at depths up to 1.8 mbrb in samples using SEM, and therefore S0 4 2 " reduction could be a present or previous TEAP at the Braid Street site. Furthermore, microorganisms can oxidize the sulfur component of pyrite with the reduction of Fe 3 + to produce S0 4 2 " that could potentially continue the sulfur redox-cycle in systems with high Fe 2 + concentrations. Sulfide oxidation is an acid-producing reaction and can have a profound effect on the pH of the system (127). 4 9 It should be noted that nutrients are difficult to measure in samples with a high concentration of salts. The presence of salts can increase the detection limit to above a significant concentration of NO3", NOV or SO42". In the samples at the Braid Street site a detection limit up to 2 ppm was observed in some of the samples collected along the . regions of the contaminant plume that are expected to be mixing with regional groundwater. Dissolved Gases CO2 can be produced from heterotrophic oxidation of PAHs both aerobically and anaerobically in the aquifer and from the fermentation of smaller organic compounds (30). Furthermore, it is well known that CO2 can be produced from the aerobic oxidation of CH4, and by anaerobic oxidation of CH4 coupled to SO42" reduction (137) and more recently to NO3" reduction (117). CH4 can be produced by heterotrophic degradation or fermentation of organic carbon, as well as by methanogens that use CO2 as a T E A producing CH4 (30). The consumption of C 0 2 is often correlated to autotrophic and mixotrophic metabolism by microorganisms in a system. A n increase in the amount of CO2 can have an effect on pH. An excess of CO2 will lower the pH while increasing the buffering capacity (alkalinity) of the system. Conversely, the consumption of CO2 will decrease the buffering capacity and the system will be more susceptible to changes in pH (32). At the Braid Street site dissolved CO2 was found in concentrations of 80 to 146 ppm in the HZ/BHZ and from 123 to 238 ppm in the A Z (Table 1). Concentrations of dissolved C H 4 ranged from 11 to 15 ppm in the HZ/BHZ and from 6 to 11 ppm in the A Z (Table 1). Dissolved gases were not measured in the SZ. CO2 inversely corresponded to C H 4 in the A Z where they intersected at about 9.5 mbrb, where C H 4 corresponded positively with iron above this depth and negatively with 50 iron below this depth. In the HZ, CO2 and CH4 are more difficult to relate. It appears that the concentration of CH4 increases while CO2 decreases in the HZ, whereas both decrease in the BHZ portion of the profile. A reasonable explanation for the decrease of CO2 corresponding to the increase in total iron in the A Z and HZ is chemoautolithotrophy or mixotrophy by lithotrophs who oxidize Fe 2 + and consume C 0 2 at these points in the aquifer (140). The elevated levels of CH4 here could be explained by the influx of groundwater to these points, which originates from the contaminant source location where methanogenesis is more likely a predominant mechanism (19). Furthermore, the negative relationship of C H 4 to CO2 with a corresponding increase of Fe 2 + in the A Z could indicate the presence of a symbiotic consortium of methanotrophs and dissimilatory Fe reducers, a mechanism proposed by Daniel et. al (36). In the BHZ portion, the decrease in CH4 and CO2 is easily explained by the hypothesized convergence of groundwater from the contaminant source with river water ingression at this point. Additionally at 4 mbrb, where groundwater from the regional discharge is expected to dominate the pore water, C H 4 , CO2 and iron begin to rise in concentration. N2 and Ar were also measured at the Braid Street site (Table 1). Ar is not reactive whereas microbes can fix N 2 for a nutrient source or in contribution to a symbiotic relationship; most notably with plants (139). Concentrations of Ar and N 2 were proportional throughout the aquifer, this is indicative that N2 was not reactive in this system. 1 Dissolved organic carbon Measuring dissolved organic carbon (DOC) gives an indication of how oligotrophic the waters are, and is particularly useful for systems contaminated with 51 heavy metals (e.g. Moser et al). In general, DOC was around 8 ppm in the HZ/BHZ and 14 in the A Z (Table 1). Stratigraphy and 100-ml flow test The 100 ml-flow test is an indication of hydraulic conductivity. There was no consistent pattern relating this time with depth, but two peaks should be noted: one in the HZ, which is linked to where indane peaks and corresponds to pebbly sand at 1 mbrb in 2004 and 1.83 mbrb in 2005, and another in the A Z , which also is linked to where indane peaks and corresponds to pebbly sand with a clay and wood-debris mix at around 12 mbrb (Table 1, Table 4). An increase in time in the 100-ml flow test might be due to degassing of the profiler tip. Applying this concept would mean that there is an increase in porosity at the peaks observed in the HZ and A Z , and the preferential flowpath of groundwater through the system could be highest at these peaks. This would explain the peaks observed in high iron concentrations corresponding to peaks in contamination in the aquifer that occur at points with these peaks in the 100-ml flow test. Microbial Biomass Microbial biomass measurement on sediment samples gives insight to the productivity of communities in the aquifer, where more productive communities should yield a higher biomass. Chloroform fumigation extractions Chloroform fumigation extractions (CFE) measure the amount of labile carbon in a sample compared to a control of the same sample where biomass carbon is first fumigated. CFE on sediment from Braid Street site revealed an average microbial biomass carbon of 51.7 ug/g sediment from depths ranging from 1 to 2 mbrb. 52 This method is not widely used on sediment samples and therefore was only comparable to a study where CFE was used to measure biomass on tidal-flat sediments by researchers in Germany (69). The biomass numbers from Braid Street site were an order of magnitude lower than those found in tidal-flat sediments, where 903 u.g/g of microbial biomass carbon was found in the aerobic layer and 476 (ig/g was found in the anaerobic layer. The discrepancy between measurements could be due in part to extraction efficiency coefficients normally used to calculate microbial biomass, whereas none were used in my calculations. The researchers in Germany did not correlate labile-carbon to the textural classes of sediment that were analyzed (69). This concept could be extrapolated to the Braid Street site where the differences in stratigraphy did not yield an extensive difference in CFE biomass carbon (data not shown). Phospholipid fatty acid analysis Phospholipid fatty acid (PLFA) analysis is an alternative measure of viable biomass in sediment samples. PLFAs were used to analyze the biomass at the Braid Street site for reasons including: their rapid decomposition upon cell death denoting that only viable microorganisms will be measured, specific PLFAs are signatures for different groups and therefore this analysis yields insight to physiological processes occurring in the aquifer; and lastly PLFAs can theoretically be extracted from any sample matrix with a viable biomass. However, it is widely recognized that the high sorption capacity often associated with subsurface sediments makes extracting microbial cells and cell components for biomass or phylogenetic analyses a cumbersome procedure. This was a significant factor that limited the number of samples analyzed for the microbial community characterization in this thesis. From the sediment samples at the Braid Street site, PLFAs were successfully extracted and analyzed from four out of twenty samples located within the HZ and A Z . 53 Following numerous steps of concentrating and diluting samples, the signal to noise ratio on the mass spectrum of sixteen samples proved to be too high to accurately integrate P L F A peaks. Furthermore, whereas a compilation of twelve P L F A peaks is normally used to measure microbial biomass, only one peak was used in this analysis. Frostegard & Baath (1996) hypothesized that a single P L F A could be used to determine the biomass of a sample, but this would mean that all bacteria would have to contain the same relative amount of this PLFA. Furthermore, the proportions of individual PLFAs differ between groups of bacteria and could in turn differ between sediments. Therefore, the use of one P L F A is likely to yield variable results (50). The biomass of four sediment samples was determined from the integration of the P L F A cyclopropane hexadecanoic acid (cyl7:0) peak from the mass spectrum following the extraction of PLFAs using the method provided by Frostegard & Baath (50). This peak was chosen for biomass analysis based on its dominance in the four samples and absence from the negative control (refer to Appendix). The P L F A cyl7:0 generally accounted for 40% of the total amount of PLFAs in forest soils analyzed by Frostegard et al. Frostegard & Baath (1996) used a consistent group of 12 different PLFAs, including cyl7:0, to represent the bacterial biomass of soil samples (50). Two samples were obtained from the HZ at depths 1.06 and 1.26 mbrb, and two samples were obtained from the A Z at depths 9.67 and 10.18 mbrb. The P L F A peaks yielded concentrations of 0.13, 0.28, 0.08 and 0.18 nmol PLFA/g sediment respectively. The equations 1.4xl0"8nmol bacterial PLFA/cell (50) and 1.69xl0"8nmol bacterial . PLFA/cell (9) were used to convert the nmol of P L F A calculated from cyl7:0 to cells/g of sediment, where an average of 1.4x107 and 9.64x106 cells/g of sediment was determined using the respective equations. Frostegard & Baath (1996) state that only a fraction of the total biomass is found in the bacterial solution after the homogenization/centrifugation steps, and determined an 54 average recovery of 8.7%, which likely differs depending on the matrix and method used for recovery (50). The Bai et al (9) conversion assumes that bacteria contain a constant portion of their biomass as phospholipids under natural conditions, where Balkwill verified this theory (9, 10). Based on Frostegard's method and Bai's assumption, a liberal estimation can be made that the cells found in my sediment are 2 orders of magnitude greater than that calculated here. This would mean an average of l x l O 9 cells/g of dry sediment. This value is comparable to a study that Moser et al (2003), did on hyporheic sediments where they determined a concentration of 107 to 109cells/g (conversion factor 2.5xl0 4 cell/pmol of PLFA), and was roughly comparable to direct microscopic counts reported for other hyporheic systems where 5xl0 5 to 2xl0 8 , 2x l0 7 to 8 8 2.5x10 and 6.3 to 6.8x10 cells/g sediment were determined (95). Furthermore, values observed at the Braid Street site were slightly higher than those found in an uncontaminated subsurface aquifer sediment where 7xl0 6 cells/g dry sediment was calculated, and was comparable to a study done on P L F A analysis of anaerobic rice paddy soil, where 2.59 to 5.31xl0 9 cells/g dry weight was calculated from total P L F A analysis (9). Additionally, Webster et al (2006) determined 106 to 10 1 0 cells cm"3 by acridine orange direct counts from deep subsurface samples found in the Peru margin basin. Webster and his colleagues observe concentrations of high-molecular-weight D N A extracted from subsurface samples from up to 200 m below the seafloor in the Peru margin to be very low (<4 ng/g sediment), and were comparable to the concentration of D N A from Braid Street site samples (data not shown), where it appears that difficulties in extracting D N A can be correlated to difficulties in extracting cells for biomass (143). The average biomass carbon measured by CFE in my sediments correlated to a P L F A concentration of 2.3 nmol/g of total P L F A according to the equation determined by Leckie et al on humus soils (81). This further supports the assumptions made in 55 calculating microbial biomass from cyl7:0, where i f this P L F A were to make up 10% of the total P L F A concentration, this would equate to an average of 1.7 nmol/g of sediment. This number is comparable to the biomass found using CFE. The highest biomass was observed in the sample closest to the HZ interface and could be indicative of a more productive region of the aquifer, whereas the lowest biomass is seen in the sample at the fringe of contaminant plume in the A Z , where more oligotrophic conditions appear to be present (Table 1 and Fig. 4). Machaughton and colleagues observed the effect of the addition of a mixture of metal salts to a prokaryotic community microcosm consisting of sandy-loam soil. When the PLFAs of the microcosm were analyzed they observed that cyl7:0 made up 1.5% of two distinguishable Burkholderia-related cultures, and that 16:0 made up 20% of the P L F A profile (89). In the analysis performed at the Braid Street site, a 16:0 peak was dominant in all samples (refer to Appendix), but was also present in the control sample and therefore not used to calculate biomass. Additionally it has been suggested that the cyclic PLFAs cyl7:0 and cyclopropane nonadecanoic acid (cyl9:0) were useful markers for estimation of biomass of Burkholderia cepacia. The P L F A composition of cells from the pure culture of five B. cepacia strains revealed contents of the PLFAs cyl7:0 and cyl9:0 in the range 15 to 28% and 8 to 24% of total content of PLFAs (118). It should be noted that cyclopropane PLFAs have been shown to increase in relative abundance concomitantly with culturing (59). In the mass spectrum of Braid Street site samples, cyl9:0 was another marker seen throughout the spectra excluding the control (refer to Appendix for GC-MS data). This P L F A was observed at a lower proportion in relative to the abundance of cyl7:0, except at the depth 9.67 mbrb, where the cyl9:0 P L F A made up a higher proportion of this sample in relative to cyl7:0. This could be attributed to a different composition of microorganisms found here. 56 In another study, cyl7:0 was found in a SO4 "-reducing microcosm from sediment taken from a monitoring well of a petroleum-hydrocarbon contaminated aquifer. Here aquifer microcosms were incubated under SOV'-reducing conditions to follow the flow of 1 3 C from toluene into biomarker FAs. Biomarker FAs characteristic for the genera Desulfobacter and Desulfobacula were studied, but the authors also noted that 1 3 C enriched several uncommon PLFAs such as cyl7:0. They therefore suggested that this P L F A is a specific biomarker for the different groups of SO4 "-reducing bacteria (112). At the Braid Street site, the enrichment of the P L F A cyl7:0 in the sediment could indicate a viable community capable of P A H degradation. Bacterial Community Composition and Structure Previous studies have determined that there is an increase in microbial community activity in areas of an aquifer with higher river water ingression, which correspond to an increase in diversity in the HZ (95). Additionally, studies have indicated that microbial community composition and abundance in the HZ can vary temporally (45). However, no study has compared this composition and increase in diversity to the other microbial communities found along the same groundwater flowpath. It was hypothesized that the bacterial composition in the HZ at the Braid Street site would not only increase in diversity, but also differ in composition to the communities found deeper in the BHZ, upgradient in the A Z , and further offshore in the SZ (Fig. 2). This composition should increase in diversity in relation to the steeper gradients in redox solutes that were observed in the HZ, e.g. Fe 2 + and Fe 3 + concentrations, compared to the lower gradients in redox solutes observed in the B H Z and A Z (Table 1). Additionally, bacterial community composition should correlate to the presence of PAHs along the groundwater flowpath originating from the contaminant source. 57 Bacterial 16S rRNA gene PCR products were obtained from twenty depths along four zones in the aquifer, where the samples were collected from sediment cores that corresponded to pore water profiles (Fig. 3). The hypotheses were then tested using DGGE analysis followed by the construction of clone libraries on three representative samples: one from the HZ, one from the BHZ and one from the AZ. Denaturing-Gradient Gel Electrophoresis DGGE is an example of a genetic fingerprinting technique; it is an electrophoretic method used to identify single base changes in DNA. In a denaturing-gradient acrylamide gel, double-stranded DNA is subjected to an increasing denaturant environment until partially denatured DNA halts in the gel. The melting temperature of the DNA is dependent on its specific sequence. The banding pattern is an image of the whole bacterial community, whereas a single band represents a phylotype (101). A phylotype is defined as a population of microorganisms with a high degree of similarity in 16S rDNA sequences. Four samples were chosen from the HZ (depths: 0.96, 1.06, 1.16, and 1.26 mbrb), four from the BHZ (depths: 3.76, 3.86, 3.96 and 4.06 mbrb), eight samples from the AZ (depths 9.36, 9.67,10.18,10.38,11.24,11.54, 12.03 and 12.25 mbrb), and four samples from the SZ (0.13, 0.23, 0.33 and 0.43 mbrb). Partial 16S rRNA gene PCR products from these depths were run on a DGGE and dominant bands were excised from the gel and sequenced (Fig. 5a, Table 3). Pearson correlation A similarity matrix was created based on the location and intensity of bands in the gel. The similarity matrix was then used to correlate the different banding patterns via the Pearson correlation clustering algorithm, where the clustering was illustrated in a dendogram (Fig. 5b). The dendogram horizontal branches represent the similarity of the banding patterns, where branch length is proportional to dissimilarity. 58 Bacterial communities from the A Z and the HZ at depths 0.96, 1.26, 10.18,10.38, 11.24,11.54,12.05 and 12.25 mbrb clustered at greater than 90% similarity. Furthermore, communities from the HZ at depths 1.16 and 1.06 mbrb clustered at 75% and 85% similarity respectively with these communities, and communities from the A Z at depths 9.36 and 9.67 mbrb both clustered at less than 50% similarity with these communities (Fig. 5b). A l l of these communities represent samples taken within the contaminant plume, and are found along the same groundwater flowpath as was determined through geochemical analysis of the corresponding pore water profiles (Fig. 4). It is a strong possibility that the communities from depths 9.36 and 9.67 mbrb are taken from the fringe of the plume where the ratios of total iron and Fe 2 + , and Fe 3 + and indane are similar to those found in the HZ. The communities at 3.76, 3.86, 3.96 and 4.06 mbrb tightly clustered at nearly 100%) similarity. These samples were taken from the BHZ where the groundwater from the contaminant plume, the regional groundwater and river water are thought to converge, yielding a region of relatively high salt concentrations along with the presence of contamination and a pH similar to that of the river water (Table 1). The communities found in the SZ could represent an area of high river water ingress based on the brown color of their stratigraphy (Table 3), which is indicative of iron oxidation. The sediments from the SZ were taken along the same transect as its corresponding pore water profile, but were collected further offshore in relation to the pore water profile (Fig. 3). The SZ communities were no greater than 85% similar in relation to each other, and they clustered at less than 10% similarity to the communities found along the contaminant flowpath. The Pearson correlation clustering of the denaturing gradient gel indicates that communities observed in the HZ fingerprints are more diverse than that observed in the A Z and BHZ, but less diverse than those found in the SZ. Furthermore, fingerprint 59 patterns along the same groundwater flowpath are more similar than that from alternate flowpaths. The depths of the communities analyzed in the SZ, HZ and BHZ represent 10 cm intervals, whereas 30 cm intervals are represented in the A Z . Therefore spatial variability was also a factor in interpreting the clustering analysis. Dominant bands Dominant bands in the denaturing gradient gel likely represent abundant and functionally important populations in their respective zones. RDP II was used to make the best possible taxonomic identification of the populations represented by bands, whereas B L A S T - N was used to find the closest relatives to try to infer function of the populations. The dominant bands cut and sequenced from the D G G E can be found in Fig. 5 a, whereas their classification in RDP II along with their closest match by B L A S T -N and percent sequence identity can be found in Table 4. The band sequenced from the SZ (band 7) closest match (100% sequence identity) was to a p-proteobacterium isolated from arsenic-discharged mine water, where the mine water was rich in dissolved Fe 2 + with precipitates rich in Fe 3 + (108). This environment is similar to what we might expect at the SZ, where the profile shows a high concentration 2+ of Fe that is likely to be oxidized rapidly when it reaches the oxygen rich river water (Fig. 3). The p-proteobacteria group contains several genera that have been recognized to both oxidize and reduce iron (140). Band 12 was dominant in all the fingerprints from communities analyzed in the HZ and A Z , where its closest match (100% sequence identity) was to the P-proteobacterium Burkholderia ferrariae. This bacterium was originally isolated from an iron-ore mine located in Brazil; iron-ores are known to be rich in iron oxides (138). B. ferrariae was initially isolated on the basis of its potential to solubilize highly insoluble phosphatic minerals under aerobic conditions. B. ferrariae 60 assimilates a wide range of carbon sources and can reduce NO3" to NO2". As well, the three most dominant PLFAs found in its PLFA profile are cyl7:0 at 18.9%, a cyl9:0 at 18.8%), and 16:0 at 18.0%. These PLFAs correspond to those observed to be dominant in the P L F A analysis of sediments at 1.06, 1.16, 9.67 and 10.18 mbrb. PLFAs degrade rapidly in the environment, and therefore the combination of the dominance of this DGGE band and the presence of cyl7:0 and cyl9:0 in the Braid Street site sediments suggests an active population of B.ferrariae or a close relative in these communities. In general Burkholderia is a remarkably diverse genus that includes groups able to aerobically degrade PAHs and plant symbionts capable of fixing N 2 (106, 139) It is the diversity of Burkholderia's metabolism that could give it a competitive advantage over other bacteria at the Braid Street site enabling it to be competitive along the contaminant groundwater flowpath. For example, many members of Burkholderia can grow as a facultative anaerobe alternating their TEAP from oxidative respiration to denitrification (106). This would enable Burkholderia to dominate in a microaerophillic environment. Furthermore, facultative anaerobes can use NO3" as a T E A while conserving oxygen for enzymatic reactions used for carbon assimilation (79). 'It is therefore a possibility that Burkholderia can degrade PAHs using a cytochrome P450 monooxygenase, while using NO3" as a T E A in a microaerophillic ecosystem. A research study conducted by Roden et al (2004), determined that in a microaerophillic ecosystem at the aerobic - anaerobic interface, P-proteobacteria can bind Fe 3 + using biogenic ligands that promote rapid microscale iron redox cycling, which limits Fe precipitation in the environment (123). These conditions could be extrapolated to Braid Street site where sediment extractions did not reveal a net increase in Fe 3 + precipitates where Fe 2 + oxidation was expected (125). As well, Weber et al. (2006) suggested that P-proteobacteria could potentially alternate between anaerobic Fe 2 + oxidation using NO3" as a TEA and Fe 3 + reduction (141). Since Burkholderia have not 61 been shown to be Fe 3 + reducers, this is an unlikely process by Burkholderia at the Braid Street site. However, along the groundwater flowpath B. ferrariae may be anaerobically oxidizing Fe 2 + using NO3" as a TEAP in juxtaposition to Fe 3 + being reduced by an alternate group of microorganisms. The presence of NO3" and the potential for Fe 2 + oxidation is not considered to be inhibitory to Fe 3 + reduction (141). The metabolic capability of B. ferrariae to anaerobically and aerobically oxidize Fe 2 + would reasonably justify the relative abundance of B. ferrariae in the HZ and A Z at the Braid Street site. Band 8 from the HZ was an unclassified bacterium with its closest match (91% sequence identity) to D N A isolated from nitrobenzene-contaminated river sediment. N -containing PAHs are significant in creosote contamination and amines were prevalent in the 2005-pore water profiles from Braid Street site. The presence of amines in conjunction with aerobic N H 4 + oxidation could account for the availability of NO3" for anaerobic oxidation of Fe 2 + . There were two additional sequenced bands from the communities located on the fringe of the contaminant plume in the A Z (bands 10 & 11). The first (band 10) corresponds to the genus Microbacterium, some of which are known to be capable of heavy-metal reduction, and its closest match (100% sequence identity) was to D N A isolated from a microcosm whose matrix was collected from the rhizosphere of Poplar trees (57). The sequence of the second band (band 11) corresponds to the order Rhizobiales whose closest match (98% sequence identity) was to a bacterial D N A isolated from the arctic permafrost soil in Spitsbergen. The Rhizobiales are an order of cc-proteobacteria where the Rhizobia, who are well recognized for their ability to fix N 2 , appear in several of the families here. They are typically considered to be oligotrophic and are commonly found in symbiosis with plants, where Rhizobia mutually convert N 2 from the atmosphere into the plant viable form N H 4 + and the plant supplies the bacteria with a usable carbon source (139). 62 Bands 10 and 11 are from sediment corresponding to the fringe of the plume where conditions are more oligotrophic. The dominance of these two phylotypes in the fingerprints suggests the importance of N-cycling at this location in the aquifer. Fixing N2, which is present in the aquifer, followed by further nitrification will yield NO3" for anaerobic Fe 2 + oxidation. Band 13 originated from near the bottom of the A Z profile and could not be classified. Its closest match (96% sequence identity) was to D N A isolated from column experiments conducted to deduce the microorganisms involved in uranium reduction and reoxidation (25). The researchers conducted column experiments where the matrix was harvested from a uranium-contaminated site rich in C a 2 + and DOC at concentrations that are thought to drive the thermodynamic equilibrium towards the oxidation of uranium even under reducing conditions. At the Braid Street site where dissolved Ca readily increases below 12 mbrb there is a corresponding peak in total iron at 13.5 mbrb, suggesting the oxidation of Fe 2 + here (Table 1, Fig. 4). Lastly, band 9 was predominant in the fingerprints of the communities analyzed in the B H Z sediment. This band, classified as Rhodococcus, matched (100%) sequence identity) to a Rhodococcus erythropolis strain isolated from a plant root soil sample. Rhodococcus erythropolis has the ability to desulfurize benzothiophene, whereas other Rhodococcus species have been determined to transform indane (28, 37). Therefore, mass loss of indane and benzothiophene in the contaminant plume might be partly attributed to the presence of Rhodococcus here. A recent study has proposed that the major environmental determinant of microbial community composition is salinity rather than extremes of temperature, pH, or other physical and chemical factors (88). The B H Z is hypothesized to be a mixture of contaminant groundwater and regional groundwater, where a concomitant increase in dissolved Ca and Cl" is seen here, with river water ingression to this point (Table 1). 63 Rhodococcus spp. have been observed to tolerate high concentrations of Cl" up to 31500 ppm under aerobic conditions (21). Furthermore, strains of Rhodococcus erythropolis are considered weak halophiles, and grow better under microaerophillic conditions in the presence of relatively high salt concentrations than do moderate halophiles (8). This competitive advantage of R. erythropolis in saline environments combined with their capability to degrade PAHs could explain why we see this phylotype predominate in the BHZ. Burkholderia band It is important to quantify results of a phylogenetic study with the geochemistry of a system in order to validate a hypothesis. The intensity of the Burkholderia band extracted from the DGGE fingerprint of sediment communities from the HZ and A Z significantly correlated to the concentration of indane measured in the corresponding pore water profiles (Fig. 6). The intensity of the Burkholderia band increases with an increase in indane concentration. The increase in Burkholderia with indane is likely attributed to the preferential flowpath of the groundwater; i.e. higher indane concentrations dissolved in groundwater are found in the system at a specific point correlating to higher growth of Burkholderia at that point. Therefore the increase in intensity of the Burkholderia band at these points could also correlate to a high concentration of iron, which is detected along with a high concentration of indane (Fig. The difference between the slopes of the regression in Figure 6, where the HZ slope is greater than the A Z slope, could be explained by the differing growth rates of Burkholderia in the HZ relative to the A Z . For example growth under aerobic conditions in addition to anaerobic conditions would increase the growth rate of Burkholderia. Furthermore the abundance of Burkholderia appears to be consistent in both zones 64 despite the increase in indane concentration in the A Z , a phenomenon supported by the P L F A profiles from these zones. This could imply that the growth of Burkholderia is limited to the availability of a nutrient, electron donor, or electron acceptor. For example lithotrophic Fe 2 + oxidation by Burkholderia coupled to NO3" reduction would mean that the availability of NO3" might limit this process. The oxidation of Fe 2 + coupled to N 0 3 " reduction is energetically favorable at neutral pH and should yield enough energy to support carbon fixation and microbial growth (140). Two characteristics of the aquifer at the Braid Street site support the hypothesis that Burkholderia oxidizes Fe 2 + coupled to NO3" reduction here: one, indane is a recalcitrant organic carbon and therefore its presence even at high concentrations may still yield oli go trophic-like conditions in the aquifer; two, it is expected that N 0 3 " is present in the aquifer and that it is less abundant along the groundwater flowpath towards the contaminant source as the aquifer becomes more anaerobic. Additionally, it is a possibility that Burkholderia is (co-)metabolically hydroxylating indane to indanol along the contaminant flowpath. This would reasonably explain the increase in the indanol peak in the mass spectra (data not shown). Enrichment Cultures Bieber (2003) determined Fe 3 + reduction to be the dominant TEAP in the contaminant plume (19). With this premise, enrichment cultures were created in 2005 under Fe -reducing conditions where cultures were amended with indane, benzothiophene and naphthalene. The Fe 3 + was present in cultures as either amorphous FeOOH (Fig. 5a; HZ-Fe (III), HZ-2-Fe (III), AZ-Fe (III)) or as F 3 + chelated by N T A (Fig. 5a; AZ-Fe (III)+NTA), and were inoculated with homogenized sediment into their respective cultures from the HZ, HZ-2, and A Z (Fig. 2). Following the time for an expected lag phase (>426 days, ref) 16S rRNA gene PCR products were obtained from 65 each eririchment culture and run on a D G G E to compare the community composition to unamended aquifer sediment from the site. When correlated using the Pearson correlation clustering algorithm, enrichment culture fingerprints were less than 60% similar to those of the unamended sediment (Fig. 5b). Six dominant bands from the DGGE gel of the enrichment cultures were sequenced (Fig. 5a, Table 3). A l l culture band sequences matched in the NCBI database to those of uncultured bacteria (98-99% sequence identity), and did not closely match those of the dominant bands that were sequenced from the unamended aquifer sediment. Furthermore, three of the six bands were from the group 8-proteobacteria according to the RDP-II classification algorithm. Two of these bands were related to Geobacter (90 and 100%) similarity).that are well known for Fe3+-reducing capabilities (85),,and the third band was related to Smithella, which is a syntrophic bacterium known for its ability to degrade propionate coupled to the reduction of SO42" (136). Past molecular studies of anaerobic bioremediation suggest that microorganisms involved in anaerobic P A H biodegradation that can be recovered in pure culture are closely related to populations in situ (84). Therefore the deviation seen here in the community structure of enrichment cultures versus unamended aquifer sediment could mean that Fe 3 + reduction is not a dominant process at the Braid Street site. Clone Libraries Clone libraries are frequently used to characterize the phylotype abundance of a microbial community. Three clone libraries were constructed to ascertain whether the diversity is greater in the HZ compared to the B H Z and A Z along the contaminant flowpath at the Braid Street site, and to improve on understanding the composition of the bacterial communities at these points. In this study a library consists of 100 randomly selected 16S rRNA gene PCR products that were cloned. The libraries were created from 66 the bacterial sediment communities at depths 1.16, 3.96 and 10.38 mbrb taken from the HZ, BHZ and A Z respectively. Whole 16S rRNA gene PCR products from clones were sequenced and assembled from individual phylotypes (defined as <97% similarity) for this analysis, where the number of clones sequenced from each depth can be found in Fig. 7. Furthermore, the classification of the individual populations in RDP II, their relative abundance, and their closest match in the B L A S T - N database with corresponding percent sequence identities, can all be found in Tables 5, 6 and 7. Relative abundance and group distribution The phylotypes determined from the clone libraries along with their relative abundance and possible function will be discussed in the following section. The relative abundance of bacterial groups determined from each clone library is illustrated in Fig. 7. These data clearly reveal the extensive diversity of phylotypes found in the HZ relative to the A Z . The six most abundant groups determined in the HZ, accounting for over 60% of the community's population, are P-proteobacteria at 20%, Firmicutes at 16.61%, Acidobacteria at 12.2%, Bacteroidetes at 6.67% and Planctomycetes at 5.56%. In the B H Z Actinobacteria accounted for 82.02% of the community and Firmicutes followed comprising 10.11% of the population. Lastly in the A Z , P-proteobacteria accounted for 63.22%) followed by Firmicutes at 22.99%. Unclassified bacteria were abundant in the HZ making up 26.67%> of the bacterial population. Conversely, unclassified bacteria made up only 1.12% and 6.9% of the BHZ and A Z bacterial communities respectively. The abundance of bacterial groups can give insight to the geochemical processes that might be occurring in their respective zones. Basedon the hypothesis that iron cycling is limited by nitrogen cycling in the aquifer, the possible relevance of these groups in contribution to these geochemical processes will be addressed. 67 Some genera from the p-proteobacteria are capable of anaerobic and microaerophillic Fe 2 + oxidation and Fe 3 + reduction, with the majority of those representatives linked to Fe 2 + oxidation (140). Additionally, the majority of N H 3 -oxidizing strains that have been isolated from terrestrial and freshwater environments belong to the group P-Proteobacteria (78). Some P-proteobacteria are well known for their capability of chemolithoautotrophic NH3 oxidation, where N H 3 oxidation is the primary step in the nitrification of N H 3 to NO3". This process is generally considered to be aerobic, but these bacteria have been isolated from microaerophillic and anaerobic environments (78). Firmicutes have only demonstrated Fe3+-reducing capabilities in regards to iron cycling, where no representatives have been shown to be Fe 2 + oxidizers (140). As well, some genera from this phylum are capable of N2 fixation. However, members of this group are not known to play a direct role in other denitrification or nitrification processes (139). Two genera from the ubiquitously distributed but poorly characterized phylum Acidobacteria are capable of either iron reduction or oxidation where only a single representative has been noted for each process (140). Even less is known regarding Acidobacteria's role in nitrogen cycling. The phylum Bacteroidetes has not demonstrated the capability of iron oxidation or reduction, however some are able to participate in denitrification (139). In addition Planctomycetes have also not shown any capabilities of iron reduction or oxidation, but some are well known for their ability to anaerobically oxidize NH3 plus NO2" to N2 (47) Lastly, two genera of Actinobacteria have demonstrated Fe 2 + oxidation capabilities, but no Fe 3 + reduction has been noted (140). As well, some Actinobacteria can fix N2, but have not been shown to partake in other nitrification or denitrification processes (139). 68 Nitrification, denitrification, Fe 2 + oxidation and Fe 3 + reduction processes are all linked to autotrophic, heterotrophic or mixotrophic growth. This link could implicate these processes to P A H degradation by phylotypes identified in the clone libraries from the Braid Street site. Phylogenetic Tree Phylogenetic trees are branching diagrams that depict the evolutionary relationship of phylotypes. A phylogenetic tree was constructed from the three clone libraries obtained from the Braid Street site samples: HZ, BHZ and A Z samples contained phylotypes that comprised at least 2% of the respective community composition (Fig. 8, Tables 5, 6 & 7). Furthermore, type strains, which link taxonomy to phylogeny were also included in the tree. The phylogenetic tree was used to assess the overall bacterial community composition of the aquifer based on the dominant groups observed. Furthermore, these groups were compared to the dominant bands observed in the denaturing-gradient gel. The individual populations of the phylogenetic tree and their possible function in the aquifer at the Braid Street site are discussed on a group basis in regards to their relationship to phylotypes in the B L A S T - N database (Table 5, 6 and 7). Assessing the bacterial community composition of an ecosystem in this manner yields insight to the biogeochemical processes that are likely occurring at the Braid Street site. In summary, the bacterial phylotypes analyzed in the representative clone libraries consistently matched phylotypes in the NCBI database whose 16S rDNA were isolated from environmental samples. Furthermore, the environments associated with the dominant phylotypes typically related to biogeochemical conditions assumed to be present at the Braid Street site, and support the hypothesis that iron and nitrogen redox cycling are important and dominant processes occurring along the flowpath of the 69 contaminant plume. In addition to this, the representative Fe3+-reducing strains such as Geobacter sp., which have been isolated from hydrocarbon-contaminated aquifers in past studies do not dominate at the Braid Street site (84, 124). This in conjunction with the analysis of enrichment cultures could indicate that Fe 3 + reduction is no longer a dominant biogeochemical process here. Lastly, the bands that dominated in the fingerprints of the denaturing-gradient gel matched the same sequences in the NCBI database as the most abundant phylotypes of corresponding clone libraries. Actinobacteria Four distinct populations were observed to be dominant belonging to the Actinobacteria phylum. Two were from the BHZ portion of the plume. BHZ-02 dominated 55% of the community with a 100% sequence identity to a cyclotrimethylenetrinitramine (RDX) degrading bacteria, and is the same (less than 3% sequence difference) as the type strain Rhodococcus erythropolis. This was the same population that was determined to dominate the DGGE in the BHZ, but matched to a different sequence in the B L A S T - N database. The discrepancy is attributed to the different lengths of the sequences queried. The second Actinobacterium, BHZ-01, comprising 26% of the B H Z community was also related to Rhodococcus. This population was most closely related to (100% sequence identity) a carbendazim-degrading Rhodococcus strain (68). The most notable connection between these two populations is their ability to degrade amines, and this may indicate their role as amine degraders in the BHZ. The degradation of amines generally leads to an increase in N H 4 + in a system, which might be oxidized to NO3" in the microaerophillic zones of the aquifer and to N2 in the anaerobic zones. Therefore these bacteria might play an important role in nitrogen cycling. 70 There were two additional Actinobacteria observed in the aquifer that were distant in their relationship to the Rhodococcus populations observed in the BHZ: AZ-01 dominating 7% of the A Z community and HZ-05 dominating 2% of the HZ community. AZ-01 matched (83% sequence identity) to 16S rDNA harvested from a deep-sea mud volcano that releases C H 4 . Whereas HZ-05 belonging to the Coriobacteriaceae family matched (96% sequence identity) to a phylotype whose D N A was isolated from the ridge flank crustal fluids in the Jaun de Fuca, which have been previously determined to the support the growth of Fe 2 +-oxidizing bacteria (63). 5-Proteobacteria Two populations of 8-proteobacteria are seen in the HZ, HZ-15 and HZ-14, making up 4% and 2% of the HZ community respectively. HZ-15 matches (97% sequence identity) to 16S rDNA isolated from a sedimentary and granite rock aquifer (93). Whereas HZ-14, belonging to the family Desulfobacteraceae, matches (96% sequence identity) to 16S rDNA isolated from a ZnS-producing biofilm dominated by SO4 "-reducing bacteria. Furthermore HZ-14 matches most closely to the SO4 reducing type strain Desulfosarcina cetonica. This may indicate the presence of S042"-reducing bacteria in the HZ and account for the FeS precipitates observed at the Braid Street site by Roschinski (2007) (125). Firmicutes Different Firmicutes populations were readily observed in the clone libraries, making up one-third of the dominant phylotype populations. Five Firmicutes phylotypes were seen in the HZ clone library: HZ-21, HZ-24, HZ-20, HZ-17 and HZ-25, comprising 3, 2, 2, 2, and 2% of the community population respectively. HZ-21 appears to relate more closely to the group 8-proteobacteria than Firmicutes. This phylotype belonged to the anaerobic class Clostridia and matched (97% sequence identity) to a JS1 candidate 71 division bacterium whose 16S rDNA was isolated from a cold-seep area of the Japan Trench (83). Cold seeps are an area of the ocean floor where H 2 S , CH4 and hydrocarbon-rich fluid seepage occurs, which may link them to P A H degradation (83). The HZ-24 phylotype also belonged to the obligate or facultative aerobic genus Bacillus and matched (97% sequence identity) to the sequence obtained from a Bacillus endospore isolated from near-subsurface granite (42). The HZ-20 belonged to the genus Laceyalla and matched (92% sequence identity) to 16S rDNA isolated from urban aerosol collected over a city in Texas (26). This phylotype was most closely related to the type strain Thermoactinomyces vulgaris, whose optimum growth is at higher temperatures as denoted by their name and are capable of utilizing a wide variety of substrates (3). The HZ-17 belonged to the genus Clostridium and matched (97% sequence identity) to the peptide-fermenting strain Clostridium proteolyticum (37). Lastly, HZ-25 belonged to the genus Acetivibrio and matched (93% sequence identity) to 16S rDNA isolated from a gas-degrading community in an industrial biofilter (48). This phylotype was the closest match to the type strain Clostridium josui, known for its production of cellulase (74). Four Firmicutes phylotypes were seen in the A Z clone library: AZ-11, AZ-07, AZ-09 and AZ-08, making up 2, 11,2 and 2% of the community population respectively. AZ-11 belonged to the genus Cryptanaerobacter and matched (93 % sequence identity) to 16S rDNA isolated from sediment capable of anaerobic dechlorination of tetrachlorobiphenyl to biphenyl. This phylotype related to the type strain Cryptanaerobacter phenolicus, an anaerobe that transforms phenol into benzoate (73). AZ-07 made up a significant portion of the community population and belonged to the order Clostridiales and matched (94% sequence identity) to 16S rDNA isolated from a column experiment studying uranium reduction and reoxidation, where net oxidation occurred even under methanogenic conditions (25). This is a phenomenon that might be occurring in the A Z at the Braid Street site, where the oxidation of Fe 2 + could be 72 occurring under the reducing condition observed here. Furthermore this phylotype appears to represent band 13 in the DGGE. AZ-09, which was closely related to AZ-01, belongs to the class Clostridia and matches (97% sequence identity) a benzene mineralizing consortium clone whose 16S rDNA was isolated from a SO^'-reducing culture (113). This suggests the PAH-degrading reducing possibilities of AZ-01109. Lastly AZ-08 belonging to the genus Clostridium matches (98% sequence identity) to Clostridium chromoreductans, an anaerobic Cr (Vl)-reducing bacterium isolated from soil. This phylotype was closely related to the type strain Clostridium acetobutylicum, which is a commercially valuable bacterium that is used for its fermentative capabilities (107). There were two Firmicutes phylotypes found in the BHZ clone library: BHZ-14 and BHZ-10, where they each made up 2% of the clone library population. The BHZ-14 phylotype belonged to the genus Cryptanaerobacter and matched (96% sequence identity) to 16S rDNA isolated from sediment capable of anaerobic dechlorination of tetrachlorobiphenyl to biphenyl. In addition to AZ-11, this phylotype was related to the anaerobic type strain Cryptanaerobacter phenolicus that transforms phenol to benzoate. Lastly, the BHZ-10 belonged to the order Clostridiales and matched (96% sequence identity) to 16S rDNA isolated from fault-bordered aquifers (83). A notable characteristic of the phylum Firmicutes is their ability to produce endospores, which are related to their longevity in an environment (147). This could mean that the presence of Firmicutes in the aquifer might proceed their functional contribution to the ecosystem. However, this could also give them an advantage to survive in an environment that fluctuates between reducing and oxidizing conditions. As observed from the genera discussed here, Firmicutes are quite diverse and many genera are capable of thriving in the conditions observed at the Braid Street site, and include the ability to reduce Fe 3 + and degrade PAHs. 73 Acidobacteria HZ-01 was the only Acidobacteria found comprising 9% of the HZ community population. This phylotype belonged to the Family Acidobacteriaceae and matched (95% sequence identity) to 16S rDNA isolated from the rhizosphere of Phragmites, a grass that grows in aquatic environments. As well, this phylotype was most closely related to the type strain Geothrix fermentans, a bacterium well known for its ability to oxidize aromatics and reduce Fe 3 + (104) which could indicate the role of HZ-01 as an Fe 3 + -reducer at the Braid Street site. Planctomycetes HZ-27 and HZ-29 were Planctomycetes phylotypes that individually made up 2% of the HZ community population. HZ-27 belonged to the family Planctomycetaceae, and matched (88% sequence identity) to 16S rDNA isolated from C H 4 hydrate sediment in the Peru Margin (66). Whereas the HZ-29 belonged to the genus Isosphaera, a multicellular filamentous bacteria, and matched (91% sequence identity) to 16S rDNA isolated from mangrove soil. Planctomycetes are well known for their ability to anaerobically oxidize N H 3 to N 2 (47). However, the Planctomycetes found here were not closely related to cultured representatives or to bacteria in the NCBI database and therefore it is difficult to predict their role in the aquifer at the Braid Street site. Bacteroidetes HZ-08 comprised 3% of the HZ community population and belonged to the order Bacteroidales. This phylotype matched (96% sequence identity) to 16S rDNA isolated from an autotrophic denitrifying anaerobic sludge reactor. The sludge reactor system might be similar to the dynamic in this area of the aquifer, where denitrification is expected and there is a decline in C 0 2 measured here. 74 Nitrospira Nitrospira are well known for their ability to aerobically oxidize NO2" to NO3" (128). The HZ-26 phylotype belonged to the phylum Nitrospira and the genus Magnetobacterium, where this phylotype matched (93% sequence identity) to 16S rDNA isolated from the sediment of a reservoir. Magnetobacterium are notable for their biomineralization capability, where they are thought to reduce Fe 3 + in vivo to form the Fe -Fe -oxide mineral magnetite (14). In many freshwater habitats the highest concentration of magnetotatic bacteria occur at the microaerophillic region, where they grow aerobically or using N2O as a T E A (46). The optimal conditions for the growth of Magnetobacterium depict the conditions assumed to be present in the HZ at the Braid Street site. Therefore the presence of this phylotype supports the assumptions concerning the geochemistry here. For example, the HZ is microaerophillic and denitrification is a biogeochemical process dominating here, which could lead to the production of N2O. p-proteobacteria There were seven P-proteobacteria phylotypes found in the HZ and A Z . The A Z -04, AZ-05, AZ-03 and AZ-02 comprised 2, 2, 8 and 50% of the population respectively. AZ-04 belonged to the genus Ralstonia and matched (96% sequence identity) to 16S rDNA isolated from an Fe3+-reducing enrichment of estuary sediment. Furthermore, this phylotype was closely related to the plant pathogen Ralstonia syzygii. AZ-05 belonged to the genus Alcaligenes and matched (99% sequence identity) to an Alcaligenes species that is a phenol-degrading isolate from an activated sludge system (150). Phenols are a known component of creosote contamination and therefore phenol degradation could be the metabolic role of this phylotype at the Braid Street site. Additionally, this phylotype was closely related to the type strain Alcaligenes faecalis, an obligate aerobe that is commonly found in the environment. 75 AZ-03 and 02 belonged to the genus Burkholderia and matched (96% and 99% sequence identity respectively) to the bacterium Burkholderia ferrariae (138). Furthermore, HZ-13 was the same sequence as AZ-02 and comprised 13% of the community population in the HZ. These sequences match the same bacterium as the D G G E band predominant in the A Z and HZ profiles. In the phylogenetic tree AZ-02 and HZ-13 closely relate to the type strains Burkholderia tropica and Burkholderia sacchari, where both species were isolated from the rhizosphere of sugar cane plants (23, 119). The rhizosphere of these agronomical plants could correspond with conditions present at the Braid Street site. That is an anaerobic environment that is rich in iron where iron reduction and oxidation are occurring in the same zone. For example, both Fe 3 + reduction and Fe 2 + oxidation has been associated with the roots of wetland plants, where a P-proteobacteria has been isolated as the Fe 2 + oxidizer from this environment (41). In addition to HZ-13, the phylotypes HZ-12 and 11 comprised 2 and 3% of the HZ population respectively. HZ-12 belonged to the Aquabacterium genus and matched (98% sequence identity) to 16S rDNA harvested from a hotspring's pink mat layer, which are rich in purple sulfur bacteria. This phylotype was closely related to the type strain Aquabacterium parvum, a bacterium capable of denitrification isolated from a drinking water system (75). Lastly, the HZ-11 phylotype belonged to the genus Roseateles and . matched (98% sequence identity) to a Proteobacterium characterized for its chitosanase production. Unclassified bacteria Relatively few unclassified phylotypes dominated the clone libraries. Furthermore, all dominant unclassified phylotypes originated from the HZ. The phylotypes HZ-30, 37, 44 and 34 comprised 4, 2, 2 and 3% of their libraries, respectively. HZ-30, whose closest relative in the tree belonged to the phylum 8-proteobacteria, 76 matched (95% sequence identity) to 16S rDNA extracted from uranium mill tailings. HZ-37 and 44 were most closely related to the Bacteroidetes phylum and matched (91% sequence identity) to 16S rDNA of populations active in metabolism of CI-compounds in lake sediment (102), and (91% sequence identity) to D N A from the anaerobic hypolimnion of a lake. Lastly HZ-34 closest relative was the P-proteobacteria phylum and matched (89% sequence identity) to a group of thermophillic bacteria (135). Bacterial Community Analysis Summary The analysis of phylotypes present at Braid Street signifies the potential for different redox processes, which increases in the HZ. In particular, it appears that both Fe 3 + reduction and Fe 2 + oxidation are occurring in the HZ and the A Z at the Braid Street site, and that the diverse phylotypes observed in these communities represented in the clone libraries and denaturing-gradient fingerprints are responsible for this cycling. Furthermore, I propose that p-proteobacteria are largely responsible for anaerobic Fe 2 + oxidation and Firmicutes for Fe 3 + reduction. Although these processes are speculative and the bacterial dynamics observed at the Braid Street site do not fully support direct microbial involvement in iron redox cycling, the data consistently support this theory and therefore do not eliminate this possibility. Roling et al (2001) observed a similar community structure in a landfill leachate-polluted aquifer that went from Fe -reducing to N 0 3 " -reducing in nature along the contaminant flowpath. Roling observed that P-proteobacteria phylotypes dominated upstream of the landfill whereas gram-positive bacterial phylotypes dominated beneath the landfill (124). Additionally, Weber et al (2006) conducted a study where freshwater sediments were used to inoculate an enrichment culture under NO3"-reducing conditions (141). Here NO3" reduction commenced immediately followed by Fe 3 + reduction after NO3" was 77 depleted. The researchers then amended the cultures with additional NO3", which resulted in immediate oxidation of Fe 2 + coupled to the reduction of NO3" to NBLi + . The microbial community dynamics were analyzed at each step. The researchers observed an increase in P-proteobacteria phylotypes from 3% under NO3" reduction to 60% under Fe 3 + reduction. On addition of N 0 3 " and the concomitant Fe 2 + oxidation they observed a similar community to that in the Fe 3 + reduction phase; that is p-proteobacteria increased only slightly from 60 to 65%. Here the authors propose that the same communities are responsible for Fe 2 + oxidation as Fe 3 + reduction. However, additional studies have shown that P-proteobacteria can bind Fe 3 + to promote rapid microscale iron redox cycling where little or no net oxide deposition will occur (123). This process could explain why Weber did not see a change in the P-proteobacteria phylotype community from the Fe 3 + reduction phase, where Fe 2 + oxidation was not ruled out. At the Braid Street site there is a shift observed in P-proteobacteria phylotypes from 20% in the HZ (where net Fe 2 + oxidation is assumed) to 63%> in the A Z (where net Fe reduction is assumed). Furthermore, the P-proteobacteria phylotype related to Burkholderia ferrariae shifted from 13 to 58%) in the respective zones. Additionally, a shift in Firmicutes from 17 to 23% was observed in the A Z relative to the HZ; where there was a concomitant shift from 1 to 11% of the phylotype related to the clone isolated form an experiment where uranium oxidation occurred over reduction in a highly reduced column. The overall richness of P-proteobacteria appears to be greater in the HZ relative to the A Z , and the overall richness of Firmicutes appears greater in the A Z relative to the HZ. I propose that p-proteobacteria in addition to other microorganisms from diverse phyla are responsible for Fe 2 + oxidation in the HZ and largely responsible for anaerobic Fe oxidation in the A Z . Furthermore, I propose that under NO3"-reducing conditions the p-proteobacteria related to Burkholderia ferrariae can optimally grow and oxidize 78 Fe 2 + in the presence of Fe3+-reducers, where Firmicutes might represent the dominant phyla of Fe3+reducers. Weber et al (2006) suggest that when inputs of organic carbon are relatively high compared to NO3", organotrophic NO3" reduction may exhaust available N0 3 " , thus allowing microbial Fe 3 + reduction and associated production of aqueous Fe 2 + (141). During subsequent periods of reduced organic carbon, loading, rates of NO3" re-supply may exceed rates of organic NO3" reduction resulting in the availability of N 0 3 " for lithotrophic, NCV-dependent Fe 2 + oxidation. Where reducing equivalents stored in the form of aqueous Fe 2 + may serve as a significant source of energy for microbial metabolism during periods of reduced carbon input (141). , Weber suggests that such coupling is likely significant at the interface between N 0 3 " and Fe 3 + reduction zones in sediments influenced by periodic fluctuations in the inputs of organic carbon and oxidants (141). These fluctuations could describe the circumstance at the Braid Street site where a high river stage elevation during freshet and high tide along with the onshore pumping well will likely lead to the influx of NO3" and oxidants into the aquifer from the river water and will therefore lead to a net Fe 2 + oxidation. Conversely low river stage at low tide and during the rainy season will reverse the groundwater gradients and organic carbon from the contaminant source would yield a net Fe 3 + reduction in the aquifer. Shannon Diversity Index The Shannon diversity index and its associated equitability were calculated from the relative abundances of populations found in the clone libraries and corresponding DGGE profiles (Table 8). The Shannon index is one of several diversity indices used to measure diversity in categorical data (49). This index accounts for the abundance and evenness of the phylotypes present, where a higher number indicates a richer and even 79 population distribution compared to a lower number. The Shannon's equitability assumes a value between 0 and 1, with 1 being complete evenness. The values calculated from the clone libraries versus the DGGE profiles were nearly identical in all three zones with the greatest deviation seen in the A Z (Table 8). This indicates a precise observation of bacterial population abundances from the three communities using 16S rDNA analyses. This agreement gives greater confidence in the results and compensates for the lack of replication. For example, both the DGGE and the clone library indicate that the HZ is higher in diversity compared to the A Z followed by the BHZ. Therefore, we can infer that the diversity increases in the HZ along a contaminant groundwater flowpath, and that the steeper gradient in redox solutes-is likely the attributing factor to this increase in diversity. 80 Chapter 4 THESIS S U M M A R Y A N D FUTURE DIRECTION Thesis Summary Frederick M . Cohan, a renowned evolutionary microbiologist, postulates that bacteria occupy discrete niches and that these "ecotypes" will correspond to the sequence clusters within a DNA-based phylogeny (116). More clearly stated, molecular diversity should relate to ecological diversity. This theory was the basis for deducing bacterial biogeochemical processes at the Braid Street site, where the bacterial community structure of four zones (the SZ, HZ, BHZ and AZ) was expected to harbor distinct ecotypes. It was concluded that the HZ harbors a more diverse population of bacterial phylotypes in relation to the BHZ and A Z , where this is likely attributed to a steeper redox gradient yielding a greater diversity of niches for growth of different ecotypes. Additionally, a single population dominated a similar niche along the contaminant flowpath that went from aerobic to anaerobic in nature. Here the overall diversity of the bacterial community phylotypes decreased in the A Z compared to the HZ. Lastly, the B H Z niche was distinct from the A Z and HZ, where the ingress of river water is predicted to yield a more consistent redox gradient and the regional groundwater makes the ecosystem saline. It was postulated that the saline groundwater here accounted for the dominance of a single phylotype. Where this hypothesis is strongly supported by a recent study indicating the importance of salinity on phylotype selection in ecosystems (25). 81 Initially it was predicted that the physico-chemical processes occurring at the Braid Street site have lead to spatially variable hyporheic flowpaths yielding a zone (BHZ) of a higher partial pressure of oxygen at a greater depth below the riverbed in relation to a zone of steeper redox gradients (HZ). Furthermore, this hypothesis is supported by the idea of convective flowpaths presented by Bianchin (2004), and the geochemistry and bacterial communities analyzed in this thesis (17). A greater fraction of river water appears to be present in the B H Z compared to the HZ that is closer to the river bottom in the aquifer. Nutrients introduced to the B H Z from the river water could move up in the aquifer towards the HZ, where groundwater in the aquifer was determined to move up below the riverbed (1). It is still likely that the B H Z and HZ support a lower partial pressure of oxygen than further offshore in the SZ where the highest partial pressure of oxygen is expected. This could reasonably explain why no Fe 3 + precipitation is seen in the HZ/BHZ. Furthermore it is hypothesized that lithotrophic Fe 2 + oxidation is a dominant process in the aquifer. An organism related to Burkholderia ferrariae could be largely responsible for Fe 2 + oxidation in the aquifer along the contaminant flowpath, and might be transforming indane to indanol. The rate of natural attenuation of a plume has invariably exceeded the rate of attenuation of its source. Therefore as the plume progresses through its life cycle, it should recede back towards the source, through the comparison of contaminant concentrations from 1996-2005 at the Braid Street site, it appears the plume is receding back to its source. Therefore in future analysis of contaminant plumes where Fe 3 + reduction is determined to be the dominant process of natural attenuation of the organic contaminant, the observation of Fe 2 + oxidation may indicate that the plume is receding. 82 Although difficulties in extracting microbial cells and their components such as D N A from the Braid Street site sediment limited the number of samples analyzed, fewer samples allowed for a comprehensive analysis of intrinsic biogeochemical processes. Future Direction The research in this thesis lays the groundwork for future experiments that can be conducted to further elucidate the presence of Fe 2 + oxidation and biodegradation of recalcitrant organic contaminants in the contaminant plume at the Braid Street site. Seasonal variation It was hypothesized in this thesis that Fe 2 + oxidation dominates along the groundwater flowpath that goes from anaerobic to aerobic in nature. Furthermore that this is correlated with freshet, when the aquifer is likely to be more oligotrophic. Therefore collecting cores in the aquifer during the rainy season when the net groundwater flow is directed towards the discharge point should elucidate a microbial community that varies seasonally, where Fe3+-reducers are expected to dominate the profiles. However, there has not been a microbial community characterization conducted at the contaminant source. Therefore conducting a microbial characterization at a point closer to the contaminant source, where communities are expected to be degrading naphthalene under Fe 3 + reducing conditions, would be a simpler approach. Here communities determined in the enrichment cultures such as Geobacter are expected to dominate the profile. Burkholderia Fe 2 + oxidation Since it was hypothesized that anaerobic Fe 2 + oxidation is a dominant process occurring in the aquifer largely by a bacterium closely related to B.ferrariae via NO3" reduction, then experiments to determine this should be initialized. Culturing would be 83 necessary since the metabolic pathway of Fe 2 + oxidation has yet to be elucidated. For example a culture could be set up inoculated with B. ferrariae where Fe 2 + is the only electron donor, furthermore an alternate culture could be set up with Fe 2 + and indane as a available electron donor and energy source to observe possible co-metabolic indane transformation by B. ferrariae, and lastly two control cultures where B. ferrariae has not been inoculated. A l l cultures should contain NO3" as a readably available electron acceptor. Samples collected at sequential time points from the three cultures and analyzed for N0 3 " , N 0 2 \ N H 4 + , Fe 2 + and total iron could elucidate anaerobic Fe oxidation by B. ferrariae. Rhodococcus degradation under moderately saline conditions Rhodococcus phylotypes determined to be present in the aquifer at the Braid Street site could be degrading the recalcitrant compounds indane and benzothiophene, and amines that are determined to be present in the water profiles at this point. Strains of the Rhodoccocus erythropolis have been studied extensively for their capability to biodesulfurize benzothiophene, where the cluster of genes encoding desulfurization enzymes are found on a plasmid contained within this bacterium (28). Primers have been developed to amplify these genes and initial experiments have been conducted on this project to assess the potential of biodesulfurization at the Braid Street site. 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Geochemica l data from pore water sample/mbrb Manganese D-Mn Iron D-Fe Barium D-Ba Calcium D-Ca Arsenic D-As Magnesium D-Mg Silicon D-Si Zinc D-Zn Sodium D-Na lOOml/flow test Surface - 2006 river 0.01 0.11 0.01 13.90 0.000 3.63 2.76 0.13 2.90 0.91 2.76 62.40 0.04 33.70 0.009 16.30 17.90 0.10 6.10 1.22 4.55 85.70 0.09 70.40 0.009 27.40 19.10 0.18 16.10 1.52 6.65 90.10 0.28 155.00 0.010 42.90 17.60 0.08 110.00 1.83 6.56 71.70 0.41 200.00 0.013 48.80 17.60 0.03 342.00 HZ/BHZ - 2004 River 0.01 0.00 0.00 13.10 0.000 2.36 ' 2.13 0.02 0.00 0.61 3.04 89.60 0.18 55.00 . 0.000 13.6.0 18.80 0.04 8.90 90.00 0.91 3.07 94.30 0.16 69.60 0.000 15.70 21.60 0.27 9.90 65.00 1.83 2.54 82.70 0.04 28.60 0.000 14.80 21.40 0.04 8.90 90.00 2.44 2.65 80.30 " 0.06 28.50 0.000 14.70 21.60 0.03 7.90 35.00 3.05 3.77 76.70 0.05' 39.00 0.000 15.10 21.50 . 0.02 . 6.80 32.00 3.66 3.69 51.40 0.05 55.10 0.000 17.10 22.10 0.03 6.30 32.00 4.27 2.43 32.70 0.05 85.60 0.000 21.40 23.70 0.48 7.30 36.00 HZ/BHZ - 2005 river 0.07 1.23 0.00 10.20 0.000 2.25 1.70 0.03 0.00 23.00 0.69 0.96 19.90 0.03 13.10 0.005 , 4.55 5.32 0.08 3.70 45.00' 0.99 1.36 32.90 0.04 17.40 0.009 7.99 11.00 0.50 7.10 104.00 1.30 2.29 55.90 0.04 25.70 0.010 12.90 17.60 0.09 9.40 36.00 1.60 0.90 21.10 0.01 11.80 0.004 5.67 7.15 0.02 3.60 30.00 1.91 1.13 25.90 0.02 15.80 0.004 7.61 8.17 0.03 4.20 31.00 2.21 0.70 14:80 • 0.01 11.70 0.003 . . 4.84 5.38 0.02 2.20 28.00 2.51 0.97 19.30 0.02 20.20 0.004 7.28 7.99 0.05 3.20 30.00 2.82 0.55 9.98 0.00 13.40 0.002 4.32 4.69 0.01 0.00 41.00 3.12 0.97 16.30 0.02 26.50 0.004 8.36 8.48 0.03 3.10 29.00 3.43 0.87 13.60 0.02 • 27.50 0.004 8.62 8.54 0.01 2.80- 33.00 3.73 0:49 7.56 0.01 17.00 0.002 5.41 4.89 0.01 29.00 4.04 1.05 16.00 0.03 39.90 0.005 13.10 9.85 0.03 3.80 30.00 4.34 1.52 21.70 0.04 62.30 0.007 20.70 11.90 0.01 5.70 29.00 4.65 1.77 . 23.80 0.05 66.70 0.006 22.30 10.30 0.03 9.90 34.00 104 Table 1. corit. Manganese Iron Barium Calcium Arsenic Magnesium Silicon Zinc Sodium lOOml/flow sample/mbrb D-Mn D-Fe D-Ba D-Ca D-As D-Mg D-Si D-Zn D-Na test AZ - 2005 river 0.02 0.07 0.00 11.60 0 .000 1.88 1.50 0.01 0 .00 29.00 7.32 0.41 17.60 0.05 14.60 0 .000 9.54 13.10 0.05 10.20 80.0.0 7.62 0.70 31.90 0.08 22.90 0 .000 14.90 19.20 0.03 16.20 7.92 0.75 34.20 0.08 21.20 0 .000 14.60 17.50 0.01 15.70 45.00 8.23 0.80 35.50 0.08 21.00 0 .000 14.40 18.20 0.01 16.60 43.00 8.53 0.43 18.30 0.04 11.10 wO.000 7.65 9.18 0.01 8.90 47.00 9.14 0.98 42.80 0.09 21.90 0 .000 15.10 17.90 0.01 17.70 45.00 10.67 1.51 69.30 0.07 24.90 0 .000 13.40 16.90 0.01 18.00 43.00 10.97 1.77 72.10 0.05 25.80 0 .000 13.50 17:20 0.01 17.80 39.00 11.58 1.93 72.40 0.04 29.00 0 .000 13.10 17.30 0.03 16.70 40.00 11.89 2.26 71.50 0.05 30.20 0 .000 14.60 18.90 0.02 19.40 65.00 12.50 3.71 53.10 0.04 47.10 0 .000 21.70- 22.40 0.01 10.30 41.00 12.80 . 2.40 46.60 0.04 62.50 0 .000 28.60 25.20 0.02 10.10 40.00 13.41 0.64 23.40 0.03 71.50 0 .000 28.70 25.30 0.03 10.50 36.00 14.02 0.56 22.40 0.03 74.20 0 .000 25.70 22.50 0.02 8.50 35.00 105 Table 1. cont. sample/mbrb PH Temp CaCOj N 0 3 N0 2- so42 Fe 2 + T-Fe Cond. cr Br Surface - 2006 river 7.20 0.16 0.00 5.99 0.00 1.13 97.10 1.78 0.00 0.91 6.57 11.80 0.01 0.00 0.00 55.61 62.20 466.00 14.00 0.00 1.22 6.72 11.70 0.00 0.00 0.00 90.02 86.50 861.00 136.00 0.41 1.52 6.75 12.80 0.00 0.00 3.00 91.91 90.10 1966.00 544.00 1.68 1.83 6.85 0.03 0.00 . 78.39 74.70 3470.00 1010.00 3.27 HZ/BHZ - 2004 River 7.89 0.00 0.00 5.50 0.00 0.72 0.00 0.61 6.73 191.00 0.00 0.00 0.00 1.33 5.01 0.00 0.91 6.64 224.00 0.35 0.00 0.00 73.72 4.76 0.00 1.83 6.78 120.00 0.00 0.00 0.00 65.90 5.19 0.00 2.44 6.74 191.00 0.00 0.00 0.00 64.22 4.89 0.00 3.05 6.75 215.00 0.00 0.00 0.00 59.75 4.51 0.00 3.66 6.94 0.00 0.00 0.00 36.86 6.10 0.00 4.27 7.01 269.00 0.00 0.00 0.00 24.01 32.90 0.06 HZ/BHZ - 2005 river 7.67 23.60 0.12 0.00 6.23 0.00 0.00 104.00 0.57 0.00 0.69 6.33 28.10 0.11 0.00 0.00 50.00 56.00 401.00 5.85 0.00 0.99 6.12 26.20 2.53 0.00 0.00 52.00 171.00 382.00 7.58 0.00 1.30 6.18 25.30 0.00 0.00 0.00 49.00 163.00 392.00 7.19 0.00 1.60 6.18 25.00 0.00 0.00 0.00 49.00 77.00 400.00 12.90 0.00 1.91 6.24 23.80 0.42 0.00 0.00 54.00 no:oo 419.00 12.50 0.00 2.21 6.29 21.40 0.00 0.00 0.00 39.00 170.00 430.00 6.96 0.00 2.51 6.31 19.30 0.00 0.00 0.00 42.00 47.00 440.00 11.40 0.00 2.82 6.36 19.50 0.00 0.00 0.00 457.00 4.83 0.00 3.12 6.37 17.50 0.00 0.00 0.00 32.00 39.00 464.00 5.18 0.00 3.43 7.07 15.30 0.00 0.00 0.00 29.00 475.00 14.10 0.00 3.73 7.14 15.80 0.00 0.00 0.00 519.00 24.80 0.08 4.04 7.15 . 16.20 0.00 0.00. 0.00 29.00 32.00 629.00 66.40 0.18 4.34 7.07 16.80 0.00 0.00 0.00 866.00 104.00 0.28 4.65 7.03 18.40 0.00 0.00 0.00 48.00 52.00 1281.00 210.00 0.61 106 Table 1. cont. sample/mbrb pH Temp CaC0 3" N 0 3 N0 2- so42 Fe 2 + T-Fe Cond. cr Br AZ - 2005 river 7.10 14.80 0.00 0.00 6.09 0.23 0.68 88.90 0.80 0.00 7.32 6.65 15.40 0.00 0.00 6.09 25.00 37.00 309.00 2.94 0.00 7.62 6.34 15.40 0.00 0.00 0.00 31.00 369.00 2.95 0.00 7.92 6.54 15.40 0.00 0.00 0.00 30.00 66.00 396.00 2.87 0.00 8.23 6.60 15.40 0.00 0.00 0.00 31.00 372.00 3.02 0.00 8.53 6.58 15.50 o.oo •> 0.00 0.00 30.00 58.00 387.00 2.97 0.00 9.14 6.60 16.30 0.00 0.00 0.00 36.00 50.00 408.00 3.04 0.00 10.67 6.54 16.80 0.00 0.00 0.00 68.00 78.00 531.00 3.62 0.00 10.97 6.52 17.20 0.00 0.00 0.00 71.00 533.00 3.80 0.00 11.58 6.47 17.80 0.00 0.00 0.00 65.00 71.00 559.00 4.37 0.00 11.89 6.46 18.40 0.00 0.00 0.00 70.00 535.00 4.38 0.00 12.50 6.52 20.20 0.00 0.00 0.00 52.00 58.00 624.00 3.69 0.00 12.80 6.63 18.70 0.00 0.00 0.00 41.00 657.00 5.49 0.00 13.41 6.59 18.00 0.00 0.00 0.00 19.00 72.00 634.00 5.41 0.00 14.02 7.25 17.10 0.00 0.00 0.00 21.00 30.00 691.00 6.30 0.00 107 Table 1. cont. sample/mbrb D O N 2 A r co2 C H 4 D O C indane benzothiophene naphthalene Surface - 2006 river 0.00 0.00 0.00 0.91 0.00 0.00 0.00 1.22 0.00 0.00 0.00 1.52 0;00 0.00 0.00 . 1.83 0.00 0.00 0.00 H Z / B H Z - 2004 River 8.60 0.00 0.00 0.00 0.61 8.30 22.70 0.00 0.00 0.91 1.80 61.60 0.00 0.00 1.83 1.80 72.00 9.50 0.00 2.44 0.45 64.70 0.00 0.00 3.05 17.60 0.00 0.00 3.66 0.29 6.00 0.00 0.00 4.27 0.41 4.70 o:oo 0.00 H Z / B H Z - 2005 river 7.50 15.11 0.56 1.14 0.01 3.58 0.00 0.00 0.00 0.69 0.35 6.32 0.31 146.15 13.09 36.50 0.00 0.00 0.99 0.65 7.34 0.34 128.50 10.78 33.60 4.70 0.00 1.30 0.21 7.49 0.36 118.78 13.05 77.80 6.20 0.00 1.60 0.12 8.20 0.38 120.50 14.08 79.10 5.40 0.00 1.91 0.10 8.48 0.39 125.91 14.52 58.10 0.00 0.00 2.21 0.21 9.79 0.44 113.78 14.96 8.86 39.10 0.00 0.00 2.51 0.19 9.47 0.41 108.62 14.36 9.76 19.60 0.00 0.00 2.82 8.96 15.30 0.00 0.00 3.12 0.09 11.39 0.48 109.51 14.32 8.69 9.60 0.00 0.00 3.43 0.15 10.34 0.42 86.88 11.40 8.92 5.20 0.00 0.00 3.73 10.26 0.42 . 83.68 11.17 8.28 4.10 o.oo • 0.00 4.04 9.64 0.40 79.61 11.42 7.93 3.60 0.00 0.00 4.34 8.94 0:39 80.39 12.99 3.20 0.00 0.00 4.65 , 7.62 0.34 88.09 13.80 3.20 0.00 0.00 108 Table 1. cont. sample/mbrb DO N 2 Ar co2 C H 4 DOC indane benzothiophene naphthalene AZ - 2005 river • 5.00 19.24 0.66 3.72 0.07 11.60 2.10 0.00 7.32 0.20 10.43 0.49 131.30 8.12 6.00 0.60 0.00 7.62 0.30 14.64 0.59 127.94 11.12 7.20 3.20 0.00 7.92 0.53 14.91 0.60 139.56 10.25 31.10 13.90 0.50 8.23 0.43 •14.84 0.59 129.13 11.40 33.10 13.00 0.50 8.53 15.83 0.60 138.48 10.90 40.80 12.30 0.60 9.14 0.79 14.64 0.57 141.41 10.82 13.60 45.90 11.90 0.60 10.67 10.70 0.42 183.60 6.39 12.90 161.00 33.80 1.90 10.97 15.46 0.61 208.83 7.05 14.20 378.70 81.90 41.90 11.58 12.06 0.53 237.51 • 6.64 14.60 385.30 80.20 20.00 11.89 10.03 0.47 234.20 7.28 477.20 63.80 10.90 12.50 0.00 14.20 640.50 62.30 34.00 12.80 10.56 0.40 209.02 10.08 , 13.70 230.20 35.00 7.60 13.41 13.41 _ 0.56 196.95 10.26 14.90 275.50 28.80 3.90 14.02 13.68 0.57 161.15 11.06 14.60 263.30 17.30 3.20 Note : A l l values are g iven i n p p m ( m g / L ) aside f rom the fo l l owing : contaminants (indane, benzothiophene and naphthalene) i n ppb (pg/1), conduct iv i ty i n uS/s , p H , temp i n ° C and lOOml / l fow test i n seconds. B l a n k cel ls indicate that the parameter was not measured, whereas zero values indicate values b e l o w the detection l imi t . i 109 Table 2. Sediment PAHs from HZ-2004 Sample/mbrb Indane/ppb Naphthalene/ppb Benzothiophene/ppb HZ/BHZ - 2004 1.65 330.70 13.94 82.97 1.68 617.70 41.49 165.95 1.73 359.90 29.86 93.83 1.78 402.70 47.03 106.85 1.83 429.00 54.16 111.66 1.88 419.90 73.28 107.29 1.93 507.70 112.41 138.95 1.98 483.40 161.35 133.93 2.03 460.80 152.09 131.70 2.08 533.90 156.21 157.39 2.13 758.39 189.36 158.58 2.18 725.55 224.83 167.00 2.24 825.92 216.25 200.67 2.29 781.86 200.62 195.66 2.34 906.85 217.61 220.90 2.39 958.55 230.06 227.28 2.44 925.91 243.66 231.70 2.49 1051.87 284.67 255.56 2.54 2524.62 289.86 634.58 2.59 951.54 238.31 222.45 2.64 846.79 200.82 183.41 2.69 883.92 203.17 187.87 2.74 844.92 175.77 165.60 2.79 791.92 ' 145.66 139.83 2.90 920.85 148.76 150.53 2.95 989.32 155.33 156.17 3.00 843.03 135.21 132.14 Table 3. Stratigraphy description of samples used for Bacterial community composition Sample/mbrb Description of stratigraphy 0.13 Dark brown sand 0.23 Dark brown sand 0.33 Dark brown sand 0.43 Dark brown sand 0.96 pebbly sand (pebbles 1mm to 1cm) 1.06 pebbly sand (pebbles 1mm to 1cm) and clam shells 1.16 pebbly sand (pebbles 1mm to 1cm) and clam shells 1.26 dark grey coarse sand 3.76 dark grey coarse sand 3.86 dark grey coarse sand 3.96 dark grey coarse sand 4.06 black medium to coarse grained sand 9.36 coarse sand with pebbles (1mm to 1cm) 9.67 coarse homogeneous sand 10.18 medium to fine grained sand with visible mica 10.38 medium grained sand with some silt 11.24 medium to coarse grained dark grey homogeneous sand 11.54 pebbles, coarse sand, clay and wood debris mix 12.05 medium to coarse grained homogeneous sand with visible white mica 12.25 medium to coarse grained homogeneous sand with visible white mica Table 4. Closes t 16S r R N A sequence match to exc ised bands us ing the B L A S T N search tool Band Group sequence Closest match by RDPII Closest match by BLASTN search (gene length/nt classification name and sequence identity) Publication and Environmental location of nearest sequence match Unclassified Bacteria 228 Bacteria [99%] Uncultured bacterium gene for 16S ribosomal RNA, partial sequence, isolate: DGGE band B08 - 98% Microb. Ecol. 52 (4), 765-773 (2006) Associated with sheaths of freshwater and brackish thioploca species. Deltaproteobacteria 481 Smithella [96%] Uncultured bacterium gene for 16S rRNA, Unpublished: Up flow anaerobic sludge partial sequence, clone: HsB48fl - 99% blanket granular sludges. 3 Deltaproteobacteria 478 Geobacter [100%] Uncultured delta proteobacterium clone AKYH1470 16S ribosomal RNA gene, partial sequence - 98% Science 308 (5721), 554-557 (2005) Farm soil adjacent to a silage storage bunker. 4 Deltaproteobacteria 393 Geobacter [90%] Uncultured Geobacter sp. 16S rRNA, partial sequence, clone: FH-28 -98% Unpublished: Amendment with crystalline iron(III) oxides in anoxic rice field soil. 5 Actinobacteria 477 6 Firmicutes 389 Betaproteobacteria 490 Coriobacteriaceae [94%] Clostridium [89%] Betaproteobacteria [100%] Uncultured actinobacterium gene for 16S rRNA, partial sequence, clone: UH-30 -98% Uncultured Clostridium sp. clone Fl-20 16S ribosomal RNA gene, partial sequence - 99% Uncultured beta proteobacterium gene for 16S ribosomal RNA, partial sequence, clone:f- 98% Appl. Environ. Microbiol. 73 (1), 101-109 (2007) Anoxic rice field soil. FEMS Microbiol. Ecol. 57 (3), 396-408 (2006) Fermentative guilds in eutrophic soils of the Florida Everglades Chem.-Geol. 212, 279-290 (2004) Discharged arsenic mine water. Unclassified Bacteria 508 Bacteria [100%] Uncultured bacterium clone C20 16S ribosomal RNA gene, partial sequence 91% Unpublished: nitrobenzene-contaminated river sediment. 9 Actinobacteria 494 Rhodococcus [100%] Rhodococcus erythropolis isolate Ml4-4 16S ribosomal RNA gene, partial sequence - 100% Unpublished: Plant root soil sample. 112 Table 4. cont. Band Group sequence Closest match by RDP II length/nt classification Closest match by BLASTN search (gene name and sequence identity) Publication and Environmental location of nearest sequence match 10 Actinobacteria 496 Microbacterium [100%] Uncultured bacterium partial 16S rRNA gene, clone 27RHU39 - 100% FEMS Microbiol. Ecol. 53 (3), 401-415 (2005) Poplar tree microcosm, rhizosphere, unflooded. 11 Alphaproteobacteria 487 Rhizobiales [100%] Uncultured bacterium clone oB02 16S Unpublished: Arctic permafrost soil ribosomal RNA gene, partial sequence - 98% from Spitsbergen. 12 Betaproteobacteria 493 Burkholderia [100%] Burkholderia ferrariae strain FeGlOl 16S ribosomal RNA gene, partial sequence -100% Int. J. Syst. Evol. Microbiol. 56 (PT 10), 2421-2425 (2006) Iron ore in Brazil. Unclassified ' 4$Q Bacteria Bacteria [100%] Appl. Environ. Microbiol. 72 (9), Uncultured bacterium clone AKAU3954 16S 6288-6298 (2006) ribosomal RNA gene, partial sequence - 96% Uranium Reduction and Reoxidation. 113 Table 5. A comparison of phylogenetic groups from HZ clone library classified in RDP II database to sequences in the NCBI database using the B L A S T N search tool . Group Identifier Fraction Closest match by RDP Closest match by BLASTN search (gene of library classification name and sequence identity) Publication and Environmental location of nearest sequence match Acidobacteria HZ-01 8/90 HZ-02 1/90 HZ-03 1/90 HZ-04 1/90 Actinobacteria HZ-05 2/90 Bacteroidetes HZ-06 3/90 HZ-07 1/90 HZ-08 1/90 HZ-09 1/90 Betaproteobacteria HZ-10 12/90 Family -Acidobacteriaceae [100%] Family -Acidobacteriaceae [100]% Family -Acidobacteriaceae [100%] Family -Acidobacteriaceae [99%] Family -Coriobacteriaceae [82%] Order - Bacteroidales [100%] Order - Bacteroidales [97%] Order - Bacteroidales [91%] Order - Bacteroidales [97%] Genus - Burkholderia [100%] Uncultured bacterium gene for 16S rRNA, partial sequence, clone: RB355 -95% Uncultured Acidobacteria bacterium partial 16S rRNA gene, clone SS LKC22 UA66 - 96% Uncultured Acidobacteria bacterium partial 16S rRNA gene, clone SS_LKC22_UA66 - 90% Uncultured candidate division OP8 bacterium clone HS9-30 16S ribosomal RNA gene, partial sequence - 87% Uncultured bacterium clone FS118-23B-02 16S ribosomal RNA gene, partial sequence - 96% Uncultured bacterium clone Eb89 16S ribosomal RNA gene, partial Sequence -95% Uncultured bacterium clone zEL21 16S ribosomal RNA gene, partial Sequence -89% Uncultured bacterium clone CI75cm.2.10 16S ribosomal RNA gene, partial sequence - 92% Uncultured bacterium clone Eb89 16S ribosomal RNA gene, partial Sequence -96% Burkholderia ferrariae strain FeGlOl 16S ribosomal RNA gene, partial sequence -99% Unpublished: Rhizosphere of Phragmites. Environ. Microbiol. 9 (6), 1523-1534 (2007) Novel Acidobacteria in microbial mats from a chemolithoautotrophically based cave ecosystem. Environ. Microbiol. 9 (6), 1523-1534 (2007) Novel Acidobacteria in microbial mats from a chemolithoautotrophically based cave ecosystem. Microb. Ecol. 51 (2), 209-219 (2006) Metal and hydrocarbon contaminated soil. Environ. Microbiol. 8 (1), 88-99 (2006) Ridge flank crustal fluid. Unpublished: ecology of an autotrophic denitrifying UASB reactor Appl. Environ. Microbiol. 72 (8), 5596-5609(2006) Frasassi sulfidic cave stream biofilm. Unpublished: sandy carbonate sediment. Unpublished: autotrophic denitrifying UASB reactor. Int. J. Syst. Evol. Microbiol. 56 (PT 10), 2421-2425 (2006) Iron ore in Brazil. 114 Table 5. cont. Group Identifier Fraction Closest match by RDP of library classification Closest match by BLASTN search (gene name and sequence identity) Publication and Environmental location of nearest sequence match Betaproteobacteria HZ-11 3/90 HZ-12 2/90 HZ-13 1/90 Deltaproteobacteria HZ-14 4/90 HZ-15 2/90 Chloroflexi Firmicutes HZ-16 1/90 HZ-17 2/90-HZ-18 1/90 HZ-19 1/90 HZ-20 2/90 HZ-21 3/90 Genus - Roseateles [99%] Genus -Aquabacterium [100%] Class - Burkholderiales [80%] Class -Deltaproteobacteria [100%] Family -Desulfobacteraceae [100%] Genus - Anaerolinea [100%] Genus - Clostridium [100%] Order - Clostridiales [88%] Genus -Desulfosporosinus [81%] Genus - Laceyella [85%] Class - Clostridia [88%] Beta proteobacterium KNU3 16S ribosomal RNA gene, partial sequence -98% Uncultured Aquabacterium sp. clone YJQ-2 16S ribosomal RNA gene, partial sequence - 98% Uncultured bacterium clone KIST-JJY029 16S ribosomal RNA gene, partial sequence - 81% Uncultured bacterium gene for 16S rRNA, partial sequence, clone: KNA6-EB08 - 97% Uncultured delta proteobacterium clone 44a-Bl-10 16S ribosomal RNA gene, partial sequence - 96% Uncultured bacterium clone Elev_16S_966 16S ribosomal RNA gene, partial sequence - 90% Clostridium proteolyticum 16S rRNA gene, strain DSM 3090 - 97% Uncultured bacterium clone AKAU3954 16S ribosomal RNA gene, partial Sequence - 94% Uncultured Desulfosporosinus sp. clone CC06 16S ribosomal RNA gene, partial sequence - 98% Uncultured bacterium clone AKIW911 16S ribosomal RNA gene, partial Sequence - 92% Uncultured candidate division JS1 bacterium gene for 16S ribosomal RNA, partial sequence, clone:JTB243 - 97% Unpublished: characterization of novel chitosanase from beta proteobacterium KNU3. Unpublished: pink mat from the spectacles hot spring in Rehai, Tengchong, China. Unpublished: mixed granule from anammox reactor. Appl. Environ. Microbiol. 71 (2), 1084-1088 (2005) Groundwater, 0.2m-captured fraction Unpublished: natural ZnS-producing biofilm dominated by sulfate-reducing bacteria in a subsurface acid mine drainage system. Unpublished: trembling aspen rhizosphere under elevated C02 conditions. Unpublished: Clostridium proteolyticum. Appl. Environ. Microbiol. 72 (9), 6288-6298 (2006) Population Dynamics during Uranium Reduction and Reoxidation. Unpublished: paper pulp column. Proc. Natl. Acad. Sci. U.S.A. 104 (1), 299-304 (2007) Urban aerosol. Mar. Biotechnol. 1 (4), 391-400 (1999) Deepest cold-seep area of the Japan Trench. 115 Table 5. cont. Group Identifier Fraction of Closest match by RDP Closest match by BLASTN search (gene name Publication and Environmental library classification and sequence identity) location of nearest sequence match Firmicutes HZ-22 1/90 Class - Clostridia [83%] Uncultured bacterium clone Tommeliten_BAC55FL 16S ribosomal RNA gene, partial sequence - 90% Unpublished: marine hydrocarbon seep sediment HZ-23 1/90 Genus - Acetivibrio [100%] Uncultured Clostridiales bacterium clone SRB16 16S ribosomal RNA gene, partial - sequence - 94% Unpublished: deep subsurface microorganisms in a mafic sill. Appl. Environ. Microbiol. 72 (4), HZ-24 2/90 Genus - Bacillus [100%] Bacillus sp. WN613 16S ribosomal RNA gene, partial sequence - 97% 2856-2863 (2006) Bacillus spp. from Endolithic and Extreme Environment. Environ. Microbiol. 4 (11), 721-734 HZ-25 2/90 Genus - Acetivibrio Uncultured firmicute 16S rRNA gene, clone (2002) [100%] BM47 - 93% Waste gas-degrading community in an industrial biofilter. Nitrospira HZ-26 2/90 Genus -Magnetobacterium [100%] Uncultured bacterium clone 35-52 16S ribosomal RNA gene, partial Sequence - 93% Unpublished: sediment of Guanting Reservoir. Planctomycetes HZ-27 2/90 HZ-28 1/90 Unclassified bacterium HZ-29 HZ-30 HZ-31 HZ-32 2/90 4/90 1/90 1/90 HZ-33 1/90 Family -Planctomycetaceae [82%] Family -Planctomycetaceae [89%] Genus - Isosphaera [81%] Bacteria [100%>] Bacteria [100%] Bacteria [100%] Bacteria [100%] Uncultured bacterium gene for 16S rRNA, clone: ODP1230B31.20 - 88% Uncultured bacterium gene for 16S rRNA, partial sequence, clone: SK27B-16 - 89% Uncultured bacterium clone MSB-2F12 16S ribosomal RNA gene, partial Sequence - 91% Uncultured bacterium partial 16S rRNA gene, clone Sh765B-TzT-35 - 95% Unidentified bacterium clone TK-SH22 16S ribosomal RNA gene, partial sequence - 96% Uncultured bacterium partial 16S rRNA gene, clone S15B-MN72 -97% Uncultured soil bacterium clone HS9-75 16S ribosomal RNA gene, partial sequence - 95% Proc. Natl. Acad. Sci. U.S.A. 103.(8), 2815-2820 (2006) Methane hydrate bearing subseafloor sediment at the Peru margin. Unpublished: Holocene mud sediment. Unpublished: mangrove soil. Unpublished: uranium mill tailings, soil sample. Unpublished: Lake Tanganyika anoxic hypolimnion. Unpublished: ground water from a monitoring deep-well at a radioactive waste disposal site. Microb. Ecol. 51 (2), 209-219 (2006) Metal and hydrocarbon contaminated soil. 116 Table 5. cont. Group Identifier Fraction of library Closest match by RDP classification Closest match by BLASTN search (gene name and sequence identity) Publication and Environmental location of nearest sequence match Unclassified bacterium HZ-34 3/90 HZ-35 1/90 HZ-36 1/90 HZ-37 HZ-38 HZ-39 HZ-40 HZ-41 HZ-42 HZ-43 HZ-44 2/90 2/90 1/90 1/90 1/90 1/90 1/90 1/90 AF402980 1439 bp DNA linear ENV 29-APR-Bacteria 2004 DEFINITION Uncultured bacterium clone [100%] bacteriap35 16S ribosomal RNA gene, partial sequence - 89% Bacteria Uncultured bacterium clone 655966 16S ribosomal [ 100%] RNA gene, partial. Sequence - 79% OP 11 r 100"/1 Uncultured bacterium clone WM61 16S ribosomal RNA gene, partial Sequence - 98%> Bacteria Uncultured bacterium clone pLW-102 16S [100%] ribosomal RNA gene, partial Sequence - 91% Bacteria Uncultured bacterium clone MSB-2F12 16S [100%] ribosomal RNA gene, partial Sequence - 90% Bacteria Uncultured bacterium clone SMIl-GC205-Bac58 [100%] 16S ribosomal RNA gene, partial sequence - 90%) Bacteria Uncultured bacterium partial 16S rRNA gene, [100%] clone C5LKS28 - 92% Bacteria Uncultured bacterium clone 656068 16S ribosomal [ 100%] RNA gene, partial Sequence - 86% Bacteria Uncultured Acidobacteria bacterium partial 16S [100%] rRNA gerie, clone SSLKC22JJB76 - 92% Bacteria Uncultured bacterium partial 16S rRNA gene, [100%] clone S15B-MN72 - 96% Bacteria Unidentified bacterium clone TK-NH7 16S [100%] ribosomal RNA gene, partial Sequence - 91% Extremophiles 7 (1), 63-70 (2003) Novel extremely thermostable 1,4-beta-xylanase isolated directly from an environmental DNA sample. Appl. Environ. Microbiol. 72 (5), 3291-3301 (2006) Contaminated sediment. Appl. Environ. Microbiol. 72 (8), 5596-5609 (2006) Frasassi sulfidic cave stream biofilm. Appl. Environ. Microbiol. 71 (11), 6885-6899 (2005) Populations active in metabolism of c 1 compounds in the sediment of Lake Washington. Unpublished: mangrove soil. Appl. Environ. Microbiol. 72 (11), 7218-7230 (2006) Anaerobic Methane-Oxidizing Community of ANME-lb Archaea in Hypersaline Gulf of Mexico Sediments. Syst, Appl. Microbiol. 30 (3), 239-254 (2007) Profundal lake sediment. Appl. Environ. Microbiol. 72 (5), 3291-3301 (2006) Contaminated sediment. Environ, Microbiol. 9 (6), 1523-1534 (2007) Microbial mats from a chemolithoautotrophically based cave ecosystem. Unpublished: groundwater from a monitoring deep-well at a radioactive waste disposal site. Unpublished: Lake Tanganyika anoxic hypolimnion. 117 Table 5. cont. Group Identifier Fraction of library Closest match by RDP classification Closest match by B L A S T N search (gene name and sequence identity) Publication and Environmental location of nearest sequence match Unclassified bacterium HZ-45 HZ-46 1/90 1/90 Bacteria [100%] Bacteria [100%] Uncultured bacterium clone Kasl39B 16S ribosomal R N A gene, partial Sequence - 87% Unidentified bacterium clone TK-NH7 16S ribosomal R N A gene, partial Sequence - 91% Unpublished: sediments of Lake Kastoria, Greece. Unpublished: Lake Tanganyika anoxic hypolimnion. 118 Table 6. A comparison of phylogenetic groups from BHZ clone library classified in RDP II database to sequences in the NCBI database using the B L A S T N search tool " Group T J Fraction Closest match by RDP Identifier .. . . , . J of library classification Closest match by BLASTN search (gene name and sequence identity) Publication and Environmental location of nearest sequence match Actinobacteria Bacteroidetes BHZ-01 23/89 BHZ-02 49/89 BHZ-03 1/89 BHZ-04 1/89 Betaproteobacteria BHZ-05 1/89 Deltaproteobacteria BHZ-06 1/89 Gammaproteobacteria BHZ-07 1/89 BHZ-08 . 1/89 BHZ-09 1/89 Genus - Rhodococcus [100%] Genus - Rhodococcus [100%] Genus -Propionibacterium [100%] Genus - Chitinophaga [99%] Betaproteobacteria [100%] Deltaproteobacteria [99%] Genus- Pseudomonas [100%] Gammaproteobacteria [99%] Gammaproteobacteria [100%] Rhodococcus sp. djl-6 16S ribosomal RNA gene, partial sequence -100% Rhodococcus erythropolis strain HS8 16S ribosomal RNA gene, partial Sequence -99% Uncultured Propionibacterineae bacterium clone PH-B24N 16S ribosomal RNA gene, partial sequence - 99% Uncultured bacterium clone 655011 16S ribosomal RNA gene, partial Sequence -100% Uncultured bacterium clone pLW-9 16S ribosomal RNA gene, partial Sequence — 97% Uncultured bacterium SHA-207 16S rRNA gene - 93% Pseudomonas sp. Nj-55 partial 16S rRNA gene, strain Nj-55 - 99% Uncultured proteobacterium clone R7C24 16S ribosomal RNA gene, partial sequence - 96% Uncultured eubacterium WD2124 partial 16S rRNA gene, clone WD2124 - 96% Curr. Microbiol. 53 (1), 72-76 (2006) Isolation and Characterization of a Carbendazim-Degrading Rhodococcus sp. djl-6. Unpublished: RDX degrading bacteria. Unpublished: Hypersaline Lake Laysan and a brackish pond on Pearl and Hermes AtolL Appl. Environ. Microbiol. 72 (5), 3291-3301 (2006) Contaminated sediment. Appl. Environ. Microbiol. 71 (11), 6885-6899 (2005) Metabolism of cl compounds in the sediment of lake Washington. Int! J. Syst. Evol. Microbiol. 50 PT 4, 1505-1511 (2000) Anaerobic 1,2-dichloropropane-dechlorinating mixed culture. Unpublished: Antarctica. Appl. Environ. Microbiol. 72 (10), 6560-6568 (2006) Coastal lagoon of the southwestern Atlantic ocean. Appl. Environ. Microbiol. 67 (4), 1874-1884 (2001) Polychlorinated biphenyl-polluted soil. 119 Table 6. cont. Group T, i . _ Fraction of Closest match by RDP Identifier , ^ library classification Closest match by BLASTN search (gene name and sequence identity) Publication and Environmental location of nearest sequence match Firmicutes BHZ-10 2/89 BHZ-11 1/89 Unclassified bacterium BHZ-12 1/89 BHZ-13 1/89 BHZ-14 2/89 BHZ-15 1/89 BHZ-16 1/89 BHZ-17 1/89 Order - Clostridiales [100%] Genus - Propionispora [82%] Family -Acidaminococcaceae [100%] Order - Clostridiales [100%] Genus -Cryptanaerobacter [96%] Genus - Streptococcus [100%] Genus - Streptococcus [100%] Bacteria [100%] Uncultured bacterium gene for 16S rRNA, partial sequence, clone: HDBW-WB61 -96% Uncultured bacterium clone AKAU3872 16S ribosomal RNA gene, partial Sequence - 98% Unidentified eubacterium from anoxic bulk.soil 16S rRNA gene (clone BSV90) - 98% Uncultured bacterium clone PL-35B5 16S ribosomal RNA gene, partial Sequence - 98% Uncultured bacterium clone A8 16S ribosomal RNA gene, partial Sequence - 97% Streptococcus thermophilus gene for 16S ribosomal RNA, strain:DLl -99% Streptococcus mitis strain Sm91 16S ribosomal RNA gene, partial Sequence - 93% Uncultured bacterium gene for 16S rRNA, partial sequence, clone: RB313 -82% Geobiology4, 147-223 (2006) Fault-bordered aquifers in the Miocene formation of northernmost Japan. Appl. Environ. Microbiol. 72 (9), 6288-6298 (2006) Uranium Reduction and Reoxidation. Appl. Environ. Microbiol. 65 (11), 5050-5058(1999) . Anoxic rice paddy soil. FEMS Microbiol. Ecol. 54 (3), 427-443 (2005) Production waters of a low-temperature biodegraded oil reservoir. Sediment Capable of Complete Anaerobic 2,3,4,5-Tetrachlorobiphenyl Dechlorination to Biphenyl. Unpublished: Georgia yogurt. J. Clin. Microbiol. 41 (7), 3051-3055 (2003) Toxic shock-like syndrome caused by Streptococcus mitis. Unpublished: Rhizosphere of Phragmites. 120 Table 7. A comparison of phylogenetic groups from SZ clone library classified in RDP II database to sequences in the NCBI database using the B L A S T N search tool Group Fraction c l o s e s t match by RDP Identifier ..,., classification - cut off 0 f l l b r a r y above 80% Closest match by BLASTN search (gene name and sequence identity) Publication and Environmental location of nearest sequence match Actinobacteria AZ-01 6/87 Betaproteobacteria AZ-02 44/87 AZ-03 7/87 AZ-04 2/87 AZ-05 2/87 Firmicutes AZ-07 2/87 AZ-08 2/87 AZ-09 2/87 Class - Actinobacteria [83%] Genus - Burkholderia [100%] Genus - Burkholderia [100%] Genus - Ralstonia [100%] Genus - Alcaligenes [100%] AZ-06 10/87 Clostridiales [84%] Genus - Clostridium [100%] Genus - Cryptanaerobacter [89%] Family - Clostridiaceae [100%] Uncultured bacterium clone Napoli-1B-12 16S ribosomal RNA gene, partial sequence - 83% Burkholderia ferrariae strain FeGlOl 16S ribosomal RNA gene, partial sequence - 99% Burkholderia ferrariae strain FeGlOl 16S ribosomal RNA gene, partial sequence. - 96% Iron-reducing enrichment clone Cl-A6 clone C1-A6 16S ribosomal RNA gene, partial sequence - 96% Alcaligenes sp. IS-67 16S ribosomal RNA gene, complete sequence - 99% Uncultured bacterium clone AKAU3954 16S ribosomal RNA gene, partial sequence - 94% Clostridium chromoreductans 16S ribosomal RNA gene, partial sequence - 98% Uncultured bacterium clone A8 16S ribosomal RNA gene, partial sequence - 93% Clostridium septicum 16S ribosomal RNA gene, partial sequence - 88% Unpublished: deep-sea mud volcanoes. Int. J. Syst. Evol. Microbiol. 56 (PT 10), 2421-2425 (2006) Isolated from an iron ore in Brazil. Int. J. Syst. Evol. Microbiol. 56 (PT 10), 2421-2425 (2006) Isolated from an iron ore in Brazil. Unpublished: iron-reducing enrichments of the Scheldt estuary sediment. FEMS Microbiol. Lett. 237 (2), 369-375 (2004) phenol hydroxylase genes among phenol-degrading isolates of Alcaligenes sp. from an activated sludge system. Appl. Environ. Microbiol. 72 (9), 6288-6298 (2006) Uranium Reduction and Reoxidation. Unpublished: novel anaerobic Cr(VI)-reducing bacterium isolated from soil. Unpublished: Sediment Capable of Complete Anaerobic 2,3,4,5-Tetrachlorobiphenyl Dechlorination to Biphenyl. Int. J. Syst. Bacteriol. 46 (4), 1174-1176 (1996) Clostridium septicum. 121 Table 7. cont. Group Fraction c l o s e s t match by RDP Identifier .. . . classification-cut off above ofhbrary g Q % Closest match by BLASTN search (gene name and sequence identity) Publication and Environmental location of nearest sequence match Unclassified bacterium AZ-10 1/87 AZ-11 1/87 AZ-121 1/87 AZ-13 1/87 AZ-15 1/87 AZ-16 1/87 AZ-17 1/87 AZ-18 1/87 AZ-19 1/87 Class - Clostridia [86%] Class - Clostridia [92%] Family -Acidaminococcaceae [97%] Family -Acidaminococcaceae [100%] AZ-14 1/87 , Order-Clostridiales [90%] Bacteria [100%] Bacteria [100%] Bacteria [100%] Bacteria [100%] Bacteria [100%] Benzene mineralizing consortium clone SB-15 16S small subunit ribosomal RNA gene, complete sequence - 97% Uncultured candidate division JS1 bacterium gene for 16S ribosomal RNA, partial sequence, clone: JTB243 - 96% Uncultured bacterium clone AKIW621 16S ribosomal RNA gene, partial sequence -96% Pelosinus fermentans strain UFO 1 16S ribosomal RNA gene, partial sequence -97% Uncultured bacterium SJA-136 16S rRNA gene, clone SJA-136 - 91% Uncultured bacterium clone G55_D25_M_B_G02 16S ribosomal RNA gene, partial sequence - 84% Uncultured bacterium clone FS117-62B-02 16S ribosomal RNA gene, partial sequence - 82% Unidentified bacterium clone TK-NH1 16S ribosomal RNA gene, partial sequence -98% Uncultured bacterium clone RL178_aah51d07 16S ribosomal RNA gene, partial sequence - 88% Uncultured bacterium clone 661230 16S ribosomal RNA gene, partial sequence -85% FEMS Microbiol. Ecol. 27 (3), 269-279 (1998) Benzene mineralizing consortium clone SB-15. Mar. Biotechnol. 1 (4), 391-400 (1999) Deepest Cold-Seep Area, the Japan Trench. Proc. Natl. Acad. Sci. U.S.A. 104 (1), 299-304 (2007) Urban aerosol. Unpublished: transformation of AQDS, Fe(III), Cr(VI), and U(VI) by Pelosinus fermentans strain UFOl. Appl. Environ. Microbiol. 65 (1), 283-286(1999) Anaerobic, trichlorobenzene-transforming microbial consortium. Unpublished: organic metabolism in thermophilic anaerobic solid waste degradation. Environ. Microbiol. 8(1), 88-99 (2006) Ridge flank crustal fluids. Unpublished: Lake Tanganyika anoxic hypolimnion. Nature 444 (7122), 1022-1023 (2006) Human gut microbes associated with obesity. Appl. Environ. Microbiol. 72 (5), 3291-3301 (2006) Contaminated sediment. 122 Table 8. Comparison of Shannon index (H) and associated Shannon equitability (EH) calculated from the relative abundance values of phylotypes in clone libraries and corresponding D G G E fingerprints Clone library/27F-1492R DGGE profile/357F-907R Sample: H E H H E H HZ-1.16mbrb 3.51 0.92 3.53 0.92 BHZ-3.86mbrb 1.50 0.53 . 1.50 0.53 AZ - 10.38mbrb 1.93 0.65 1.69 0.58 Figure 1. Geographic location of Braid Street site. B r a i d Street site is located o n the north a r m o f the Fraser R i v e r i n N e w Westmins ter i n B r i t i s h C o l u m b i a , Canada. The site has been an active w o o d preserving fac i l i ty since the 1920's . 124 -20 3 .25 q , , , , , ,—I 0 20 40 60 80 100 120 D i s t a n c e f r o m S h o r e [ m ] Figure 2. Seasonal Groundwater gradients and diurnal river water ingression General groundwater gradients change seasonally at B r a i d Street site, (a) D u r i n g freshet the net groundwater f l o w is directed towards the shore, (b) In the rainy season the groundwater is directed towards the discharge point, (c) A d d i t i o n a l l y , groundwater penetrates the aquifer d iurna l ly w i t h h igh r iver stage e levat ion dur ing h i g h tide and is discharged w i t h l o w r iver stage elevat ion dur ing l o w tide. 125 5 : 0 20 40 60 80 100 120 Distance from Shore Tml Figure 3. Location of sampling in Braid Street site Four zones were ana lyzed at B r a i d Street site dur ing freshet i n 2004, 2005 or 2006. Sediment cores and pore water prof i les were taken f rom each zone: (a) The S Z sampled i n 2006 located outside the contaminant p lume; the pore water and sediment profi les here were col lec ted a long the same transect, however they do not correspond i n depth and locat ion, (b) The hyporheic zone/bot tom o f hyporheic zone ( H Z / B H Z ) sampled i n 2004 and 2005 located inside the contaminant p lume; the H Z is represented by depths d o w n to 2.5 mbrb and the B H Z at depths b e l o w 2.5 mbrb . (c) T h e anaerobic zone ( A Z ) sampled i n 2005 located inside the core o f the contaminant plume, where previous studies determined through geochemica l analysis the condi t ions to be F e 3 + r educ ing a long w i t h possible methanogenesis (6, 16, 19). 126 (a) 0 25000 .. 50000 75000 (b) r 2 50000 100000 150000. 0 • Indane, Dissolved iron/ppb A Total iron, Fe2+ iron/ppb A Naphthalene, Benzothiophene/ppb ( Figure 4. Dissolved iron, Fe , total iron, indane, benzothiophene and naphthalene and location of sediment samples corresponding to profiles Selected geochemical data f rom the 2005 B r a i d Street site pore water profiles col lected at the (a) Surface zone ( S Z ) , (b) H y p o r h e i c zone /Bot tom o f hyporheic zone ( H Z / B H Z ) , and (c) Anae rob ic zone ( A Z ) . The ful l analysis o f geochemical data i nc lud ing the above data can be found i n Table 1. Dashed l ines represent where corresponding sediment samples were obtained for the bacterial communi ty characterization. 127 (a) ••~*TnB Y Enrichments Surface zone Hyporheic zone Bottom of Hyporheic zone Anaerobic zone (b) 50" E CO 100" \r <V ^ 7 Figure 5. DGGE analysis of bacterial 16S rRNA gene using primer 357F-907R D G G E analysis o f bacterial part ial 16S r R N A genes isolated f rom var ious sediment depths at the B r a i d Street site and F e 3 + - r e d u c i n g enrichment cultures. L a b e l s indicate the depths i n mbrb that correspond to pore water profi les and the different F e 3 + reducing enrichment cultures, (a) B a n d i n g patterns seen i n enr ichment cultures and sediment communi t ies are represented here. M a r k e d D G G E bands were exc ised and sequenced, where band identities can be found i n Table 4. (b) The Pearson correla t ion c lus ter ing o f banding patterns is depicted by the dendogram. T h e ver t ical length o f the l ines i n the dendogram represent the percent s imi la r i ty o f banding patterns connected v i a hor izonta l l ines. 128 10 n 7 -J — , , , , = , , 0 100 200 300 400 500 600 indane/ppb Figure 6. Burkholderia band intensity versus indane concentration The natural logarithm of the B. ferrariae band from the sediment communities significantly correlated (<0.01 and <0.001 in the HZ (a) and A Z (b) respectively) to indane concentrations measured in the corresponding pore water profiles. The slope of the regression line correlating B. ferrariae to indane in the HZ was 6x the slope of the regression line determined from the AZ. 129 27F-1492R HZ-1.86 BHZ-3.86 AZ-10.38 ^ Acidobacteria H Actinobacteria H Bacteroidetes S b-proteobacteria • d-proteobacteria g-proteobacteria • Chloroflexi ICE Firmicutes 11 Nitrospira &2 Planctomycetes • unclassified bacteria 0% 50% 100% Figure 7. Distribution and relative abundance of clone libraries The distribution of bacterial 16S r R N A gene sequences obtained using the primer set 27F-1492R from the sediment samples along with the relative abundance of different groups is illustrated here. Clone libraries were constructed from a representative sample in the H Z at 1.86 mbrb (a), in the B H Z at 3.86 mbrb (b) and lastly in the A Z at 10.38 mbrb (c). Numbers of clones analyzed are shown in parentheses. 130 131 Rhodococcus erythropolis BHZ-02, 0.55 ' BHZ-01,0.26 AZ-01,0.07 - HZ-05. 0.02. -£ HZ-30, 0.04 — Desulfosarcina cetonicS" a:' HZ-14. 0.02 HZ-15. 0.04 . _ _ _ _ _ _ _ — HZ-21. 0.03 , AZ-11, 0.02 — Cryptanaerobacter phenolicus BHZ-14, 0.02 C AZ-07, 0.11 AZ-09, 0.02 HZ-24. 0.02 Thermoactinomyces vulgaris — HZ-20, 0.02 Clostridium histolyticum HZ-17, 0.02 Clostridium acetobutylicum AZ-08, 0.02 — BHZ-10, 0.02 — Clostridium josui HZ-25,0.02 Geothrix termentas HZ-01, 0.09 HZ-27, 0.02 HZ-29, 0.02 HZ-08, 0.03 HZ-37, 0.02" HZ-44, 0 .0_ HZ-26, 0.02 • HZ-34. 0.03 AZ-04, 0.02 Ralstonia syzygii r Alcaligenes faecalis AZ-05, 0.02 HZ-12, 0.02 Aauabacterium oarvum HZ-11.0.03 AZ-03, 0.08 Burkholderia tropica Burkholderia sacchari AZ-02, 0.5 HZ-10, 0.13 Actinobacteria "JUnclassified Deltaproteobacteria Firmicutes _~J Acidobacteria Planctomycetes Bacteroidetes Unclassified Nitrospira 3 Unclassified Betaproteobacteria Scale :1 I Figure 8. Phylogenetic tree of clones that dominated at least 2% of the clone libraries Phylogenetic tree of representative bacterial 16S rRNA gene sequences found in Tables 5, 6 and 7., The tree is based on the Weighbour algorithm and is derived from a distance matrix created from the Jukes-Cantour distance model of positions aligned by RDP's aligner in the RDP II database. The tree is based on a maximum of 1465 bases of aligned 16S rRNA gene sequences. Bootstrap values are shown. Type strains whose sequence identity's were 90% that of the Braid Street site phylotypes were included. A P P E N D I X P H O S P H O L I P I D F A T T Y A C I D ( P L F A ) M A S S S P E C T R A Fi le : D:\MSDCHEM\1\DATA\K070510A\0201002.D Operator : Gord Acquired : 10 May 2007 12:03 using AcqMethod PLFA2 Instrument : Biotech5 9 Sample Name: 6 Misc Info Via l Number: 2 Abundance 1.1e+07 TIC: 0201002.D 1e+07 9000000 8000000 7000000 6000000 5000000 4000000 3000000 2000000 1000000 Time-> •4 I ' ' ' ' I ' ' ' f I ' ' ' ' I ' ' ' ' I ' ' ' .' I ' ' ' ' I | • i • i i i • i i 1 1 1 1 I 1 1 1 1 I 1 1 1 1 I 1 1 1 1 T 1 v 1 1 I 1 1 1 1 I 1 1 1 1 I 1 1 1 1 I M 1 1 I 1 1 1 1 I 1 1 1 1 I 1 .' ' 1 I 1 1 1 1 r 1 1 1 r 1 1 1 i 6.00 8.00 10.00 12.00 14.00 16.00 18.00 20.00 22.00 24.00 26.00 28.00 30.00 32.00 34.00 36.00 38.00 40.00 42.00 44.00 46.00 48.00 50.00 52.00 54.00 F i l e Operator Acquired Instrument Sample Name Misc Info V i a l Number D:\MSDCHEM\1\DATA\K070510A\0301003.D Gord 10 May 2007 13:08 using AcqMethod PLFA2 Biotech59 7 Abundance 1.2e+07 TIC: 0301003.D i:ie+07 1e+07 9000000 8000000| 7000000 6000000 5000000 4000000 3000000 2000000 1000000 T i m e " > — . - 6 - i » - i i k _ 1 0 | g g ^ 38!og' 4oloO 42I0O 44^ 00 46:00 •&0Q Soldo' SZOO 54^ 0 F i l e D : \ M S D C H E M \ 1 \ D A T A \ K 0 7 0 5 1 0 A \ 0 5 0 1 0 0 5 . D O p e r a t o r G o r d A c q u i r e d 10 May 2007 15:18 u s i n g AcqMethod PLFA2 I n s t r u m e n t B i o t e c h 5 9 Sample Name 14 M i s c I n f o V i a l Number 5 Abundance 1.2e+07 TIC: 0501005.D 1.1e+07 1e+07 9000000 8000000 7000000 6000000 5000000 4000000 3000000 2000000 1000000 I I 1 1 1 1 I ' 1 ' 1 I 1 ' ' 1 1 1 1 1 1 I ' 1 1 1 I 1 ' 1 1 I ' 1 1 1 I ' I I I I ' ' ' I I I ' ' I I I I I I I I I I I I • I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I • I I I I I [ Time-> 6.00 8.00 10.00 12.00 14.00 16.00 18.00 20.00 22.00 24.00 26.00 28.00 30.00 32,00 34.00 36.00 38.00 40.00 42.00 44.00 46.00 48.00 50.00 . 52.00 54.00 " 135 F i l e O p e r a t o r A c q u i r e d I n s t r u m e n t S a m p l e Name M i s c I n f o V i a l N u m b e r D : \ M S D C H E M \ l \ D A T A \ K 0 7 0 5 1 0 A \ 0 4 0 1 0 0 4 . D G o r d 10 M a y 2 0 0 7 1 4 : 1 3 u s i n g A c q M e t h o d P L F A 2 B i o t e c h 5 9 12 Abundance 1.2e+07 1.1e+07 TIC: 0401004.D 1e+07 9000000 8000000 7000000 6000000 5000000 4000000 3000000 2000000 1000000 T i m e ~ > M0._AQQ„jmQ^lM0 ulo'o 16loO isloo 20I0O 22I0O 24.'oO 26.00 28^0 30I0O 32^0 3 4 I 0 0 ' 36. 00 38.00 40.00 42.00 44.00 46.00 48.00 50.00 52X)Q 54.00 136 Fi l e Operator Acquired Instrument Sample Name Misc Info V i a l Number D:\MSDCHEM\1\DATA\K070510A\0601006.D Gord 10 May 2007 16:24 using AcqMethod PLFA2 Biotech59 15 Abundance TIC: 0601006.D 1e+07 9000000 8000000 7000000 6000000 5000000 4000000 3000000 F i l e D:\MSDCHEM\1\DATA\K070510A\0701007.D Operator Gord Acquired 10 May 2007 17:29 using AcqMethod PLFA2 Instrument Biotech59 Sample Name 3-2 • - . Misc Info V i a l Number 7 Abundance 2000000 1900000 1800000 1700000 1600000 1500000 1400000 1300000 1200000 1100000 1000000 900000 800000 700000 600000 500000 400000 300000 200000' 100000 TIC: 0701007.D J Time-> i 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i •' 1 1 1 i 1 1 1 1 i 1 1 1 1 I 1 1 1 1 I 1 1 1 1 I 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i 1 ' 1 1 I 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i 1 1 1 1 i 1 1 1 1 I 1 1 6.00 8.00 10.00 12.00 14.00 16.00 18.00 20.00 22.00 24.00 26.00 28.00 30.00 32.00 34.00 36.00 38.00 40.00 42.00 44.00 46.00 48.00 50.00 52.00 54.00 138 


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