Applying Innovative Diagnostic Tools at New Jersey Publicly Funded Sites

Home , Acidovorax, Acidovorax defluvii, Acidovorax facilis, Acidovorax temperans, Anaerolinea, Aquincola

Final Report New Jersey Department of Environmental Protection

Project Title: Applying Innovative Diagnostic Tools at New Jersey Publicly Funded Sites

Investigators: Donna E. Fennell1 and Max Häggblom2

1 Department of Environmental Sciences 2 Department of Biochemistry and Microbiology Rutgers University

Mailing Address: Department of Environmental Sciences 14 College Farm Road New Brunswick, NJ 08901 Phone: 848-932-5748 Fax: 732-932-88644 Email: [email protected]

Project Manager: Robert Mueller Total budget Year 1: $83,500 Project Duration: 30 Jun 2011 to NCE requested to 31 Dec 2013 Reporting Period: 15 July 2012 – 31 December 2013

Contract number: SR11-025

Acknowledgements

This study was funded by the New Jersey Department of Environmental Protection (NJDEP). This project supported the research of Dr. Weimin Sun, Post-Doctoral Associate, in the Department of Environmental Sciences at Rutgers University. We thank Ms. Lora McGuinness and Professor Lee J. Kerkhof of Rutgers University for assistance with stable isotope probing analyses. We acknowledge Mr. Shuai Luo, a Rutgers University undergraduate student, who assisted with laboratory work. We thank Mr. Eric Raes of Engineering and Land Planning Associates, Inc. for enabling access to contaminated sites and for sharing information. We thank Mr. Robert (Bob) Mueller for his support of this project and for establishing site contacts.

1 Table of Contents

Acknowledgements 1

Executive Summary 3

Introduction/Problem Statement 4

Project Objectives 4

Project Design and Methods 5

Quality Assurance 8

Results and Discussion 9

Conclusions and Future Work 20

Recommendations and Application and Use by NJDEP 21

Expenditures 22

References 24

Project Publications and Presentations 27

2 Executive Summary

This project demonstrated use of environmental molecular diagnostics (EMDs) for detecting microbial biodegradation of contaminants and identifying bacteria responsible for contaminant biodegradation or biotransformation at three contaminated sites in New Jersey.

Three site studies were performed. First, sediment cores were collected from the contaminated aquifer at the former Pagan-Martinez site (Jersey City, NJ) to aid an on-going bioremediation effort. The former Pagan-Martinez site has perchloroethene and trichloroethene contamination along with fuel oil contamination. A work plan was developed for the site and microcosms were established using the aquifer core materials. After six months, no dechlorination of PCE was detected in the microcosms from the site. The site has since undergone bioaugmentation and biostimulation. An ongoing effort involves community analysis of the original aquifer sediments using terminal restriction fragment length polymorphism and sequencing.

Methods to target metabolically active bacteria have been developed using isotope incorporation (e.g., 13C) into newly biosynthesized macromolecules. This is termed stable isotope probing (SIP). The 13C DNA can be recovered separately from the unlabeled DNA from other organisms. In the second project, stable isotope probing (SIP) was used to identify dehalorespiring bacteria in sediments from the organochloride-contaminated Hackensack River, NJ. Certain dehalorespiring bacteria do not assimilate carbon from the chlorinated compounds they use as electron acceptors. To label DNA from dechlorinators using SIP, site materials were exposed to 13C-based carbon sources that are assimilated by the bacteria (i.e., 13C-acetate and 13C-labeled lactate which is fermented to 13C-labeled acetate and hydrogen). The bacterial community in active Hackensack River microcosms included Dehalococcoides, Dehalogenimonas, Desulfobacterium, Geobacteraceae and Desulfuromonas as identified in the clone library derived from 13C-heavy fraction. Dehalococcoides was detected when 13C-labeled lactate was added to microcosms, but not when 13C-labeled acetate alone was added. This indicated that both hydrogen and acetate were needed to stimulate dechlorination by Dehalococcoides. This method could also be used to guide and monitor the bioremediation effort at the Pagan-Martinez site after bioaugmentation (since no dechlorination was observed in the native sediments).

Finally in the third effort supported by this project, bacteria involved in the biodegradation of aniline at a chemical industrial contaminated site were identified. Results from microcosms operated over many months as part of an earlier project indicated loss of aniline in site sediments. However, the biodegradation of aniline was not conclusively demonstrated because intermediates were not identified and loss from sorption could not be ruled out. Results using SIP with 13C-labeled aniline conclusively demonstrated that aniline biodegradation occurred and the specific bacteria responsible for biodegradation under different redox conditions were identified. The use of this molecular method therefore provided evidence for bioremediation and specifically identified bacterial targets for monitoring further bioremediation efforts.

This study applied EMD techniques at contaminated sites in New Jersey. Results have greatly enhanced enriched the findings that would normally be obtained using traditional site investigation tools.

3 Introduction/Problem Statement Management of contaminated sites encompasses everything from site characterization to ultimate selection, implementation, and completion of a site remedial action plan and then finally, site case closure. The investigation begins with characterization that includes specifying the contaminants and media (e.g., groundwater aquifer, sediments or soils) of interest, assessing exposure pathways, and determining remediation goals. The estimated volumes or areas of the site to be remediated must be defined. Finally, detail concerning the equipment, methods, and locations to be evaluated for each possible site remediation approach alternative (e.g., potential natural recovery processes, potential remediation plan, need for monitoring and/or institutional controls, etc.) must be determined. To the extent possible the time frame(s) in which alternatives are expected to achieve cleanup levels are also estimated. Remediation of contaminated sites may be performed using physical-chemical processes such as pump and treat with sorption or volatilization, dredging and landfilling, or chemical oxidation. More sustainable remediation options include bioremediation approaches to degrade or transform pollutants to benign end products. The Environmental Protection Agency’s definition of bioremediation is “a managed or spontaneous process in which (micro)biological processes are used to degrade or transform contaminants to less toxic or nontoxic forms, thereby remedying or eliminating environmental contamination” (USEPA). To assess the potential for application of bioremediation to a site or to select the type of bioprocess to be used in bioremediation, a variety of assessment protocols have been developed that provide guidance for monitored natural attenuation (MNA) or enhanced bioremediation based on geochemical indicators such as pH, prevailing redox conditions as indicated by the presence of various electron acceptors, and presence of transformation products of the original contaminants (for example see rating protocols from USEPA [http://www.afcee.af.mil/shared/media/document/AFD-071210-020.pdf; accessed online April 2014] and the Department of Defense [www.serdp.org/content/download/3728/59365/file/Rabitt_Protocol.pdf; accessed online April 2014]). To directly assess potential natural or enhanced biological processes for contaminant removal, powerful new environmental molecular diagnostics (EMDs) that detect and quantify the microorganisms, genes and enzymes responsible for specific processes are also recommended for use in site monitoring and management [http://www.itrcweb.org/teampublic_EMD.asp; accessed online April 2014]. These tools provide critical information to support other field data.

Project Objectives The overarching goal of this project was to demonstrate applications for innovative environmental molecular diagnostics (EMDs) for informing bioremediation at contaminated sites in New Jersey. The specific objectives were:

1. Utilize emerging environmental molecular diagnostics (EMDs) (in-house and commercially available) to facilitate and guide site assessment to direct a bioremediation approach. 2. Assess a site that is or has undergone unsuccessful or stalled bioremediation and use diagnostic tools for re-assessment to re-screen and re-initiate remedial actions.

4 3. Provide data through EMD analyses to support site closure at one or more sites that have been successfully remediated or which are undergoing monitored natural attenuation. 4. Perform research in support of development and application of EMDs for novel bioremediation approaches for emerging contaminants or for those with no commercially available EMDs.

In this final report we report results and findings from three NJ contaminated sites: 1. Microbial investigation of the former Pagan-Martinez site, Jersey City, NJ 2. Stable isotope probing to identify dechlorinating bacteria in the highly contaminated sediments of the Hackensack River, NJ 3. Stable isotope probing investigation of aniline

degraders under different redox conditions at a large Figure 1. Sample collection at industrial site in NJ the Pagan-Martinez site in Jersey

City, NJ. Project Methods Collection of Environmental Samples Aquifer materials. Sediment cores and groundwater samples were collected from the former Pagan Martinez site in Jersey City, NJ in March 2013 under the direction of Land Planning Associates, Inc. (see Figure 1). At this site a leaking fuel oil tank was excavated and the excavation was back-filled with soil containing a variety of chlorinated solvents. After the tank was removed, ground water analysis detected tetrachloroethene (PCE) and trichloroethene (TCE) and other chlorinated ethanes. The sediment cores collected for this study include those from the contaminated groundwater plume and from uncontaminated portions of the aquifer as controls. In all, four sediment cores from different locations and two groundwater samples from different monitoring wells were collected. Samples were transported immediately to the Department of Environmental Sciences at Rutgers University and stored at 4°C until use. Hackensack River sediments. Surficial grab samples were obtained from the Hackensack River (Figure 2). Samples were placed in sterile jars, transported immediately to the Department of Environmental Sciences and stored at 4°C until use.

Microcosm protocols

Dechlorinating microcosms. Anaerobic Figure 2. Hackensack River, NJ. microcosms were used to identify the interacting and interdependent microorganisms involved in dechlorinating

5 communities with tetrachloroethene (PCE) used as a model chlorinated compound. Organohalide contaminated river sediments (Hackensack River) and PCE contaminated aquifer sediments (Pagan-Martinez site, Jersey City, NJ) were studied. For the Pagan-Martinez site, 24 microcosms seeded from sediment cores were set up to test in situ dechlorination capability. Briefly, a top surface sediment and bottom sediment from each core were used as inoculum to set up triplicate microcosms. For the Hackensack sediments triplicate microcosms were established from several locations. Microcosms were amended with PCE as the electron acceptor and either acetate or lactate was added as electron donors/carbon sources or electron-hydrogen donors/carbon sources, respectively, for the dechlorinators. 13C- acetate and 13C-lactate were provided to parallel dechlorinating microcosms in order to label the population of interest for further analysis. All original sediments and digester sludge were sub- sampled for later molecular analysis and stored at -80ºC. Aniline-degrading microcosms. Aerobic and anaerobic microcosms prepared from environmental media collected for a previous study from a contaminated site in New Jersey were used for development of protocols for this project and for scientific study. These sediments and microcosms were part of an existing project in our laboratory. Briefly, sediment slurries of 100 mL containing 20 mL of sediment and 80 mL of site water or minimal media (Fennell et al., 2004) were constructed under sterile conditions. Microcosms seeded with contaminated sediments (freshwater canal or groundwater aquifer) were prepared under aerobic, denitrifying or methanogenic conditions. The SIP experiment incorporated nine microcosms for each treatment. 13 Among them, triplicate microcosms were amended with labeled aniline (1µl, ring- C6 aniline, Cambridge Isotope Laboratories, Inc. Andover, U.S.) or unlabeled aniline (1µl, Sigma Aldrich, St. Louis, U.S.). Triplicate abiotic autoclaved controls amended with unlabeled aniline were also included. The abiotic controls were sterilized in an autoclave for 45 min for three consecutive days. The microcosms were incubated at room temperature (20°C) with reciprocal shaking.

Chemical analytical methods PCE and its daughter products vinyl chloride (VC) and ethene, and methane, were determined using an Agilent Technology 6890N (Agilent Technologies, Inc. Santa Clara, CA, USA) gas chromatograph (GC) with a flame ionization detector (FID). 250 μL headspace samples were injected into the GC-FID equipped with a GS-GasPro capillary column (30 m × 0.32 mm I.D.; Agilent Technologies, Inc. Santa Clara, CA, USA) with helium as the carrier gas. The injector and detector temperatures were 250°C. The constant flow of helium (carrier gas) was 1.0 mL min-1. The oven temperature was programmed to initially hold at 50°C for 2 min, increase to 180°C at 15°C/min, and then hold at 180°C for 2 min. Retention times of individual compounds were 12.08, 10.624, 9.809, 8.507, 5.393, 1.349, and 0.862 min for PCE, TCE, cis-DCE, trans- DCE, VC, ethene, methane, respectively. Microcosms were sampled for aniline, initially weekly and thereafter, periodically. For aniline analysis, one mL of well-mixed slurry was withdrawn from each serum bottle using a sterile 1 mL plastic syringe equipped with an 18-gauge needle that had been pre-flushed with sterile nitrogen. Samples were placed in 1.5 mL Eppendorf tubes and either frozen or extracted immediately. Extractions were performed using slight modification of previously described methods (Struijs and Rogers 1989) by adding 1 mL of acetonitrile, mixing, and centrifuging at 10,000 rpm for 3 min. The supernatant was removed using a plastic syringe, filtered through a 0.45 µm nylon filter, and placed in a glass sample vial sealed with a Teflon-lined butyl rubber

6 septum. Analysis was by an Agilent 1100 high performance liquid chromatography (HPLC) system equipped with a diode array detector operating at 244 nm for detection of aniline and PCA. Isocratic separations were made on a Luna 5µ C18 (2) 120 column (250*2 mm) (Phenomenex, Torrance, CA). The column was held at 40°C. A water: acetonitrile (ACN) mixture (45:55 volume:volume) was supplied at a flow rate of 0.33 mL min-1 as the mobile phase. Aniline eluted at 2.8 min and was identified by comparison of the retention time to a known standard. Aniline concentrations in samples were quantified using a five point calibration. The microcosms were amended with approximately 1500 µM aniline and the concentration of the standards ranged from 100 µM to 2000 µM.

Molecular methods Methods to target metabolically active bacteria have been developed using isotope incorporation into newly biosynthesized macromolecules. This is termed stable isotope probing (SIP). For example, active bacteria can be detected in environmental samples by following 13C substrates into 13C DNA, RNA or proteins and is termed DNA-SIP, RNA-SIP and protein-SIP, respectively. We used stable isotope probing (SIP) of DNA and sequencing of this labeled DNA as a major means to identify bacteria responsible for biodegradation at contaminated sites in NJ. The overall work flow for SIP is shown in Figure 3 and each methodological step is described in the following sections. Nucleic acid extraction. Slurry samples were obtained from microcosms using pre- assembled sterile plastic syringes with stainless-steel needles and pre-flushed with sterile anoxic nitrogen. DNA was extracted from 0.3 g original sediment or 1 mL microcosm slurry samples using the PowerSoil™ DNA Isolation Kit according to the manufacturer’s instructions (MoBio

Laboratories Inc., Carlsbad, Figure 3. Work flow for stable isotope probing to identify CA). DNA extracts were microbes responsible for contaminant degradation stored at −20°C prior to analysis. Stable Isotope Probing (SIP). SIP was performed using the principles first presented by Radajewski et al. (2000). Total genomic DNA was subjected to ultracentrifugation to separate the 12C-DNA from DNA labeled with 13C. PCR amplification was performed on the separated DNA and was analyzed by TRFLP analysis (Gallagher et al., 2005) to compare the 13C-labeled microbial communities with the 12C background. SIP for identification of dechlorinating bacteria was described previously (Kittelmann and Friedrich, 2008; Yamasaki et al., 2012). As discussed earlier, to label DNA in dehalorespiring bacteria, 13C-labeled acetate or lactate was amended to the microcosms exhibiting dechlorination.

7 Terminal restriction fragment length polymorphism (T-RFLP). PCR with universal bacterial primers, 5'-end 6-FAM labeled 27F (forward) and 1100R (reverse), was used to amplify 16S rRNA genes from genomic DNA (Knight et al., 1999). PCR products were digested with MnlI at 37 °C for 3 hours. Digested samples were precipitated and resuspended in formamide with a ROX standard and denatured at 95 °C (Gallagher et al., 2005). Samples were analyzed on an ABI 310 automated sequencer (Applied Biosystems Instruments, Foster City, CA), which generated the T-RFLP fingerprint for each community. A minimum peak height cut-off value of 50 fluorescence units was used for identifying T-RFs. Bacterial identification and phylogenetic analysis. 16S rRNA genes were amplified by PCR with 27F and 1100R primers (Lane, 1991). PCR products were cloned into pCR4-TOPO vector using the TOPO TA Cloning Kit (Invitrogen, Life Technologies, Grand Island, NY), according to the manufacturer’s instructions. Plasmids were transformed in One Shot TOP10 chemically competent Escherichia coli (Invitrogen, Life Technologies, Grand Island, NY). Transformed cells were plated on LB agar plates containing 75 mg/mL ampicillin with 40 µL X- gal per plate. White colonies were picked and grown overnight in LB- media containing 100 µg/mL ampicillin. Plasmid DNA was extracted from 1 mL of culture using the Purelink Quick Plasmid miniprep kit (Invitrogen, Life Technologies, Grand Island, NY). PCR amplification of the 16S rRNA gene insert was performed using 5'-end 6-FAM labeled 27F and 1100R primers for T-RFLP analysis, as described above, to identify unique clones. The complete insert from unique 16S rRNA gene sequences was amplified using M13F and M13R primers (Invitrogen, Life Technologies, Grand Island, NY). The cloned PCR products were sequenced by Genewiz Inc. (North Brunswick, NJ). Homology searches were performed using BLAST (http://www.ncbi.nlm.nih.gov/BLAST). The programs CLUSTAL X (1.64b) and GeneDoc (2.0.004) were used to align the sequences and determine their homologies to known bacterial 16S rRNA gene sequences. The Ribosomal Database Project II SeqMatch tool (http://rdp.cme.msu.edu) was used to characterize rRNA gene sequences obtained in this study and to identify closely related genes. All 16S rRNA sequences were aligned using the ClustalX software (Thompson et al., 1997).

Quality Assurance Sampling in the field followed guidance available from recent Department of Defense funded projects that examined QA/QC procedures for sampling and analysis in support of bioremediation (e.g., Ritalahti et al., 2010; Lebrón et al. 2008) and using NJ DEP guidance documents. Sample custody procedures. Samples were collected in the field and transported in anoxic, sealed glass vessels or core sleeves (soils, sediment, groundwater). Samples were labeled and a list of samples was kept on a Chain of Custody Form and in a computer file (Excel spreadsheet). Calibration procedures and preventive maintenance. All molecular biology applications included positive and negative controls to assure purity of reagents. Documentation, data reduction and reporting. Data collected through the course of the investigation was recorded in laboratory notebooks; collected real-time via computer interface; or recorded directly into computer software such as Microsoft Excel. Laboratory notebooks and raw data including replicate results will be stored indefinitely in the PI’s office in a secure cabinet for up to 20 years following the publication of manuscripts or reports resulting from this work. Data recorded in electronic files will be reviewed by the PI and uploaded to a secure

8 server site and made accessible to all project personnel. The Department of Environmental Sciences at Rutgers has in house IT support with full-time staff and extensive storage capacity on secure servers. The Department of Environmental Sciences at Rutgers University maintains a policy of long-term data storage (decades) with ample redundancy and back-up. [Note that this capacity is intrinsic because of the presence of a large and active atmospheric science group who has extensive long term data storage needs. This capacity is freely available to all faculty members and funded by F&A return.] Data Analysis. Following data collection, the values of triplicate measurements were averaged as appropriate and standard deviations were computed using Microsoft Excel. Tables comparing phylogenetic matches or phylogenetic trees were established to illustrate site data.

Results and Discussion Microbial Investigation of the Former Pagan-Martinez Site A work plan was developed in conjunction with Engineering and Land Planning Associates, Inc. (E&LP) for remediation of the Pagan-Martinez site. For the Pagan-Martinez site both microcosms for PCE dechlorination and community analysis of the aquifer was planned. Dechlorination results. No PCE dechlorination was detected in the aquifer microcosms after six months of monitoring. E&LP sent additional samples to a commercial laboratory (Microbial Insights) for a variety of molecular analyses and the results confirmed low numbers of known dechlorinating bacteria at the site. Additional work at the site was performed and a bioaugmentation pilot test was initiated. Although our project ended prior to this, we will maintain contact with E&LP to seek opportunities for additional collaboration. Microbial community analysis. To determine whether in situ dechlorinating bacteria exist and to understand how microorganisms in the aquifer respond to contamination, a depth-resolved microbial community analysis for the different portions of the site was commenced. Genomic DNA was extracted for six different points with various depths within the heavily contaminated plume (denoted P2) and for five different points with different depth at a pristine sampling point (denoted P1). Ground water samples from two different monitoring wells were also collected. Phylogenetic information from the background microbial community analysis will be performed. A remediation effort involving injection of Dehalococcoides bacteria into the contaminated plumes has commenced. We will maintain our connection with E&LP to leverage any future efforts with samples from this site.

Identification of Dechlorinating Communities in Hackensack River, NJ We used DNA-SIP to identify active bacteria during PCE dechlorination in contaminated sediments from the Hackensack River, NJ. The sediments of the Hackensack River are impacted by nearby sites contaminated by chlorinated benzenes, dioxins and PCBs, among other contaminants (ATSDR, 2005). We used PCE was used as a model compound to stimulate the dehalogenating microbial community to obtain more rapid results for activity. The process of complete dechlorination of PCE to ethene is currently known only to be carried out by Dehalococcoides (Löffler et al., 2013). These organisms exclusively use hydrogen as an electron donor and chlorinated compounds as electron acceptors. Many other dehalorespiring bacterial genera are known which can dechlorinate the higher chlorinated ethenes to intermediate daughter products (Holliger et al., 2003). The growth of Dehalococcoides is dependent on interaction with other microbial groups such as fermenters, acetogens and methanogens. The Dehalococcoides only use hydrogen as an

9 electron donor and short chain organic compounds such as acetate as a carbon sources and thus are dependent on these other community members to produce these substances through fermentation reactions of more complex organic matter (e.g., in this study, lactate) (Figure 4). Thus, by adding labeled acetate, which may taken up in an assimilatory (biosynthesis) process, dechlorinators may be identified by SIP (Kittelmann

and Friedrich, 2008; Figure 4. Interactions between chloroethene dehalorespiring Yamasaki et al., 2012). The bacteria and other microbes in the environment. dechlorinating groups in a real contaminated site may be more complicated than some established laboratory-scale dehalogenating enrichments, and SIP may be able to capture this complexity. To our knowledge this is the first time this technique has been used to characterize NJ sites. Microcosms seeded from the Hackensack River showed rapid dechlorination. PCE was degraded to cis-1,2-dichloroethene in less than one month and was subsequently dechlorinated to ethene during a five-month incubation (data not shown). We added either 13C-acetate or 13C-lactate to active Hackensack River dechlorinating microcosms. DNA was extracted five days after addition of labeled acetate or lactate and T-RFLP analysis was performed. 13C-TRFLP profiles were compared to the 12C-TRFLP profiles (Figure 5). Many TRFLP fragments were obviously enriched in the panel with red lines (13C samples). Notably, a T-RF matching Dehalococcoides was only directly observed in the

microcosms amended with lactate and not with those Figure 5. TRFLP profiles retrieved from Hackensack River, NJ amended with acetate. This PCE-dechlorinating microcosms. Black lines: 12C DNA. Red may be because addition of lines: 13C DNA. acetate would not result in direct

10 release of hydrogen, but lactate would (Figure 4). Thus, an important finding at this location is that both hydrogen and acetate were needed to stimulate Dehalococcoides in these sediments. In addition to direct observation of the T-RFs, clone libraries were set up from amplicons produced from PCR of the DNA from the 13C bands retrieved from SIP. The phylogenetic information obtained in clone libraries was used to link with TRFLP analysis and identify the phylotypes of dechlorinating bacteria. In addition to Dehalococcoides, a number of sulfate- reducing bacterial phylotypes, sulfur-oxidizing bacterial phylotypes and other dechlorinating bacteria were identified, indicating that in situ dechlorination is a complex microbial interaction between different bacterial groups. Sulfate-reducing bacteria, Desulfobacteraceae, Desulfobacterium, Geobacteraceae and Desulfuromonas were identified in the clone library derived from 13C-heavy fraction. Notably, a phylotype related to genus Dehalogenimonas was also enriched in 13C-labeled clone library. Dehalogenimonas is a relatively newly identified dehalorespiring bacterium and has not been widely identified in environmental samples (Moe et al. 2009). Therefore, we propose a diverse and complex dechlorinating food web for the Hackensack river sediments. One or several bacterial groups are directly responsible for dehalorespiration while other bacteria with different physiological traits were also involved. For example, sulfate-reducing bacteria are also known to switch to fermentation of organic compounds (including lactate) when sulfate is depleted. This may have promoted the dechlorination by providing acetate and hydrogen. [Note that as mentioned above, during the course of this NJDEP-funded project dechlorination was not observed in the microcosms from the Pagan-Martinez site. We will continue to monitor these microcosms to use for other experiments, if the opportunity arises.]

SIP Investigation of Aniline Degraders from a NJ Contaminated Site under Different Redox Conditions Aniline is a widely-detected groundwater pollutant at chemical industrial sites because of its use as a precursor compound in chemical manufacturing. Very little is known about the microorganisms responsible for aniline biodegradation under anoxic/anaerobic conditions. To date, only a single sulfate-reducing bacterium has been thoroughly characterized for its ability to degrade aniline (Schnell et al., 1989). This lack of information is perhaps a reflection of the delicate syntrophic chains involved in anaerobic aniline metabolism and the resulting long incubation times needed to establish suitable microbial populations with this functional ability. The in-situ metabolism of aniline-degrading bacteria is likely to be controlled and shaped by the availability of electron acceptors. Aniline biodegradation was investigated in sediment microcosms from a chemical manufacturing site. To investigate the diversity of bacteria involved, the microcosms were established under a variety of redox conditions. Results of microcosm studies. Aerobic and anaerobic aniline-degrading microcosms were established from a contaminated aquifer and sediments from an adjacent contaminated canal were set up in August, 2007. Aniline degradation experienced a more than 200 day lag period in the aquifer anaerobic methanogenic microcosms. Loss of aniline experienced a 50-day lag phase in the canal sediment methanogenic treatments. Aniline depletion was seen in both treatments but only canal sediments demonstrated substantial methane production during the pre-SIP incubation (these results were recently summarized in Li, Y. 2014). In 2012, after a five-year incubation, parent microcosms were distributed to daughter microcosms and SIP investigation was applied to these daughter microcosms (canal sediment aerobic microcosms (SO), canal sediment denitrifying microcosms (SN), canal sediment methanogenic microcosms (SM).

In addition, as part of this project additional triplicate microcosms were set up in 2012 under aerobic (AO), denitrifying (AN) and methanogenic (AM) conditions using the contaminated aquifer sediments and under methanogenic conditions in the canal sediments (SM). The microcosms were operated by re-amending aniline and exogenous electron acceptors (i.e., - NO3 ) as needed to assure activity. AO and AN demonstrated relatively rapid aniline degradation (Figure 6). Figure 7. TRFLP profiles from aerobic aniline- degrading microcosms (SO) from freshwater canal Complete biodegradation occurred over sediments. Black lines: 13C DNA. Blue lines: 12C DNA. approximately 19 days (AO) and 23 Panel 1: 13C ~30% biodegradation days (AN). In contrast to aerobic and Panel 2: 13C ~100% biodegradation denitrifying treatments, aniline was Panel 3: 12C ~100% biodegradation degraded more slowly under Panel 4: 13C ~50% biodegradation methanogenic conditions (AM and SM) Red arrows indicate phylotypes with 13C labeled (Figure 6). Aniline was depleted after DNA—active bacteria. more than four months in AM and SM.

Little methane was produced in AM while SM generated substantial methane during this SIP incubation (data not shown). Canal microcosms. Figure 7 shows TRFLP analysis of 16S rRNA genes amplified from DNA extracted from aerobic microcosms seeded from canal sediments. 13C DNA was extracted and analyzed when ~30%, ~50% and ~100% of the added aniline was degraded. 12C DNA was extracted and analyzed when aniline was completely depleted. Four TRFLP fragments (T-RF) 13 were enriched in the C samples, 137 bp, Figure 8. TRFLP profiles from aniline-degrading 164 bp, 190 bp, 207 bp and 234 bp. microcosms from freshwater canal sediments Notably, these T-RFs were not enriched established under denitrifying conditions (SN). in 12C-fractions, indicating successful Red lines: 12C DNA. Blue Lines: 13C DNA 13 carbon assimilation. Figure 8 shows the Panel 1: C ~30% biodegradation Panel 2: 13C ~50% biodegradation TRFLP profiles of 16S rRNA genes 13 amplified from DNA extracted from Panel 3: C ~100% biodegradation Red arrows indicate phylotypes with 13C labeled denitrifying microcosms seeded from DNA—active bacteria. canal sediments. DNA was also extracted

0.7 0.5 0.45

0.45 0.4 0.6 0.4 0.35 0.5 0.35 0.3 0.3 0.4 0.25 0.25 0.3 0.2 0.2 0.15 0.2 0.15 Aniline ConcentrationAniline (mM) Aniline ConcentrationAniline (mM) Aniline ConcentrationAniline (mM) 0.1 0.1 0.1 0.05 0.05

0 0 0 0 5 10 15 20 25 30 0 20 40 60 80 100 120 140 0 20 40 60 80 100 120 140 Time (Day) Time (Day) Time (Day)

A B C

Figure 6. Aniline concentration in microcosms amended with labeled aniline, abiotic controls in AO and AN (A), AM (B) and SM (C). Symbols: (A) aniline-degrading aerobic aquifer microcosms (AO); aniline-degrading aquifer microcosms under denitrifying conditions (AN);  abiotic microcosms. (B)  aniline-degrading aquifer microcosms under anaerobic conditions (AM); abiotic microcosms. (C)  aniline-degrading canal sediment microcosms under anaerobic conditions (SM); abiotic microcosms.

at three different time points representing ~30%, ~50% and ~100% loss of added aniline. Five T-RFs were enriched, indicating five bacterial phylotypes may be responsible for aniline biodegradation under nitrate-reducing conditions. The enriched T-RFs are indicated by red arrows (Figure 8). In SN, the Figure 9. TRFLP profiles from methanogenic aniline- 207 bp T-RF was also enriched degrading microcosms from canal sediments (SM). 13 Black lines: 12C DNA. Red Lines: 13C DNA in the C TRFLP profiles along 12 with the dominance of a 120 bp Panel SM-T1: C ~25% biodegradation 13 Panel SM-T1: 13C ~25% biodegradation T-RF in the C-fractions. Panel SM-T2: 12C ~100% biodegradation Another T-RF, 301 bp, was also Panel SM-T2: 13C ~100% biodegradation detected in the heavy fractions. Black arrows indicate phylotypes with 13C labeled DNA— In Figure 9, the TRFLP profiles active bacteria. for the methanogenic canal sediment (SM) microcosms are shown. T-RFs of 120, 208 and 274 were dominant and indicated similarity to SO and SN. Aquifer microcosms. Figure 10 and Figure 11 show TRFLP profiles of amplified 16S rRNA genes from DNA extracted from oxic (AO) and denitrifying (AN) microcosms from contaminated aquifer sediments, respectively. DNA was also extracted at three different time points representing different aniline degradation stages. In both cases, T-RFs 178 bp and 274 bp were enriched, indicating that two bacterial phylotypes may be responsible for aniline biodegradation in the aquifer. Figure 10. TRFLP profiles from aerobic aniline-degrading In contrast to oxic and microcosms from aquifer sediments (AO). denitrifying conditions, the Black lines: 12C DNA. Red Lines: 13C DNA methanogenic aniline-degrading Panel 1: 12C ~25% biodegradation microcosms (AM) from aquifer Panel 2: 13C ~25% biodegradation 12 sediments demonstrated a Panel 3: C ~65% biodegradation Panel 4: 13C ~65% biodegradation different scenario. As shown in 12 Figure 12, a 243-bp T-RF was Panel 5: C ~100% biodegradation Panel 6: 13C ~100% biodegradation enriched under those conditions, Red arrows indicate phylotypes with 13C labeled DNA— indicating a single bacterial active bacteria. phylotype associated with 243-bp

T-RF may be responsible for aniline biodegradation under methanogenic conditions. Identification of aniline- degrading bacteria. Clone libraries were set up from amplicons produced from PCR of the DNA from the 13C bands retrieved from SIP. The phylogenetic information obtained in clone libraries was to linked to the TRFLP analysis and used to identify the phylotypes of Figure 11. TRFLP profiles retrieved from aniline-degrading aniline-degrading bacteria. Figure microcosms from aquifer sediments under denitrifying 13 shows the phylogenetic tree of conditions (AN). aniline-degrading bacteria identified Black lines: 12C DNA. Red lines: 13C DNA. from SIP in microcosms seeded 12 Panel 1: C ~35% biodegradation from canal sediments. Figure 14 Panel 2: 13C ~35% biodegradation 12 shows the phylogenetic tree of Panel 3: C ~60% biodegradation aniline-degrading bacteria identified Panel 4: 13C ~60% biodegradation Panel 5: 13C ~100% biodegradation from SIP in microcosms seeded Panel 6: 12C ~100% biodegradation from aquifer sediments. Table 1 Red arrows indicate phylotypes with 13C labeled DNA shows the phylogenetic information of the bacterial phylotypes obtained in the current study. The 13C-fraction clone librar ies for the canal sediments compromised five phyla: Proteobacteria, Firmicutes, Actinobacteria, Verrucomicrobia and Chlorobia. The oxic and denitrifying canal sediment treatments showed some similarity in 13C-TRFLP profiles and clone libraries. T-RF 207 bp enriched in 13C-TRFLP profiles of SO and SN represented Figure 12. TRFLP profiles retrieved from aniline-degrading a phylotype belonging to the methanogenic microcosms from aquifer sediments (AM). Black lines: 12C DNA. Red lines: 13C DNA. genus Magnetospirillum within 12 the family of Rhodospirillaceae Panel 1: C ~25% biodegradation Panel 2: 13C ~25% biodegradation (Figure 13). The predominance 12 13 Panel 3: C ~55% biodegradation of the Magnetospirillum in C- Panel 4: 13C ~55% biodegradation TRFLP profiles and clone Panel 5: 12C ~100% biodegradation libraries suggested that they were Panel 6: 13C ~100% biodegradation directly correlated with carbon Red arrows indicate phylotypes with 13C labeled DNA— uptake from aniline and hence active bacteria. aniline degradation. In the oxic

15 treatment, T-RF 163 bp was related to the genus Pelomonas. A 234 bp T-RF was affiliated with the genus Ferribacterium. Acidovorax-related bacteria (274 bp) demonstrated predominance in 13C-fractions of all treatments, especially in the more reduced environments (denitrifying and methanogenic conditions) of microcosms under anaerobic conditions (SM); abiotic microcosms. In the methanogenic microcosms (SM) seeded from lightly contaminated canal sediments, substantial methane was produced. 150 clones were screened and eight OTUs were identified in the clone library but only three T-RFs, 120 bp, 208 bp and 274 bp, were highly enriched in the 13C-TRFLP profiles, indicating bacterial associated with these T-RFs might be correlated with label assimilation. In the SM microbial community, phyla Chloroflexi (33%) and β-proteobacteria (50%) predominanted in the clone library constructed from 13C-fractions. A cluster of uncultured relatives of the family Anaerolineaceae (genus unclassified) were represented by the 120 bp T-RF, which dominated the 13C-TRFLP profiles. T-RF 208 bp was represented by a phylotype belonging to the family of Burkholderiales. The closet pure isolates were Caldimonas sp. JK14, Ideonella sp. O-1 (96% sequence similarity) and Methylibium petroleiphilum PM1 (95% sequence similarity). Again, the Acidovorax-related sequences were numerous in the clone library (24% of clones) and were represented by a major T-RF, 274 bp. The 301 bp T-RF was of uncertain classification. In the gene clone library constructed from the 13C-fractions of AO (121 clones), 9 OTUs were identified at 97% similarity level. The bacterial community derived from AO was widespread among five different phyla and was dominated by β-proteobacteria (55%) and Actinobacteria (34%) phyla. Three T-RFs, 164 bp, 178 bp and 274 bp, became the dominant T- RFs in 13C fractions with relative abundances greater than 10%. Notably, 178 bp was only enriched in 13C-TRFLP profiles while the other two T-RFs were enriched in both 12C and 13C bands. The 178 bp T-RF was affiliated with genus Rhodococcus, showing 99% sequence similarity to Rhodococcus sp. Q5. Intriguingly, Rhodococcus spp. were found to be able to incorporate 13C carbon under denitrifying conditions, indicating these taxa could degrade aniline with nitrate as the electron acceptor. T-RF 274 bp, was identified as genus Acidovorax, and was less enriched in 13C-fractions than those of 164 bp and 178 bp. This phylotype was highly enriched in more reduced environments such as AN and AM. Sequences affiliated with the order of Burkholderiales were responsible for the 13C-enriched TRFLP fragment, 164 bp. Very similar 13C-TRFLP profiles in all sampling points in AO and AN were observed, indicating these phylotypes are likely metabolically versatile with electron acceptors and could likely use both oxygen and nitrate as electron acceptors. A pronounced shift in bacterial community structure was observed in AM compared to the AO, which was seeded from the same sediments. These results emphasize that the redox conditions exerted a strong selective pressure on the active functional bacterial species. The AM was characterized by a microbial community (136 clones) dominated by Chlorobi (42%), β- proteobacteria (39%) and Actinobacteria (7%) phyla. Firmicutes (7%) and δ-proteobacteria (6%) were also present in AM with a relatively small proportion. Two T-RFs, 239 bp (labeled 243 on Figure 12) and 274 bp, became dominant in 13C TRFLP profiles of AM with relative abundances greater than 20%. The 239 bp T-RF was solely enriched in 13C fractions, whereas 274 bp T-RF showed high abundance in both 13C and 12C fractions. The 239 bp T-RF fell within the cluster of genus Ignavibacterium, a novel genus proposed in 2010 (Iino et al., 2010) (Table 1, Figure 14). Genus Ignavibacterium branches deeply in the phylum of Chlorobi. Unlike other

16 autotrophic members of phylum Chlorobi, a recent discovery suggested that Ignavibacterium album, the only cultivable representative of genus Ignavibacterium, was a “strictly anaerobic, moderately thermophilic, neutrophilic and obligately heterotrophic bacterium” (Liu et al., 2012). Complete genome analysis suggested I. album could live under both oxic and anoxic conditions. Another closely related cultivated relative, Melioribacter roseus, were isolated from a microbial mat under the flow of hot water (46°C) coming out of a 2775-m-deep oil exploration well (Podosokorskaya et al., 2013). M. roseus was identified as a facultatively anaerobic and obligately organotrophic bacterium. The Ignavibacterium-related phylotypes identified in this study were most closely related to these two cultivable strains, with 92.7% similarity to I. album and 87.5% to M. roseus. Thus, two anaerobic aniline-degrading microbial communities seeded from adjacent areas (canal sediments and groundwater aquifer) demonstrated distinct primary aniline-degrading bacteria, inferring a wide diversity of in situ aniline-degrading bacteria. Notably, sampling sites for AM experienced higher aniline concentrations while sampling sites for SM were located upstream of the highly-contaminated area. In addition, sediments for AM were taken from 0-10 inches below groundwater surface, which might be anoxic conditions, but sediments for SM were obtained from the bottom of the canal, which could include aerobic or micro-aerobic environments. It is apparent that contamination history and redox conditions may shape different

Figure 13. Phylogenetic tree of 16S rRNA gene sequences of major aniline-degrading bacteria (denoted by terminal restriction fragment (T-RF)) amplified from 13C-labeled DNA from the canal sediments and the closest matches within GenBank, as well as the closest type strain, constructed with MEGA 5.0 software using the neighbor-joining method. * Indicates aniline-degrading isolates identified in previous studies. indigenous active microbial populations. Comamonadaceae are involved in aniline degradation. Three phylotypes enriched in 13C fractions, represented by T-RFs 164 bp (enriched in AO), 208 bp (enriched in SM) and 274 bp (enriched in AO, AN, AM and SM), belong to the family of Comamonadaceae, suggesting these bacteria may play an important role in aniline degradation. Members of Comamonadaceae have been frequently associated with biodegradation of anthropogenic substances, especially hydrocarbon degradation. For example, Comamonadaceae were identified as the most active microorganisms in many environments contaminated by polycyclic aromatic hydrocarbons (PAHs) (e.g., Herbst et al., 2013). In addition, members of Comamonadaceae were correlated

17 with toluene and m-xylene degradation under denitrifying and aerobic conditions (Xie et al., 2010 and Sun et al., 2012). T-RF 274 bp 99 Uncultured bacterium clone Bac5 (HQ603933.1) Uncultured Acidovorax sp. clone 5_0_B1_b (JQ087024.1) 86 Uncultured beta proteobacterium clone GC0AA4ZC04PP1 (JQ919609.1) Acidovorax sp. W20 (KF560400.1) 92 75 Iron-reducing bacterium enrichment culture clone FeC_1_H5 (FJ802381.1) 65 Flectobacillus lacus strain R73022 (HM032866.1) 97 Acidovorax defluvii strain BSB411 (NR 026506.1) Diaphorobacter sp. J5-51 (GU017974.1) 96 Delftia sp. AN3 (AY052781.1)* 100 Aniline-degrading bacterium HY99 (AF210313.1)* Beta proteobacterium JPPB B32 (GU368378.1) 51 100 Rhodoferax sp. Asd M2A1 (FM955857.1) Mitsuaria sp. H24L1C (EU714910.1) 39 Aquabacterium sp. A7-R (KF061035.1) 41 83 Burkholderiales bacterium YT0099 (AB362826.1) Caldimonas sp. JK14 (KF206367.1) 97 79 T-RF 164 bp 99 Beta proteobacterium HTCC333 (AY429712.1) 100 Uncultured bacterium clone WPUB078 (FJ006757.1) 32 Aquincola tertiaricarbonis strain C93 (JF304446.1) 61 Aquatic bacterium R1-B35 (AB195767.1) 64 75 Aquincola sp. HME6814 (JF308947.1) T-RF 208 bp 100 Uncultured bacterium clone AN162 (GQ859926.1) 78 Uncultured proteobacterium clone BWET3cm1 (JQ793295.1) 73 Uncultured bacterium clone LaP15L15 (EF667620.1) Erwinia amylovora strain HSA 6 (GQ222272.1)* 99 Anaerolinea thermophila strain UNI-1 (NR_074383.1) Anaerolinea thermolimosa strain IMO-1 (NR_040970.1) 97 Uncultured Chloroflexi bacterium clone Aug-VN107 (JQ795368.1) 100 Uncultured bacterium clone d0-9 (AM409808.1) 100 Uncultured bacterium clone NY-33 (KC290433.1) 100 Uncultured Chloroflexi bacterium clone 4.26 (GQ183434.1) 72 T-RF 120 bp 65 Uncultured Chlorobi clone Aug-CD246 (JQ795235.1) 72 Uncultured bacterium clone UA_55 (JX120425.1) 99 Uncultured bacterium clone BSTSG-72 (JN104894.1) 100 T-RF 239 bp 100 Ignavibacterium album strain JCM 16511 (NR_074698.1) 100 Ignavibacterium album gene for 16S rRNA (AB478415.1) Melioribacter roseus P3M-2 (NR_074796.1) 82 Rhizobium borbori strain DN316 (EF125187.1) * T-RF 178 bp 78 Rhodococcus qingshengii strain IARI-JR-58 (KF055006.1) Nocardiaceae bacterium Ben-13 isolate KFC-53 (EF459533.1) 100 Rhodococcus sp. K5 (KF790905.1) Nocardiaceae bacterium Ben-13 (EF028121.1) 5 Rhodococcus sp. Q5 (JN644309.1) 15 Nocardiaceae bacterium KVD-1700-17 (DQ490437.1)

0.02

18 Table 1. Selected bacterial phylotypes identified in clone libraries generated from 13C-fractions from microcosms fed aniline under different redox conditions.

No. of clones retrieved

Aquifer sediments, Aquifer sediments, Canal sediments, Mnl I digested Phylogenetic group Oxic (AO) Methanogenic (AM) Methanogenic (SM) clone fragment size (bp) Acidobacteria Gp3 3 10 132 (AO), 170(SM) Actinobacteria Rhodococcus 41 9 12 179 Alphaproteobacteria Xanthobacteraceae 8 63 Betaproteobacteria Burkholderiales 26 14 168 Acidovorax 25 39 37 277 Neisseriaceae 3 174 Nitrosospira 6 168 Comamonadaceae 7 282 Rhodocyclaceae 11 146 Alcaligenaceae 30 211 Gammaproteobacteria Arenimonas 2 101 Deltaproteobacteria Syntrophaceae 8 259 Chlorobi Ignavibacterium 57 243,285 Chloroflexi Anaerolineaceae 51 54,120 Firmicutes Peptococcaceae 6 130 Fusibacter 3 388 OP8 2 N/A Unclassified Bacterium 2 N/A Total 121 136 155

The 164 bp T-RF was only enriched in labeled heavy fractions of aerobic and denitrifying treatment. The phylotype corresponding to 164 bp T-RF was closely related to Methylibium petroleiphilum PM1 (96% sequence identity) and Aquincola tertiaricarbonis (95% sequence identity). Strain PM1 is the first and currently only known species of bacteria capable of using MTBE as a sole source of carbon (Schmidt et al., 2008). This phylotype was also closely related to Aquincola tertiaricarbonis, a tertiary butyl moiety-degrading bacterium, originally obtained from MTBE groundwater and from a wastewater treatment plant (Lechner et al., 2007). In the study preceding this one, we attempted to characterize the intermediates during aniline degradation (Li, 2014). However, due to the complex nature of the microcosms, we failed to identify biodegradation intermediates. It is difficult to determine whether this phylotype primarily degraded aniline or biodegradation intermediates. Based on the enrichment in 13C fractions (>20%), we propose phylotypes related to 164 bp T-RF might have directly incorporated labeled carbon atoms from 13C-aniline. Acidovorax-related bacteria (274 bp) demonstrated predominance in 13C-fractions of all treatments, especially in the more reduced environments (denitrifying and methanogenic conditions). Acidovorax-related bacteria were closely related to Acidovorax sp. W20 (98% sequence identity) that was isolated from hydrocarbon-contaminated soil (unpublished study). Members of Acidovorax were also responsible for chlorobenzene (Monferrán et al., 2005), polycyclic aromatic hydrocarbon, nitrobenzene (Singleton et al., 2009) and mono-nitrophenol degradation (Zhao et al., 1999). Some strains of Acidovorax, such as Acidovorax sp. strain NO1, Acidovorax delafieldii and Acidovorax facilis were shown to be facultative bacteria (Huang et al., 2012; Willems et al., 1990). Because AM and SM were constructed under strict anaerobic conditions, it is hypothesized that the Acidovorax detected might be facultative aniline degraders. Bacteria related to 208 bp T-RF were closely related to Ideonella sp. 0-0013 (96% sequence identity) and Ideonella sp. O-1(96% sequence identity). Ideonella sp. 0-0013 was isolated from the soil (Tanaka et al., 2011). I. dechloratans grew anaerobically with chlorate as electron acceptor (Malmqvist et al., 1994), coinciding with the observations that phylotypes affiliated with 208 bp fragments were enriched in methanogenic microcosms. These observations suggest Ideonella species may be metabolically versatile toward electron acceptors. Phylotypes closely related to Aquincola, Ideonella and Acidovorax demonstrated high abundance in both 13C and 12C fractions. The enrichment in light fractions suggested not all of these bacteria use the 13C carbon and some of them consumed unlabeled carbon sources, probably indigenous organic matter or other contaminants. The enrichment in both resident and active fractions suggested that phylotypes were metabolically versatile and showed broad degradation capability toward labeled aniline and other organic compounds present in the microcosms. The prolonged incubation periods and the oxygen-limited environments likely helped these phylotypes outcompete other bacteria for competition of carbon sources. In the light of earlier and our own findings, we hypothesize that these Comamonadaceae-related populations might play an important role in aniline degradation.

Conclusions and Future Work We have established in-house protocols and a work flow for environmental samples from contaminated sites. We have successfully characterized active dechlorinating bacteria from contaminated Hackensack River sediments and active aniline degraders from a NJ industrial contaminated site by using SIP. Unfortunately, the lack of dechlorinating activity in microcosms at the Pagan-Martinez site and the nature of the timeline for further site work did not allow

additional followup at this time. However, we will maintain contact with E&LP to leverage any future resources for performing additional research with this remedial effort. One of the major findings to emerge from this work is that EMDs such as SIP can not only identify microbes responsible for biodegradation and biotransformation, but it can also help elucidate important pathways. For example, in the Hackensack River sediments, we observed that adding lactate [presumably fermented to hydrogen (electron donor) and acetate (carbon source)] stimulated Dehalococcoides, but addition of acetate alone did not. While this is not entirely surprising, it does show that SIP could also be used as a rapid means to assess success of different amendment strategies, in addition to identifying active microbes to be tracked for remedial efforts. In the case of the aniline-contaminated site, SIP conclusively proved that aniline was biodegraded and incorporated into cellular biomass. Further, there appear to be many bacteria involved in anaerobic degradation at the site and candidate bacteria were identified. In particular, the identification of Ignavibacterium as a likely aniline degrader under anaerobic conditions is novel. These are critical pieces of information supporting the effectiveness of bioremediation that had been difficult to prove by tracking the metabolites or matching electron acceptor consumption. Compounds such as aniline may undergo sorption or polymerization, which are difficult to quantify. SIP could be used to confirm biodegradation is occurring rather than physical-chemical sequestration.

Recommendations and Application and Use by NJDEP The EMDs used in this project allowed us to:

• Identify active bacteria involved in the biodegradation or biotransformation of contaminants at a variety of New Jersey sites • Identify which amendments would be most effective for stimulating this activity • Confirming that contaminants are undergoing biodegradation

Monitored natural attenuation and enhanced bioremediation are increasingly implemented as a treatment options for contaminated aquifers and sediments. The key challenge, however, is to accurately assess the effectiveness of the remediation techniques in situ. Bioindicators allow better understanding of important microbial processes. Once the responsible bacteria are known, testing of amendments to improve the degradation conditions will lead to methods for optimizing biodegradation in situ. Monitoring is important for gaining an understanding of different microbial processes and how these processes, and thus the remediation, is affected by different amendments and other engineering approaches. In the case of dechlorination of PCE we observed that SIP could distinguish between populations stimulated by addition of lactate (provides acetate and hydrogen upon fermentation) versus acetate alone. This is a very interesting observation because it highlights the potentially different outcomes depending on what electron donors are added during biostimulation. In this study, DNA-SIP was applied to a number of aniline-degrading microcosms under different redox conditions. This is the first time that we realized the diversity and complexity of aniline degraders and scavengers in this complex microbial system as elucidated by DNA-SIP. A significant finding of this study was that all taxa groups associated with 13C-enriched T-RFs in anaerobic treatments were different from any known aniline degraders, especially the Ignavibacterium-affiliated bacteria that have not previously been correlated with biodegradation and bioremediation. Future experiments are needed to explore the significance and distribution

of these strains that might be used as biomarkers in monitoring natural attenuation of aniline, halogenated anilines and aromatic amines contamination. Further, after several previous years attempting to prove biodegradation of aniline in microcosms of a contaminated site, we showed that by using SIP, it could be ascertained rapidly not only that biodegradation was occurring, but also which organisms are involved.

Expenditures Our budget expenditures up until the beginning of July 2013 are shown in Table 2.

Table 2. Budget expenditures and commitments through December 31, 2013.

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Project Publications and Presentations

Sun, W.; McGuiness, L.; Kerkhof, L.; Häggblom, H.; Fennell, D.E. 2013 Applying innovative environmental molecular diagnostics (EMDs) for informing bioremediation at contaminated sites. poster presented at the 113th General Meeting of the American Society for Microbiology. May 18 – 21, 2013. Denver CO.

Sun, W. 2014. Investigation of bioremediation potential in New Jersey contaminated sites. A talk presented at the Theobald Smith Society, American Society for Microbiology, Meeting in Miniature. Rutgers University, New Brunswick, NJ, April 6, 2014.

Raes, E.; Fennell, D.E.; Sun, W.; Mueller, R.; Biernacki, A.; Ogles, D.; Baldwin, B.R.; and Sublette, K. 2014. Using a variety of microbial analyses in different media to quantify remedial effectiveness, from qPCR, to qPCR/RT-PCR to QuantArray. Abstract submitted to the Ninth International Conference on Remediation of Chlorinated and Recalcitrant Compounds, May 19- 22, 2014, Monterey, California.

Sun, W.; Luo S.; Li Y.; McGuiness, L.; Kerkhof, L.; Häggblom, H.; Fennell, D.E. Identification of in situ aerobic and anaerobic aniline-assimilating microbial populations in aniline contaminated sites by DNA-stable isotope probing. (in preparation).

Sun, W.; Luo S.; McGuiness, L.; Kerkhof, L.; Häggblom, H.; Fennell, D.E. Identification of in situ aniline-degrading bacteria from contaminated canal sediments by stable isotope probing. (in preparation).