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Natural attenuation of non-volatile contaminants in the capillary fringe Zohre Kurt, Elizabeth Erin Mack, and Jim C Spain Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b02525 • Publication Date (Web): 12 Aug 2016 Downloaded from http://pubs.acs.org on August 16, 2016

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Environmental Science & Technology

Natural attenuation of non-volatile contaminants in the capillary fringe

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ZOHRE KURT1,2*, E. ERIN MACK3 AND JIM C. SPAIN1,4

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1

6

Atlanta Georgia 30332-0512

7

2

8

Panama

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3

School of Civil and Environmental Engineering, Georgia Institute of Technology,

Institute of Scientific Research and High Technology Services, Calle Pullpn, Panamá,

DuPont, Corporate Remediation Group, P.O. Box 6101, Glasgow 300, Newark, DE

10

19714-6101

11

4

12

Pensacola Florida 32514-5751

13 14 15 16 17

*

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Abstract

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Center for Environmental Diagnostics and Bioremediation, University of West Florida,

Corresponding Author Phone: +(507) 317-1745 Fax: +(507) 507-0020 E-mail: [email protected]

When anoxic polluted groundwater encounters the overlying vadose zone an

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oxic/anoxic interface is created- often near the capillary fringe. Biodegradation of volatile

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contaminants in the capillary fringe can prevent vapor migration. In contrast, the

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biodegradation of non-volatile contaminants in the vadose zone has received

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comparatively little attention. Non-volatile compounds do not cause vapor intrusion, but

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they still move with the groundwater and are major contaminants. Aniline (AN) and

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diphenylamine (DPA) are examples of toxic non-volatile contaminants found often at dye

ACS Paragon Plus Environment

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and munitions manufacturing sites. In this study, we tested the hypothesis that bacteria

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can aerobically biodegrade AN and DPA in the capillary fringe and decrease the

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contaminant concentrations in the anoxic plume beneath the vadose zone. Laboratory

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multiport columns that represented the unsaturated zone were used to evaluate

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degradation of AN or DPA in contaminated water. The biodegradation fluxes of the

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contaminants were estimated to be 113 ± 26 mg AN • m-2 • hr-1 and 76 ± 18 mg DPA • m-

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2

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contaminant profiles along with enumeration of bacterial populations indicated that most

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of the biodegradation took place within the lower part of the capillary fringe. The results

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indicate that bacteria capable of contaminant biodegradation in the capillary fringe can

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create a sink for non-volatile contaminants.

• hr-1 in the presence of bacteria known to degrade AN and DPA. Oxygen and

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Introduction

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A considerable portion of microbial biodegradation at subsurface contaminated

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sites occurs at plume fringes where electron donors and electron acceptors intersect 1-4.

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For example, oxic/ anoxic interfaces are often found at the leading edges of plumes. The

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vadose zone often includes an oxic/anoxic interface in or near the capillary fringe when

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anoxic contaminated plumes interact with the overlying air 5. In principle, capillary

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fringes are similar to sediment/water interfaces6 and plume fringes7, 8 because microbial

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activity creates steep redox gradients, resulting in high rates of contaminant

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biodegradation.

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The fate of volatile contaminants in the vadose zone has been of great interest

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because they pose a risk of vapor intrusion9. As a result, biodegradation of petroleum


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hydrocarbons in the vadose zone has been studied extensively. Field measurements 10, 11

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and laboratory studies revealed rapid biodegradation of hydrocarbons in the presence of

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oxygen and the processes have been described by several models 5, 12-14. Such models

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along with the field data related to dissolved hydrocarbons clearly reveals that

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biodegradation in the lower part of the unsaturated zone near the capillary fringe is

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sufficient to prevent vapor migration and intrusion 11. Furthermore, studies performed

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with laboratory columns15, 16 and field measurements 17, 18 recently established the

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potential for aerobic biodegradation of chlorinated compounds in the vadose zone.

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Although considerable effort has been devoted to understanding the

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biodegradation of volatile compounds in the vadose zone, biodegradation of non-volatile

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compounds has received little attention. Some non-volatile compounds are known to

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biodegrade in unsaturated soil 19. Pesticide biodegradation in the vadose zone was

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demonstrated by microcosm studies 20, 21, field measurements 20, 22 and by isotope

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fractionation 20. Non-volatile and semi-volatile petroleum hydrocarbons 23 and some

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explosives 24, 25 have also been studied in laboratory systems designed to simulate the

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vadose zone. These contaminants were typically introduced from the surface. In contrast,

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biodegradation of dissolved phase non-volatile contaminants in the vadose zone is poorly

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understood.

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Example of non-volatile toxic contaminants found in soil and groundwater are

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aniline (AN) and diphenylamine (DPA) 26, 27, which are precursors of industrial chemicals

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such as dyes, pharmaceuticals and explosives 28-30. . The dimensionless Henry’s Law

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constants for AN and DPA at 20 0C are 5.82x10-5 and 1.83x10-4, but both are ionized in

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water at circumneutral pH. Both AN and DPA are readily biodegradable and used as


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carbon sources under aerobic conditions. Furthermore, aniline is biodegradable under

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anoxic conditions29, 31-34. The aerobic biodegradation pathway of DPA was reported to

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involve AN as an intermediate 29, but its biodegradation in soil was not established. The

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goal of this study was to determine the extent to which non-volatile compounds in AN or

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DPA contaminated groundwater are transported to and aerobically biodegraded in the

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capillary fringe of laboratory columns. The columns were inoculated with Burkholderia

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sp. JS667 and Ralstonia sp. JS668 29 to catalyze aerobic degradation of either AN or DPA

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in the capillary fringe. The results revealed rapid biodegradation in the lower portion of

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the capillary fringe and enhanced diffusive transport of the contaminants to the active

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zone.

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Materials and Methods

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Chemicals. AN (> 99.5 %) and DPA (> 99 %) were from Sigma-Aldrich. All other

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chemicals were reagent grade.

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Analytical methods. AN and DPA were analyzed in aqueous samples by high

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performance liquid chromatography (HPLC) by the method of Shin and Spain 29.

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Absorbance was monitored at 265 nm and detection limits were below 0.5 mg/L. The AN

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and DPA concentrations in the unsaturated region above the capillary fringe were

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analyzed with 45 µm Supelco solid phase microextraction (SPME) C18 fiber probes in 22

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Ga needles (catalog number 57281-U). The probes were inserted through the ports into

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the sand and held for 10 minutes. Desorption was performed for 20 minutes in a 50:50

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mixture of acetonitrile-water. Probes were calibrated with known amounts of AN and


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DPA in duplicate samples of sand with 5% w/w moisture content. AN or DPA used for

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calibration was prepared by adding to dry sand appropriate concentrations (0.2- 50 mg/L.

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Variability among duplicate samples in standards was less than 1%. Gas phase samples

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(100 µL) were analyzed for oxygen with an Agilent Technologies 6850 Network GC

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equipped with a thermal conductivity detector (TCD) as described previously 35. All the

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gas samples were analyzed in duplicate except the first port above the saturated layer of

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the multiport columns where there was only enough volume for single samples.

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Growth conditions. Burkholderia sp. strain JS667 and Ralstonia sp. strain JS668 that

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degrade both AN and DPA were grown as described by Shin and Spain 29 with either AN

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or DPA as the sole carbon source. When the cells were in the exponential growth phase

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they were centrifuged, washed and suspended in minimal salt medium (MSB) 36 to an

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OD600 of 0.2. Equal volumes of the suspensions were combined to obtain the mixture

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used to inoculate the columns and microcosms. One-quarter strength trypticase soy agar

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(1/4-TSA) was used to check the purity of the cultures based on colony morphology.

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Microcosms. Microcosms designed to determine whether bacteria could degrade DPA in

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unsaturated soil were constructed in 35 mL serum bottles sealed with Teflon lined

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stoppers. Soil used for the study was from the subsurface of a field site heavily

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contaminated for decades mainly with DPA. The bottles contained 5 g of homogenized

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DPA contaminated soil or contaminated soil that was mixed with potting soil. All the soil

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mixtures were dried for 24 hours at room temperature, inoculated with 1 mL of the

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Burkholderia sp. strain JS667 and Ralstonia sp. strain JS668 mixture described above to

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a final moisture content of 20% w/w and sieved (0.42 mm). Control microcosms included


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20 mg of sodium azide/g of soil. The microcosms were incubated at room temperature

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with no stirring. At appropriate intervals microcosms were sacrificed and the contents

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extracted with acetonitrile-water (50:50). The initial biodegradation rates were estimated

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from the slopes of the concentrations between the first and second data points in Figs.

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1A, B and C.

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The microcosms prepared to evaluate the effect of moisture content on

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biodegradation rates and microbial growth were prepared using 15 mL scintillation vials

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sealed with Teflon lined caps. Each vial contained 10 g of sand (0.6 to 0.42 mm sieve

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size) and 0.2 mg of DPA. Moisture was adjusted with ¼ strength MSB. Inoculated vials

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contained 0.5 mL of the inoculum described above. Vials were vortexed for 30 sec and

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then centrifuged at 500 rpm for 10 min to pack the sand uniformly. Individual vials were

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sacrificed to measure DPA and protein concentrations at appropriate intervals. DPA was

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extracted with 3 mL of acetonitrile after addition of sufficient ¼ MSB to bring the

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moisture content to saturation (30 % w/w). Suspensions were mixed by vortexing for 30

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sec and incubated for 10 hours to allow proper desorption at 40 C prior to HPLC analysis

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of the liquid phase for DPA. Biodegradation rates were calculated from the initial slopes

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of the DPA depletion curves. Bacterial biomass was estimated by adding ¼ strength MSB

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to a final volume of 5 mL of liquid per vial, mixing for 1 min and using 1 mL of

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supernatant for protein measurements. Protein was quantified with a BCA assay reagent

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kit (Rockford, IL, USA) after suspending the cells in 0.1 N sodium hydroxide and

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incubation at 95 0C for 10 min. Controls included vials with no inoculum or with no

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DPA.

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Multiport columns. Multiport columns (30 cm × 2.2 cm) 15 described previously were

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packed with sand that was baked overnight at 550 0C (0.6 to 0.42 mm sieve size) (Fig.

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S1). Deionized water containing either AN or DPA prepared under nitrogen was pumped

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in the bottom of the column and exited 1 cm above the feed port to simulate a water table.

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The aqueous feed and air flows in the column were maintained at 1 mL • h-1 to ensure

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that oxygen concentrations were sufficient. Autoclaved sand was used for packing all

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columns but no attempt was made to maintain sterility after columns were set up.

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Uninoculated control (abiotic) columns were first operated until steady state for AN and

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DPA concentrations were established (8 days of constant concentration profile of DPA

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and AN) and then inoculated with 25 mL of Burkholderia sp. JS667 and Ralstonia sp.

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JS668 mixture. Effluent ports were closed and the column was filled with the bacterial

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suspension from the bottom port of the columns to create the active columns. The liquid

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used for inoculation was drained immediately from the lowest ports of the column by

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gravity. The capillary fringe in the columns was estimated to be 12 to 17 cm thick based

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on the mesh size of the sand 12 and 13 cm as estimated visually. At the end of the study

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the moisture content in each section of the column was measured by scooping the sand

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out of the column and then weighing samples before and after drying at 105 0C for 12

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hours.

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Quantitative polymerase chain reaction (qPCR). A PowerSoil DNA Isolation Kit (MO

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BIO Laboratories, Carlsbad, CA) was used to extract DNA from 2 cm segments of the

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column packing. An ABI 7500 Fast Real-Time PCR System equipped with SDS v. 2.0.3

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software with default SYBR Green cycling parameters was used for qPCR with a


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previously established protocol 37. Total bacterial 16S rRNA, aniline dioxygenase (tdnA1)

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and diphenylamine dioxygenese (dpaAa) genes were amplified using BacF/BacR 38,

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ANDO-F/ANDO-R 29 and DPADO-F/DPADO-R 29 primers respectively. Calibrations

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were performed in triplicate with serial 10-fold dilutions (from approximately 3 × 108

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copies down to approximately 3 copies per reaction 39) of the 16S rRNA, aniline

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dioxygenase and diphenylamine dioxygenase genes of Burkholderia sp. JS667 cloned in

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the pGEM-T Easy Vector System. Standard curves had a linear slope of 3.1–3.4 and an

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R2 value higher than 0.90 where 33 copies per reaction was the minimum detection limit.

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It was assumed that there was a single AN and DPA oxygenase gene in each organism

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and that there were 3 and 6 copies of 16S rRNA gene per cell of Ralstonia and

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Burkholderia respectively 40.

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Results and Discussion

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Microcosms with contaminated soil

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Even though field samples were mainly contaminated with DPA there were also

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minor amounts of other uncharacterized contaminants present in the samples. There was

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no detectable biodegradation of DPA in the microcosms prepared with undiluted field

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samples (10,000 mg DPA/kg) during the course of the experiment as indicated by the

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similarity between concentrations in active microcosms and sterile controls (Fig. 1A).

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The results suggested that the field samples contained toxic concentrations of DPA or

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other contaminants. In contrast, DPA was biodegraded rapidly in inoculated microcosms

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containing diluted soil at DPA concentrations up to 2000 mg /kg dry soil (Fig. 1B-C).

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The biodegradation rates were 7.2 ± 1.4 µg • h -1 and 0.8 ± 0.2 µg • h -1 per g of soil for


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10x and 100x diluted soil, respectively. The results indicate that bacteria in the

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unsaturated soil could biodegrade DPA at loading rates far above the water solubility (50

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mg/L). It is also clear that the DPA degrading bacteria can degrade DPA under the

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unsaturated conditions expected to be encountered in the capillary fringe of the vadose

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zone.

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Figure 1. Degradation of DPA in microcosms containing inoculated soil (blue) and

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uninoculated soil with sodium azide as a control (red). Field sample (A), field sample


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diluted 1:10 with potting soil (B) and field sample diluted 1:100 with potting soil (C)

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(Data represent means of duplicate microcosms).

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Biodegradation in the vadose zone columns

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Sand columns designed to represent the capillary fringe in the vadose zone above

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contaminated subsurface plumes were used to evaluate the biodegradation of AN or

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DPA. The uninoculated columns reached steady state within 66 hours (Fig. 2A and Fig.

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3A). The aqueous feed of the columns contained 59 ± 6 mg of AN/L or 37 ± 4 mg of

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DPA/L, and the effluent of the columns was 56 ± 7 mg of AN/L or 36 ± 5 mg of DPA/L

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which indicated negligible flux to the atmosphere or losses due to abiotic transformation.

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Contaminants were not detected above the capillary fringe (16 cm above the saturated

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layer) which is consistent with their low vapor pressure. The inoculated columns reached

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steady state after 58 hours (AN) and 62 hours (DPA) and were operated subsequently for

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201 hours (Fig. 2B and Fig. 3B) with a feed of 60 ± 5 mg of AN/L or 38 ± 4 mg of

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DPA/L. Effluent concentrations of 17 ± 3 mg of AN/L or 7 ± 2 mg of DPA/L indicated

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substantial flux due to biodegradation. DPA and AN were not detected 6 cm above the

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saturated layer. When air pumped into the top of the columns was switched to nitrogen

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biodegradation stopped within 36 hours, and the system remained stable for a total of 92

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hours (Fig. 2C and Fig. 3C). The activity in both columns recovered within 46 (AN) and

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52 (DPA) hours after air was reintroduced.

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Figure 2. AN concentration profiles in the multiport column. Uninoculated control

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column after 100 hours (A), active column 201 hours after inoculation with bacteria (B),

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and anoxic column 92 hours after the air feed was replaced with nitrogen (C).

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Figure 3. DPA concentration profiles in the multiport column. Uninoculated control

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column after 100 hours (A), active column 201 hours after inoculation with bacteria (B),


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and anoxic column 92 hours after air feed was replaced with nitrogen (C).

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Columns were designed so that the difference between the influent and effluent

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contaminant concentrations could be used to determine the mass balance and

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biodegradation flux in the capillary fringe. The non-volatile contaminants move

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throughout the columns only via diffusion in the water; therefore, there was no net flux of

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the contaminants to the unsaturated zones of the abiotic columns from the saturated zones

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after steady state was established. In the inoculated columns the difference between

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influent and effluent concentrations indicated the extent to which biodegradation of AN

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and DPA causes mass removal from the contaminated water. The fluxes were calculated

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as described in Kurt et al. 15, where the horizontal plane is the intersection between the

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top of the saturated zone and the base of the capillary fringe. The measured

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biodegradation fluxes were 113 ± 26 mg AN • m-2 • hr-1 and 76 ± 18 mg DPA • m-2 • hr-1.

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The conditions in the laboratory columns were favorable for biodegradation and the

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short distance from the capillary fringe to the oxygen source would be unusual in the

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field. In subsurface situations where there are no permeability issues oxygen will not be

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limiting because its diffusion in air would be faster than diffusion of non-volatile

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chemicals through the saturated zone. To estimate whether the biodegradation fluxes in

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the capillary fringe would remove appreciable amounts of contaminant mass from the

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underlying plumes, the maximum diffusion fluxes of the contaminants were determined

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based on the solubility of the contaminants in water at room temperature (36,000 mg/L

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for AN and 40 mg/L for DPA). Fick’s law of diffusion in Eq. (1) was applied assuming

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that biodegradation created a sink for the contaminants (3 cm above the saturated zone)


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where the effective diffusion coefficient was calculated based on Millington and Quirk 41

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by assuming that the sand is saturated with water.

250 251

𝐽 = −𝐷

!×!! !

(1)

!

252 253

Where J is the flux, D is the diffusion coefficient of the compounds in the water (8.3 x 10-

254

6

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the distance above the saturated layer (3 cm) and n is the water-filled porosity measured

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for the sand (0.42). The calculated diffusion fluxes were 19 ± 2 mg AN • m-2 • hr-1 and 16

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± 4 µg DPA • m-2 • hr-1. Comparing these diffusion fluxes to the biodegradation fluxes

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above clearly indicated that biodegradation in the columns was faster than the diffusion

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of AN and DPA therefore the process is diffusion limited. Biodegradation in the columns

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increases the diffusion flux of contaminants by creating a sink in the capillary fringe.

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Because the columns were filled with sand, they are likely to have more void volume

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than the field materials, and therefore the fluxes would likely be lower in the field. The

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high biodegradation rates at the oxic/ anoxic interface are consistent with numerous

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previous reports related to biodegradation of volatile compounds in the vadose zone 5, 14-

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16, 18

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Microbial distribution in the columns

cm2/s for AN and 6.3 x 10-6 cm2/s for DPA42), C is the contaminant concentration, Y is

.

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Microbial biomass in the vadose zone increases with increasing biodegradation

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fluxes 43, 44, and it has been established with similar columns that biomass is concentrated

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within the capillary fringe when volatile contaminants are the electron donors and oxygen


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is the electron acceptor 15, 16. Recent reports 45-47 indicated that bacterial growth and

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attachment to sand in the capillary fringe is concentrated in the transition zone when

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lysogeny broth was the carbon source for Escherichia coli and Pseudomonas putida

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strains. To determine the location of the bacteria that degrade AN and DPA, the microbial

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biomass was measured by qPCR throughout the column at the end of the experiments.

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Total bacterial numbers based on 16S rRNA genes were similar to the numbers

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based on the dioxygenase genes, indicating that most of the bacteria in the column were

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derived from the inocula (Fig. 4). The bacteria grew throughout the capillary fringe but

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were most concentrated near the saturated zone at the oxic/ anoxic interface (Fig. 4, 5).

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The results were similar to previously established microbial profiles in the unsaturated

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zone of columns designed to study biodegradation of volatile contaminants 15, 16. The

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initial inoculum that adhered to the sand (dashed line in Fig. 4) is highly overestimated as

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indicated by the brief contact time and the turbidity of the suspension (data not shown)

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that drained from the column after inoculation.

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Figure 4. Estimated microbial biomass in the AN (A) and DPA (B) multiport columns at

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the end of the experiments. Dashed lines represent the average initial bacterial inoculum

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calculated based on an assumption of 100% adherence of the bacteria to the sand during

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the brief exposure to the inoculum.

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Figure 5. Measured moisture content in the AN (A) and DPA (B) multiport columns at

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the end of the experiments described above.

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The correlation between moisture content and bacterial activity is well established

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48-51

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80% saturation 47 as would be found in the lower part of the capillary fringe where gas

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transport is greater than in the saturated zone. The results are consistent with previous

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observations that the microbial biomass accumulates at the transition zone in the capillary

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fringe 45-47. Similarly, a recent review concludes that maximum microbial activity in soil

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occurs where the water content and oxygen availability intersect 48. To investigate the

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effect of moisture content on biodegradation of contaminants, DPA was used as the

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electron donor for the microcosms prepared with different water saturation levels. The

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results revealed maximum biodegradation rates at 60-80 % saturation (Fig. 6), which

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corresponds to the lower level of the capillary fringe in the columns where most of the

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biomass was located.

308

and recent studies demonstrated that microbial activity is maximum between 50 and

Oxygen diffusion is the key parameter that causes the microbial biomass to

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accumulate in the transition zone above the saturated layer. Oxygen diffusion in the

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capillary fringe is correlated with the volumetric air content because the diffusion of

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oxygen in air is faster than in the water. Therefore, there is a sigmoidal decrease of

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oxygen in the capillary fringe above anoxic plumes in the absence of bacteria 52. When

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bacteria are present, biodegradation creates an oxygen sink in the capillary fringe and the

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rate of oxygen transfer to the saturated layers below is further diminished. Another

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parameter that affects oxygen diffusion to the water in the unsaturated layer is the air-

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water interfacial area which decreases as the water content increases 53. Because aerobic


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microbes require water and oxygen to function there is a narrow region where oxygen

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and moisture content are optimal for microbial activity 54. In addition the microbial

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growth is limited by the rate of diffusive transport of the carbon and energy source from

320

below. Thus, a transition zone in the capillary fringe supports microbial activity better

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than the saturated zone or the zone of maximum air-water interfacial area.

322 323

324 325 326

Figure 6. Growth (solid circle) and biodegradation rates (empty circle) of DPA degraders

327

at different water saturation levels. Data represent the means of duplicate analyses.

328 329

The rapid oxygen-dependent AN and DPA biodegradation in the capillary fringe

330

along with moisture and microbial profiles throughout the columns indicated that the

331

biodegradation takes place in a narrow zone where the moisture content is high, but not

332

saturated, and where electron donors and acceptors intersect. The results are consistent

333

with previous findings for volatile compounds 5, 13-16, 18 and with the predicted enrichment


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of microbial biomass at the interface between sources of electron donors and acceptors in

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plume fringes 7. Field data for many plumes of volatile dissolved hydrocarbons are

336

consistent with the above conclusions in that such plumes do not cause vapor intrusion

337

because of the high biodegradation rates immediately above the saturated zone 11. Even

338

though the design of the columns we used is a good representation of the lower part of

339

the vadose zone in an idealized system it has limitations that reduce the ability to

340

extrapolate to field situations. Diffusion rates of both contaminants and oxygen would

341

typically be lower in the field. Temperature, pH, nutrient concentrations, porosity or other

342

factors would also affect transport and microbial activity in the field. Therefore, the

343

findings should be confirmed at appropriate field sites before they are integrated into

344

conceptual site models or risk assessments.

345

The average concentration of DPA at a contaminated site was reported to be

346

between 1400 to 2900 mg/kg of soil 28 and near solubility concentrations (around 35

347

mg/L) in contaminated water 28. Similarly, previous treatability studies have considered

348

that industrial wastewaters could contain AN concentrations approaching 1000 mg/L 55,

349

56

350

the capillary fringe could be sufficient to remove considerable amounts of contaminant

351

mass from the plumes.

352

The column study established that biodegradation removed DPA and AN in the capillary

353

fringe. The microcosm studies established that natural attenuation of DPA could be

354

achieved in the unsaturated zone as long as DPA degraders are present. Although the

355

column was not designed to reflect field conditions at the contaminated site, the overall

356

results clearly indicated that natural attenuation in the capillary fringe can remove

. Calculations based on the columns used here indicate that the biodegradation flux in


19


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substantial mass of AN and DPA from an underlying groundwater plume and it seems

358

likely the findings can be extrapolated to other aerobically biodegradable non-volatile

359

contaminants.

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Models developed for biodegradation of hydrocarbons in the vadose zone

361

describe microbial activity either by instantaneous degradation or by first-order or Monod

362

kinetics considering oxygen availability as a limiting factor for biodegradation 5, 12-14, 57.

363

Biodegradation of chlorinated compounds (chlorobenzene) in the capillary fringe was

364

also modeled by including biomass concentration in addition to Monod kinetics 58. The

365

model revealed that increase in contaminant flux increases oxygen flux and

366

biodegradation rate resulting in instantaneous degradation in the capillary fringe when

367

there are no physical limitations. Because the diffusivity of volatile contaminants is faster

368

in air than in the water, modeling volatile contaminants relies only on the air

369

concentrations, therefore application of the volatile compound models is not appropriate

370

for non-volatile compounds. Development of a model including biomass and describing

371

biodegradation of non-volatile compounds in the capillary fringe will be required to

372

predict natural attenuation and its impact on plume concentrations in the field.

373 374

375

Acknowledgment

376

Funding was provided by the DuPont Corporate Remediation Group.

377 378

Supporting Information Available

379

Figure S1 is in supplementary figure section.

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References

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1.

384

resolved quantification of anaerobic toluene degraders and aquifer microbial community

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