Stimulating In Situ Hydrogenotrophic Denitrification

Stimulating In Situ Hydrogenotrophic Denitrification with Membrane-Delivered Hydrogen under Passive and Pumped Groundwater Conditions Brian P. Chaplin1; Matthew R. Schnobrich2; Mark A. Widdowson3; Michael J. Semmens4; and Paige J. Novak5 Abstract: A technology was developed to stimulate autotrophic biological denitrification by supplying hydrogen 共H2兲 to groundwater via gas-permeable membranes. The purpose of this project was to investigate this technology at field scale, determining whether it could be successfully scaled up from the laboratory. The field site was located in Becker, Minnesota and contained high levels of NO−3 共22.8⫾ 2.0 mg/ L-N兲 and dissolved oxygen 共DO兲共7 ⫾ 1 mg/ L兲. Membranes installed in groundwater wells were successful in delivering H2 to the groundwater over the two-year operating period. Hydrogen stimulated microbial reduction of DO and NO−3 , degrading up to 6 mg/L DO and converting up to 10.0 mg/L NO−3 -N to NO−2 -N when operated passively. When recirculation pumps were installed performance in the field did not improve significantly because of mixing with more oxygenated water. However, complementary modeling studies showed that complete DO reduction and denitrification to N2 was possible but the zone of influence and total H2 demand were limiting factors. Water was recirculated in the field from downgradient to upgradient membrane-containing wells to increase the H2 delivery through the membrane by an increase in water velocity. The depth to groundwater 共⬃13.7 m兲 caused some water reoxygenation during recirculation, which may preclude the use of this technology at deep sites, as this makes it more difficult to install sufficient wells and control recirculation. DOI: 10.1061/共ASCE兲EE.1943-7870.0000021 CE Database subject headings: Denitrification; Hydrogen; Membranes; Groundwater; Simulation; Biological treatment.

Introduction Nitrate 共NO−3 兲 is a common pollutant in groundwater, often as a result of septic tank discharge and the overapplication of fertilizers. The consumption of NO−3 -contaminated drinking water has been directly linked to methemoglobinemia 共blue-baby syndrome兲 in infants 共Comly 1945兲, and prompted the United States Environmental Protection Agency to enforce maximum contaminant levels 共MCL兲 for NO−3 and NO−2 of 10.0 mg/L and 1.0 mg/L as N, respectively, 共U.S. Environmental Protection Agency 1995兲. Studies estimate that approximately 5% 共U.S. Environmental Protection Agency 1990兲 to 11% 共Squillace et al. 2002兲 of wells in 1

Research Associate, The Dept. of Chemical and Environmental Engineering, Univ. of Arizona, 1133 E. James E. Rogers Way, Tucson, AZ 85721. 2 Project Engineer, ARCADIS G&M, Inc., 6 Terry Drive, Suite 300, Newton, PA 18940. 3 Professor, The Dept. of Civil and Environmental Engineering, Virginia Tech Univ., Patton Hall 220A, Blacksburg, VA 24061. 4 Professor, The Dept. of Civil Engineering, Univ. of Minnesota, 122 Civil Engineering Building, 500 Pillsbury Dr. S.E., Minneapolis, MN 55455. 5 Associate Professor, The Dept. of Civil Engineering, Univ. of Minnesota, 122 Civil Engineering Building, 500 Pillsbury Dr. S.E., Minneapolis, MN 55455 共corresponding author兲. E-mail: [email protected] Note. This manuscript was submitted on December 14, 2007; approved on November 11, 2008; published online on April 3, 2009. Discussion period open until January 1, 2010; separate discussions must be submitted for individual papers. This paper is part of the Journal of Environmental Engineering, Vol. 135, No. 8, August 1, 2009. ©ASCE, ISSN 0733-9372/2009/8-666–676/$25.00.

the United States contain NO−3 concentrations that exceed the MCL. Conventional methods for NO−3 removal, including ion exchange, reverse osmosis, and electrodialysis, are effective but have the disadvantage that they produce to concentrated brines that require disposal or additional treatment. Biological denitrification is a destructive method for NO−3 removal 共Lee and Rittmann 2002; Schnobrich et al. 2007兲 and is a promising solution that can convert NO−3 and NO−2 to nitrogen gas 共N2兲. It is also ideal for in situ applications where it can utilize indigenous microorganisms for NO−3 reduction and aquifer material for subsequent filtration of biomass and organic by-products 共Schnobrich et al. 2007兲. A variety of organic electron donors have been used 共e.g., ethanol, methanol, sucrose, acetate, and formate兲 that stimulate heterotrophic denitrifiers 共Mercado et al. 1988; Janda et al. 1988; Smith et al. 2001兲. High biomass production associated with heterotrophic denitrifiers and the resulting high organic content of the denitrified water may result in clogging and/or requires additional treatment 共e.g., filtration, chlorination, ultraviolet radiation兲, which can substantially increase the cost of drinking water treatment and can contribute to the formation of toxic disinfection by-products during disinfection 共Mercado et al. 1988兲. Inorganic electron donors such as hydrogen 共H2兲 gas, reduced iron compounds, and reduced sulfur compounds can also be used to stimulate autotrophic denitrifiers 共Batchelor and Lawrence 1978; Kurt et al. 1987; Gantzer 1995; Till et al. 1998; Lee and Rittmann 2002; Haugen et al. 2002; Schnobrich et al. 2007兲 while producing less biomass 共Rutten and Schnoor 1992兲. The use of H2 to stimulate autotrophic denitrification is desir-

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N Flow Direction

1.5 m

M= Monitoring Well H=Hydrogen Injection Well M2 M3 M1 M4

H6

H1 H2 H3 H4 H8 M6

H7

H5 M8

M7 M11

M5

M9

M10

M12

Fig. 2. Membrane module designs constructed in this study: 共a兲 initial open-ended module design run in series; 共b兲 sealed-end module design used in Experiments 1 and 2; and 共c兲 sealed-end module design with water recirculation. The MEMA configuration is shown in the expanded image.

Fig. 1. Field site map showing well placement

able because it is nontoxic, produces low biomass, and its low water solubility 共1.6 mg/L at 20° C兲 results in little residual. When delivered to water by hollow fiber membranes, H2 diffuses efficiently into the aqueous phase, preventing the bubbling of gas that can cause an explosive environment 共Gantzer 1995兲. In addition, hollow fiber membranes can support biofilm growth, allowing direct contact with the electron donor 共H2兲 and the electron acceptor 共NO−3 兲共Lee and Rittmann 2002兲. Hollow fiber membrane reactors have been tested successfully in several lab-scale studies 共Gantzer 1995; Lee and Rittmann 2002; Haugen et al. 2002; Chung et al. 2007; Schnobrich et al. 2007兲 but to our knowledge have not been applied at the field scale. The objective of this research was to test the feasibility of using hollow fiber membranes in situ to stimulate hydrogenotrophic denitrification at the field scale, upgradient of a production well. Specifically, the ability of H2 to stimulate denitrification when added passively or in conjunction with groundwater recirculation was tested. Impacts on downgradient water quality were assessed and a mathematical model was used to help identify optimum membrane placement and required surface area to achieve complete denitrification.

共Northern States Power 共NSP兲 1991兲. The well casings consisted of 5.1-cm outside diameter 共OD兲 polyvinyl chloride 共PVC兲 pipes and were screened over a length of 4 m, 14–18 m below the land surface. The aquifer material consisted of poorly graded sand and gravel with a mean grain diameter of 0.4 mm and porosity of 0.3. The groundwater was aerobic 共7 ⫾ 2 mg dissolved oxygen 共DO兲/L, average ⫾ 95% confidence interval兲 and contained 21.2⫾ 8.1 mg NO−3 -N/ L. A bromide tracer test was conducted to determine the direction and velocity of the groundwater flow and the local longitudinal dispersion coefficient in the aquifer. Water was pumped from Well M2 into two 20-L carboys and mixed with NaBr to create a stock solution of 2,500 mg/L bromide 共Br−兲. This solution was immediately pumped into Well H6 共approximately 1 m below the water table兲 at a rate of 1 L/min using a peristaltic pump. A Br− selective electrode 共Cole Parmer, model #27502-04, Vernon Hills, IL兲 equipped with a data logger 共WTW, pH340I, Germany兲 was positioned downgradient in Wells H2, H7, and M7 共approximately 1 m below the water table兲 to record hourly changes in the Br− concentration of the water.

Experimental Methods Membrane Module Design Site Description and Characterization The field site was located in Becker, Minnesota, approximately 100 km northwest of Minneapolis, Minn. The aquifer was a surficial sand and gravel aquifer confined at the base by a dense glacial till approximately 18–46 m below land surface 共Xcel Energy 2000; Northern States Power 共NSP兲 1991兲. The primary groundwater flow direction was from the northeast to southwest and the main discharge area was the Mississippi River 共located approximately 200 m downgradient of the field site兲共Xcel Energy 2000兲. The depth to groundwater at the site was approximately 15 m below the land surface. The well placement at the field site is shown in Fig. 1. This well placement was chosen based on the general groundwater flow direction being perpendicular to the Mississippi River

Two membrane module designs were tested at the field site 共Fig. 2兲. Both types of modules used 1-mm model # ⫻1.6-mm OD nonporous silicone-coated reinforced fiberglass membrane fibers 共Varflex Corporation, #20RC1, Rome, N.Y.兲. The first membrane module was designed to be operated in series 关Fig. 2共a兲兴 and each module consisted of 20 membrane fibers 共3.0-m length兲. The top and bottom of the fibers were sealed 共3M, Scotch Weld DP-125, St. Paul, Minn.兲 into 5.1-cm lengths of 2.5-cm OD stainless steel pipe. When operated in series, the exhaust gas from the one module was connected to the influent gas line of another module via Tygon tubing. Gas from the last module was vented to the atmosphere. This membrane module design was tested in the field under constant H2 flow but was plagued by water condensation within the membranes, which

JOURNAL OF ENVIRONMENTAL ENGINEERING © ASCE / AUGUST 2009 / 667


Table 1. Summary of Operating Parameters for Field Experiments Passive denitrificaton experiments

Groundwater recirculation experiment

Duration 共days兲

Membrane surface area 共m2 / well兲

Hydrogen delivery wells

Hydrogen pressure 共atm兲

Days of operation

Extraction well; flow rate 共L/min兲

1

111

0.12

H1, H2, H3, H4

0–62a

M7; 4.0

H1, H3, H4, H6

1.68

2

42

0.12

1.68 共days 0–82兲, 2.36 共days 83–96兲, 0 共days 97–11兲 1.68 共days 0–27兲, 2.36 共days 27–42兲

63–91

M7; 4.0

H3, H4, H6

1.68

Experiment number

H1, H2, H3, H4, H6, H7, H8 a On Day 58 a flow distributor was added to each injection well.

prevented gas flow 共results not shown兲. To overcome these limitations, a second membrane design was developed. The second type of membrane module was designed with a sealed end 关Fig. 2共b兲兴. A single membrane fiber, 24.0 m in length, was coiled to form a 3-m-long module 共approximately 1 loop/ cm兲. The influent gas line was connected to one end of the membrane and the other end was attached to a moisture escape membrane apparatus 共MEMA兲. The MEMA was made with six polyethylene microporus membranes 共5.1 cm long, 0.30-mm OD, 0.26-␮m ID兲共Mittsubishi Rayon, MHF 300, New York, N.Y.兲. The polyethylene membranes were sealed at the terminal end with DP-125 epoxy and wetted with methanol immediately before they were installed in the well. When the coiled membrane was pressurized, any condensed water inside the membrane was transported to the surrounding groundwater. A 4.4-cm solid stainless steel weight was also attached to the bottom of the membrane to compensate for the buoyancy of the membrane module and to keep the membrane fully extended 共⬃3 m兲 within the well. Field Experiments Passive Denitrification Experiments The membrane modules were placed in the wells and positioned approximately 0.5–1.0 m below the water table. Industrial grade H2 gas was supplied to the inside of the silicone-coated membranes 共0–2.36 atm兲. Two replicate field experiments were conducted with this setup and the operating parameters are summarized in Table 1. Experiment 1 lasted 111 days with sealed–end membrane modules placed in Wells H1, H2, H3, and H4, all connected in parallel. Experiment 2 lasted 42 days with sealed–end membrane modules in Wells H1, H2, H3, H4, H7, H8, and H6 with the modules connected in parallel. Groundwater Recirculation Experiment A 91-day groundwater recirculation experiment was also conducted on site and the operating parameters are summarized in Table 1. A Sta-Rite Convertible Deep Well Jet Pump 共Sta-Rite Industries, Delavan, Wis.兲 was installed in Well M7 to draw water from 1.8 m below the water table at a rate of 4 L/min. At the land surface, the pump discharge was connected to a four-way PVC manifold and the total flow volume was split evenly to the injection wells, and was delivered at the top of the membrane module. Each injection well contained a membrane module connected in parallel and pressurized with 1.68 atm of H2. On Day 58, a flow distributor was added to each membrane-containing well in an attempt to improve the water flow past the membrane. The flow distributor was a 2.4-m length of 1.3-cm 共1/2 in.兲 PVC pipe that was sealed at the bottom and contained 4-mm holes along its length. It was mounted in the center of the coiled membrane to

Injection wells

Hydrogen pressure 共atm兲

distribute pumped water radially outward through the membrane and thus promote a more uniform H2 delivery to the surrounding aquifer 关Fig. 2共c兲兴. Groundwater Sampling Wells were sampled approximately every one to two weeks for NO−3 , NO−2 , DO, pH, temperature, total organic carbon 共TOC兲, alkalinity, and water level. Samples were collected using a submersible 4.6-cm diameter pump 共Grundfos, Redi-flow2 pump, Fresno, Calif.兲 into 250-mL Teflon-lined polyurethane bottles 共Nalgene兲. Dissolved H2 was not quantified at the site because it was determined that the submersible Grundfos pump produced H2 by electrolysis during sampling; this has been observed by others using similar submersible pumps 共Chapelle et al. 1997兲. NO3− and NO2− Microcosm Experiments Microcosm experiments were conducted in the laboratory to determine kinetic parameters for the reduction of NO−3 and NO−2 for use in the model. Sediment was obtained from the upper 2 m of the saturated zone at the Becker site and was stored under aerobic conditions at 10° C prior to use. Microcosms were prepared in triplicate in 715-mL glass bottles 共Wheaton, Millville, N.J.兲 containing 30 g wet sediment, 297-mL synthetic groundwater 关per 1 L Milli-Q 共Millipore, Bedford, Mass.兲 water兴: 23 mg MgSO4, 167-mg MgCl2 · 6H2O, 247-mg NaHCO3, 13-mg KHCO3兲, 3 mL of a trace element stock solution 共per 1 L Milli-Q water: 8.5-mg K2HPO4 · 3H2O, 13.4-mg MnCl2 · 6H2O, and 5-mg FeCl2 · 4H2O兲, and varying amounts of NaNO3 or NaNO2 共1–18 mg/L as N兲. The liquid in all microcosms was sparged with ultrahigh purity N2 for approximately 2 min to remove DO. The headspace was flushed with 10% H2 / 90% N2 gas for approximately 30 min and the electron donorfree biological controls were flushed with 100% N2. The bottles were sealed with screw-on caps fitted with butyl rubber septa for sample collection 共Wheaton, Millville, N.J.兲. Sterile controls were prepared by autoclaving microcosms containing a 10% H2 / 90% N2 headspace for 30 min. The microcosms were placed on their sides on a shaker table 共New Brunswick Scientific, Model R-2, Edison, N.J.兲 and shaken at 250 rpm at 10° C for the duration of the experiment. A 3.5-mL liquid sample was withdrawn to quantify NO−3 and NO−2 concentrations every 1–3 days. Analytical Methods The DO concentrations were measured in the field using 1–12 mg/L Chemet ampules 共Chemetrics Inc., Calverton, Va.兲. During the groundwater recirculation experiment, DO was measured with



a DO probe 共MI-730 Dip-type O2 microelectrode兲 and meter 共OM-4, oxygen meter兲 from Microelectrodes Inc. 共Bedford, N.H.兲. Additionally, a YSI Model 58 共YSI Inc., Yellow Springs, Ohio兲 DO meter and probe with a 30.5-m lead were used to obtain in situ DO concentration profiles. Temperature and pH were measured in the field using a portable pH meter 共Oakton, model-pH 6兲 and electrode 共Oakton, model-WD-35801-00, Vernon Hills, Ill.兲. Water levels were measured using an electronic water level finder 共Slope Indicator Co., Model 51453, Seattle, Wash.兲 before pumping the groundwater. Quantification of alkalinity 共mg/L as CaCO3兲 was determined by Standard Method 2320 B 共American Public Health Association 共APHA兲 et al. 1998兲. Nitrate and nitrite concentrations were determined by ion chromatography 关Metrohm 761 Compact Ion Chromatograph 共Herisau, Switzerland兲; Metrosep A Supp 5 column; 3.2-mM Na2CO3 and 1.0-mM NaHCO3 eluent; 0.7 ml/min eluent flow rate; 100 mM H2SO4 regenerant兴. The method detection limits 共MDLs兲 for NO−3 and NO−2 were 0.035 mg/L NO3-N and 0.033 mg/L NO2-N, respectively, as determined by Standard Method 1030 C 共American Public Health Association 共APHA兲 et al. 1998兲. The TOC was analyzed in triplicate using a Dohrmann Phoenix 8000 TOC analyzer equipped with a STS 8000 autosampler 共Dohrmann, Mason, Ohio兲. The MDL for TOC was 0.12 mg/L as carbon.

Mathematical Model

source utilization 共M / L3 / T兲; Rflux,H accounts for the mass flux of H2 k from the membranes 共M / L3 / T兲; Rsin bio,E represents the mass loss of electron acceptor due to biological utilization 共M / L3 / T兲; and − Rsource bio,E accounts for the production of NO2 , according to the stoichiometry in Eq. 共9兲.

H2 Source Term The source term for the mass transfer of H2 to the aqueous phase source 共Rflux,H 兲 is expressed as the mass transfer rate per unit volume of liquid source Rflux,H =

Transport Equations The general form of the 2D equations of mass balance for electron donor and acceptor transport and biodegradation are − vx

⳵H ⳵ 2H ⳵ 2H ⳵H sin k source + Dx 2 + Dy 2 − Rbio,H + Rflux,H = ⳵x ⳵x ⳵y ⳵t

共1a兲

⳵E ⳵E ⳵E ⳵E k source + Dx 2 + Dy 2 − Rsin bio,E + Rbio,E = ⳵x ⳵x ⳵y ⳵t

共1b兲

2

˙ = KA共Hⴱ ␻ − H兲 M H L

− vx

respectively, where H = aqueous phase concentration of hydrogen 共M / L3兲; E = electron acceptor concentration 共M / L3兲; x and y are distances along and across the direction of flow, respectively 共L兲; vx = average groundwater velocity 共L/T兲; Dx and Dy are the dispersion coefficients in the x- and y-directions, respectively sin k 共L2 / T兲; Rbio,H represents the mass loss of H2 due to biological

共3兲

where K = overall H2 mass transfer coefficient 共L/T兲; A = surface area of the membrane 共L2兲; ␻ = molecular weight of H2 共M/mole兲; and HLⴱ = H2 liquid saturation concentration 共mol/ L3兲, which is related to the H2 partial pressure by the Henry’s Law constant. The value for K can be predicted using the following dimensionless relationship 共Fang et al. 2002兲: K = 0.824

冉冊冉冊冉冊 ␷ wd ␯

D d

0.39

␯ D

0.33

共4兲

where D = diffusivity of H2 in water 共L2 / T兲; d = outside membrane diameter 共L兲; ␷w = water velocity across the membranes 共L/T兲; and ␯ = kinematic viscosity of water 共L2 / T兲. Reaction Terms sin k 兲, expressed as the sum The reaction term for H2 utilization 共Rbio,H − of utilization as a result of DO, NO3 , and NO−2 reduction, is

sin k Rbio,H =

M ␪

M

兺 ␯i = ␪ 共␯H,O + ␯H,N

1

+ ␯H,N2兲

共5兲

where M = autotrophic biomass per volume of porous medium 共M / L3兲; ␪ = aquifer porosity; and ␯H,O, ␯H,N1, and ␯H,N2 are the H2 utilization rates by the autotrophic biomass for DO, NO−3 , and NO−2 reduction, respectively 共M/M per T兲. The DO, NO−3 , and k NO−2 mass loss rates Rsin bio,E, resulting from their utilization as an electron acceptor, are linked to H2 utilization by the stoichiometry of the biological reduction k Rsin bio,E = ␥E

2

共2兲

where M H = mass transfer rate of H2 out of the membrane 共M/T兲; ␪w = porosity of the mass transfer zone 共L3 / L3兲; and VT = total volume of the mass transfer zone 共L3兲共i.e., one model cell兲. The H2 mass transfer rate under steady state conditions is expressed as

Conceptual Model The mathematical model used in this study was an extension of an existing two-dimensional 共2D兲 model that incorporated advectivedispersive transport and biodegradation of one electron donor 共formate兲 and two electron acceptors 共NO−3 and NO−2 兲 coupled with the growth and decay of a single microbial population 共Killingstad et al. 2002兲. The Killingstad et al. 共2002兲 model was modified to simulate constituent concentrations in a real, 2D domain in which concentrations were assumed to be uniform with depth, and to account for the transfer of H2 through hollow-fiber membranes, according to the gas transfer correlation developed by Fang et al. 共2002兲. The model was also modified to model three electron acceptors 共O2, NO−3 , and NO−2 兲. The microbial population was a single facultative population attached to sediment particles. The population occupied no physical space within the model domain.

˙ M H V T␪ w

M ␯H,E ␪

共6兲

where ␥E represents the mass of electron acceptor reduced per unit mass of hydrogen 共M/M兲 and E represents either DO, NO−3 , or NO−2 . Utilization rates in Eqs. 共5兲 and 共6兲 are represented by dual Monod kinetics of the electron donor and acceptor. Switching functions prevented the uptake of NO−3 and NO−2 until DO was below a set level 共␬O,N1 and ␬O,N2, respectively兲, and prevented the degradation of NO−2 until NO−3 was below a set level 共␬N1,N2兲共Killingstad et al. 2002兲.



Table 2. Values for Model Input in Field Site Simulation Variables initial condition

Value

Units

Source

O N1 N2 H M

6.00 24.10 0.00 — 0.22

g/m g / m3 g / m3 g / m3 g / m3

Sampling data Sampling data Sampling data Fang et al. 共2002兲 Microcosm experiment

Parameters ␷x ␣x ␣y ␪ ␪w ␥O ␥N1 ␥N2 YO Y N1 Y N2 kd max ␷H,O max ␷H,N1 max ␷H,N2

Value 0.31 0.02 0.002 0.30 0.90 7.62 6.72 4.45 0.19 0.18 0.18 0.05 3.50 1.37 0.19

Units m/day m m

g O/g H g N1/g H g N2/g H g bio/g H g bio/g H g bio/g H day−1 g H/g bio/day g H/g bio/day g H/g bio/day

KH,O

0.02

g / m3

KH,N1

0.02

g / m3

KH,N2

0.02

g / m3

␬O,N1

2.00

g / m3

␬O,N2 ␬N1,N2

2.00 2.0⫻ 10−5

g / m3 g / m3

Source Tracer test Tracer test Killingstad et al. 共2002兲 Experimental Calculated McCarty 共1972兲 McCarty 共1972兲 McCarty 共1972兲 McCarty 共1972兲 McCarty 共1972兲 McCarty 共1972兲 McCarty 共1972兲 Estimated Microcosm experiment Microcosm experiment Average of values reported from Killingstad et al. 共2002兲 and Kurt et al. 共1987兲 Average of values reported from Killingstad et al. 共2002兲 and Kurt et al. 共1987兲 Average of values reported from Killingstad et al. 共2002兲 and Kurt et al. 共1987兲 Interstate Technology and Regulatory Cooperation Work Group 共ITCR兲共2000兲 Interstate Technology and Regulatory Cooperation Work Group 共ITCR兲共2000兲 Experimental

3

Microbial Densities Changes in the microbial densities are simulated by a mass balance equation that accounts for biomass growth and decay terms 1 ⳵M = 共Y O␯H,O + Y N1␯H,N1 + Y N2␯H,N2兲 − kd M ⳵t

共7兲

where Y O, Y N1, and Y N2 are the yield coefficients 共M/M兲, the ratio of microbial biomass produced per mass of electron donor consumed for DO, NO−3 , and NO−2 respiration, respectively, and kd is the biomass decay coefficient 共M/M per T兲. Eq. 共7兲 is solved directly and is updated at each time step.

represented as no-flow boundaries, and the downgradient boundary was set as a zero gradient condition. Hydrogen source nodes were added at the well locations to represent H2 addition via membrane modules. Parameter Estimation

Model Setup

Input parameters for the model were obtained from the field site tracer study, microcosm and laboratory experiments, stoichiometric relationships, and from the literature. No parameters were fit to the field data. Table 2 is a complete listing of input parameters used in the model. Model simulations included a site-wide analysis showing the entire field site study area and model runs focused on the apparent flow path through Wells H3 and M5.

A 2D finite-difference grid was used with a 0.0508-m node spacing. Initial conditions for concentrations of NO−3 , NO−2 , and DO in the model were determined by sampling data collected just prior to H2 addition and were assumed to be uniform throughout the model domain. The initial condition for H2 concentration was assumed to be zero throughout the model domain. Concentrations of NO−3 , NO−2 , DO, and H2 were assumed to be constant with time and uniform over the width of domain at the inflow boundary. The lateral, upper, and lower boundaries were

Transport Groundwater velocity, direction of flow, and the longitudinal dispersion coefficient were obtained from the analysis of the Br tracer study 共Table 2兲. The values for ␷x and ␣x were obtained by analyzing the Br data between Wells H2 and M7 using the method of moments 共Valocchi 1989兲. The model was aligned to the direction of flow. A value for Dy was obtained by assuming that Dy = 0.1Dx 共Killingstad et al. 2002兲. Numerical dispersion



Parameter

Units

mg/L NO3-N mg/L NO2-N DO mg/L pH pH units Temperature °C TOC mg/L as C Alkalinity mg/L as CaCO3

Average

95% confidence Number of interval samples

21.2 0 7.0 7.2 11.4 0.6 182

8.1 0 2 0.4 2.4 0.4 14

97 97 51 84 66 59 25

Concentra ation (mg/L)

Table 3. Background Water Quality Parameters and Values

30 25 20 Nitrate (as ( N))

15

Nitrite (as N)

10

DO

5 0 0

20

40

60

80

Days

was limited by requiring the Courant number to approach unity. Aquifer porosity and sediment density were set at constant values of 0.30 and 2.0⫻ 10−6 g / m3, respectively. Stoichiometric Coefficients The electron acceptor utilization coefficients 共␥O, ␥N1, and ␥N2兲 and cell yields 共Y O, Y N1, and Y N2兲 for DO, NO−3 , and NO−2 reduction, respectively, were derived from McCarty 共1975兲. A cell residence time of 1,000 days was used to account for stationary biomass and a decay coefficient of 共0.05 day−1兲 was assumed. The following stoichiometric relationships were calculated according to the methods presented by McCarty 共1975兲: O2共aq兲 + 0.01NO3 + 0.009H+− + 0.046CO2 + 2.13H2 → 2.11H2O + 0.009C5H7O2N

共8兲

NO−3 + 0.0044H+ + 0.021CO2 + 1.06H2 → 1.05H2O + 0.0044C5H7O2N + 0.9956NO−2

共9兲

NO−2 + 0.007NO−3 + 0.007H+ + 0.033CO2 + 1.59H2 → 2.08H2O + 0.00657C5H7O2N + 0.5N2

共10兲

where C5H7O2N represents biomass. Utilization Rates and Initial Biomass Concentration The NO−3 and NO−2 microcosm experiments were used to obtain a max value for M and the Monod parameters ␯H,E , and KE. The average value for M was converted to a representative site-wide value based on density and porosity at the field site.

Results and Discussion

Fig. 3. Field results for NO−3 , NO−2 , and DO at Well M5 collected during experiment 1 共error bars represent 2% sampling error兲. Hydrogen at 1.68 atm.

tions observed in Experiment 1 in Well M5 are shown in Fig. 3. On Day 3 the DO concentration in Well M5 began to decrease and reached a concentration of approximately 2 mg/L by Day 13, significantly below the background DO concentration. Concentrations of DO were relatively constant from Day 13 to 41, after which time they increased gradually to background levels 共6 mg/L兲 by Day 111. The increase in the lumen H2 pressure to 2.36 atm on Day 83 did not appear to affect the DO level in Well M5. The concentration of NO−3 at Well M5 was approximately 25 mg-N/L at Day 0 and remained relatively constant until Day 10 共Fig. 3兲. The NO−3 concentration decreased steadily from Day 10 to 19, was approximately 15 mg-N/L from Day 19 to 48, and began to increase after Day 48 to background levels 共Fig. 3兲. Nitrite was observed in Wells H5, M5, H7, and M8, all downgradient of H2-addition wells; no NO−2 was detected upgradient of the H2-addition wells. The largest concentration of NO−2 was detected in Well M5. The NO−2 concentration profile in Well M5 appeared to mirror that of NO−3 in the same well 共Fig. 3兲, suggesting that most of the NO−3 was reduced to NO−2 . Nitrite was not observed in Well M5 after Day 83. The TOC was measured both the upgradient and downgradient of Well H3. The TOC levels in all of the monitoring wells were similar 共e.g., 0.4⫾ 0.2 mg/ L as C in Well M2 and 0.6⫾ 0.2 mg/ L as C in Well M5兲. A TOC sample taken on Day 60 in Well H3 where H2 was added; it was significantly higher than the value observed in the other wells, 2.29 mg/L as C. The elevated TOC level in Well H3 was attributed to increased biomass production within the well as a result of H2 addition.

Background Water Quality Average values for each water quality parameter measured from all wells at the site before H2 addition are shown in Table 3. The average NO−3 and DO concentration before H2 addition was 21.2⫾ 8.1 mg/ L as N and 7.0⫾ 2.0 mg/ L, respectively 共⫾ 95% confidence interval兲. Nitrite was never detected in the field before the H2 addition. The pH, temperature, TOC, and alkalinity were 7.2⫾ 0.4, 11.4⫾ 2.4° C, 0.6⫾ 0.4 mg/ L as C, and 182⫾ 14 mg/ L as CaCO3, respectively. Results from Passive Denitrification Experiments Experiment 1 The membranes in Wells H1, H2, H3, and H4 were pressurized with 1.68 atm H2 on Day 0. The DO, NO−3 , and NO−2 concentra-

Experiment 2 A replicate H2-addition experiment was performed approximately two weeks after Experiment 1 ended. Membranes in Wells H1, H2, H3, H4, H7, H8, and H6 were initially pressurized with 1.68 atm H2 on Day 0. On Day 14, the membrane in Well H1 was removed because of leaks. Nitrogen and DO data for Experiment 2 are shown in Fig. 4. The largest amount of NO−2 production and DO reduction was once again observed in Well M5. The DO in Well M5 remained constant at 6 mg/L until Day 10 when it dropped to 5 mg/L and remained around this level. The maximum concentrations of NO−2 observed in Wells H5, M5, M7, and M8 throughout Experiment 2 were 0.55, 2.81, 0.34, and 2.17 mg-NO−2 / L, respectively. On Day 10, NO−2 was first observed in Well M5 and concentrations between 0.32 and 1.69 mg-N/L were observed for the remainder of the second experi-



30

Concentra ation (mg/L)

25 20

Nitrate (as N) Nitrite (as N)

15

DO Model

10 5 0 0

10

20

30

40

Days Fig. 4. Nitrate, nitrite, and DO concentrations and model results at Well M5 during Experiment 2. Hydrogen at 1.68 atm.

ment. As in Experiment 1, NO−3 and NO−2 profiles approximately mirrored one another. In addition, the total nitrogen concentrations in Wells M5 共downgradient兲 and M2 共upgradient兲 were 26.3⫾ 0.2 and 25.4⫾ 0.4 mg-N/ L, respectively, during the experiment. This observation indicates that NO−3 was reduced to NO−2 without complete denitrification to N2. On Day 28, the H2 pressure in the membranes was increased to 2.36 atm but no changes in NO−3 , NO−2 , and DO concentrations were observed 共Fig. 4兲. At the end of the experiment, the membrane modules were removed and the membrane-containing wells 共H2, H3, H4, H7, H8, and H6兲 were immediately sampled. The average concentrations of NO−3 , NO−2 , and DO were 19.2⫾ 3.3 mg-N/ L-N, 5.1⫾ 2.6 mg-N/ L, and 3.0⫾ 1.0 mg/ L, respectively. Well H3 showed the highest levels of denitrification with NO−3 , NO−2 , and DO concentrations on Day 42 of 16.7 mg/L-N, 7.0 mg/L-N, and 1.0 mg/L, respectively. The TOC levels measured in upgradient and downgradient wells at M2 共0.4 mg/L as C兲, M5 共0.6 mg/L as C兲, and M8 共0.5 mg/L as C兲 were similar. Nevertheless, the TOC values in the membrane-containing wells on Days 18 and 42 were higher 共2.1⫾ 0.4 mg/ L as C in Well H3 and 2.5⫾ 1.1 mg/ L as C in other membrane-containing wells兲. The elevated TOC levels in the membrane wells were again attributed to biomass production on or near the membrane fibers. No significant increase in TOC concentration was observed downgradient of Well H3, demonstrating that the aquifer effectively filtered out the biomass. There was no change observed in pH or alkalinity in the field as a result of denitrification. Discussion of Passive Denitrification Experiments Although oxygen was observed in samples containing NO−2 , it is assumed that this was a result of oxygen mixing into the sample during sample withdrawal rather than the simultaneous reduction of oxygen and NO−3 . The sequential reduction of oxygen followed by NO−3 was confirmed in the Becker aquifer material by Schnobrich et al. 共2007兲 and by the microcosm studies. The greatest quantity of denitrification was observed in Well M5 during Experiment 1 in which concentrations of DO and NO−2 were 2.0 mg/L and 8.0 mg/L-N, respectively. According to the stoichiometry shown in Eqs. 共8兲–共10兲, approximately 1.74 mg/L H2 is needed for this observed reduction. This quantity of H2 is

comparable to the solubility of H2 at 10° C 共1.73 mg/L兲 and the concentration of H2 共⬃1.5 mg/ L兲 that is predicted by the clean water gas transfer correlation developed by Fang et al. 共2002兲. Numerous studies have shown that biofilm growth on a membrane surface results in a gas transfer rate as high as 14 times that predicted by clean water gas transfer correlations 共Semmens and Essila 2001; Haugen et al. 2002; Edstrom et al. 2005; Schnobrich et al. 2007兲. The H2 needed to reduce DO and NO−3 and NO−2 below their respective MCLs at the site is approximately 6.6 mg/L H2, only approximately four times higher than the clean water gas transfer correlation. The residence time within the H2-delivery wells 共⬃1.0– 3.0 h, depending on the well capture zone兲, should be sufficient for microbial consumption to increase the overall H2 mass transfer out of the membrane. In fact, laboratory results from Schnobrich et al. 共2007兲, using aquifer material from the Becker site with a nearly identical flow velocity, slightly shorter residence time 共1.45 h兲, and slightly more 共3.5 to 10 times greater兲 membrane surface area per groundwater flow rate, demonstrated that it was possible to add sufficient H2 to meet the electron donor demand at the Becker field site. Indeed, the flow velocity, and therefore the Reynolds number, in the well bore was expected to be approximately three times that in the laboratory experiment by Schnobrich et al. 共2007兲. We would therefore expect enhanced H2 transfer from the membrane to the surrounding water in the well, resulting in an increased extent of denitrification in the well itself 共Fang et al. 2002兲. Nevertheless, the lack of more significant denitrification within the membrane wells themselves 共H2, H3, H4, H7, H8, and H6兲, indicated that other factors at the site, such as low phosphorus concentrations 共Schnobrich et al. 2007兲, may have contributed to incomplete denitrification. Because the zone of influence of each membrane-containing well was relatively narrow 共Haugen et al. 2002; Clapp et al. 2004兲, lower levels of denitrification would be observed downgradient as treated water mixed with untreated water from the surrounding area, resulting in higher observed NO−3 concentrations. In addition, slight shifts in the direction of flow could have profound impacts on observations at downgradient wells. Therefore, modeling studies were used to help determine how to improve performance in the field. Modeling Results Microcosm Studies Microcosm experiments were conducted to characterize NO−3 and NO−2 reduction kinetics for use in the model. All experiments showed an initial lag period of 10–24 days, which was not considered in the determination of the reduction kinetics. No observable degradation of NO−3 or NO−2 occurred in either the sterile or N2 headspace controls in any of the experiments. The microcosm experiments yielded an average value for M of 0.22 g / m3 and values for ␯H,N1 and ␯H,N2 of 1.37⫾ 0.51 and 0.19⫾ 0.04 g H2 / g biomass/day, respectively. Results indicated that NO−3 and NO−2 reduction were following zero order kinetics at field concentrations. This finding is supported by the literature Kurt et al. 共1987兲. In the NO−3 microcosm experiments, NO−3 was converted to NO−2 but no NO−2 was reduced until NO−3 was completely degraded. For this reason, the value of ␬N1,N2 was set to 2.0⫻ 10−05 g / m3. Field Experiments A site-wide model simulation was conducted to illustrate the overall effectiveness of the passive remediation strategy. A two-



Fig. 5. Steady state NO−2 concentration from the site wide model

Normalized Nitrogen Concentration

dimensional plan view of steady state concentrations of NO−2 downgradient of the H2 addition wells is shown in Fig. 5. The model predicts that denitrification occurred in narrow bands of approximately 0.15 m 共⬃3 times the well diameter兲. The model also showed that NO−2 concentrations should be detected at Wells M7, M8, H5, and M5, with the largest NO−2 concentrations detected at Well M5. This result is consistent with the field sampling data from both Experiments 1 and 2 共Figs. 3 and 4兲. In addition, if denitrification occurred over a large transverse direction, depleted DO or accumulation of NO−2 should have been observed in Wells M6 and M11 but this was not the case. Simulated and observed NO−3 and NO−2 concentration data are shown in Fig. 6共a兲 for the first 80 days of Experiment 1. The NO−3 and NO−2 data were normalized to total nitrogen 共NO−3 + NO−2 兲 to

1 0.8 0.6

Nitrate

a)

Nitrite

0.4

Model

0.2 0 0

20

40

60

80

Normalized Nitrogen Concentration

Days

1 0.8

b)

0.6

Nitrate Model

0.2 0 0

20

40

60

Design Considerations The ability of the model to simulate experimental results from the field site by using only the microcosm data and literature values to obtain kinetic parameters suggested that it could be useful for predicting the performance of different design scenarios at the field site. Model simulations were therefore used to determine the total membrane module surface area and the number of membrane wells needed to reduce DO and both NO−3 and NO−2 below the regulatory MCLs. According to Eq. 共3兲, the mass transfer rate of H2 can be controlled by the surface area of the membrane and the H2 partial pressure in the membrane. Using experimental results for membrane spacing 共Fang et al. 2002兲, the membrane fiber can be coiled 600 coils/m to produce a total membrane surface area of 0.77 m2 per well, which yielded a membrane surface area to fluid volume ratio of 131 m−1. Pressurizing these membranes with 2.36 atm of H2 would result in DO of 1.0 mg/L within 6 days, NO−3 from 25.0 mg-N/L to less than 5.0 mg-N/L within 35 days, and NO−2 of 10 mg-N/L within 60 days. Although a marked improvement in performance, this is still inadequate for meeting the NO−2 MCL of 1 mg-N/L. Simulations were performed with an additional membrane-containing well placed between Wells H3 and M5. Results showed that both NO−3 and NO−2 were reduced below the MCLs at Well M5 in approximately 90 days. Nevertheless, this well placement produced a treated swath of water only 0.15-m wide. Therefore, a successful application of this treatment strategy would require the following: 共1兲 two membrane-containing wells along the direction of flow with a specific membrane surface area of 131 m−1; 共2兲 a well spacing of three times the well diameter perpendicular to flow; 共3兲 a hydrogen pressure of 2.68 atm in the membranes; and 共4兲 approximately 90 days for biomass growth. Recirculation Experiment

Nitrite

0.4

account for variable NO−3 concentrations observed in Well M5 共Fig. 6兲. For the first 48 days of operation, the model simulated the changes in concentrations of NO−3 and NO−2 quite well and it also predicted the pseudosteady state conditions that were reached in Experiment 1. After Day 48 the model results did not agree with field observations. Because of the narrow bands of treated water that result from this system, small changes in the groundwater flow direction can cause dramatic changes in the observed concentration at downgradient monitoring wells 共Fig. 5兲. Model results observed at Well M5 after a 5° shift in groundwater flow direction are shown in Fig. 6共b兲. As seen, a small change in the groundwater flow direction could have a significant impact on our ability to observe the effect of H2 addition. Model simulations for DO, NO−3 , and NO−2 at Well M5 for Experiment 2 based on the flow direction in Fig. 6共b兲 are shown in Fig. 4 with the collected field data.

80

Days

Fig. 6. Model results 共Experiment 1兲 and field data normalized to total nitrogen concentrations at Well M5. 共a兲 Original groundwater flow direction; 共b兲 shift of groundwater flow direction by 5°.

Although the membranes installed at the site were successful in delivering H2 to the groundwater during Experiments 1 and 2, the zone of influence was limited and the overall mass of H2 delivered was insufficient to stimulate complete denitrification; this was confirmed by the modeling simulations. To address these problems, water recirculation was initiated to increase H2 gas transfer by increasing the velocity of the water flowing past the membranes 关see Eq. 共4兲兴. During the 91 days of pumping operation, NO−3 , NO−2 , and DO concentrations fluctuated, as shown in Fig. 7. Total nitrogen



DO and d Nitrogen conc. (mg/L)

30

25

20

15

Nitrate Nitrite DO

10

5

0

00

20 20

40 40

60 60

80 80

100 10

Days

Fig. 7. DO and nitrogen concentrations observed at Well M5 during the groundwater recirculation experiment

The membrane modules were removed on two occasions, Days 58 and 91. On Day 58 approximately 5.5 mg NO−2 -N/ L and 1.1 mg/L DO were observed in Well H3. On Day 91, approximately 8.5 mg NO−2 / L was observed in Well H4. These results confirm that membrane delivery of H2 was still able to reduce DO and NO−3 in these wells despite the reintroduction of O2 and NO−3 -rich groundwater. Agarwal et al. 共2005兲 tested the effects of pumping on membrane gas 共O2兲 transfer and the zone of influence in a laboratory tank reactor filled with fine sand. Under passive conditions, with a flow velocity of 1 m/day, the distance between the 1 mg/L O2 contour was 7.3 times the diameter of the membrane module well 共Agarwal et al. 2005兲. At pumping rates of 1, 20, and 40 mL/min into the membrane well, however, the zone of influence was found to be approximately 2.4, 11.0, and 15.7 times greater than passive operation, respectively 共Agarwal et al. 2005兲. These results suggest that the dispersion of gas from a hollow fiber membrane should increase with groundwater recirculation if it is performed carefully to avoid reoxygenation.

Technology Applications concentrations between the upgradient Well M2 共23.4 ⫾ 0.9 mg-N/ L兲 and Well M5 共23.4⫾ 0.5 mg-N/ L兲 were not significantly different. Nitrite accumulation reached a maximum concentration of approximately 8.0 mg NO−2 -N/ L on Day 62, four days after installing the flow distributor. Profiles of DO concentrations taken in situ are shown in Fig. 8. The low DO concentrations observed on Day 62 are indicative of microbial activity. Between Days 62–73, however, air was observed entering the manifold through one of the delivery tubes 共Well H1兲, and as a result, pumping to Well H1 was stopped. Once pumping to H1 stopped, the DO profiles obtained for M5 on Day 78 showed that O2 entrainment had ceased. The reduction of 4.0 mg NO−3 -N/ L to NO−2 was observed in Well M5 at this time. In general, the pump installed at Well M7 appeared to draw a significant amount of water from untreated regions of the aquifer. The mixing of DOand NO−3 -rich groundwater with treated water prevented consistent observation of NO−3 reduction at, and downstream of, membrane-containing wells H6, H3, and H4. As a result, the H2 required for reducing O2 and NO−3 at the site increased.

Feet Dow wn from Surface

0

2

DO (mg/L) 4

6

8

48 50 52 54 56

M2 (mean) Day 62

58

Day 73

60

Day 78

62

Day 83

Fig. 8. DO readings taken in Well M5 with 50⬘ probe extension. Error bars on M2 value represent the standard deviation in background locations. The membrane delivery depth in aquifer was between 50–60 ft below grade.

The field site contained background NO−3 and DO levels that were extremely high; this presented a worst-case scenario in which to challenge our technology. Delivering H2 by discreet membranecontaining wells produced only a small zone of influence. Therefore, closer well spacing 共⬃3 times the well bore diameter兲, water recirculation, or trenching and the use of membrane fiber cloth must be considered to achieve complete denitrification in the field. Because of cost considerations, these strategies are only possible at a site where the depth to groundwater is relatively shallow, approximately ⬍8 m. The distance to groundwater at the Becker site was relatively deep, 14.7⫾ 0.2 m. This resulted in a number of problems—共1兲 increased well installation costs, which caused fewer wells to be installed, 共2兲 mandated use of more expensive positive displacement pumps for water sampling and later water recirculation experiments, and 共3兲 the possibility of gas loss during recirculation. Therefore, site selection has important ramifications, particularly with respect to costs, the ability to place wells more closely, and the effectiveness of water recirculation. In addition, laboratory experiments performed with the aquifer material from the Becker site 共Schnobrich et al. 2007兲 indicated that phosphorous addition was required for complete denitrification. If a site is close to a phosphorous-limited water body, such as the Mississippi River in this case, the addition of this nutrient might not be feasible. The model developed in this study was able to predict the trends observed at the field site without fitting or adjusting parameters to match the field data. It was determined that understanding the exact flow direction of the groundwater was critical for observing plumes of treated groundwater downgradient because of the narrow zone of influence of the membranes. The model also was a useful tool for determining what changes could be made at the site that would decrease the NO−3 and NO−2 concentrations below their respective MCLs. Namely, increasing the membrane surface area and number of wells in the direction of flow and decreasing well spacing to three times the well bore diameter perpendicular to flow. The use of such a model is highly recommended for the application of in situ denitrification technologies such as this one.



Summary and Conclusions In this study, hollow-fiber membranes were used to deliver H2 in situ at a NO−3 -contaminated site in Becker, Minn., which was the first field-scale demonstration of this technology. The complete degradation of NO−3 was not achieved in the field because of a limited zone of influence. This small zone of influence is due to microbial consumption of H2 at the point of delivery, the limited transverse dispersion of H2 under groundwater flow conditions, and its low solubility. Attempts to increase the dispersion of H2 by groundwater recirculation were compounded by a slight reoxygenation of recirculated water, which increased the demand for H2. The use of in-well recirculation of groundwater accompanied by packers within the well bore may overcome these problems. The downgradient water quality at the site was not negatively impacted with respect to organic content by the stimulation of autotrophic denitrifiers. Although TOC levels increased in wells where H2 was delivered, elevated TOC levels were not observed downgradient. The ability of the aquifer sediment to filter biomass shows that hollow fiber membrane technology holds promise for aquifer restoration. However, possible biomass clogging of aquifer pores or well screens needs to be considered. A mathematical model 共Killingstad et al. 2002兲 was modified to account for H2 transfer through hollow-fiber membranes. The model was used to predict NO−3 , NO−2 , and DO concentrations at the field site. Model results indicated that in order for the complete removal of 6.0 mg/L of DO and conversion of 25.0 mg/L of NO−3 -N to N2, membrane well spacing of three times the well diameter 共normal to flow兲 and two deep 共parallel to flow兲 was required. In addition, for this technology to be used, appropriate field sites where the depth to groundwater is relatively shallow 共⬍8 m deep兲 need to be chosen.

Acknowledgments The project, Implementing Denitrification Strategies for Minnesota’s Contaminated Aquifers, was funded by Environment and Natural Resources Trust Fund as recommended by the Legislative Commission on Minnesota Resources. Additionally, gratitude is extended to Steve Blumm and Chuck Donkers at Xcel Energy for allotting land for the field site. Additionally, special thanks to MNDOT for the donation of the traffic cabinet for field supplies.

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