Stimulating In Situ Hydrogenotrophic Denitrification with Membrane-Delivered Hydrogen under Passive and Pumped Groundwater Conditions
Publication: Journal of Environmental Engineering
Volume 135, Issue 8
Abstract
A technology was developed to stimulate autotrophic biological denitrification by supplying hydrogen 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 and dissolved oxygen (DO) . Membranes installed in groundwater wells were successful in delivering to the groundwater over the two-year operating period. Hydrogen stimulated microbial reduction of DO and , degrading up to 6 mg/L DO and converting up to 10.0 mg/L to 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 was possible but the zone of influence and total demand were limiting factors. Water was recirculated in the field from downgradient to upgradient membrane-containing wells to increase the delivery through the membrane by an increase in water velocity. The depth to groundwater 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.
Introduction
Nitrate is a common pollutant in groundwater, often as a result of septic tank discharge and the overapplication of fertilizers. The consumption of -contaminated drinking water has been directly linked to methemoglobinemia (blue-baby syndrome) in infants (Comly 1945), and prompted the United States Environmental Protection Agency to enforce maximum contaminant levels (MCL) for and of 10.0 mg/L and 1.0 mg/L as N, respectively, (U.S. Environmental Protection Agency 1995). Studies estimate that approximately 5% (U.S. Environmental Protection Agency 1990) to 11% (Squillace et al. 2002) of wells in the United States contain concentrations that exceed the MCL.
Conventional methods for 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 removal (Lee and Rittmann 2002; Schnobrich et al. 2007) and is a promising solution that can convert and to nitrogen gas . It is also ideal for in situ applications where it can utilize indigenous microorganisms for reduction and aquifer material for subsequent filtration of biomass and organic by-products (Schnobrich et al. 2007).
A variety of organic electron donors have been used (e.g., ethanol, methanol, sucrose, acetate, and formate) that stimulate heterotrophic denitrifiers (Mercado et al. 1988; Janda et al. 1988; Smith et al. 2001). 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 (e.g., filtration, chlorination, ultraviolet radiation), which can substantially increase the cost of drinking water treatment and can contribute to the formation of toxic disinfection by-products during disinfection (Mercado et al. 1988). Inorganic electron donors such as hydrogen gas, reduced iron compounds, and reduced sulfur compounds can also be used to stimulate autotrophic denitrifiers (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) while producing less biomass (Rutten and Schnoor 1992).
The use of to stimulate autotrophic denitrification is desirable because it is nontoxic, produces low biomass, and its low water solubility (1.6 mg/L at ) results in little residual. When delivered to water by hollow fiber membranes, diffuses efficiently into the aqueous phase, preventing the bubbling of gas that can cause an explosive environment (Gantzer 1995). In addition, hollow fiber membranes can support biofilm growth, allowing direct contact with the electron donor and the electron acceptor (Lee and Rittmann 2002). Hollow fiber membrane reactors have been tested successfully in several lab-scale studies (Gantzer 1995; Lee and Rittmann 2002; Haugen et al. 2002; Chung et al. 2007; Schnobrich et al. 2007) 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 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.
Experimental Methods
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 (Xcel Energy 2000; Northern States Power (NSP) 1991). The primary groundwater flow direction was from the northeast to southwest and the main discharge area was the Mississippi River (located approximately 200 m downgradient of the field site) (Xcel Energy 2000). 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 (Northern States Power (NSP) 1991). The well casings consisted of 5.1-cm outside diameter (OD) polyvinyl chloride (PVC) 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 ( dissolved oxygen (DO)/L, average ± 95% confidence interval) and contained .
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 . This solution was immediately pumped into Well H6 (approximately 1 m below the water table) at a rate of 1 L/min using a peristaltic pump. A selective electrode (Cole Parmer, model #27502-04, Vernon Hills, IL) equipped with a data logger (WTW, pH340I, Germany) was positioned downgradient in Wells H2, H7, and M7 (approximately 1 m below the water table) to record hourly changes in the concentration of the water.
Membrane Module Design
Two membrane module designs were tested at the field site (Fig. 2). Both types of modules used 1-mm model # OD nonporous silicone-coated reinforced fiberglass membrane fibers (Varflex Corporation, #20RC1, Rome, N.Y.).
The first membrane module was designed to be operated in series [Fig. 2(a)] and each module consisted of 20 membrane fibers (3.0-m length). The top and bottom of the fibers were sealed (3M, Scotch Weld DP-125, St. Paul, Minn.) 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 flow but was plagued by water condensation within the membranes, which prevented gas flow (results not shown). To overcome these limitations, a second membrane design was developed.
The second type of membrane module was designed with a sealed end [Fig. 2(b)]. A single membrane fiber, 24.0 m in length, was coiled to form a 3-m-long module (approximately 1 loop/cm). The influent gas line was connected to one end of the membrane and the other end was attached to a moisture escape membrane apparatus (MEMA). The MEMA was made with six polyethylene microporus membranes (5.1 cm long, 0.30-mm OD, ID) (Mittsubishi Rayon, MHF 300, New York, N.Y.). 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 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 gas was supplied to the inside of the silicone-coated membranes (0–2.36 atm). 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.Table 1. Summary of Operating Parameters for Field Experiments
Passive denitrificaton experiments | Groundwater recirculation experiment | |||||||
---|---|---|---|---|---|---|---|---|
Experimentnumber | Duration(days) | Membranesurface area | Hydrogendelivery wells | Hydrogen pressure(atm) | Days ofoperation | Extraction well; flow rate (L/min) | Injectionwells | Hydrogen pressure(atm) |
1 | 111 | 0.12 | H1, H2, H3, H4 | 1.68 (days 0–82),2.36 (days 83–96),0 (days 97–11) | 0–62a | M7; 4.0 | H1, H3, H4,H6 | 1.68 |
2 | 42 | 0.12 | H1, H2, H3, H4,H6, H7, H8 | 1.68 (days 0–27),2.36 (days 27–42) | 63–91 | M7; 4.0 | H3, H4, H6 | 1.68 |
a
On Day 58 a flow distributor was added to each injection well.
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 (Sta-Rite Industries, Delavan, Wis.) 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 . 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 (1/2 in.) 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 distribute pumped water radially outward through the membrane and thus promote a more uniform delivery to the surrounding aquifer [Fig. 2(c)].
Groundwater Sampling
Wells were sampled approximately every one to two weeks for , , DO, pH, temperature, total organic carbon (TOC), alkalinity, and water level. Samples were collected using a submersible 4.6-cm diameter pump (Grundfos, Redi-flow2 pump, Fresno, Calif.) into 250-mL Teflon-lined polyurethane bottles (Nalgene). Dissolved was not quantified at the site because it was determined that the submersible Grundfos pump produced by electrolysis during sampling; this has been observed by others using similar submersible pumps (Chapelle et al. 1997).
and Microcosm Experiments
Microcosm experiments were conducted in the laboratory to determine kinetic parameters for the reduction of and 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 prior to use.
Microcosms were prepared in triplicate in 715-mL glass bottles (Wheaton, Millville, N.J.) containing 30 g wet sediment, 297-mL synthetic groundwater [per 1 L Milli-Q (Millipore, Bedford, Mass.) water]: 23 mg , 167-mg , 247-mg , 13-mg ), 3 mL of a trace element stock solution (per 1 L Milli-Q water: 8.5-mg , 13.4-mg , and 5-mg ), and varying amounts of or (1–18 mg/L as N). The liquid in all microcosms was sparged with ultrahigh purity for approximately 2 min to remove DO. The headspace was flushed with 10% gas for approximately 30 min and the electron donor-free biological controls were flushed with 100% . The bottles were sealed with screw-on caps fitted with butyl rubber septa for sample collection (Wheaton, Millville, N.J.). Sterile controls were prepared by autoclaving microcosms containing a 10% headspace for 30 min. The microcosms were placed on their sides on a shaker table (New Brunswick Scientific, Model R-2, Edison, N.J.) and shaken at 250 rpm at for the duration of the experiment. A 3.5-mL liquid sample was withdrawn to quantify and concentrations every 1–3 days.
Analytical Methods
The DO concentrations were measured in the field using 1–12 mg/L Chemet ampules (Chemetrics Inc., Calverton, Va.). During the groundwater recirculation experiment, DO was measured with a DO probe (MI-730 Dip-type microelectrode) and meter (OM-4, oxygen meter) from Microelectrodes Inc. (Bedford, N.H.). Additionally, a YSI Model 58 (YSI Inc., Yellow Springs, Ohio) 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 (Oakton, model-pH 6) and electrode (Oakton, model-WD-35801-00, Vernon Hills, Ill.). Water levels were measured using an electronic water level finder (Slope Indicator Co., Model 51453, Seattle, Wash.) before pumping the groundwater. Quantification of alkalinity (mg/L as ) was determined by Standard Method 2320 B (American Public Health Association (APHA) et al. 1998). Nitrate and nitrite concentrations were determined by ion chromatography [Metrohm 761 Compact Ion Chromatograph (Herisau, Switzerland); Metrosep A Supp 5 column; 3.2-mM and 1.0-mM eluent; 0.7 ml/min eluent flow rate; 100 mM regenerant]. The method detection limits (MDLs) for and were 0.035 mg/L and 0.033 mg/L , respectively, as determined by Standard Method 1030 C (American Public Health Association (APHA) et al. 1998). The TOC was analyzed in triplicate using a Dohrmann Phoenix 8000 TOC analyzer equipped with a STS 8000 autosampler (Dohrmann, Mason, Ohio). The MDL for TOC was 0.12 mg/L as carbon.
Mathematical Model
Conceptual Model
The mathematical model used in this study was an extension of an existing two-dimensional (2D) model that incorporated advective-dispersive transport and biodegradation of one electron donor (formate) and two electron acceptors ( and ) coupled with the growth and decay of a single microbial population (Killingstad et al. 2002). The Killingstad et al. (2002) 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 through hollow-fiber membranes, according to the gas transfer correlation developed by Fang et al. (2002). The model was also modified to model three electron acceptors ( , , and ). The microbial population was a single facultative population attached to sediment particles. The population occupied no physical space within the model domain.
Transport Equations
The general form of the 2D equations of mass balance for electron donor and acceptor transport and biodegradation arerespectively, where phase concentration of hydrogen ; acceptor concentration ; and are distances along and across the direction of flow, respectively (L); groundwater velocity (L/T); and are the dispersion coefficients in the - and -directions, respectively ; represents the mass loss of due to biological utilization ; accounts for the mass flux of from the membranes ; represents the mass loss of electron acceptor due to biological utilization ; and accounts for the production of , according to the stoichiometry in Eq. (9).
(1a)
(1b)
Source Term
The source term for the mass transfer of to the aqueous phase is expressed as the mass transfer rate per unit volume of liquidwhere transfer rate of out of the membrane (M/T); of the mass transfer zone ; and volume of the mass transfer zone (i.e., one model cell). The mass transfer rate under steady state conditions is expressed aswhere mass transfer coefficient (L/T); area of the membrane ; weight of (M/mole); and liquid saturation concentration , which is related to the partial pressure by the Henry’s Law constant. The value for can be predicted using the following dimensionless relationship (Fang et al. 2002):where of in water ; membrane diameter (L); velocity across the membranes (L/T); and viscosity of water .
(2)
(3)
(4)
Reaction Terms
The reaction term for utilization , expressed as the sum of utilization as a result of DO, , and reduction, iswhere biomass per volume of porous medium ; porosity; and , , and are the utilization rates by the autotrophic biomass for DO, , and reduction, respectively (M/M per T). The DO, , and mass loss rates , resulting from their utilization as an electron acceptor, are linked to utilization by the stoichiometry of the biological reductionwhere represents the mass of electron acceptor reduced per unit mass of hydrogen (M/M) and represents either DO, , or . Utilization rates in Eqs. (5) - (6) are represented by dual Monod kinetics of the electron donor and acceptor. Switching functions prevented the uptake of and until DO was below a set level ( and , respectively), and prevented the degradation of until was below a set level (Killingstad et al. 2002).
(5)
(6)
Microbial Densities
Changes in the microbial densities are simulated by a mass balance equation that accounts for biomass growth and decay termswhere , , and are the yield coefficients (M/M), the ratio of microbial biomass produced per mass of electron donor consumed for DO, , and respiration, respectively, and is the biomass decay coefficient (M/M per T). Eq. (7) is solved directly and is updated at each time step.
(7)
Model Setup
A 2D finite-difference grid was used with a 0.0508-m node spacing. Initial conditions for concentrations of , , and DO in the model were determined by sampling data collected just prior to addition and were assumed to be uniform throughout the model domain. The initial condition for concentration was assumed to be zero throughout the model domain.
Concentrations of , , DO, and 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 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 addition via membrane modules.
Parameter Estimation
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.Table 2. Values for Model Input in Field Site Simulation
Variables initial condition | Value | Units | Source |
---|---|---|---|
6.00 | Sampling data | ||
24.10 | Sampling data | ||
0.00 | Sampling data | ||
— | Fang et al. (2002) | ||
0.22 | Microcosm experiment | ||
Parameters | Value | Units | Source |
0.31 | m/day | Tracer test | |
0.02 | m | Tracer test | |
0.002 | m | Killingstad et al. (2002) | |
0.30 | Experimental | ||
0.90 | Calculated | ||
7.62 | g O/g H | McCarty (1972) | |
6.72 | g N1/g H | McCarty (1972) | |
4.45 | g N2/g H | McCarty (1972) | |
0.19 | g bio/g H | McCarty (1972) | |
0.18 | g bio/g H | McCarty (1972) | |
0.18 | g bio/g H | McCarty (1972) | |
0.05 | McCarty (1972) | ||
3.50 | g H/g bio/day | Estimated | |
1.37 | g H/g bio/day | Microcosm experiment | |
0.19 | g H/g bio/day | Microcosm experiment | |
0.02 | Average of values reported from Killingstad et al. (2002) and Kurt et al. (1987) | ||
0.02 | Average of values reported from Killingstad et al. (2002) and Kurt et al. (1987) | ||
0.02 | Average of values reported from Killingstad et al. (2002) and Kurt et al. (1987) | ||
2.00 | Interstate Technology and Regulatory Cooperation Work Group (ITCR) (2000) | ||
2.00 | Interstate Technology and Regulatory Cooperation Work Group (ITCR) (2000) | ||
Experimental |
Transport
Groundwater velocity, direction of flow, and the longitudinal dispersion coefficient were obtained from the analysis of the Br tracer study (Table 2). The values for and were obtained by analyzing the Br data between Wells H2 and M7 using the method of moments (Valocchi 1989). The model was aligned to the direction of flow. A value for was obtained by assuming that (Killingstad et al. 2002). Numerical dispersion was limited by requiring the Courant number to approach unity. Aquifer porosity and sediment density were set at constant values of 0.30 and , respectively.
Stoichiometric Coefficients
The electron acceptor utilization coefficients ( , , and ) and cell yields ( , , and ) for DO, , and reduction, respectively, were derived from McCarty (1975). A cell residence time of 1,000 days was used to account for stationary biomass and a decay coefficient of was assumed. The following stoichiometric relationships were calculated according to the methods presented by McCarty (1975):where represents biomass.
(8)
(9)
(10)
Utilization Rates and Initial Biomass Concentration
The and microcosm experiments were used to obtain a value for and the Monod parameters , and . The average value for was converted to a representative site-wide value based on density and porosity at the field site.
Results and Discussion
Background Water Quality
Average values for each water quality parameter measured from all wells at the site before addition are shown in Table 3. The average and DO concentration before addition was as N and , respectively (± 95% confidence interval). Nitrite was never detected in the field before the addition. The pH, temperature, TOC, and alkalinity were , , as C, and as , respectively.Table 3. Background Water Quality Parameters and Values
Parameter | Units | Average | 95% confidenceinterval | Number ofsamples |
---|---|---|---|---|
mg/L | 21.2 | 8.1 | 97 | |
mg/L | 0 | 0 | 97 | |
DO | mg/L | 7.0 | 2 | 51 |
pH | pH units | 7.2 | 0.4 | 84 |
Temperature | °C | 11.4 | 2.4 | 66 |
TOC | mg/L as C | 0.6 | 0.4 | 59 |
Alkalinity | mg/L as CaCO3 | 182 | 14 | 25 |
Results from Passive Denitrification Experiments
Experiment 1
The membranes in Wells H1, H2, H3, and H4 were pressurized with 1.68 atm on Day 0. The DO, , and concentrations 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 (6 mg/L) by Day 111. The increase in the lumen pressure to 2.36 atm on Day 83 did not appear to affect the DO level in Well M5.
The concentration of at Well M5 was approximately 25 mg-N/L at Day 0 and remained relatively constant until Day 10 (Fig. 3). The 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 (Fig. 3).
Nitrite was observed in Wells H5, M5, H7, and M8, all downgradient of -addition wells; no was detected upgradient of the -addition wells. The largest concentration of was detected in Well M5. The concentration profile in Well M5 appeared to mirror that of in the same well (Fig. 3), suggesting that most of the was reduced to . 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 (e.g., as C in Well M2 and as C in Well M5). A TOC sample taken on Day 60 in Well H3 where 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 addition.
Experiment 2
A replicate -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 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 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 observed in Wells H5, M5, M7, and M8 throughout Experiment 2 were 0.55, 2.81, 0.34, and 2.17 mg- , respectively. On Day 10, 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 experiment. As in Experiment 1, and profiles approximately mirrored one another. In addition, the total nitrogen concentrations in Wells M5 (downgradient) and M2 (upgradient) were and , respectively, during the experiment. This observation indicates that was reduced to without complete denitrification to . On Day 28, the pressure in the membranes was increased to 2.36 atm but no changes in , , and DO concentrations were observed (Fig. 4).
At the end of the experiment, the membrane modules were removed and the membrane-containing wells (H2, H3, H4, H7, H8, and H6) were immediately sampled. The average concentrations of , , and DO were , , and , respectively. Well H3 showed the highest levels of denitrification with , , 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 (0.4 mg/L as C), M5 (0.6 mg/L as C), and M8 (0.5 mg/L as C) were similar. Nevertheless, the TOC values in the membrane-containing wells on Days 18 and 42 were higher ( as C in Well H3 and as C in other membrane-containing wells). 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 , 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 . The sequential reduction of oxygen followed by was confirmed in the Becker aquifer material by Schnobrich et al. (2007) and by the microcosm studies.
The greatest quantity of denitrification was observed in Well M5 during Experiment 1 in which concentrations of DO and were 2.0 mg/L and 8.0 mg/L-N, respectively. According to the stoichiometry shown in Eqs. (8) - (10), approximately 1.74 mg/L is needed for this observed reduction. This quantity of is comparable to the solubility of at (1.73 mg/L) and the concentration of that is predicted by the clean water gas transfer correlation developed by Fang et al. (2002). 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 (Semmens and Essila 2001; Haugen et al. 2002; Edstrom et al. 2005; Schnobrich et al. 2007). The needed to reduce DO and and below their respective MCLs at the site is approximately 6.6 mg/L , only approximately four times higher than the clean water gas transfer correlation. The residence time within the -delivery wells ( , depending on the well capture zone), should be sufficient for microbial consumption to increase the overall mass transfer out of the membrane. In fact, laboratory results from Schnobrich et al. (2007), using aquifer material from the Becker site with a nearly identical flow velocity, slightly shorter residence time (1.45 h), and slightly more (3.5 to 10 times greater) membrane surface area per groundwater flow rate, demonstrated that it was possible to add sufficient 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. (2007). We would therefore expect enhanced transfer from the membrane to the surrounding water in the well, resulting in an increased extent of denitrification in the well itself (Fang et al. 2002). Nevertheless, the lack of more significant denitrification within the membrane wells themselves (H2, H3, H4, H7, H8, and H6), indicated that other factors at the site, such as low phosphorus concentrations (Schnobrich et al. 2007), may have contributed to incomplete denitrification.
Because the zone of influence of each membrane-containing well was relatively narrow (Haugen et al. 2002; Clapp et al. 2004), lower levels of denitrification would be observed downgradient as treated water mixed with untreated water from the surrounding area, resulting in higher observed 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 and 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 or occurred in either the sterile or headspace controls in any of the experiments.
The microcosm experiments yielded an average value for of and values for and of and biomass/day, respectively. Results indicated that and reduction were following zero order kinetics at field concentrations. This finding is supported by the literature Kurt et al. (1987). In the microcosm experiments, was converted to but no was reduced until was completely degraded. For this reason, the value of was set to .
Field Experiments
A site-wide model simulation was conducted to illustrate the overall effectiveness of the passive remediation strategy. A two-dimensional plan view of steady state concentrations of downgradient of the addition wells is shown in Fig. 5. The model predicts that denitrification occurred in narrow bands of approximately 0.15 m ( times the well diameter). The model also showed that concentrations should be detected at Wells M7, M8, H5, and M5, with the largest concentrations detected at Well M5. This result is consistent with the field sampling data from both Experiments 1 and 2 (Figs. 3 and 4). In addition, if denitrification occurred over a large transverse direction, depleted DO or accumulation of should have been observed in Wells M6 and M11 but this was not the case.
Simulated and observed and concentration data are shown in Fig. 6(a) for the first 80 days of Experiment 1. The and data were normalized to total nitrogen to account for variable concentrations observed in Well M5 (Fig. 6). For the first 48 days of operation, the model simulated the changes in concentrations of and 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 (Fig. 5). Model results observed at Well M5 after a 5° shift in groundwater flow direction are shown in Fig. 6(b). As seen, a small change in the groundwater flow direction could have a significant impact on our ability to observe the effect of addition. Model simulations for DO, , and at Well M5 for Experiment 2 based on the flow direction in Fig. 6(b) are shown in Fig. 4 with the collected field data.
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 and below the regulatory MCLs.
According to Eq. (3), the mass transfer rate of can be controlled by the surface area of the membrane and the partial pressure in the membrane. Using experimental results for membrane spacing (Fang et al. 2002), the membrane fiber can be coiled 600 coils/m to produce a total membrane surface area of per well, which yielded a membrane surface area to fluid volume ratio of . Pressurizing these membranes with 2.36 atm of would result in DO of 1.0 mg/L within 6 days, from 25.0 mg-N/L to less than 5.0 mg-N/L within 35 days, and of 10 mg-N/L within 60 days. Although a marked improvement in performance, this is still inadequate for meeting the 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 and 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: (1) two membrane-containing wells along the direction of flow with a specific membrane surface area of ; (2) a well spacing of three times the well diameter perpendicular to flow; (3) a hydrogen pressure of 2.68 atm in the membranes; and (4) approximately 90 days for biomass growth.
Recirculation Experiment
Although the membranes installed at the site were successful in delivering to the groundwater during Experiments 1 and 2, the zone of influence was limited and the overall mass of delivered was insufficient to stimulate complete denitrification; this was confirmed by the modeling simulations. To address these problems, water recirculation was initiated to increase gas transfer by increasing the velocity of the water flowing past the membranes [see Eq. (4)].
During the 91 days of pumping operation, , , and DO concentrations fluctuated, as shown in Fig. 7. Total nitrogen concentrations between the upgradient Well M2 and Well M5 were not significantly different. Nitrite accumulation reached a maximum concentration of approximately 8.0 mg 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 (Well H1), 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 entrainment had ceased. The reduction of 4.0 mg to 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 DO- and -rich groundwater with treated water prevented consistent observation of reduction at, and downstream of, membrane-containing wells H6, H3, and H4. As a result, the required for reducing and at the site increased.
The membrane modules were removed on two occasions, Days 58 and 91. On Day 58 approximately 5.5 mg and 1.1 mg/L DO were observed in Well H3. On Day 91, approximately 8.5 mg was observed in Well H4. These results confirm that membrane delivery of was still able to reduce DO and in these wells despite the reintroduction of and -rich groundwater.
Agarwal et al. (2005) tested the effects of pumping on membrane gas 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 contour was 7.3 times the diameter of the membrane module well (Agarwal et al. 2005). 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 (Agarwal et al. 2005). 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
The field site contained background and DO levels that were extremely high; this presented a worst-case scenario in which to challenge our technology. Delivering by discreet membrane-containing wells produced only a small zone of influence. Therefore, closer well spacing ( times the well bore diameter), 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 . The distance to groundwater at the Becker site was relatively deep, . This resulted in a number of problems—(1) increased well installation costs, which caused fewer wells to be installed, (2) mandated use of more expensive positive displacement pumps for water sampling and later water recirculation experiments, and (3) 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 (Schnobrich et al. 2007) 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 and 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 in situ at a -contaminated site in Becker, Minn., which was the first field-scale demonstration of this technology. The complete degradation of was not achieved in the field because of a limited zone of influence. This small zone of influence is due to microbial consumption of at the point of delivery, the limited transverse dispersion of under groundwater flow conditions, and its low solubility. Attempts to increase the dispersion of by groundwater recirculation were compounded by a slight reoxygenation of recirculated water, which increased the demand for . 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 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 (Killingstad et al. 2002) was modified to account for transfer through hollow-fiber membranes. The model was used to predict , , 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 to , membrane well spacing of three times the well diameter (normal to flow) and two deep (parallel to flow) was required. In addition, for this technology to be used, appropriate field sites where the depth to groundwater is relatively shallow ( deep) 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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© 2009 ASCE.
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Received: Dec 14, 2007
Accepted: Nov 11, 2008
Published online: Apr 3, 2009
Published in print: Aug 2009
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