| HOME | HELP | FEEDBACK | SUBSCRIPTIONS | ARCHIVE | SEARCH | TABLE OF CONTENTS |
|
|
|
|||||||
|
|||||||||
School of Natural Resource Sciences, University of Nebraska-Lincoln, Lincoln, NE 68583-0915
* Corresponding author (scomfort{at}unl.edu)
Received for publication December 1, 2000.
 |
ABSTRACT |
|---|
 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
|---|
Abbreviations: HPLC, high-performance liquid chromatography
|
INTRODUCTION |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
|---|
The realization that many pollutants normally considered persistent in aerobic environments may be less persistent under anaerobic conditions has generated considerable interest in engineering a reducing environment in soils, sediments, and aquifers for remediation purposes. Under reducing conditions, detoxification of many contaminants can occur through reductive dehalogenation reactions. Although there are exceptions, as when vinyl chloride, a human carcinogen, is produced from reduction of more highly chlorinated compounds (Suflita et al., 1982), there is ample evidence to indicate that reducing or removing electron-withdrawing moieties from parent structures can result in more biodegradable products (Hundal et al., 1997; Singh et al., 1998a; Fathepure and Tiedje, 1999). Based on this premise, one technology gaining widespread acceptance is the use of zerovalent metals for remediating ground water contaminated with chlorinated solvents. Zerovalent iron (Fe0, Eoh = -409 mV; Weast, 1978) is an avid electron donor and has a high capacity to reduce a wide array of organic compounds. Under aerobic conditions, oxygen is the usual electron acceptor, while in anoxic environments, electron release from the reaction of Fe0 with water can be coupled to the reduction of chlorinated and nitroaromatic compounds (Gillham and O'Hannesin, 1994; Agrawal and Tratnyek, 1996).
Current research has focused on the use of zerovalent metals for in situ treatment of contaminated ground water, with less research on using zerovalent metals for treating contaminated soils. Earlier work with zerovalent zinc demonstrated the utility of metals to treat soils contaminated with dichlorodiphenyltrichloroethane (DDT) (Staiff et al., 1977), methyl parathion (Butler et al., 1981), and polychlorinated biphenyls (Cutshall et al., 1993). More recent research indicates the tremendous potential of Fe0 to remediate soils contaminated with 2,4,6-trinitrotoluene (TNT), hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) (Hundal et al., 1997; Singh et al., 1998a, 1999), and atrazine (Singh et al., 1998b). Despite encouraging results, this work has yet to evolve into routine large-scale treatment of contaminated spill sites.
Our objective was to use zerovalent iron to remediate a metolachlor-contaminated field site. We report laboratory observations demonstrating the capacity of Fe0 to dechlorinate metolachlor in aqueous solution, then demonstrate the effectiveness of Fe0 to decrease metolachlor concentrations in static soil microcosms and at the field scale in soil windrows.
|
MATERIALS AND METHODS |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
|---|
Laboratory Solution Experiments
Aqueous solutions of metolachlor were prepared from the commercial product Dual 8E (Syngenta, Greensboro, NC) and spiked with [14C-U-phenyl] metolachlor (28.6 µCi mg-1) to produce a final concentration of 32 Bq mL-1. Experimental units consisted of 250-mL Erlenmeyer flasks filled with 150 mL of metolachlor solution. All flasks were covered with Parafilm-coated stoppers and agitated on a reciprocating shaker at ambient temperature (23°C).
Two sets of companion batch experiments were conducted with zerovalent iron. The first set compared the destructive capacity of Fe0 with and without Al2(SO4)3; the second used Fe0 with and without acetic acid. Initial metolachlor concentration was 1.09 mM for the Fe0–Al2(SO4)3 experiment and 1.00 mM for the Fe0–acetic acid experiment. Each treatment (including controls) was replicated three times. Each treatment flask (containing 150 mL of metolachlor solution) received 18.75 g of unannealed Fe0, with and without 0.75 g commercial-grade Al2(SO4)3 or 0.75 mL glacial acetic acid. At 0, 2, 4, 6, 8, 12, 16, 24, and 48 h, 1.5-mL aliquots were removed and transferred to 1.7-mL polypropylene microcentrifuge tubes, centrifuged at 13000 x g for 10 min, and analyzed by high-performance liquid chromatography (HPLC) for metolachlor and dechlorinated metolachlor. Carbon-14 in the aqueous solution was also monitored using the same sampling scheme by mixing 1 mL of uncentrifuged sample with 6 mL of Ultima Gold cocktail (Packard, Meriden, CT) and analyzed by liquid scintillation counting (LSC) using a Packard 1900TR liquid scintillation counter (Packard Instrument Co., Downers Grove, IL). Chloride analysis was also conducted on the samples taken for HPLC analysis with a Dionex (Sunnyvale, CA) DX-120X ion chromatograph using an AS14 IonPac column and a sodium carbonate (3.5 mM)–sodium bicarbonate (1.0 mM) eluent at a flow rate of 1.2 mL min-1.
To determine differences in biodegradability between metolachlor and the Fe0–treated metolachlor products (primarily dechlorinated metolachlor), 25 mL of 0.35 mM metolachlor solution (spiked with 14C-metolachlor) was treated with and without 4.5 g Fe0 and mixed with 24 mL of double strength Pseudomonas minimum media (without glucose; Hundal et al., 1997). Solutions were transferred to sterilized glass jars, then inoculated with 1 mL of a consortium obtained from the metolachlor-contaminated soil. This inoculum was extracted by shaking 4 g soil with 20 mL of saline solution for 48 h. The saline solution contained (g L-1) NaCl (8.5), KH2PO4 (0.3), Na2HPO4 (0.6), and peptone (0.1). Evolved 14CO2 was captured by placing 10 mL of 0.5 M NaOH in a 20-mL scintillation vial and suspending it inside the sealed glass jar and over the inoculated media. Carbon dioxide traps were changed approximately every 5 d. Captured 14CO2 was determined by removing 0.5 mL from the CO2 traps, mixing with scintillation cocktail, and determining total 14C activity by liquid scintillation counting.
Laboratory Soil Incubation Experiments
Because solution experiments provided evidence that small additions of CH3COOH or Al2(SO4)3 facilitated Fe0–mediated destruction of metolachlor, we subsequently conducted batch studies with the metolachlor-contaminated soil to determine optimum concentrations needed for efficient metolachlor destruction in static soil microcosms. This was accomplished by incubating 20 g (oven dry) soil with 5% Fe0 (w/w) in 40-mL Teflon centrifuge tubes at 30°C and a soil water content of 0.40 kg kg-1. Aluminum sulfate additions were between 0 and 5% (w/w) [0–1.0 g Al2(SO4)3 to 20 g soil]. Acetic acid additions were between 0 and 2.5% (0–0.5 mL to 20 g soil). Soil treated with Fe0 + Al2(SO4)3 was incubated for 2 d; soil treated with Fe0 + CH3COOH was incubated for 7 d. Following incubation, the soil was extracted with 20 mL of CH3CN and analyzed by HPLC (procedure described in Soil Analysis, below).
Field Experiment
The stockpiled soil was moved with a front-end loader into seven windrows, five of which were used in the field experiment. Windrow dimensions were 3.35 to 3.65 m wide by 1.07 to 1.37 m high and 25.91 to 60.66 m in length. The soil in each windrow was mixed three times within 24 h by using a tractor-pulled, high-speed soil mixing and fractionation implement (Frontier Industrial Corp., Salem, OR), sold under the trade name Microenfractionator (H&H Eco Systems, North Bonneville, WA). This implement is similar in appearance to a conventional composter but differs in that its components have been augmented and redesigned to handle windrows containing 100% soil. This is facilitated by a John Deere (Moline, IL) 6068T 170-horsepower diesel engine that propels a large 32-cm (diam.) stainless steel rotating drum with 50 fan-knife blades (30.8-cm length). This implement also allows simultaneous injection of liquids (i.e., water and CH3COOH in this experiment) into the mixing tunnel via pressurized lines connected to a water tank, which is pulled along with the Microenfractionator (Fig. 1).
|
Once the initial samples were obtained, five treatments were initiated, one per windrow. Treatments included: (i) control (no Fe0), (ii) Fe0–only, (iii) Fe0 + CH3COOH, (iv) Fe0 + Al2(SO4)3, and (v) Fe0 + CH3COOH + Al2(SO4)3. Treatments were added as a percentage of the oven-dry soil mass, which was estimated by multiplying the volume of each windrow times a soil bulk density of 1.4 g cm-3. The Fe0 was added at 5% (w/w), CH3COOH at 0.5% (v/w), and commercial grade Al2(SO4)3 at 2% (w/w). Expressing these percentages in the more conventional unit of mass of chemical added per cubic yard (1 yd3 = 0.765 m3) of soil (1070 kg or 2360 lb, assuming a 1.4 g cm-3 bulk density), our treatment rates were equivalent to 53.5 kg (118 lb) of Fe0, 21.4 kg (47.1 lb) of Al2(SO4)3, and 5.4 L (1.42 gal) of acetic acid per 0.765 m3 (yd3) of contaminated soil.
The Fe0 was unannealed iron purchased from Peerless Metal Powders (Detroit, MI) in 22.7 kg unlined paper bags. This Fe0 had a specific surface area of 2.55 m2 g-1 (Micromeritics, Norcross, GA). Aluminum sulfate (commercial grade) was purchased in 22.7-kg bags from Van Waters & Rogers (Omaha, NE). Glacial acetic acid was obtained in 208-L barrels from Celanese Chemicals (Dallas, TX), diluted with H2O and added during the mixing operations.
The required amounts of iron and Al2(SO4)3 (in 22.7-kg unlined paper bags) were placed on top of the windrows and directly mixed in with the Microenfractionator a minimum of three times. Water and acetic acid were also added during the mixing process until the soil gravimetric water content was between 0.35 and 0.40 kg kg-1, which was determined at the field site by weight loss following repeated cycles of heating in a microwave oven. An added benefit of adding H2O during mixing was that it greatly reduced the amount of Fe0 dust released into the atmosphere during the first pass.
Once the desired soil water content was obtained, the windrows were covered with clear plastic that was held in place with wooden pallets. Windrows were sampled as described above at 1, 14, 28, 42, 60, 75, and 90 d. After sampling at 14, 28, 42, 60, and 75 d, additional water was applied to the top of the windrows (without mixing) to rewet the surface soil. Two sets of 90-d samples were taken, one before and one after mixing the windrows.
Soil Analysis
Soil samples obtained from the windrows were inventoried and stored at 4°C. For each sample, we determined soil water content, pH, and metolachlor concentration. Soil water content was determined on three 10-g subsamples by determining weight loss after drying in a microwave oven. Soil pH was determined on 20-g soil samples (oven-dry basis) using a 1:1 soil to H2O ratio.
Metolachlor was extracted from 4.75 to 5.25 g soil (oven-dry basis) in a 40-mL Teflon centrifuge tube by adding 20 mL CH3CN and shaking overnight (
8 h) on a reciprocating shaker at ambient temperature. The tubes were then centrifuged at 5000 x g for 10 min and 1.5 mL was removed and microcentrifuged at 13000 x g for an additional 10 min. After centrifuging, 1 mL of supernatant was stored in a glass HPLC vial at 4°C until analysis.
Metolachlor and its dechlorinated product were measured by HPLC by injecting 20 µL of the CH3CN extract into a 4.6- by 250-mm Keystone Betasil NA column (Keystone Scientific, Bellefonte, PA) connected to a Shimadzu (Kyoto, Japan) UV detector or photodiode array detector. The mobile phase was 50:50 acetonitrile and water at 1.0 mL min-1 with quantification at 220 nm. Under these conditions, typical retention times were 12 min for metolachlor and 8 min for the dechlorinated metolachlor.
Standard soil nutrient and metal analyses (Table 1) were conducted by Midwest Analytical Laboratories (Omaha, NE) on initial (t = 0 d) and t = 90 d (mixed) samples.
|
= 0.05). |
RESULTS AND DISCUSSION |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
|---|
|
|
Our results are consistent with previous reports of metolachlor dechlorination by iron metal in water (Eykholt and Davenport, 1998). Mass balance experiments using 14C-metolachlor, with measurement of the dechlorinated compound and recovery of Cl- released from metolachlor, indicated that the conversion is essentially complete, without the initial formation of other products.
Because dechlorination was the main destructive mechanism of the Fe0 treatment, we compared the biodegradability of metolachlor with the Fe0–treated metolachlor products in a mixed culture. The Fe0–treated products were five times more utilizable as a sole carbon source than metolachlor (cumulative 14CO2: 4.6 vs. 0.89%, Fig. 4), indicating that they are more biodegradable than the parent metolachlor. Singh et al. (1998b) similarly found in a 120-d study that adding Fe0 to atrazine-contaminated soil increased cumulative mineralization from 4.1 to 11.2%.
|
|
|
With time, diffusion of oxygen into the windrow was inevitable and the brown oxidation layer at the surface gradually increased in size. Whereas the oxidation of green rust to goethite may also provide a source of electrons for reduction reactions, formation of a Fe(III) oxide layer passivates the iron surface. At the 14-d sampling, we took a vertical core in the Fe0 + CH3COOH + Al2(SO4)3–treated windrow and segmented it into three 10-cm increments. The top (surface) segment, which was brown, had a metolachlor concentration of 34 mg kg-1 while the middle (brown-green) and bottom (green) sections (10–20, 20–30 cm) had concentrations 
9 mg kg-1. When the windrows were finally mixed at the end of the experiment (t = 90 d), the center and bottom of the windrows still exhibited a green rust color. Based on these observations, it is likely that anoxic conditions were maintained in the center and bottom of the windrows and by mixing the windrows, we exposed soil that had greater rates of metolachlor destruction to areas where the windrows were routinely sampled. This probably explains why metolachlor concentrations were lower after the post–90 d mixing.
Comparisons among treatments at each sampling date revealed that the Fe0–only treatment significantly decreased metolachlor concentrations from the control (Table 2). Adding Al2(SO4)3, acetic acid, or both with Fe0 significantly increased metolachlor destruction. This was particularly apparent in treatments containing Al2(SO4)3, which yielded the lowest metolachlor concentrations. Reasons for enhanced destruction by the addition of Al2(SO4)3 and acetic acid are complex but can in part be attributed to alterations in soil solution pH, redox potential, and soil solution composition during corrosion of the Fe0 surface. Decreasing pH will increase destruction rates by providing protons for reductive transformations and slow down Fe(II) oxidation and passivation of the iron surface. Although the Al2(SO4)3 amendment significantly decreased soil pH in the windrows, the acetic acid treatment only slightly lowered the pH after application and residual effects were relatively short-lived (Table 3). After the first few samplings, pH tended to be higher in windrows receiving acetic acid than the comparable treatment [Fe0 or Fe0 + Al2(SO4)3; Table 3].
|
Aside from lowering the pH, it is likely that the Al2(SO4)3 played other roles in enhancing metolachlor destruction. An abundance of aluminum during Fe0 oxidation promotes its incorporation into the oxidized iron structure (Schwertmann and Cornell, 1991). This is important because Al3+ has a smaller ionic radius that disrupts crystallization and favors formation of ferrihydrite (Fe5HO8) (Stucki et al., 1988). Ferrihydrite has a large surface area (>200 m2 g-1; Schwertmann and Cornell, 1991) and can serve as a reservoir for reduced iron (Baltpurvins et al., 1996). Klausen et al. (1995) demonstrated that Fe(II) bound to iron hydroxide surfaces or surface coatings plays an important role in reductive transformation of nitroaromatic compounds. Experiments conducted in our laboratory indicated that the presence of Fe(II), Fe(III), or Al(III) during Fe0 corrosion can facilitate metolachlor destruction in aqueous solution. In these experiments, Al(III) was readily sorbed by, and/or incorporated into, the oxidizing iron and this corresponded with a release of Fe(II) into solution. Therefore, an indirect effect of adding Al2(SO4)3 was likely its influence on Fe(II) concentration in the soil solution during the corrosion of Fe0. Because sulfate was added as part of the Al2(SO4)3, its presence must also be considered. Sulfate has been shown to sustain higher rates of iron corrosion (Reardon, 1995), apparently by dissolving the oxide film that coats the surface during oxidation (Gu et al., 1999). Under reducing conditions, sulfate also promotes formation of green rust II [FeII4FeIII2(OH)12SO4 · nH2O] (Refait et al., 1999).
Changes in Soil Chemical Properties
Initial analysis of the contaminated soil indicated that, in addition to metolachlor, this soil had very high concentrations of NO-3, NH+4, and P as well as diethylenetriaminepentaacetic acid (DTPA)–extractable metals (Zn, Mn, Fe, Cu) (Table 1). Following 90 d of treatment, some notable increases and decreases in soil chemical properties were observed (Table 1). Phosphorus concentrations decreased in windrows receiving Al2(SO4)3. Aluminum sulfate has been traditionally used to remove phosphate from wastewater streams by forming AlPO4 (Snoeyink and Jenkins, 1980). Hsu (1976) indicated the optimum pH for P removal by Al was 5.5 to 8.0, well within the range observed in the soil windrows (Table 3). Under acidic conditions (pH < 6), AlPO4 predominates, whereas at pH 6 to 8, an Al(OH)3 forms, which removes P from solution by sorption of inorganic phosphate and entrapment of organic particles containing P (Cooke et al., 1986).
Nitrate concentrations were also reduced substantially by all Fe0 treatments, corroborating previous research showing transformation of NO-3 to NH+4 upon addition of Fe0 (Cheng et al., 1997; Till et al., 1998). Other notable changes include an increase in Cl-, which is probably attributable to dechlorination of metolachlor, as well as increases in SO2-4 for treatments receiving Al2(SO4)3 (Table 1). Although adding Al2(SO4)3 with Fe0 increased metolachlor destruction, the high SO2-4 concentrations observed after treatment are a potential concern, especially in terms of salinity and subsequent influence on plant growth. It is also noteworthy that despite the large additions of Fe0 and aluminum added to the soil, changes in DTPA-extractable iron increased only about eightfold (40 to ca. 300 mg kg-1) and no increases in extractable aluminum were observed (Table 1).
The ability of aluminum to complex with acidic functional groups of soil humus has been well documented (McBride, 1994). This property was manifested in the soil extracts from windrows receiving Al2(SO4)3, which were transparent, whereas the others were dark brown. Binding of the dissolved organic fraction by the Al oxides probably explains the decreases in extractable organic matter observed in the Al2(SO4)3 treatments and corresponding increases in cation exchange capacity (CEC) (Table 1).
Treatment Costs
In many U.S. states, pesticide-contaminated soils are usually handled in one of three ways: the contaminated soil is reapplied to farmland at label rates or the soil is excavated and shipped to a certified landfill or incinerated (Paulson, 1998). Land-spreading is relatively easy if access to appropriate acreages is obtained and pesticide concentrations are low, but problems arise when the soil contains more than one pesticide (limiting cropping options) or is grossly contaminated and requires dilution. Incineration or landfill options are often too costly and neither treat the soil on-site. Given the multitude of pesticide-contaminated sites on individual farms and cooperatives throughout the USA, additional treatment options are needed that can be readily implemented and are inexpensive for end-users.
An important factor in evaluating remediation technologies is cost. Although factoring in labor, capital outlays, and equipment depreciation is complicated, listing chemical expenditures per mass of soil treated is relatively straightforward. Given that the soil was treated with 5% Fe0 (w/w), and the unit cost of Fe0 (bagged and delivered) was $0.63 per kg, we calculated the cost of Fe0 at $44 per m3 ($34 per yd3). Using a similar approach for the Al2(SO4)3 (unit cost: $0.37 per kg) and acetic acid ($0.90 per L), the cost of these additions per cubic meter (m3) of soil was $10.35 ($8 per yd3) for Al2(SO4)3 and $6.33 ($4.84 per yd3) for acetic acid. Collectively, total chemical expenditures varied between $44 and $61 per m3 ($34–$47 per yd3). Additional costs were incurred for soil mixing, plastic sheeting, and analytical sampling. Unless personnel in charge of treating a contaminated site have access to their own mixing equipment, custom soil mixing would need to be factored into the overall costs. Soil mixing costs are often price-quoted on the volume of soil treated. H&H EcoSystems, the vendor of the soil mixing implement we used, estimated custom mixing charges of approximately $33 per m3 ($25 per yd3) for the volume of soil treated (T. Horn, H&H EcoSystems, personal communication, 2000). Therefore, chemical amendments plus mixing charges for treatment of the contaminated soil ranged between $77 and $94 per m3 ($59–$72 per yd3). This exceeds typical charges for land-spreading (ca. $39 per m3 or $30 per yd3) but is considerably less than what would be charged for removal and disposal by incineration ($261–$1961 per m3, $200–$1500 per yd3; USEPA, 1990). The quote from the local incinerator in Nebraska was $790 per m3 ($604 per yd3). Although land-spreading appears to be the most economical treatment currently available, this option may not be allowed for soils contaminated with the numerous pesticides that have been suspended or banned. Therefore, when land-spreading is not an option, results from our field trial support the use of Fe0 as a possible alternative for on-site treatment of pesticide-contaminated soil.
|
ACKNOWLEDGMENTS |
|---|
|
REFERENCES |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
|---|
This article has been cited by other articles:
|
|
 |
|
  T. Lee and C. H. Benson Sorption and Degradation of Alachlor and Metolachlor in Ground Water Using Green Sands J. Environ. Qual., September 1, 2004; 33(5): 1682 - 1693. [Abstract] [Full Text] [PDF]
|
|
|
|
|
|
D. H. Phillips, D. B. Watson, Y. Roh, and B. Gu Mineralogical Characteristics and Transformations during Long-Term Operation of a Zerovalent Iron Reactive Barrier J. Environ. Qual., November 1, 2003; 32(6): 2033 - 2045. [Abstract] [Full Text] [PDF]
|
|
|
|
|
|
S. D. Comfort, P. J. Shea, T. A. Machacek, and T. Satapanajaru Pilot-Scale Treatment of RDX-Contaminated Soil with Zerovalent Iron J. Environ. Qual., September 1, 2003; 32(5): 1717 - 1725. [Abstract] [Full Text] [PDF]
|
|
|
|
|
|
T. Satapanajaru, S. D. Comfort, and P. J. Shea Enhancing Metolachlor Destruction Rates with Aluminum and Iron Salts during Zerovalent Iron Treatment J. Environ. Qual., September 1, 2003; 32(5): 1726 - 1734. [Abstract] [Full Text] [PDF]
|
|
|
|
|
|
H.M. Gaber, S.D. Comfort, P.J. Shea, and T.A. Machacek Metolachlor Dechlorination by Zerovalent Iron during Unsaturated Transport J. Environ. Qual., May 1, 2002; 31(3): 962 - 969. [Abstract] [Full Text] [PDF]
|
|
| ||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||
| HOME | HELP | FEEDBACK | SUBSCRIPTIONS | ARCHIVE | SEARCH | TABLE OF CONTENTS |
| The SCI Journals | Agronomy Journal | Crop Science | |||
| Journal of Natural Resources and Life Sciences Education |
Vadose Zone Journal | ||||
| Soil Science Society of America Journal | Journal of Plant Registrations | The Plant Genome | |||
