Evaluation of MTBE Remediation Options

Transcription

1 Evaluation of MTBE Remediation Options A Report Written for: The California MTBE Research Partnership By: Daniel N. Creek, P.E. James M. Davidson, P.G. Alpine Environmental, Inc. April 2004 I

2 Published by the Center for Groundwater Restoration and Protection National Water Research Institute NWRI Ellis Avenue P.O. Box Fountain Valley, California (714) Fax: (714) II

3 Limitations This document was prepared by Malcolm Pirnie, Inc. and Alpine Environmental, Inc. and is intended for use by members of the California MTBE Research Partnership (Partnership) pursuant to the Partnership agreement. Malcolm Pirnie, Alpine Environmental, and the Partnership do not warrant, guarantee, or attest to the accuracy or completeness of the data, interpretations, practices, conclusions, suggestions, or recommendations contained herein. Use of this document, or reliance on any information contained herein, by any party or entity other than members of the Partnership, is at the sole risk of such parties or entities. The purpose of this document is to provide a general understanding of the fate, transport, and remediation of MTBE in the environment. The authors also hope to provide a reasonably comprehensive summary of the state of research and practice for MTBE issues, albeit a rapidly evolving field of study. This document is not intended to be a document for prescribing the development of remedial strategies, technology selection, or remediation design for MTBE-contaminated sites. i

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5 Acknowledgements This report was prepared by Rula A. Deeb, Ph.D., Amparo E. Flores, Andrew J. Stocking, P.E., Scott E. Thompson, P.E., and Michael C. Kavanaugh, Ph.D., P.E., of Malcolm Pirnie, Inc.; Daniel N. Creek, P.E., of Alpine Environmental, Inc.; and James M. Davidson, P.G., formerly of Alpine Environmental, Inc. The authors would like to thank the California MTBE Research Partnership and the National Water Research Institute (NWRI) for sponsoring this work. The authors are grateful to the many members of the Partnership s Research Advisory Committee who provided valuable support and review of this work. We also extend our appreciation to the many engineers and scientists, especially Richard Sloan of Lyondell Chemical Company (Houston, Texas), Richard Woodward of Sierra Environmental Services (Houston, Texas), and David Ramsden, Ph.D., of URS Greiner Woodward Clyde (Houston, Texas), who shared their documents, data, time, and knowledge to improve this report. iii

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7 California MTBE Research Partnership The California MTBE Research Partnership is comprised of the following individuals: Steering Committee Mark Beuhler, Chair Metropolitan Water District of Southern California Research Advisory Committee Walter Weber, Jr., Chair University of Michigan at Ann Arbor Rey Rodriguez, Chair, Source Water Protection Subcommittee H 2 O R 2 Consulting Engineers, Inc. Scott Tenney, Chair, Treatability Subcommittee Exxon Mobil Richard Atwater, Inland Empire Utilities Agency Bruce Bauman, American Petroleum Institute Ivo Bergsohn, South Tahoe Public Utilities District Tim Buscheck, Chevron Texaco David Camille, Tosco Robert Cheng, Long Beach Water Department Chi-Su Chou, Consultant Krista Clark, Association of California Water Agencies Mike Cooper, Placer Co. Daniel Creek, Alpine Environmental, Inc. James Crowley, Santa Clara Valley Water District James Davidson, Exponent Marshall Davis, Metropolitan Water District of Southern California Rula Deeb, Malcolm Pirnie, Inc. Shahla Farahnak, State Water Resources Control Board Jack Fraim, Chevron Texaco John Gaston, CH2M Hill Don Gilson, Chevron Texaco John Gustafson, Equilon, Inc. Elliot Heide, McClintock, Weston, Bens Roy Hodgen, Lyondell Chemical Company Tracy Hemmeter, Santa Clara Valley Water District Dave Smith, BP Mike Wang, Western States Petroleum Association Rick Hydrick, South Tahoe Public Utilities District Steven Inn, Alameda County Water District Michael Kavanaugh, Malcolm Pirnie, Inc. Charles Kish, Riverview Water District John Kneiss, Oxygenated Fuel Association Sun Liang, Metropolitan Water District of Southern California Ronald Linsky, National Water Research Institute Ernie Lory, U.S. Naval Facilities Engineering Dave McKinney, Shell Ralph Moran, BP Stephen Morse, Regional Water Quality Control Board, San Francisco Region David Pierce, Chevron Texaco Roger Pierno, Santa Clara Valley Water District Dave Ramsden, GZA, GeoEnvironmental, Inc. Melinda Rho, Los Angeles Department of Water and Power Richard Sakaji, California Department of Health Services Dick Sloan, Lyondell Chemical Company Curt Stanley, Shell Global Solutions, Inc. Paul Sun, Shell Global Solutions, Inc. Martin Varga, Kern County Water Agency Christine White, Equiva Services, LLC. Ken Williams, Regional Water Quality Control Board, Santa Ana Region Jeff Wilson, Western States Petroleum Association Dick Woodward, Sierra Environmental Services, Inc. Nira Yamachika, Orange County Water District v

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9 Contents Report Overview...xvii 1. Introduction Problem Statement History of MTBE Use Scope of Report MTBE Fate and Transport Introduction Fate and Transport of MTBE in the Vadose Zone Fate and Transport of MTBE in the Saturated Zone Vertical Transport of Dissolved-Phase MTBE Distribution Between Soil and Water Partitioning Between Cellular Organic Matter and Water Partitioning Between Air and Water Review of Plume Studies Benzene-Only Studies BTEX and MTBE Plume Studies Natural Attenuation Mechanisms for MTBE Relative to Benzene Conclusions on MTBE Fate and Transport Remedial Strategies Introduction Site Characterization Site Investigations Development of a Site Conceptual Model Establishing Cleanup Levels Regulatory Framework Risk-Based Corrective Action Evaluation of Mass Flux and Receptor Impacts Technology Selection Applicable Technologies Factors to Consider for Technology Selection Optimizing Technical Applications Source-Area Issues Conclusions...26 vii

10 4. Remediation Technologies Pump-and-Treat/Groundwater Extraction Description Effects of Contaminant and Site Characteristics Predictive Relative Effectiveness for MTBE Pump-and-Treat Case Studies Pump-and-Treat Conclusions Soil Vapor Extraction (SVE) Description Effects of Contaminant and Site Characteristics Predictive Relative Effectiveness for MTBE SVE Case Studies SVE Conclusions Multi-Phase Extraction (MPE) Description Effects of Contaminant and Site Characteristics Predictive Relative Effectiveness for MTBE MPE Case Studies MPE Conclusions In Situ Air Sparging Description Effects of Contaminant and Site Characteristics Predictive Relative Effectiveness for MTBE In Situ Air Sparging Case Studies In Situ Air Sparging Conclusions In Situ Chemical Oxidation Processes Description Effects of Contaminant and Site Characteristics Predictive Relative Effectiveness for MTBE In Situ Chemical Oxidation Case Studies In Situ Chemical Oxidation Conclusions In Situ Bioremediation Description Effects of Contaminant and Site Characteristics Predictive Relative Effectiveness for MTBE In Situ Bioremediation Case Studies In Situ Bioremediation Conclusions Natural Attenuation Description Effects of Contaminant and Site Characteristics Predictive Relative Effectiveness for MTBE viii

11 4.7.4 Natural Attenuation Case Studies Natural Attenuation Conclusions Overall Conclusions Emerging Technologies, Techniques, and Process Enhancements Optimization of Pump-and-Treat/Groundwater Extraction Pulsed Pumping Adaptive Pumping Ex Situ Treatment Advanced Oxidation Processes (AOPs) Synthetic Resin Sorbents In Situ Thermal Processes Six-Phase Heating In Situ Radio Frequency Heating Dynamic Underground Stripping In Situ Chemical Reduction Phytoremediation Remediation Case Studies Site Summaries Primary Lessons Learned Pump and Treat/Groundwater Extraction Treatment of Extracted Water Soil Vapor Extraction (SVE) and Multi-Phase Extraction (MPE) Vapor Treatment Air Sparging Systems Remediation Cost Estimates Introduction Methodology Contamination Scenarios Scenario A Young Shallow Release Scenario B Old Large Plume Scenario C Large Vadose Zone Sensitivity Analysis Conclusions Conclusions...91 References...95 Appendix: Summaries of Case Study Sites ix

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13 Tables 1-1 Oxygenated Fuels and History of Usage in the United States Solubility of BTEX and MTBE into Water from Gasoline Physical and Chemical Properties of BTEX and MTBE Description and Applications of Analytical Methods for Fuel Oxygenates State Cleanup Levels for MTBE Soil and Groundwater Remediation as of Effect of Retardation on Remediation Time Using Groundwater Extraction and Treatment Effect of Soil Moisture Content on Retardation and MTBE Remediation Time Using SVE Estimates of BTEX and MTBE Vapor-Phase Concentrations Near an LNAPL Using Raoult s Law Estimates of BTEX and MTBE Equilibrium Vapor-Phase Concentrations in Sparge Bubbles Production of Oxidizing Agents Microbial Metabolism of Organic Matter Under Representative Aerobic and Anaerobic Conditions Natural Attenuation Mechanisms and Their Effectiveness for MTBE Removal in the Subsurface Relative to Benzene Observations from Representative MTBE Natural Attenuation Field Studies Concentration Reductions with Groundwater Extraction and Treatment Systems Over Time Air Stripper Performance at Three Case Study Sites Vapor Treatment Case Studies Plume Characterization of Scenario A Plume Characterization of Scenario B Plume Characterization of Scenario C...87 xi

14 A-1 Costs for the 30-gpm Pump-and-Treat System A-2 Costs for the Moderate-Sized SVE System A-3 Costs for the Small-Scale Combined SVE and 3-gpm Pump-and-Treat System A-4 Costs for the Small-Scale Combined SVE and Pump-and-Treat System A-5 Approximate Initial Concentrations and Concentrations Measured at the End of a 24-Hour Air Sparing System Performance Test xii

15 Figures 3-1 Flow chart of applicable MTBE remediation technologies United States Environmental Protection Agency graphical representation for predicting the successfulness of SVE Types of MPE systems Effect of soil moisture on mass removal rates at varying vacuum rates and P w (atm) Influent water concentrations for Site 4 pump-and-treat system (with SVE) Scenario A LUST creating a dissolved plume 20 feet below the surface Scenario B LUST with a MTBE plume ahead of a BTEX plume Scenario C LUST with contamination in shallow and deep aquifers xiii

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17 Acronyms, Symbols, and Abbreviations AOP BTEX CARB cfm cm/sec GAC GC gpm H 2 O 2 IST LNAPL LP LUFT LUST MP MPE MS MTBE mg/l nm O 3 O&M ORC ppmv RFG SVE TBA TiO 2 UST UV µg/l Advanced oxidation process Benzene, toluene, ethylbenzene, and xylenes (o-, m-, p-xylene) California Air Resources Board Cubic feet per minute Centimeters per second Granular activated carbon Gas chromatography Gallons per minute Hydrogen peroxide Integrated Science & Technology, Inc. Light non-aqueous phase liquid Low pressure Leaking underground fuel tank Leaking underground storage tank Medium pressure Multi-phase extraction Mass spectrometry Methyl tertiary butyl ether Milligrams per liter Nanometer Ozone Operation and maintenance Oxygen Release Compound Parts per million by volume Reformulated gasoline Soil vapor extraction Tertiary butyl alcohol Titanium oxide Underground storage tank Ultraviolet Micrograms per liter xv

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19 Report Overview Recent reports of the widespread occurrence of methyl tertiary butyl ether (MTBE) in groundwater samples collected at gasoline underground storage tank (UST) sites have led to the reassessment of cleanup strategies for MTBE-impacted sites by regulatory agencies. Due to the introduction of MTBE to subsurface environments, UST owners now recognize that new remediation challenges and potentially higher damage assessments can arise. In addition, since MTBE is resistant to biodegradation relative to other gasoline components, the effectiveness of natural attenuation in cleaning up most gasoline-contaminated aquifers can be limited. MTBE presents unique remediation challenges because of its physical and chemical properties: it is highly soluble in water, does not sorb strongly to aquifer materials, and exhibits a low tendency to volatilize from groundwater. Moreover, depending on the release scenario, MTBE plumes can extend farther than benzene, toluene, ethylbenzene, and xylenes (BTEX) plumes, ultimately impacting larger volumes of groundwater; therefore, there are concerns about the economic feasibility of remediating MTBE-contaminated sites. This document was prepared under the auspices of the California MTBE Research Partnership to provide an in-depth evaluation of remediation strategies and technologies capable of addressing cleanup challenges caused by MTBE and its byproducts. Chapter 1 reviews the history of MTBE s use in gasoline as an oxygenate and summarizes the problems that have followed. Chapter 2 reviews the fate and transport of MTBE following a subsurface release of MTBEblended gasoline, with an emphasis on the relevance of fate and transport characteristics to subsurface remediation strategies. A comprehensive review of BTEX and MTBE plume studies is presented in an effort to confirm the behavior of these compounds in different geologic and hydrogeologic settings. Chapter 3 discusses the process for developing remedial strategies at MTBE-impacted sites, including site investigations, site conceptual model development, and stakeholder agreement on remedial objectives. Although traditional approaches for site investigation and corrective action at gasoline-impacted sites are applicable and useful at MTBE-impacted sites, some issues are unique to sites contaminated with MTBE. Chapter 4 discusses a number of demonstrated in situ and ex situ technologies that can be used to effectively remediate MTBE-impacted sites. Summaries of case studies are also presented for a range of technologies. Most conventional technologies and some emerging technologies have been successful in partially or fully remediating MTBE-impacted sites. The most widely used technologies include pump-and-treat, soil vapor extraction (SVE), multi-phase extraction (MPE), in situ air sparging, in situ chemical oxidation, and in situ bioremediation. In addition, natural attenuation is discussed in this section because it frequently merits consideration as a remedial solution and, in many cases, is the most economic remedial alternative. xvii

20 Chapter 5 discusses emerging technologies that have shown some promise for MTBE remediation and are potentially more efficient or cost-effective than demonstrated technologies. In addition, this section focuses on techniques and process enhancements that can potentially be implemented together with a demonstrated technology for improving efficiency and/or reducing remediation costs or the duration for achieving cleanup goals. Some of the emerging technologies for MTBE removal include ex situ technologies, such as advanced oxidation processes (AOPs) and synthetic resin sorbents, as well as in situ technologies, such as thermal processes, phytoremediation, and chemical reduction. Chapter 6 summarizes the findings from five case studies dealing with subsurface MTBE remediation. The purpose of this chapter is to present a synopsis of the primary lessons learned from each case study. Chapter 7 evaluates the relative costs associated with different remediation strategies at MTBE-impacted sites. Due to the site-specific nature of remediation operations and numerous factors involved, it is often difficult to extrapolate these costs to other sites. Some of the factors that greatly influence cost estimates include: Presence of other contaminants. Range of hydrocarbon concentrations. Location of contaminants of concern. Hydrogeological parameters at the site. Size of the impacted area. Target cleanup levels. Three release scenarios of MTBE-blended gasoline were evaluated comparatively: a young shallow release, an older large plume, and a large vadose zone release. Chapter 7 demonstrated that MTBE and BTEX plumes can be cost-effectively remediated using a number of technologies. As long as the MTBE and BTEX plumes are equivalent in size, remediation costs for MTBE-impacted sites are expected to be similar to or larger than those associated with BTEX-only sites where active remediation is required; however, if an MTBE plume has migrated past the associated BTEX constituents, site characterization and remediation costs could significantly increase because of the greater area that needs to be characterized and remediated. Finally, Chapter 8 provides key conclusions from each of the preceding sections. xviii

21 1. Introduction 1.1 Problem Statement In 1995, the Lawrence Livermore National Laboratory reported that greater than 90 percent of the groundwater plumes (defined as the 10 micrograms per liter [µg/l] benzene isoconcentration level) emanating from UST gasoline releases in California were likely to stabilize (i.e., stop increasing in size) at distances of less than 250 feet downgradient of leaking underground storage tank (LUST) releases (Rice et al., 1995). These plumes, identified primarily by one or more BTEX components, moved slowly and eventually stabilized due to natural biodegradation and retardation. The Lawrence Livermore National Laboratory concluded that many BTEX plumes might not require active remediation due to these natural attenuating processes and that monitored natural attenuation as a remedial strategy can be relied on at many UST sites in California and other states. Shortly thereafter, MTBE (an oxygenate added to gasoline to increase octane levels and to meet federal and state fuel specification requirements for oxygen content) was detected in drinking-water wells in the City of Santa Monica, California. This discovery caused an immediate reassessment of cleanup strategies at gasoline UST sites by regulatory agencies in California. Reports of the widespread occurrence of MTBE in groundwater samples collected at UST sites in California and elsewhere confirmed that UST owners were now faced with new remediation challenges. MTBE and other ether oxygenates were thought to be resistant to biodegradation during that time period, thereby limiting the use of natural attenuation for the cleanup of contaminated groundwater at UST sites. MTBE represents some unique remediation challenges because of its physical and chemical properties. MTBE is highly soluble in water, is only weakly sorbed to most soils, and exhibits a low tendency to volatilize from water. Consequently, MTBE moves approximately at the rate of groundwater flow, and (if no active remediation is undertaken) can pose a significant threat to downgradient water supply wells. Moreover, depending on the release scenario, MTBE may move farther than BTEX plumes, ultimately impacting a larger volume of groundwater compared to BTEX-only plumes. These inherent characteristics raised concerns about the extent of MTBE impacts to groundwater (Squillace et al., 1997; Davidson, 1995) and the feasibility of remediating these MTBE-impacted sites at a comparable cost to active remediation costs at BTEX-only contaminated sites (Creek and Davidson, 1998; Davidson and Parsons, 1996; Keller et al., 1998; Peargin, 1998). 1.2 History of MTBE Use MTBE has been used in gasoline since the late 1970s as an oxygenate to replace tetra ethyl lead, which is an octane enhancer/anti-knock component. Beginning in 1992, the United States Environmental Protection Agency mandated the use of oxygenates, including MTBE, to meet federal regulations for cleaner-burning gasoline. There are several types of oxygenated 1

22 fuels (Federal Oxyfuel, Federal Reformulated Gasoline [RFG], and several California Air Resources Board [CARB] fuels) with different specifications for oxygen content (Table 1-1). Each type of oxygenated fuel requires the addition of an oxygenate. Although MTBE is not specifically required in any fuel blends, it became the most commonly used oxygenate at the time because other oxygenates (e.g., ethanol, ethyl tertiary butyl ether, tertiary amyl ether, and di-isopropyl ether) were not commercially available in the quantities required, did not produce the same proven air-quality benefits as MTBE, did not have as desirable fuel blending characteristics, or were too expensive. Table 1-1 Oxygenated Fuels and History of Usage in the United States Octane- Reformulated California Enhanced Gasoline Air Resource Board Characteristics Fuel Oxyfuel (RFG) (CARB) Fuel Percent Oxygen by Weight Unknown to 2.2 Percent MTBE by Volume < Varies Date of First Use 1979 Winter January 1, 1995 January 1, 1995 (lead 1992 to replacement) 1993 Use in California 1979 Not used Year-round Year-round in ozone in all non-attainment other areas areas Use in 17 Other States 1979 Wintertime in Year-round Not used carbon monoxide in ozone non-attainment non-attainment areas areas 1.3 Scope of Report This document has been prepared under the auspices of the California MTBE Research Partnership to provide an in-depth evaluation of remediation strategies and technologies capable of addressing UST cleanup challenges created by MTBE and its breakdown products. Although ether oxygenates other than MTBE are not specifically addressed in this report, many of the strategies and technologies discussed are relevant to these compounds. This report will critically assess the effectiveness of existing and emerging remediation strategies and technologies to meet site-specific remediation objectives, such as source control, plume control, risk reduction, and the achievement of soil and groundwater cleanup criteria. The report is intended to provide a sufficient level of detail on existing and emerging technologies to allow consultants, UST owners, regulatory personnel, and other interested parties to address a wide range of UST cleanup alternatives at MTBE-impacted sites. 2

23 2. MTBE Fate and Transport 2.1 Introduction The selection of a remediation strategy to address the cleanup of a release of MTBEcontaining gasoline depends on the volume of impacted soil and groundwater, contaminant concentrations and mass present, site-specific hydrogeological characteristics, and the riskbased or regulatory-driven cleanup objectives. Understanding the relationship between the release of the gasoline and its existing and possible future impacts on soil and groundwater requires a thorough knowledge of the fate and transport of the contaminants of concern at an individual site; thus, fate and transport assessments become an essential component of the decision-making process regarding actions to be taken to protect human health and the environment. Of particular interest at MTBE-impacted sites is the unusual behavior of MTBE relative to other constituents in gasoline and the potential for MTBE to impact a much larger volume of groundwater relative to sites impacted only by BTEX compounds; thus, fate and transport issues for developing remedial strategies at these sites require a more thorough analysis compared to gasoline-only sites. A gasoline release into the subsurface can originate from several sources, including a leaking underground fuel tank (LUFT), a leaking fuel pipeline, a catastrophic tank failure, minor ongoing spillage, or a surface spill. As a component of gasoline, MTBE could enter the subsurface through any of these release scenarios. Assuming the release is sufficiently large to reach groundwater, the concentrations of MTBE in groundwater would depend on the duration and volume of the release, as well as the concentration of MTBE in the fuel. The aqueous solubility for pure MTBE is approximately 50,000 milligrams per liter (mg/l) at 25 C; however, based on chemical thermodynamics and assuming that MTBE comprises 11 to 15 percent of gasoline by volume, the maximum theoretical MTBE concentration dissolved in groundwater from free-phase gasoline is approximately 6,000 mg/l (Zogorski et al., 1997) (Table 2-1). In contrast, maximum MTBE concentrations observed in the field in the immediate vicinity of a gasoline release have rarely exceeded 400 mg/l (Buscheck et al., Table 2-1 Solubility of BTEX and MTBE into Water from Gasoline (Zogorski et al., 1997) Theoretical Approximate Predicted Solubility Aqueous Solubility Percent by Volume in Groundwater Compound (mg/l) in Gasoline (mg/l) Benzene 1, Toluene Ethylbenzene Xylene MTBE 43,000 to 54, to 15 5,280 to 7,200 3

24 1998; Crowley, 1998). This observation is due to dilution effects in the vicinity of the release, the slow rate of dissolution of MTBE into the groundwater, well siting, or well screen length. In addition to the concentrations and amount of MTBE in the groundwater, another major factor in determining a remediation strategy at these sites is the size of the source area contributing to the MTBE/BTEX plume and the time that has passed since the first and last release. Most historical releases from LUFTs have occurred prior to the introduction of MTBE and, thus, many sites were characterized by BTEX plumes that had stabilized. MTBE has only been added to gasoline in significant quantities (greater than 8 percent by volume) since the early 1990s. In general, given a scenario with historic groundwater contamination, as well as a recent gasoline release, it is typical that the leading edge of the MTBE plume has not yet passed the leading edge of the previously released BTEX plume (Happel et al., 1998; Mace and Choi, 1998). Two other fate and transport factors influence remedial decisions at MTBE-impacted sites. First, the natural biodegradation of MTBE by indigenous microorganisms is known to occur very slowly, if at all. Because of the limited natural biodegradation of MTBE relative to BTEX compounds, MTBE plumes may migrate significant distances, and this causes unacceptable impacts on community water supply wells. One recent assessment of the potential vulnerability of groundwater basins suggests that if MTBE plumes are not remediated, up to 9,000 community water supply wells in the United States could be at risk of MTBE contamination. Others dispute the probability of such widespread impacts (Einarson and Mackay, 2001). Second, contrary to common assumptions, MTBE dissolution from gasoline released into groundwater can be a slow process; thus, residual gasoline containing MTBE could behave as a long-term source of dissolved-phase MTBE (Rixey and Joshi, 2000). Because of these factors, early responses to LUFTs are essential to reduce or eliminate the impacts to public water supply systems. 2.2 Fate and Transport of MTBE in the Vadose Zone The properties of MTBE and its environmental fate and transport characteristics have been thoroughly reported (Zogorski et al., 1997; Happel et al., 1998; Malcolm Pirnie, 1998a). It is useful, however, to compare the properties of MTBE and BTEX compounds (Table 2-2) and to assess the impact of physicochemical properties on the fate and transport of these compounds in various subsurface zones (i.e., vadose zone, capillary fringe, and saturated zone). Factors that influence the relative rates of transport include both chemical-specific parameters (e.g., soil/water partition coefficient, Henry s constant, aqueous solubility, vapor pressure, and density) and hydrogeological parameters (e.g., hydraulic conductivity, porosity, hydraulic gradient, fractional organic carbon content of the soil, microbial population characteristics, and reduction-oxidation [redox] conditions). Once the extent of MTBE in the subsurface is assessed (American Petroleum Institute, 2000), a risk-based remediation strategy can be developed to meet remedial objectives (American Society for Testing and Materials, 1998). The technologies selected to achieve these objectives should take advantage 4

25 Table 2-2 Physical and Chemical Properties of BTEX and MTBE Physical and Chemical Benzene Toluene Ethylbenzene o-xylene MTBE Properties Molecular Weight (g/mole) Vapor Density at 1 atm; 10 C (g/l) Specific Gravity at 25 C 0.88 a a a a a Boiling Point ( C) 80.1 a a a a 53.6 to 55.2 a Water Solubility (mg/l) 1,730 a a 161 a 175 a 43,000 to 54,300 a Vapor Pressure [mm Hg] (at 25 C) 76, a 28.4 a 9.53 a 6.6 a 245 to 276 a [kpa] (at 100 F) 1.26 b 3.79 b 1.23 b b Henry s Law 0.23 b b b b a Constant [ ] a a a a Log K oc 1.18 to to 2.25 a to 1.83 a 1.091, 1.035, 1.50 to 2.16 a 1.98 to 3.04 a a Log K ow 2.36 b 2.73 b 3.24 b 3.10 b 1.20 a Diffusivity (m 2 /s) Liquid Gas b b b b b a Zogorski et al., atm = Atmospheres. mm Hg = Millimeters mercury. b Crittenden et al., kpa = Kilo pascals. [ ] = Does not have a unit. of the specific properties of MTBE so that the removal efficiency of MTBE from the subsurface can be optimized and remedial costs can be minimized. It should be noted, however, that when developing a remedial strategy, one should consider the effectiveness and cost for all contaminants of concern, not just MTBE. Following a release of an MTBE-blended gasoline from a LUFT, the gasoline will migrate vertically downward through the vadose zone. The rate and extent of vertical migration of the gasoline depend upon the volume of release and the hydraulic characteristics (permeability, heterogeneity) of the soil. Gasoline can become trapped in pore spaces as a residual product (i.e., it is no longer mobile), which effectively smears the downward migrating gasoline (Einarson, 1998), thereby creating a thicker vadose zone source area. The vapor pressure of MTBE is higher than that of BTEX compounds (see Table 2-2). MTBE will, therefore, volatilize more quickly than BTEX compounds from either the downward migrating gasoline or the residual product to form a vapor-phase plume in the unsaturated soil. The vapor density 5

26 of MTBE is similar to that of BTEX compounds (see Table 2-2); thus, MTBE vapor will behave similarly to BTEX vapor and can be expected to diffuse radially from the source area due to concentration, temperature, and pressure gradients (Pankow and Cherry, 1996). Infiltrating (recharge) water from precipitation will also travel vertically toward the groundwater and, in the process, dissolve gasoline components that come in contact with it. 2.3 Fate and Transport of MTBE in the Saturated Zone Because gasoline has a specific gravity less than water and is relatively immiscible in water, it is referred to as a light non-aqueous phase liquid (LNAPL). If LNAPL releases are sufficiently large at LUFT sites, the free product will reach the capillary fringe between the groundwater and vadose zone. Once at the groundwater boundary, soluble components of the LNAPL will dissolve into the groundwater, forming a plume containing gasoline constituents and dominated by the presence of the more soluble components of gasoline, namely BTEX and MTBE. As stated earlier, the maximum theoretical solubility of MTBE from gasoline containing 11- to 15-percent MTBE by volume is approximately 6,000 mg/l at ambient temperatures, whereas the maximum theoretical solubility of benzene is about 17 mg/l (i.e., 1,730 mg/l maximum solubility multiplied by its 1 percent by volume in gasoline). Hypothetically, MTBE should become depleted from LNAPL at a more rapid rate than BTEX compounds. The rate of dissolution will be limited by groundwater contact with LNAPL (Pankow, 1998). Recent theoretical and laboratory studies have clearly shown that MTBE dissolution from residual gasoline will be much slower than predicted by equilibrium calculations under most release scenarios (Rixey and Joshi, 2000). The reduction in the theoretical rate of dissolution is primarily due to limited contact between flowing water and the residual product, causing the dissolution process to be limited by mass transfer processes. Diffusion processes are inherently slow. Rixey and Joshi show that the number of pore volumes of water that must be flushed through the source zone could increase by an order of magnitude or more to reduce MTBE in the source zone to acceptable levels if the dissolution process is masstransfer limited. This slow rate of dissolution may explain why there have been few examples of MTBE plumes completely separating from BTEX plumes, as would otherwise be expected given MTBE s higher solubility. Once MTBE has entered pore water in the saturated zone, the primary mechanism for BTEX and MTBE mass transfer away from the original release area is advection. Groundwater and any dissolved-phase contaminants will flow preferentially through zones of higher hydraulic conductivity. In a heterogeneous aquifer, areas of higher hydraulic conductivity could contain higher concentrations of BTEX and MTBE than adjacent regions that are less permeable, establishing a diffusive gradient to drive the compounds into lower permeable regions (Mackay, 1998). This phenomenon impacts MTBE transport more than BTEX compounds due to its higher anticipated concentration in groundwater. The time required for diffusion out 6

27 of the low permeable regions during remediation is likely to cause the duration of remediation to increase compared to predictions made using equilibrium models (Mackay, 1998) Vertical Transport of Dissolved-Phase MTBE In most hydrogeological settings, the magnitude of horizontal hydraulic conductivity is several times greater than vertical hydraulic conductivity; thus, gasoline and MTBE will primarily migrate horizontally unless significant vertical gradients exist (American Petroleum Institute, 1994). Like all dissolved components, dissolved-phase MTBE migrates in a direction controlled by hydraulic gradients and toward areas of lower concentration; thus, the phenomenon of plunging MTBE plumes reported in the literature (e.g., American Petroleum Institute, 1994; Borden et al., 1997) occurs primarily due to the presence of strong vertical gradients. A significant vertical migration of MTBE plumes has been noted at several field sites, such as East Patchogue (New York) and LaCrosse (Kansas). Observations at these sites suggest that the vertical movement of MTBE is enhanced by the following mechanisms: MTBE exhibits minimal retardation and can migrate farther downgradient than BTEX compounds; therefore, the vertical migration of MTBE will be proportionally greater than the total vertical migration of the shorter BTEX plumes. In areas with significant infiltration, longer MTBE plumes might be driven deeper due to multiple infiltration events, causing a downward slug displacement of the MTBE plume. If axially longer than BTEX plumes, MTBE plumes may be preferentially driven deeper due to vertical gradients from groundwater extraction systems (e.g., pumping wells, dewatering systems). If axially longer than BTEX plumes, MTBE plumes may exhibit vertical migration after encountering geologic heterogeneities. In conclusion, the potential for greater vertical transport of MTBE plumes compared to BTEX plumes is real and must be evaluated as part of site assessment and remedial action activities. This phenomenon can complicate characterization and remediation efforts Distribution Between Soil and Water Once BTEX and MTBE have dissolved into groundwater, these compounds will partition between the soil and groundwater. The soil/water distribution coefficient, K d, specifies the equilibrium ratio of a contaminant s concentration in/on the solid phase to that in the aqueous phase. The soil/water distribution coefficient determines a contaminant s relative rate of movement in groundwater due to adsorption to aquifer solids. In the context of contaminant interactions with soils, the distribution coefficient is a linear estimate of sorption and usually applies only over limited concentration ranges; however, use of this term is common in the 7

28 literature and, therefore, will be employed here as a point of reference for estimating the distribution of a contaminant between water and soil. For non-ionic compounds such as MTBE, K d values are a function of the organic carbon content of the aquifer matrix and the organic carbon-based distribution coefficient, K oc. K d is the product of K oc and the fraction of organic carbon in the soil, f oc (i.e., K d = K oc f oc ). Although values of f oc are sitedependent, they are typically low (approximately 0.005) in most aquifers (Zogorski et al., 1997). Additionally, the K oc value for MTBE is low relative to values for BTEX compounds (see Table 2-2); therefore, MTBE s K d is generally low, which results in a retardation factor, R, close to unity (Zogorski et al., 1997). Retardation is the phenomenon of diminished contaminant velocity relative to actual groundwater velocity (Equation 2.1). R = ν ν c (Equation 2.1) where: R ν ν c = Retardation factor. = Groundwater velocity. = Contaminant velocity. Sorption in subsurface environments is typically quantified using a chemical-specific retardation factor. Since the retardation factor for MTBE is close to one, dissolved-phase MTBE can be expected to migrate at the velocity of the groundwater, especially in sand or gravel aquifers. This fact was first confirmed by field data from the Borden site in Canada. During this study, an MTBE plume from the release of MTBE-blended gasoline was monitored in considerable detail (Schirmer and Barker, 1998). The migration rate of MTBE was found to be comparable to that of a conservative tracer (chloride ion). MTBE migration in this study and in other studies has been shown to be much more rapid than that of BTEX compounds (Barker et al., 1990; Davidson, 1995). Aquifer materials at the Borden site have an f oc value of about Partitioning Between Cellular Organic Matter and Water The partitioning of MTBE into cellular organic matter from water can affect the potential for bioaccumulation in the environment, as well as the ability of MTBE to move through microorganisms and plants. The partitioning of a compound between organic matter and water is described by the organic matter-water partition coefficient, K om. Because n-octanol is reasonably representative of environmental and physiological organic matter, the octanolwater partition coefficient, K ow, is often used as a reference for estimating the K om of compounds. Log K om is directly proportional to log K ow (Schwarzenbach et al., 1993), so it can be estimated for various compounds using K ow. K ow values are reported in the literature for a wide variety of organic compounds. 8

29 MTBE can also be uptaken by plant roots. In fact, phytoremediation as a cleanup technology relies on either the transport of water and contaminants into and through plants or the promotion of conditions in the plant rhizosphere to affect contaminant destruction. Phytoremediation involves the uptake and degradation, bioaccumulation, or transpiration of contaminants. It can also involve sorption to filtration by the plant rhizosphere. If a contaminant is passively adsorbed, the tendency of a plant to adsorb and metabolize a contaminant can be represented using a linear adsorption isotherm where the bioconcentration of a contaminant is typically proportional to the octanol-water partition coefficient, K ow (typically presented as log K ow ). The low K ow value for MTBE compared to BTEX compounds (see Table 2-2) suggests that the potential for MTBE bioaccumulation in microorganisms and plant materials is less likely compared to that of BTEX compounds; however, the high solubility of MTBE would indicate that transpiration will occur easily (see Chapter 5 for a discussion of MTBE remediation using phytoremediation) Partitioning Between Air and Water The partitioning rate of MTBE from the dissolved phase into the vapor phase is primarily a function of the dissolved-phase concentration, the air/water contact area, and the Henry s constant. As the dissolved-phase concentration increases, a larger equilibrium driving force is established. The Henry s constant defines the ratio of a compound s gas-phase concentration to its liquid-phase concentration at equilibrium; however, a compound will only reach equilibrium if there is adequate mixing and a sufficient amount of interfacial area. Otherwise, volatilization is limited by molecular diffusion (e.g., groundwater off-gassing or surface water volatilization) (Malcolm Pirnie, 1998b). Volatilization can affect the fate of MTBE in groundwater through the transfer of MTBE to the vadose zone, thus reducing the MTBE concentration in the plume. Data from the United States Geological Survey indicate that volatilization rates of MTBE and benzene from groundwater are similar (United States Geological Survey, 1998). Since the Henry s constant of MTBE is relatively low (an order of magnitude smaller than that of benzene), the natural volatilization (or off-gassing) of MTBE from groundwater is controlled by molecular diffusion. The magnitude of this process has not been reported in literature, but in certain hydrogeological settings (e.g., shallow plumes or warmer climates), MTBE off-gassing could be a measurable mass transfer mechanism. Losses of MTBE in flowing rivers (Pankow, 1996) and lakes (Malcolm Pirnie, 1998b) have been confirmed as significant mass loss mechanisms. 2.4 Review of Plume Studies Over the past few years, several studies characterizing BTEX and MTBE plumes have been conducted to determine the distribution of BTEX and MTBE in groundwater and to estimate contaminant plume dimensions. A sufficient number of plumes have been evaluated such that inferences about BTEX and MTBE plumes at other sites can be made. These studies represent the compilation of data from several hundred benzene plumes and approximately 9

30 130 MTBE plumes, and provide a reasonably complete depiction of plume behavior for both contaminants Benzene Only Studies Study by Mace et al. (1997) Mace et al. (1997) retrieved geographical, soil, groundwater, and chemical information from Texas Natural Resource Conservation Commission files for 605 LUFTs. The authors generated a database of over 500,000 entries that included chemical and water level data for more than 4,000 monitoring wells. The authors assessed hydraulic gradients, groundwater flow directions, average benzene-plume concentrations, and benzene-plume dimensions over time for several different hydrogeological and climatic regions of Texas. The purpose of the study was to determine: Benzene plume dimensions. Relative predictability of benzene plume concentrations and lengths. Rates at which benzene plumes intrinsically attenuate. Benzene plume classification system. The authors reported that groundwater sampling at vertically discrete intervals (i.e., threedimensional plume characterization) was not performed at the sites they reviewed. Mace et al. concluded that 75 percent of benzene plumes in Texas caused by LUFTs are less than 250 feet in length and impact an area that is less than 49,000 square feet (approximately 1 acre) in size each, as defined by the 10-µg/L benzene contour. With or without remediation, the mass and size of benzene plumes pass through three predictable stages. If sources are properly controlled, plume volumes initially increase, stabilize, and then decline over time relatively rapidly. Plume lengths behave similarly, except that plume lengths decline slowly at first and then more rapidly. Only 14 percent of benzene plumes studied increased in concentration and only 3 percent increased in length at the time of the study. Study by Rice et al. (1995) Rice et al. (1995) compiled data from a master list of 5,700 California LUFT sites to determine the following: Whether plumes behave in a predictable fashion. The factors that influence plume length and mass. The extent to which plumes are affecting California groundwater. 10

31 Two-dimensional models were used to assess plume lengths. The purpose of the study was to support a revision of the California LUFT corrective action process and to identify additional data required to support risk and resource-management decisions. Rice et al. concluded the following: The average benzene plume length (based on a 10-µg/L benzene concentration) was 250 feet and changed slowly with time. Benzene plume average concentrations tended to decrease more rapidly than plume lengths. Hydrogeological variables had little predictive value regarding plume length. If sources are reduced or removed, natural attenuation will likely be effective in controlling the migration of BTEX plumes. The total volume of groundwater affected by benzene plumes in California is approximately 7,000 acre-feet, which is less than one two-thousandths ( percent) of California s groundwater resources BTEX and MTBE Plume Studies Buscheck et al. Study (1998) Buscheck et al. (1998) assessed Chevron s investigation records for 466 operating sites and 243 non-operating sites located in California, Texas, Florida, and some northeastern states. The purpose of the study was to further understand the behavior of MTBE in groundwater by means of time-series data analysis. The authors concluded the following: Benzene and MTBE concentrations in fuel hydrocarbon plumes are not correlated. Fifty to 65 percent of 55 MTBE sites (with an average of 11 quarterly rounds of timeseries data) have stable or decreasing MTBE concentrations at the leading edge of the plume. The potential for false-positive MTBE analytical results suggests that analysis by gas chromatography (GC)/mass spectrometry (MS) is warranted where total petroleum hydrocarbon concentrations are greater than 1,000 mg/l and where MTBE concentrations are less than 100 µg/l. Study by Davidson (1995) Davidson (1995) assessed MTBE groundwater concentration data available prior to 1995 from hundreds of public water supply sites in Illinois, Massachusetts, New York, New Jersey, 11

32 and Wisconsin and private water supply sites in Georgia, Minnesota, New Jersey, New Mexico, and Wisconsin. Davidson s work was performed in response to the following questions: Does MTBE cause significant co-solvency of aromatic hydrocarbons? Does MTBE inhibit natural aerobic biodegradation? Does MTBE accumulate in groundwater at concentrations greater than source-area concentrations? Davidson found that MTBE did not act as a co-solvent, did not inhibit natural biodegradation, and did not accumulate at concentrations greater than source-area concentrations. In addition, Davidson reported that at 30 MTBE-blended gasoline spill sites studied, MTBE plumes (on average) were 1.5 to 2.0 times larger than the associated benzene plumes. Study by Happel et al. (1998) Happel et al. (1998) performed a statistical analysis of MTBE and benzene data from two groups of California LUFT sites: 63 sites statewide and 29 sites in San Diego County. The San Diego data set included time series sampling data collected between 1992 and The statewide data set included the results of one sampling round. Based on the statewide (single sampling round) data set, Happel et al. indicate that the spatial extent and concentrations of MTBE and benzene plumes were not well correlated. In contrast to previous studies, however, the benzene and MTBE plume lengths (based on 10- and 20-µg/L contours, respectively) were similar. Based on 43 sites with coincident MTBE and benzene plumes, 90 percent of the MTBE plumes were less than 325 feet in length while 90 percent of the benzene plumes were less than 400 feet long. Happel et al. suggested that the similarity in plume lengths may be a function of release age and histories, rather than a consequence of the fate and transport characteristics of MTBE. Based on a statistical analysis of the co-occurrence of MTBE and benzene in the time-series data set from San Diego, the MTBE and benzene plumes at the 29 sites in San Diego appeared to disassociate over time, with MTBE plumes apparently moving away from the monitoring networks. Benzene concentrations decreased significantly relative to the change in MTBE concentrations during the assessment period. Study by Mace and Choi (1998) Mace and Choi (1998) returned to the data set used for the 1997 benzene plume study to examine MTBE concentration data at 609 leaking petroleum storage tank sites located in Texas. The purpose of this 1998 study was to quantify the temporal and spatial variability of MTBE plumes in Texas and to investigate whether MTBE plumes had naturally attenuated. Plume lengths were estimated based on pre-1997 data for MTBE at 89 sites and for benzene 12

33 at 289 sites. The change in MTBE plume lengths through time was studied at 20 sites for which plume lengths were estimated and for which time series data were available. Timeseries data sets from 471 wells with at least four MTBE monitoring events were each reviewed to determine whether MTBE concentrations were stable or changing with time. MTBE plume lengths (defined by the 10-µg/L isopleth) were found to have a geometric mean of 182 feet. Benzene plume lengths (defined by the 10-µg/L isopleth) were found to have a geometric mean of 155 feet. At several sites, the length of the MTBE plumes was found to decrease with time. Data from 50 percent of the 471 wells having time-series data indicated stable MTBE concentrations. Nine percent of the time-series data sets indicated decreasing concentrations and 7 percent of the time-series data sets indicated increasing concentrations. MTBE was not detected at the remaining 34 percent of the wells. Mace and Choi concluded that the apparent stabilization of MTBE concentrations and plume lengths, along with the generally similar mean MTBE and benzene plume lengths, constituted evidence of the natural attenuation of MTBE. While the natural attenuation of MTBE plumes may occur, the data set that was examined by Mace and Choi to understand the dynamic nature of MTBE plumes contains some apparent deficiencies. First, the data examined for this plume study were collected for the 1997 benzene plume study only a few years after MTBE was added to gasoline in high volume percentages (greater than 3 percent); thus, while the benzene plumes likely represented a range of new and older releases, most of the MTBE plumes were expected to be young releases, dated past This confounding factor may have made some of the examined MTBE plumes appear much shorter and more stable than they actually were. Second, some sites were included in the numerical analysis that clearly mixed plume-length data from older gasoline spills (without MTBE) with data from newer gasoline spills (with MTBE). As a result, the plume length conclusions from this report should be interpreted cautiously. Study by Integrated Science & Technology, Inc. (1999) Integrated Science & Technology, Inc. (IST) reviewed fuel release information from 149 BP gas stations in Florida (1999). Eighty of these sites were selected for detailed study based on the period of groundwater monitoring, the presence of MTBE and benzene, the absence of phase-separated hydrocarbons, and a minimum number (three) of monitoring wells with detectable MTBE. The study objectives included the following: Comparing MTBE and benzene plume lengths and plume areas. Assessing hydrogeological controls. Determining matrix-chemical interactions and intrinsic controls on plume behavior. Deriving empirical attenuation rates. 13

34 IST established plume lengths and areas based on hand contouring MTBE and benzene isopleths on site maps. The mean detection frequency and concentration of MTBE and benzene were similar (38.9 percent and 317-µg/L MTBE, and 42.8 percent and 365-µg/L benzene). The mean lengths of the 10-µg/L MTBE and benzene plumes were 140 and 115 feet, respectively. The mean areas of the 10-µg/L MTBE and benzene plumes were 11,985 square feet and 7,910 square feet, respectively. Based on well-specific time-series analysis, IST concluded that 89 percent of the MTBE plumes studied are decreasing in size. IST indicated that the average extent of MTBE migration (in comparison to benzene) observed at Florida BP gas stations is less than that reported in previous MTBE plume studies completed outside of Florida. Empirically derived mean attenuation-rate constants for MTBE (at 137 wells) and benzene (at 146 wells) were very similar. IST concluded that similarity in the rate constants could reflect related biological controls on each compound; however, there is a biological preference for the mono-aromatic benzene molecule over the ether MTBE molecule, leading to a competitive disadvantage for MTBE biodegradation. IST hypothesizes that the decay of Florida MTBE plumes lags behind the decay of benzene plumes because benzene is the preferential carbon source for most microorganisms. IST hypothesized that MTBE biodegradation does not begin until MTBE has migrated beyond the benzene plume or until the benzene has become depleted and MTBE is the primary carbon source. Hydrocarbon plume behavior in Florida is influenced by the subtropical climate, the nearly universal flat hydraulic gradients, shallow water tables, high rates of infiltration and recharge, and moderately to highly permeable shallow aquifers. In comparison to other regions in the United States, these conditions promote lower groundwater flow rates, more rapid flushing and groundwater oxygenation, and higher volatilization rates. 2.5 Natural Attenuation Mechanisms for MTBE Relative to Benzene Remediation by natural attenuation is achieved when natural attenuation mechanisms (including biodegradation, volatilization, sorption, advection/diffusion, and dilution) reduce the total mass of a contaminant, its concentration, or its mobility in the environment. As discussed earlier, MTBE is not readily retarded by aquifer soil matrices. In addition, MTBE does not easily volatilize from groundwater. Finally, MTBE is more resistant to microbial attack than benzene and other gasoline aromatic components; thus, for a given release of MTBE-blended gasoline, MTBE can be expected to migrate more rapidly and farther in distance than other hydrocarbons and is, therefore, more likely to impact larger volumes of the saturated zone. A number of studies have evaluated the fate of MTBE relative to benzene in subsurface environments, as discussed in detail in Section Results from a number of field studies 14

35 differed between one site and another primarily due to variations in site conditions. A limited number of field investigations indicates that the degradation of MTBE in the environment does occur, albeit slowly. The site s history of exposure to petroleum hydrocarbons can greatly affect MTBE bioattenuation rates. The natural attenuation of an MTBE release at a site could potentially proceed at accelerated rates if indigenous microbial populations at the site have had prior exposure to MTBE and have, thereby, developed the mechanisms to derive energy from MTBE biodegradation. This process is most often referred to as microbial acclimation or adaptation. Natural attenuation is less effective for MTBE relative to benzene; however, based on reported MTBE and benzene plume studies, it appears that MTBE has the potential to naturally attenuate. MTBE attenuation rates appear to be site-dependent and can be expected to vary with time and position across a single plume. In addition, bioattenuation rates can be greatly influenced by the petroleum release history at the site. 2.6 Conclusions on MTBE Fate and Transport The fate and transport of BTEX and MTBE in the subsurface can be accurately evaluated based on knowledge of the chemical, physical, and biological properties of these chemicals, as well as laboratory-scale and field studies of their behavior in various geologic and hydrogeological settings. The following conclusions and implications for selecting appropriate remedial strategies for MTBE-impacted sites can be reached: MTBE s pure component solubility (approximately 50,000 mg/l at 25 C) and mole fraction in gasoline (3 to 15 percent) are much higher than those of benzene (1,780 mg/l and greater than 1 percent, respectively). This can result in higher concentrations of MTBE in the source area relative to benzene. A rapid response to LUFTs with MTBE-containing gasoline is essential to minimize remediation costs and costs associated with off-site impacts. Rapid responses can also reduce or avoid potential impacts to community water systems. MTBE concentrations in groundwater have been observed at significantly lower levels than predicted using equilibrium relationships, indicating that conventional assumptions of instantaneous equilibrium are inappropriate; thus, residual gasoline containing MTBE is likely to remain a long-term source of MTBE to groundwater, and more a complete remediation of the residual source may be required compared to BTEX-only LUFT sites. MTBE has a much higher vapor pressure than BTEX compounds, which indicates that vapor extraction technologies will be effective in removing MTBE from soils with low to moderate soil moisture levels. Because MTBE is highly water-soluble, the removal of MTBE from soils with high moisture content could be more difficult. MTBE exhibits low sorption to soil materials and exists predominantly in the dissolved phase in aquifers. This indicates that most MTBE plumes could move 15

36 approximately at the rate of groundwater. It also means that groundwater flushing technologies should be effective for MTBE removal from hydraulically accessible soil strata. In addition, air sparging is likely to be effective for MTBE remediation because of low MTBE adsorption to aquifer materials. Evidence for the natural degradation of MTBE in subsurface environments based on plume studies is weak; however, some studies suggest that at sites with relatively low groundwater velocities, there is a strong potential for biodegradation once MTBE has migrated past anaerobic source areas. MTBE plumes could migrate past or disassociate from BTEX plumes given MTBE s higher mobility and lower tendency to naturally biodegrade compared to BTEX compounds. MTBE has a low K ow, which suggests that it is not likely to bioaccumulate in plants and other organisms. MTBE has a lower Henry s constant than BTEX, which indicates that using air stripping to remove MTBE from water will require a greater air-to-water ratio to achieve similar removal efficiencies. Because of lower adsorption onto soil and other solids, adsorption technologies may be less efficient for removing MTBE from water compared to BTEX removal. 16

37 3. Remedial Strategies 3.1 Introduction Remedial strategies for cleaning up gasoline release sites are highly site-dependent. Numerous factors play a significant role in developing remediation approaches. The most dominant factors include: The nature and extent of contamination. The nature of the hazards represented by the release. The potential or actual risks to receptors (particularly public water supply wells). Local or state regulatory requirements. Subsurface conditions. Present and future land use. Ownership of the facility where the release(s) occurred. Various federal, state, and local agencies have developed a framework for establishing an acceptable remedial cleanup strategy at these sites. The remedial strategy can be driven entirely by regulatory requirements. In other cases, remedial strategies acceptable to regulatory agencies can be developed using a risk-based approach for assessing the risks to human health and the environment at the site, as well as the need, if any, for corrective action. Remedial strategies at gasoline release sites are relatively well established. Numerous guidance documents have been prepared by various regulatory agencies and industrial groups (American Petroleum Institute, 1996); however, the widespread use of oxygenates, primarily MTBE, has caused a re-evaluation of the remediation approaches used at these sites because of the unique properties of oxygenates. Specific issues that must be reevaluated include the impact of MTBE on site investigation requirements, the appropriate cleanup levels, and the applicability of the risk-based corrective action approach developed by the American Society for Testing and Materials for application at petroleum release sites impacted with MTBE (American Society for Testing and Materials, 1994). Generally speaking, risk-based corrective action is based primarily on human health risks rather than ecological risks. The objective of this section is to provide information on the development of remedial strategies at MTBE-containing gasoline release sites with the primary focus on releases at fueling stations with LUFTs. Other release scenarios of interest affected by the presence of MTBE include releases from pipelines and/or spills at refineries or fuel terminals. A broad range of remediation challenges can arise at LUFT sites following the detection of MTBE and other ethers in the soil and groundwater at the sites. In groundwater plumes, 17

38 MTBE concentrations are typically less than 1 mg/l; however, this level is significantly higher than the California secondary maximum contaminant level of 5 µg/l. In addition, MTBE may be present in the residual product at levels as high as 15 percent by volume. Soil concentrations can be as high as 100,000 milligrams per kilogram. Concentrations in the soil are dependent on the magnitude of the release and the time that has passed since the release or releases occurred. 3.2 Site Characterization The two key activities required for an adequate definition of contamination at petroleum release sites are as follows: (1) a site investigation and (2) the development of a site conceptual model. Each of these issues is discussed briefly Site Investigations Over the past 20 years, many thousands of site investigations have been completed at petroleum release sites throughout the United States. Based on this experience, numerous guidance documents have been prepared that describe in detail the appropriate approach and tools needed to adequately characterize the nature and extent of contamination at petroleum release sites (American Petroleum Institute, 1996). In general, these guidance documents are also applicable to MTBE sites, although there are some specific issues that must be addressed when MTBE is present. The first issue is the use of appropriate analytical techniques for detecting MTBE and other oxygenates in soils and groundwater. Table 3-1 summarizes the approved analytical methods for soil and groundwater to quantify the presence of MTBE and other oxygenates in the respective media. Table 3-1 Description and Applications of Analytical Methods for Fuel Oxygenates Method EPA 8020A/21B EPA 8240B/60B ASTM D4815 Process/Detector Purge and trap, GC-PID Purge and trap, GC-MS Purge and trap, GC-FID Specificity Retention time/elution Mass-to-charge Retention time/elution TPH Yes Yes Yes $/Sample $35+ $100 to $300 $100+ Notes Not for tertiary butyl alcohol. Good for all oxygenates. Excellent for all oxygenates, Heated purge reduces Heated purge improves including methanol and signal-to-noise. tertiary butyl alcohol sensitivity. ethanol, but not for BTEX Reliable for ethers when Detection limit of oxygenates or TPH. Highly customized TPH is <1,000 µg/l. in non-aqueous phase liquid method. limited to ~2,000 mg/l. EPA = United State Environmental Protection Agency. ASTM = American Society for Testing and Materials. PID = Photoionization detector. TPH = Total petroleum hydrocarbons. FID = Flame ionization detector. 18

39 A significant issue with respect to selecting the appropriate analytical technique for MTBE is the detection limit in the presence of other petroleum hydrocarbons and the potential for false positives or negatives. This issue has been addressed by a number of laboratory investigations. A second issue for MTBE-impacted sites is the need for characterizing the vertical distribution of MTBE in preferential pathways. Because MTBE is relatively recalcitrant to biodegradation and is highly soluble, it persists longer in groundwater than other soluble constituents of gasoline. As a result, there is a greater potential for the vertical migration of MTBE at LUFT sites. Consequently, depth profiles to characterize MTBE distribution are required, which result in more extensive site investigations. More effort is also required to identify preferential pathways because small but significant MTBE plumes moving in these pathways could be missed by less intensive site investigation activities. A third issue regarding the characterization of MTBE sites is the presence of co-contaminants and/or biological breakdown products of MTBE. One compound of significance is tertiary butyl alcohol (TBA), an oxygenate in itself, which can also be found as an impurity in manufactured MTBE. In addition, TBA is a biodegradation byproduct of MTBE. TBA requires a different analytical method than MTBE. Recently, the American Petroleum Institute published an extensive document describing various site investigation techniques and their applicability to MTBE sites (American Petroleum Institute, 2000) Development of a Site Conceptual Model A site conceptual model is a key tool used to summarize the nature and extent of the problem(s) at a LUFT site and to describe the potential risks due to the release of MTBEcontaining gasoline. The focus of a site conceptual model is to define the sources of contamination at the site, specify the potential or known migration pathways, and assess the likely receptors of contamination should no corrective actions be undertaken. Various guidance documents prepared by the United States Environmental Protection Agency or state agencies describe the essential components of a site conceptual model. The key uses of a site conceptual model include the following: Identifying the source and distribution of chemicals and how this distribution has changed over time. Defining the pathways for migration. Specifying the potential current and future receptors, as well as other environmental issues that must be addressed at the site. 3.3 Establishing Cleanup Levels Remedial strategies at MTBE-impacted sites should be based on an accurate site conceptual model and a clear definition of remediation objectives. Establishing remediation objectives or 19

40 cleanup levels can be a complex and lengthy process; however, industry experience over the past 15 years at LUFT sites provides an adequate framework and foundation for establishing remediation objectives that are acceptable to all stakeholders Regulatory Framework Cleanup objectives for soil and groundwater are usually driven by the potential or actual use of groundwater. For example, soil cleanup levels are established based on the potential threat to groundwater quality due to MTBE leaching during groundwater recharge. Groundwater cleanup levels are based on classification of the aquifer or by a definition of whether the aquifer is considered a potential source of drinking water. Typically, for groundwater aquifers that are considered potential or actual sources of drinking water, cleanup standards for regulated chemicals are based on drinking-water standards. In the case of MTBE, there are currently no federal drinking-water standards, but some states have established both a secondary and primary standard. For example, in California, the secondary maximum contaminant level is 5 µg/l and the primary maximum contaminant level is 13 µg/l. Table 3-2 provides a summary of soil and groundwater cleanup objectives for states throughout the United States. These objectives are subject to change, however, and this table is included for illustrative purposes only. A key regulatory driver for setting cleanup standards is the definition of what constitutes an actual or potential source of drinking water. State and federal governments vary in their definition of what constitutes a potential source of drinking water. Typically, the definition is based on a specific level of total dissolved solids, the potential yield from the aquifer, and the likelihood of the water being used as a source of drinking water. Setting site-specific cleanup objectives to meet regulatory requirements generally requires close communication and/or negotiation with regulatory agencies. Regulatory language usually sets very conservative standards in the initial phases of site evaluation, and achieving these standards is subject to an evaluation of technical and economic feasibility. Typical language describes the goal as reaching the lowest concentrations that are technologically and economically achievable Risk-Based Corrective Action Because of the wide range of site-specific cleanup levels, various industry groups have proposed alternative approaches to establishing cleanup levels for corrective action at petroleum release sites. The best known decision process is described in a publication prepared by the American Society for Testing and Materials (1994). The guide, Emergency Standard Guide for Risk-Based Corrective Action Applied at Petroleum Release Sites, has been widely used across the United States to establish the need for and extent of corrective actions at petroleum release sites. MTBE is easily incorporated into the risk-based corrective action decision process. Key features of risk-based corrective action include a tiered evaluation, which 20

41 Table 3-2 State Cleanup Levels for MTBE Soil and Groundwater Remediation as of 1999 a Soil Cleanup Standards Groundwater Cleanup Standards State Action Level Cleanup Goal Action Level Cleanup Goal Alabama Site-specific Risk-based 20 µg/l Site-specific California Arkansas, Colorado, NA c NA Site-specific Site-specific Virginia Delaware 0.15 mg/l 0.15 mg/l 0.18 mg/l b NA Florida NA 350/6,100 mg/l c NA 35/350 µg/l c Idaho, Illinois, Site-specific Site-specific Site-specific Site-specific Pennsylvania Kansas NA NA NA 20 µg/l Maine NA NA 20 µg/l 50 µg/l Maryland >Background Site-specific or 10 mg/l 10 µg/l NA Massachusetts NA 0.3/200 mg/l c NA 70 µg/l/50 mg/l c Michigan 4,800 mg/l 4,800 mg/l NA 240/690 µg/l c Minnesota 40 mg/l Site-specific NA NA Missouri NA NA NA 40/400 µg/l c Indiana, Nevada, Arizona, Mississippi, NA NA Georgia U.S. Environmental Protection Agency maximum contaminant level/ driver for state standard New Hampshire 3 mg/l 3 mg/l NA 70 µg/l New Jersey NA NA NA 70 µg/l New Mexico, NA NA 0.1 mg/l 0.1 mg/l Connecticut (New Mexico) New York 1 mg/l Site-specific 50 µg/l 10 µg/l or Site-specific North Carolina NA Risk-based NA 200 µg/l Oregon NA NA NA 180 µg/l Rhode Island NA 390/10,000 mg/l c NA 0.04/5 mg/l c South Carolina, NA NA 40 µg/l Site-specific Vermont Utah NA NA NA 200 µg/l b Wisconsin NA NA NA 12 µg/l Wyoming NA NA >200 µg/l 200 µg/l a Simmons et al., 1998; Simmons et al., b Tier 1 screening level. c Potable/non-potable. NA = Not applicable or not available. 21

42 compares contaminant concentrations detected to screening levels for the chemicals of concern (Tier 1). If the levels exceed these screening levels, site-specific target levels are developed based on human health risk and fate and transport considerations (Tier 2). Ecological risks can also be incorporated into the analysis. If the contaminant levels exceed the site-specific target levels, corrective action is required (Tier 3). Remedial strategies based on risk-based corrective action are widely used at LUFT sites; however, some limitations have been encountered because of regulatory and stakeholder concerns over limitations in the risk assessment process. As a result, some states have rejected site-specific target levels established using the risk-based corrective action process Evaluation of Mass Flux and Receptor Impacts Einarson and Mackay (2001) describe a technique to evaluate and predict impacts to downgradient receptors based on estimating the mass flux or discharge rate from a release site. The analysis can be used to define remediation goals, prioritize sites to be remediated, and assess progress in protecting water resources. The general example presented by Einarson and Mackay portrays a source area, dissolvedphase plume, and impacted downgradient water supply well. Equation 3-1 can be used to obtain a rough estimate of the maximum potential steady-state concentration of the contaminant in the water supply well, C sw. C sw = M d /Q sw (Equation 3-1) where: C sw = Maximum potential steady-state concentration of the contaminant in the water supply well. M d = Mass discharge rate of contaminant from the release site. Q sw = Pumping rate from the supply well. Equation 3-1 will only roughly estimate the maximum possible concentration (i.e., if all of the released contaminant flowed to the well) and if steady-state conditions apply. These five conditions include the following: The aquifer flow field is constant in rate and direction. The release rate of dissolved contaminants from the source zone is constant. Mass transfer processes within the saturated zone (e.g., sorption or diffusion) are at mass equilibrium. Mass transfer from the saturated zone (e.g., volatilization, transpiration) is negligible. Biotic or abiotic mass destruction is not occurring. 22

43 This simple relationship can also be a useful tool for predicting if an existing release that has yet to impact a downgradient receptor will, at any point, impact that receptor. It can also be used to estimate the probable maximum concentration of the contaminant at the downgradient well. It provides a simple analysis of worst-case scenarios resulting from a potential release. The calculation of C sw also provides a basis of comparison to actual impact data should some of the conditions not hold for a given site. 3.4 Technology Selection Once it has been concluded that corrective action is required at a petroleum release site impacted with MTBE, the next stage is identifying and selecting applicable and appropriate technologies to achieve the cleanup goals. In this document, both established (Chapter 4) and emerging (Chapter 5) technologies for remediating soil and groundwater contaminated with MTBE are discussed. In this section, a brief overview is provided on: Applicable technologies. Factors to consider in technological selection. Approaches to optimizing the selection of technologies to minimize lifecycle costs Applicable Technologies A wide range of technologies is available to address soil and groundwater cleanup at MTBEimpacted sites. All of the technologies traditionally used for cleanup at LUFT sites can also be applied to MTBE-impacted sites; however, the unique properties of MTBE must be taken into account in assessing the suitability of these technologies for meeting cleanup objectives. Figure 3-1 provides a flowchart of applicable MTBE remediation technologies for addressing the remediation of source areas or groundwater plumes. Evaluations of some of the key technologies in Figure 3-1 are presented in Chapters 4 and Factors to Consider for Technology Selection Technology selection depends upon site conditions and a number of other factors. In some cases, physical constraints at a site limit the type of technologies that can be applied. For example, packed tower air strippers are effective at removing MTBE, but their use is limited by aesthetic concerns if a LUFT is surrounded by residences or other sensitive receptors, such as schools or hospitals. A shallow tray air stripper would be a more appropriate choice despite the fact that packed towers are often more cost effective. Economic issues are also a significant factor to consider. There is often a trade-off between capital and operation and maintenance (O&M) costs. It is important to assess the lifecycle costs of any remediation strategy. Groundwater extraction systems may remove MTBE more quickly than other constituents. This may reduce the lifecycle treatment costs relative to the treatment of other constituents that are more difficult to remove from the subsurface. 23

44 What Is the Remedial Objective? Remediate Source Area (Groundwater, Soil, and Free Product) Remediate Groundwater Plume Groundwater: Pump-and-Treat Air Sparging MPE Bioremediation Clinical Oxidation Soil and Free Product: Pump-and-Treat MPE SVE Bioventing Thermal Processes Soil Excavation Groundwater: Pump-and-Treat Air Sparging Bioremediation Chemical Oxidation MPE Phytoremediation Above Ground Air and Off-Gas Treatment Ex Situ Water Treatment: Air Stripping Granular Activated Charcoal (GAC) AOPs Synthetic Resins Biotreatment Membranes Above Ground Soils Treatment Figure 3-1. Flow chart of applicable MTBE remediation technologies. Process reliability is also a key factor in selecting technologies for MTBE cleanup. Typically, a pump-and-treat system consisting of groundwater extraction followed by a physical treatment technology can be maintained at a 95-percent operations level without significant downtime. On the other hand, a biological treatment process is inherently less reliable since living organisms are harder to control than mechanically based systems. 24

45 Another cost-related issue is the schedule. In some circumstances, rapid cleanup may be one remediation goal. If the owner is interested in transferring the property and wishes to remove contamination quickly when it is technically feasible, a more aggressive remediation program may be appropriate. In addition to considering the cleanup duration, site owners must also account for potential liability concerns. Because of the more rapid migration of MTBE from LUFT sites, rapid responses to petroleum releases at MTBE-impacted sites are advisable. The longer remediation is delayed when needed, the more expensive and difficult it becomes to execute. Other factors to consider include mass reduction, risk reduction, constructability/implementability, safety, and long-term and short-term effectiveness Optimizing Technical Applications Selecting a sequence of treatment processes is often necessary at MTBE-impacted sites. For example, at a site with significant residual product in the saturated zone, MTBE concentrations in extracted groundwater may be as high as 400 mg/l. If the discharge standard is as low as 5 µg/l, several treatment processes in series would be needed to achieve the target cleanup level. In deciding whether to use single, sequenced, or multiple technologies at a site, the same criteria discussed previously needs to be used to evaluate each possible technology, each set of sequenced technologies, and each multiple technology. The application of standard process engineering principles could provide a sound basis for optimizing remediation systems from both construction and operation perspectives Source-Area Issues Whenever discussing the possible remediation of a gasoline-contaminated site, it is critical to consider the petroleum source area. Years of petroleum hydrocarbon remediation experience have shown that it is inappropriate to treat only the mobile dissolved-phase contamination without considering the impacts of the source area and contamination. When left untreated, residual hydrocarbons in the soil can leach dissolved-phase compounds for years, decades, or longer, depending upon soil types and their degree of sorting. As discussed previously in Section 2.3, although MTBE has the potential to preferentially leach from residual LNAPL under ideal conditions, the process can be very slow in reality (Rixey and Joshi, 2000). When reviewing the subsequent sections of this report, and when designing any MTBE remediation scheme, the reader is urged to consider the impacts of this potential long-term source of MTBE. Quite often, the best remediation scheme may be one that combines technologies to address both residual and mobile gasoline constituents. In recent years, some workers have considered how much source-area remediation is necessary to adequately reduce health and environmental risk from a contaminant source area (American Society for Testing and Materials, 1998). Considering this multi-faceted 25

46 question is beyond the scope of this report. With regards to MTBE source areas, recent theoretical work has suggested that while the groundwater and soil in source areas may contain high concentrations of MTBE, the actual MTBE mass in transit is quite limited (Einarson and Mackay, 2001). If the mass flux of MTBE across a given subsurface plane is limited, and the volume of water extracted is large, actual MTBE concentrations in extracted groundwater may be quite small, especially if the groundwater extraction point is far from the source area. This means that while the source-area concentration of MTBE may be high (up to hundreds of thousands of parts per billion), the concentration of MTBE in groundwater withdrawn by a public well some distance downgradient may be quite low (tens of parts per billion). Certainly, more work is needed on this topic; however, the reader is reminded to carefully consider the impacts of residual gasoline when considering MTBE remediation options. 3.5 Conclusions In summary, traditional approaches for site investigations and corrective action at petroleum release sites are directly applicable to MTBE-impacted sites as well; however, there are some issues that must be addressed that are unique to MTBE-impacted sites. The following key conclusions can be drawn for developing remedial strategies for an MTBE-impacted site: MTBE-impacted sites could require a more extensive site investigation than BTEXonly sites due to the migration of MTBE beyond BTEX plumes and the need for the vertical characterization of MTBE distribution. The risk-based corrective action approach for establishing cleanup standards is applicable to MTBE-impacted sites. In general, there could likely be a need for greater residual product removal at MTBE sites compared to BTEX-only sites due to the potential for continuous slow releases of MTBE from the residual product. In general, MTBE sites require more active remediation compared to BTEX-only sites. Technologies applicable to BTEX-only sites are also suitable for MTBE-impacted sites; however, suitable remedial approach may be required to be more aggressive and the costs of these technologies may be higher at MTBE-impacted sites. Significant challenges exist for selecting the appropriate combination of technologies to achieve optimum cleanup. The definition of optimum is highly site-specific. Tools are available, however, to develop an optimum technical solution for removing MTBE from contaminated media. 26

47 4. Remediation Technologies This section evaluates a variety of technologies that have been demonstrated to be partially or fully effective for remediating MTBE-impacted sites. Remediation technologies can be categorized in a variety of ways. For example, technologies that can treat contamination without extracting contaminated media from the subsurface are referred to as in situ technologies while technologies that can extract contaminated media and treat it aboveground are referred to as ex situ treatment technologies. This section will discuss both ex situ (e.g., pump-and-treat/groundwater extraction) and in situ (i.e., SVE, MPE, air sparging, in situ chemical oxidation, in situ bioremediation, and natural attenuation) remediation technologies. The following are presented for each of these technologies: The effects of the contaminant and site characteristics. The predicted relative effectiveness for MTBE. Summaries of case studies. 4.1 Pump-and-Treat/Groundwater Extraction Description In conventional pump-and-treat systems, contaminated groundwater is extracted from the subsurface and transported to a treatment system located aboveground; however, the term pump-and-treat can be used to refer more generally to any remediation scheme that involves withdrawal from (and sometimes injection into) groundwater (United States Environmental Protection Agency, 1996). Pump-and-treat systems are primarily used for (1) hydraulic containment to prevent the continued expansion of the contaminated zone and/or (2) groundwater extraction and treatment to remove contaminant mass and, thus, reduce dissolved contaminant concentrations in groundwater to achieve compliance with cleanup standards (United States Environmental Protection Agency, 1996). Once extracted, contaminated groundwater can be treated using established technologies, such as air stripping and granular activated carbon (GAC), and less established technologies, such as AOPs, sorption using synthetic resins, and biological treatment. Depending on site-specific standards and conditions, the treated water is then either re-injected into the subsurface or discharged to a sewer or surface water body (National Research Council, 1994). Water reuse for industrial use is also a possibility. Pump-and-treat or groundwater extraction systems have been used for groundwater containment and remediation for over 20 years; therefore, pump-and-treat is a well-understood remediation technology (Freeze and Cherry, 1979; American Petroleum Institute, 1996; National Research Council, 1994). Although the effectiveness of pump-and-treat systems has been challenged (United States Environmental Protection Agency, 1989a,b,c; Freeze and Cherry, 1989; Travis and Doty, 1990), pump-and-treat systems remain a critical 27

48 component of remediation systems. Pump-and-treat systems have been used at approximately three-quarters of Superfund sites where groundwater is contaminated and at most sites where cleanup is required by the Resource Conservation and Recovery Act or state laws (National Research Council, 1994). Pump-and-treat systems have also been used at a number of MTBE-impacted sites since 1980 (Davidson and Parsons, 1996; Creek and Davidson, 1998). Procedures and methodologies for evaluating the effectiveness of pump-and-treat are well understood Effects of Contaminant and Site Characteristics The effectiveness of pump-and-treat systems depends on three main factors: The fraction of the contaminant present in the dissolved phase. The ability (rate) of contaminants to transfer to the aqueous phase. The degree to which the contaminated groundwater can be extracted from the subsurface. These factors are functions of both contaminant and site characteristics. Because pump-andtreat systems only remove contaminants dissolved in groundwater, it is most applicable to contaminants with high water solubility and low affinity for sorption onto soil (United States Environmental Protection Agency, 1994). Characteristics of the saturated zone, such as the type of geologic materials present, hydraulic conductivity, saturated thickness, degree of geologic heterogeneity, and porosity, can affect the ease with which groundwater can be extracted. Complex site stratigraphy can detrimentally affect the progress of pump-and-treat remediation due to poor removal from low permeability zones and zones with limited hydraulic connection to more permeable horizons (National Research Council, 1994). Contaminants with high K oc geologic materials that have high organic content in the soil matrix can limit the effectiveness of pump-and-treat systems due to sorption of the contaminant and its resulting retardation in aquifers (National Research Council, 1994). One approach for estimating the time needed for remediation with pump-and-treat systems is the use of the batch flushing model (National Research Council, 1994). The batch flushing model provides an estimate of the number of groundwater pore volumes required to flush the contaminant from the affected region. A pore volume is the total volume of water contained within the pores of the soil of the contaminated site. To remove one pore volume (and allow it to be replaced by natural flow) means to remove the equivalent volume of all water at the site (i.e., one time). If it is necessary to remove three times that equivalent volume to obtain a final concentration level, then it is said that three pore volumes must be removed. Equation 4-1, based on the batch flushing model, illustrates that as the average dissolved-phase concentration of the contaminant in the groundwater plume increases, so does 28

49 the number of pore volumes that must be removed and, thus, the time to complete remediation (Zheng et al., 1991, 1992): where: N PV = R.1n [ C o Cgoal [ (Equation 4-1) N PV = Number of pore volumes that need to be removed to achieve the cleanup goal. R = Retardation factor (defined in Section 2.3.2). C goal = Aqueous concentration cleanup goal. = Average initial aqueous concentration. C o This simplified model relies on several assumptions. The subsurface is assumed to behave as a continuous mixed flow reactor operating under the following conditions (National Research Council, 1994): No spatial variability of soil properties. No dispersion. No mass transfer limitations. Linear sorption. Local and instantaneous equilibrium. No free or residual organic liquids. From the number of pore volumes given by the batch flushing model, the length of time required for complete remediation can be estimated based on assumed or calculated extraction rates; however, because this equation represents a highly idealized situation, the equation will usually under-predict the actual time required to achieve cleanup goals. This approach is presented here primarily as a tool to compare the relative effectiveness of pumpand-treat to remove different contaminants with different values of the retardation factor or different cleanup goals. The batch flushing model accounts for aquifer properties and physicochemical properties of the contaminant through the retardation factor given Equation 4-2 (United States Environmental Protection Agency, 1994): where: R = 1+ [ K oc f oc ρ b η [ (Equation 4-2) R K oc f oc = Retardation factor. = Organic carbon-based distribution coefficient. = Fraction of organic content in the soil. 29

50 ρ b η = Bulk density of the soil. = Soil porosity. Thus, the number of pore volumes required to achieve a given cleanup goal based on contaminant concentration is directly linked to the bulk density of the soil, the fraction of organic content in the soil, soil porosity, and the contaminant s organic carbon-based distribution coefficient. The site s hydrogeology also plays a major role in determining the number of wells required to extract a given number of pore volumes from the aquifer each year. Contaminants that are trapped in low permeability zones could require either more extraction wells or a longer remediation time because flushing through these zones is limited. There are several industry-accepted models that are commonly used to simulate the complexities of fate and transport in groundwater. These models generally can be used to simulate pumping and injection scenarios in two and three dimensions. Widely used models include MT3D, RT3D, MODFLOW, and FLOWPATH Predicted Relative Effectiveness for MTBE In general, pump-and-treat is an effective technology for removing highly soluble chemicals from contaminated sites. Under favorable hydrogeological conditions, pump-and-treat is expected to successfully remove MTBE from aquifers due to MTBE s high water solubility and low retardation factor. For a gasoline release, the effective water solubility of MTBE at 11 percent by volume is approximately 5,280 mg/l, which is over 300 times that of benzene (17 mg/l at 1 percent by volume) (see Table 2-2). Based on Equation 4-1, a decrease in MTBE concentration in groundwater from 100 mg/l to 15 µg/l requires approximately 65 percent of the pore volumes needed to decrease benzene from only 10 mg/l to 15 µg/l (Table 4-1). Table 4-1 Effect of Retardation on Remediation Time Using Groundwater Extraction and Treatment Number of Remediation Remediation Retardation Pore Volumes Time Time with Correction Contaminant C o (mg/l) K oc Factor [ ] a Required b (months) c (years) d MTBE Benzene Toluene Ethylbenzene Xylenes Assumptions: a ρ b = 1.4 grams per cubic centimeter, f oc = 0.005, η = 0.4. [ ] = Does not have a unit. bc goal = 15 µg/l for all contaminants. c Pore volume residence time = 6 months per pore volume. d A correction multiplier of 6, which is based on field experience, was used to determine a more realistic remediation time. 30

51 Given appropriate source control measures, pump-and-treat systems are expected to be more effective for removing MTBE compared to BTEX compounds. Due to the properties of MTBE, however, MTBE plumes may extend greater distances than BTEX plumes; thus, the extent and capacity of the groundwater pumping system may be expanded for MTBE remediation relative to a BTEX-only system, and groundwater extraction wells may be sited downgradient of the contaminant plume s leading edge. Once contaminated groundwater has been pumped to the surface, established water treatment technologies such as air stripping and GAC can be used to remove MTBE from groundwater, provided that influent concentrations for MTBE are less than 10 mg/l. In addition, bioreactors can be used to remove MTBE from contaminated water. For high levels of MTBE (greater than 100 mg/l), a sequence of treatment processes could be required to meet low treatment targets (less than 15 µg/l). Several above-ground treatment technologies are discussed below. Emerging ex situ technologies, such as AOPs and synthetic resins, are discussed in Chapter 5. Air Stripping Air stripping is a physical process that removes volatile organic compounds from water. In general, the effectiveness of air stripping increases with a compound s Henry s constant and the operating air-to-water ratio. Because the Henry s constant for MTBE is several times lower than that for other organic compounds commonly treated through air stripping (e.g., trichloroethylene, benzene, toluene), air stripping is more difficult and more costly for MTBE than for these other compounds; however, air stripping is a proven technology that has been used successfully to remove MTBE from water. Typically, high air-to-water ratios (100-to-250) are needed to remove MTBE effectively from water. The most common design for air stripping systems is a packed tower. In addition to packed towers, other established and emerging air-stripping technologies applicable for MTBE treatment include low profile air strippers, bubble diffusion strippers, spray towers, and aspiration air strippers. Packed tower aeration has been found to be superior to other air stripping technologies from a cost perspective (California MTBE Research Partnership, 1999). Air stripping has been used to successfully remove MTBE at concentrations that are typically associated with groundwater contamination from LUFTs (California MTBE Research Partnership, 1999). An evaluation of MTBE performance data from eight case studies of low profile and packed tower air strippers showed that air stripping has the potential to be more widely used in both drinking water and remediation applications (Deeb et al., 2001b). For each of the studies evaluated, MTBE removal efficiencies exceeded 90 percent. Because state and/or local air-quality regulations sometimes require stripper off-gas treatment (GAC, thermal and catalytic oxidation, advanced oxidation), process costs can increase significantly. Granular Activated Carbon (GAC) The success of GAC in removing organic contaminants from water depends on numerous factors, including GAC characteristics (e.g., surface area, pore size distribution) and the 31

52 physical and chemical characteristics of the adsorbate. Based on the sorption isotherms currently available, coconut shell GAC appears to have better adsorption characteristics than coal-based GAC for MTBE; however, the physical and chemical characteristics of MTBE are generally considered poor for adsorption relative to other compounds routinely removed using GAC (e.g., BTEX). In particular, MTBE s high solubility and low K oc cause the compound to preferentially remain in solution rather than be adsorbed onto GAC surfaces (American Petroleum Institute, 1994). Isotherm studies of the widely used Calgon Filtrasorb 300 indicate that this particular brand of GAC is an order of magnitude less efficient for MTBE than for toluene on a mass equivalent basis (Dobbs and Cohen, 1980). The poor adsorptive characteristics of MTBE on GAC can cause early breakthrough and, thus, frequent carbon change-out requirements. Because of the low adsorption capacity of GAC for MTBE, system design may require the use of three or more vessels in series to contain the mass transfer zone and to maximize carbon usage efficiency. As such, treating MTBE-impacted water using GAC may result in higher upfront capital costs and higher O&M costs due to frequent carbon change-out relative to BTEX-only contaminated water. GAC can be used effectively to treat low flows of water from private wells with low MTBE concentrations. For treating higher concentrations of MTBE, GAC can be used as a polishing step in a treatment train following air stripping or chemical oxidation. Background water quality and co-contaminant concentrations influence MTBE removal efficiencies. High concentrations of natural organic matter and other gasoline constituents in influent water streams compete with MTBE for GAC sorption sites, thereby increasing GAC usage rates and process costs. A thorough analysis of ex situ MTBE treatment using GAC can be found in a recently completed study by the California MTBE Research Partnership (2001). Biological Treatment Alkyl ethers, such as MTBE, are relatively difficult to degrade due to the high energy required by microorganisms to cleave the ether bond and to the resistance of the branched carbon structure to microbial attack (White et al., 1996). Several early studies reported that MTBE is biologically recalcitrant under most environmental conditions (Fujiwara et al., 1984; Jensen and Arvin, 1990; Suflita and Mormile, 1993; Yeh and Novak, 1995). Recent laboratory and field studies, however, have shown that a number of bacterial and fungal cultures from various environmental sources are capable of degrading MTBE under aerobic (Deeb et al., 2000b; Stocking et al., 2000) and anaerobic (Bradley et al., 2001a,b; Finneran and Lovely, 2001; Landmeyer et al., 2001) conditions. Microorganisms can either use MTBE as a sole source of carbon and energy or can degrade it cometabolically following growth on short-chain alkanes, such as propane. 32

53 The recent isolation and characterization of pure cultures capable of degrading MTBE and TBA have allowed for the refinement of ex situ biological treatment systems. Three organisms have been instrumental in this process: Rubrivivax sp. designated PM1 (Deeb et al., 2000a, 2001a; Hanson et al., 1999). Rhodococcus sp. (Salanitro et al., 2000). Hydrogenophaga flava designated ENV735 (Steffan et al., 2001). These microorganisms are being used in both ex situ and in situ biological treatment applications. Several types of ex situ reactors can be used for MTBE treatment, including fluidized bed, porous pot, and membrane and trickling filter bioreactors. Membrane bioreactors sustain high densities of microorganisms and, therefore, are suited for treating high strength waste streams, as well as waste streams containing slowly biodegraded compounds such as MTBE. In one study, the use of a membrane bioreactor reduced MTBE from high (10 to 2,000 mg/l) to very low concentrations (40 µg/l) during normal operation (Steffan et al., 2001). Fluidized bed reactors have also shown great promise for biologically treating MTBE-contaminated water streams. In fact, the organisms PM1 and ENV735 are currently being used in commercial fluidized bed reactor applications Pump-and-Treat Case Studies The effectiveness of groundwater extraction systems for remediating MTBE-contaminated sites has been well-documented at a number of full-scale field sites. As stated previously, the properties of MTBE differ from those of BTEX compounds, which indicates that pump-andtreat will be more effective for MTBE remediation relative to BTEX; thus, it is not surprising that there are many case studies of sites that have successfully used pump-and-treat systems for the groundwater remediation of MTBE (Arulanantham et al., 1999; Creek and Davidson, 1998; Griend and Kavanaugh, Groundwater extraction was first applied at an MTBE-impacted site in 1981, where it significantly lowered contaminant concentrations in less than 2 years (McKinnon and Dyksen, 1984). Fifteen sites using pump-and-treat for MTBE remediation were described in an early American Petroleum Institute report. American Petroleum Institute (1992) also reported on six hydrocarbon remediation sites where pump-and-treat systems have been used. At most of these sites, initial influent MTBE concentrations were low (less than 300 µg/l) and decreased significantly over just a few years. These low initial concentrations suggest that most of the MTBE had been depleted from residual nonaqueous phase liquid. Similarly, Bass and Sylvia (1992) discuss a site where pumping system influent MTBE concentrations declined from 3,000 to just 50 µg/l after several years of pumping. Bass et al. (1994) also report on two additional sites in Massachusetts and New Jersey, where several years of pumping reduced MTBE concentrations from 2,600 to 2 µg/l and 15,100 to 243 µg/l, respectively. By evaluating these and other cases studies, Creek and Davidson (1998) found MTBE removal rates to be 1.4 to 7 times higher than benzene removal rates using groundwater extraction. 33

54 Ex situ treatment has been used for many years for volatile organic compounds and has been demonstrated to be effective for MTBE (California MTBE Research Partnership, 1999). In Rockaway Township, New Jersey, groundwater contaminated with several volatile organic compounds, including MTBE, was initially treated using GAC, but costs soon became prohibitive (California MTBE Research Partnership, 1999; McKinnon and Dyksen, 1984). A packed tower air stripper operating with a volumetric air-to-water ratio of 200 was then added prior to GAC. Together, the two systems (air stripping and GAC) reduced initial MTBE concentrations from approximately 96 µg/l to below 5 µg/l. In LaCrosse, Kansas, a system consisting of two packed tower air strippers operated in series (with air-to-water ratios of 175 in each tower) is being used to treat influent MTBE concentrations as high as 900 µg/l (California MTBE Research Partnership, 1999). The first air stripper typically reduces MTBE concentrations by approximately 90 percent and the second stripper consistently reduces MTBE to less than 10 µg/l. In these two cases, packed tower air strippers successfully reduced MTBE concentrations to levels suitable for drinking-water applications. Air strippers are being used to remediate MTBE-impacted sites in many locations, including Culver City (California), Somersworth (New Hampshire), and Elmira (California). In general, the volumetric air-to-water ratios required for MTBE are high (100 to 250) relative to those of other contaminants routinely treated using air strippers, such as trichloroethylene (less than 30) (California MTBE Research Partnership, 1999); however, this additional treatment system cost can be off-set by the reduced cost of extracting MTBE mass from the subsurface. Ex situ biological treatment for MTBE has been demonstrated using both suspended growth and fixed-film bioreactors (Acuna-Askar et al., 2000; Dupasquier et al., 2002; Fortin and Deshusses, 1999a,b; Hatzinger et al., 2001; Steffan et al., 2001; Vainberg et al., 2002). Experience at two sites in California and Nevada demonstrated that fixed-film bioreactors can be operated at high efficiencies and can deliver low effluent MTBE concentrations over extended periods of time (Stocking et al., 2000). Studies demonstrating MTBE removal in a suspended growth bioreactor showed that although cell yields of microorganisms were low with MTBE, the addition of co-substrates can be used to sustain high cell densities (Stocking et al., 2000). Microorganisms in this bioreactor were able to reduce MTBE concentrations from 2,400 to 1.6 mg/l, with an average removal rate of 96 percent over the course of the study Pump-and-Treat Conclusions As with any remediation approach, the efficiency of pump-and-treat remediation strategies depends on contaminant properties and site characteristics. The low sorption, low retardation, and high water solubility of MTBE make groundwater extraction more effective for removing this contaminant from the subsurface compared to BTEX compounds. Hydrogeological conditions can limit pumping rates and, therefore, hinder the effectiveness of pump-and-treat systems, but these limitations are independent of the presence of MTBE; therefore, assuming that the gasoline leak source has been stopped and that only minimal residual-phase contamination exists, dissolved-phase MTBE removal by groundwater 34

55 extraction can be a technically effective remediation approach. Established water treatment technologies, such as air stripping and GAC, have been shown to be cost-effective and reliable for removing MTBE from extracted groundwater at moderate influent levels (less than 10 mg/l). In addition, other above-ground treatment technologies, including biological treatment, show great promise for the effective removal of MTBE from water. 4.2 Soil Vapor Extraction (SVE) Description SVE is an in situ remediation technology that reduces concentrations of volatile contaminants in the unsaturated or vadose zone. With this technology, subsurface air flow which is induced through the contaminated area by vacuum pumps or blowers advectively carries vaporized contaminants to extraction wells. Extraction wells are generally screened above the water table, accounting for seasonal flow fluctuations. Extracted contaminant vapors are then either directly discharged to the atmosphere or treated using above-ground treatment systems, including vapor-phase GAC or thermal or catalytic oxidation. In some instances, SVE has an additional benefit in that it stimulates contaminant biodegradation by circulating oxygen-laden air through the contaminated zone. SVE is a proven and well-understood technology for remediating sites impacted by gasoline releases (Grasso, 1993; National Research Council, 1994; Pederson and Curtis, 1991; Suthersan, 1997; United States Environmental Protection Agency, 1995; Watts, 1998a) Effects of Contaminant and Site Characteristics In designing an SVE system, contaminant properties, hydrogeology, and contaminant distribution must be considered. Soil properties affecting SVE performance include temperature, organic carbon content, and air permeability of the contaminated media. Air permeability is affected by the heterogeneity, porosity, stratigraphy, and moisture content of the soil. Similar to groundwater extraction, the concentration of contaminants remaining in the soil is related to the number of extracted air pore volumes. It is much easier to remove soil gas pore volumes than groundwater pore volumes because induced air flow rates are generally much faster than induced groundwater flow rates; however, the short-circuiting of air flow around or through contaminated areas is a much larger issue in the design and operation of SVE systems relative to groundwater extraction systems because of inherent differences in fluid viscosities. Furthermore, the presence of underground utilities and applications at shallow sites can be additional factors that cause short-circuiting (i.e., short-circuiting to the atmosphere can occur at shallow unpaved sites). In addition, the effectiveness of SVE significantly decreases as the soil moisture content increases. As shown in Equation 4-3, when the soil moisture content (θ w ) increases, the effective contaminant soil gas-phase retardation factor (R soil ) increases (Pankow and Cherry, 1996). This is because contaminants could dissolve in the soil moisture, requiring even more air (i.e., longer air flushing times) to 35

56 re-volatilize them. The soil moisture content could increase as a result of infiltrating water or fluctuating water table elevations. As shown in Table 4-2, an increase in soil moisture content can significantly increase the time required for remediation, depending on site conditions for a given contaminant. Table 4-2 Effect of Soil Moisture Content on Retardation and MTBE Remediation Time Using SVE Soil Moisture Soil Retardation Time for Remediation Porosity Content (%) Factor [ ] a (months) b Assumptions: a ρ b = 1.4 grams per cubic centimeter; H = 0.02 (dimensionless); f oc = 0.005; K oc = 11. bc o = 1,000 µg/l; C goal = 20 µg/l; τ = 1 day per pore volume. [ ] = Does not have a unit. The soil retardation factor can be calculated using the following relationship: where: (Equation 4-3) R soil = Effective contaminant soil gas-phase retardation factor. θ w = Soil moisture content. θ a = Soil air content (air-filled porosity). H = Henry s constant. ρ b = Soil bulk density. K d = Soil/water partition coefficient. Based on Equation 4-1, Equation 4-4 predicts the estimated remediation time to achieve the cleanup goal for a given residence time of the soil vapor in the contaminated zone and initial concentration. Note that for a given residence time (typically 1 to 2 days), the remediation time increases with increasing retardation, as shown by Equation 4-4: where: R soil = 1+ T = τ R 1n θ w 1 θ a H ( C o C goal + ( ρ b K d θ a H (Equation 4-4) T = Estimated remediation time. C goal = Cleanup goal. τ _ Given residence time of soil vapor in the contaminated zone. C o = Initial contaminant concentration. 36

57 The second significant factor in the design of SVE is the radius of influence of the extraction wells, which depends on site geology, soil moisture content, and air flow rate and pressure. A more permeable soil will exhibit a larger radius of influence for each extraction well. Typical radii of influence are between 20 to 200 feet, with a maximum influence depth of approximately 20 feet below the well casing, depending on the degree of soil heterogeneity, permeability, surface seals, and well screening depth (Grasso, 1993; Johnson and Ettinger, 1997; Suthersan, 1997; Watts, 1998a;). Shallow wells in less permeable soils could result in a smaller radius of influence (Johnson and Ettinger, 1997). Similarly, shortcircuiting between the well screening and the surface could decrease the radius of influence. Preferential pathways in a heterogeneous soil could also affect the radius of influence, and contaminant removal will, therefore, be greater within these preferential pathways relative to neighboring, less permeable zones (Suthersan, 1997). The radius of influence is site-specific and must be measured for each site (Johnson et al., 1990). Once the radius of influence is determined, SVE systems can be designed with multiple extraction wells such that their respective radii of influence slightly overlap, allowing for small-scale heterogeneities in the geology and, thereby, achieving complete coverage of the contaminated area. Analysis and design guidance for SVE systems is provided by the United States Environmental Protection Agency (1994, 1997) Predicted Relative Effectiveness for MTBE If geologic conditions are amenable to SVE, the effectiveness of SVE becomes highly dependent on the chemical properties of the contaminant. The three most influential physiochemical properties are Henry s constant, vapor pressure, and water solubility (Johnson and Ettinger, 1997; Watts, 1998a). The effect of Henry s constant is discussed above in the context of soil moisture. As soil moisture increases up to saturation, the amount of MTBE that dissolves from the gasoline phase into the water phase will increase; however, when freephase and residual product are present in the unsaturated zone, MTBE is expected to remain present in high concentrations in the gasoline phase because the volume ratio of gasoline-towater in the pore spaces will be high. Consequently, the potential effectiveness of SVE for MTBE remediation is primarily a function of MTBE s vapor pressure and its volume percentage in gasoline. As shown in Table 2-1, MTBE has a much higher vapor pressure than BTEX compounds. Raoult s law of partial pressure can be used to estimate relative concentrations of gasoline constituents in the vapor phase near a gasoline spill (Table 4-3). Predictions made by applying Raoult s law indicate that MTBE should be the most concentrated gasoline constituent in the vapor phase; thus, SVE can be used to remove vapor-phase MTBE more readily than other gasoline constituents. Figure 4-1, published by the United States Environmental Protection Agency, illustrates that using SVE for MTBE remediation has a higher likelihood of success than for BTEX compounds and other potential oxygenates. This technology is most effective when MTBE is present in the unsaturated zone either as a downward migrating product or as a residual trapped nonaqueous phase liquid and, as such, is most effective for MTBE removal 37

58 Table 4-3 Estimates of BTEX and MTBE Vapor-Phase Concentrations Near an LNAPL Using Raoult s Law Vapor-Phase Concentration Volume Percent P i Near Gasoline Near Gasoline Constituent in Gasoline (mm Hg) a (g/m 3 ) b Benzene Toluene Ethylbenzene Xylenes MTBE a Raoult s law: P i = VP i Mole Fraction. Note that P i is the partial pressure of component i [mm Hg] and VP i is the vapor pressure of component i [mm Hg]). b Concentration [g/m 3 ] = P i MW/(R T). Note that MW is the molecular weight of the contaminant; R is the ideal gas law constant [m 3 mm Hg/(mol K)]; and T is temperature [K]. Vapor Pressure (mm Hg at 20 C) SVE Likelihood of Success Soil/Air Permeability Time Since Release MTBE ETBE Methanol Benzene TAME Ethanol TBA Toluene Xylene Success Very Likely Success Somewhat Likely HIGH (Gravel, (Coarse (Sand) MEDIUM (Fine Sand) Week Month Year Week Month Year Success Less Likely LOW (Clay) Week Month Year Match Point Figure 4-1. United States Environmental Protection Agency graphical representation for predicting the successfulness of SVE. Adapted from Pederson and Curtis (1991). 38

59 if system deployment occurs rapidly after a gasoline release (i.e., before MTBE leaches into groundwater) SVE Case Studies SVE is a well-understood technology commonly used for remediating gasolinecontaminated soils. Since much of the focus of MTBE-contaminated sites is on groundwater impacts, most SVE systems are implemented in combination with other technologies that address groundwater remediation. Most typically, SVE is combined with aquifer air sparging because of the similarities of system components and complementary effects. Studies by Griend and Kavanaugh (1996) and Creek and Davidson (1998) report on several sites that effectively remediated MTBE using SVE with and without groundwater treatment. In addition, there are other published and unpublished studies where SVE was used for simultaneous BTEX and MTBE removal (Arulanantham et al., 1999; Bass et al., 1994; Johnson and Ettinger, 1997) SVE Conclusions The effectiveness of SVE is highly dependent on soil and contaminant properties. Low permeability zones could impede the ability of SVE to remove contaminants; however, where site characteristics are favorable and where contaminants are present as residual or free-phase products in the vadose zone, SVE is expected to be more effective at removing MTBE relative to BTEX compounds. Once MTBE has dissolved into groundwater, other remediation technologies such as in situ air sparging are needed in combination with SVE for effective MTBE removal from soil and groundwater. The rapid application of SVE following a gasoline release is expected to be the most effective method for removing MTBE and BTEX compounds from the vadose zone. This reduces the overall time and cost of remediation. 4.3 Multi-Phase Extraction (MPE) Description MPE is a highly flexible and versatile technology that can be used for groundwater recovery, groundwater depression, SVE, bioventing (oxygen addition using above-grade mechanical systems to anaerobic subsurface environments to stimulate the microbial activity of microorganisms [United States Environmental Protection Agency, 1995]), and free-product recovery. A multiple-phase or dual-phase system applies a high vacuum (i.e., up to 28 inches of mercury or 32 feet of water) to a well for the removal of both soil vapor and groundwater (United States Environmental Protection Agency, 1997). Dewatering in the vicinity of the well allows vacuum-induced air flow to better access contaminated soils within the capillary fringe and below the previous elevation of the water table. Soil vapor and groundwater recovery rates are increased over SVE or pump-and-treat alone due to the high vacuum, with reports of up to a tenfold increase in water extraction rates (United States Environmental 39

60 Protection Agency, 1997). One advantage of MPE is that it is much more effective for freeproduct and/or residual product removal than either pump-and-treat or SVE alone. Frequently, free-product removal rates as LNAPL are comparable to skimming/pumping systems. MPE systems, however, have the added benefit of removing significantly more product in the vapor phase. MPE systems can be operated in a flexible manner to reduce remediation costs. An MPE system can be operated such that groundwater depression and recovery can be minimized and higher-purity (i.e., less water) product can be recovered through slurp tubes. Alternatively, an MPE system can be used to depress the groundwater table, expose the smear zone, and increase the depth of vadose zone, thereby allowing for the SVE of residual LNAPL and/or bioventing for remediating the vadose zone or smear zone soils. MPE is typically implemented in one of three ways: drop-tube entrainment, well-screen entrainment, or downhole-pump extraction. Drop-tube systems (Figure 4-2a) are more efficient for sites with low hydraulic conductivity (less than 10 4 centimeters per second [cm/sec]) and low groundwater yield (5 gallons per minute [gpm] or less) due to the use of a common blower for both vapor and liquid removal at the top of the well casing (United States Environmental Protection Agency, 1997). Well-screen systems (Figure 4-2b) are the simplest to implement and are most effective at sites with shallow groundwater (less than 10-feet below ground surface) (United States Environmental Protection Agency, 1997). The use of a downhole pump (Figure 4-2c) is the most efficient of the MPE systems at removing groundwater and is used at sites with higher hydraulic conductivity (greater than 10 4 cm/sec) and single well yields that are greater than 15 gpm since a separate pump is used to remove groundwater (United States Environmental Protection Agency, 1997). Vacuum-enhanced pumping significantly increases the radius of influence of pumping for dissolved-phase and free product by creating a pressure gradient toward the recovery well. a. b. c. Drop-Tube Entrainment Extraction Figure 4-2. Types of MPE systems. Well-Screen Entrainment Extraction Downhole-Pump Extraction Effects of Contaminant and Site Characteristics MPE systems are ideal for use in soils with intrinsic permeabilities of 10 1 to 10 3 Darcy and in saturated soils with hydraulic conductivities of less than 10 4 cm/sec (United States 40

61 Environmental Protection Agency, 1997). MPE can be applied to higher permeability soils, but drop-tube and well-screen entrainment systems become less cost-effective due to the increased volume of air required to maintain the vacuum for groundwater removal (United States Environmental Protection Agency, 1997). The effective air permeability, which depends in part on soil moisture content, determines the efficiency of the vapor extraction system by controlling the airflow rate. The application of the different types of MPE systems depends on the depth of the water table. In most cases, MPE systems are not effective at sites where the water table is less than 3 feet below the surface due to the short circuiting of air flow (United States Environmental Protection Agency, 1997). Downhole-pump extraction is usually more effective for remediating sites with relatively deep groundwater and is the most commonly used technique for these types of sites Predicted Relative Effectiveness for MTBE In general, the relative effectiveness of MPE for removing MTBE from soil and groundwater can be predicted by examining the relative effectiveness of pump-and-treat and SVE. As previously discussed, these technologies have been shown to be more effective at removing MTBE than BTEX compounds. As such, MPE is also expected to be effective for removing MTBE from soil and groundwater relative to BTEX compounds. The mass removal rate of MTBE using MPE is inversely proportional to soil moisture content, as illustrated in Figure 4-3. In addition, the removal of MTBE increases as a higher Mass Removal Rate (grams/day) MTBE P w = 0.60 MTBE P w = 0.85 MTBE P w = 0.95 P w = 0.60 P w = 0.85 P w = Benzene Water Porosity/Total Porosity Figure 4-3. Effect of soil moisture on mass removal rates at varying vacuum rates and P w (atmosphere). 41

62 vacuum is applied and the pressure at the wellhead decreases. The effects of increasing vacuum and decreasing soil moisture content on benzene removal rates are not as dramatic because of benzene s lower vapor pressure (see Table 2-1) MPE Case Studies Peargin (1998) identified eight UST sites in Maryland using groundwater extraction to dewater the capillary zones while SVE was used to remove nonaqueous phase liquid from smear zones. While the removal rates for MTBE were not as high as theoretically predicted, MPE was found to be at least as effective for MTBE as for BTEX compounds. In addition, there are several other reports in the literature where MPE was used successfully for MTBE remediation MPE Conclusions The combination of SVE, pump-and-treat, and biostimulation allows for the remediation of both soil and groundwater at higher efficiencies than separate applications of these systems. Furthermore, MPE has a great degree of operational flexibility. The high vacuum applied in the subsurface increases the mass removal rate from the soil by dewatering the capillary fringe and exposing more of the contaminated soil. The increased groundwater pumping rate increases the removal of dissolved-phase contaminants. MPE is highly effective for removing MTBE and BTEX compounds from the subsurface and is increasingly being used to remediate MTBE-impacted sites (Arulanantham et al., 1999). 4.4 In Situ Air Sparging Description In situ air sparging is a well-established technology for BTEX remediation (Hinchee, 1994; Suthersan, 1997). According to the United States Environmental Protection Agency, in situ air sparging was used in 1996 to remediate 13 percent of a total of 36,000 UST sites (United States Environmental Protection Agency, 1996). Traditional air sparging involves direct air injection into the saturated zone. Air bubbles contact pore water and circulate groundwater to volatilize contaminants from groundwater. The mechanism is effectively the same aeration process as that used by an aboveground aeration system, but contaminant removal is generally less efficient in the subsurface. Once extracted, soil vapors can be passively vented to the atmosphere or captured by SVE wells and treated prior to atmospheric discharge. The injection of air has the additional advantage of introducing oxygen to subsurface environments, thereby potentially enhancing aerobic contaminant biodegradation rates (Johnson, 1998). Several process variations can be used to enhance the effectiveness of in situ air sparging. For example, ozone-laden air can be injected into the aquifer chemically, oxidizing volatilized 42

63 contaminants in situ and reducing the need for vapor-phase treatment (McCulloch, 1996). Steam or heated air can also be injected into the subsurface to increase the contaminant s Henry s constant and volatility. These process enhancements will be discussed further in Chapter Effects of Contaminant and Site Characteristics The effect of site characteristics on air sparging is similar to its effect on other in situ remediation technologies. Air sparging readily removes contaminants from more permeable zones since air sparged into the aquifer tends to move through high permeability zones and through vertical and horizontal preferential pathways. As with SVE, horizontal and vertical airflow is determined by the effective porosity and grain size of the surrounding media. If vertical permeability is significantly less than horizontal permeability, vapors will tend to travel horizontally (Suthersan, 1997). In general, air will travel 1 to 2 feet horizontally for every vertical foot, resulting in a 10- to 40-foot radius of influence (Suthersan, 1997). Many researchers and practitioners have discussed in detail the impact of subsurface heterogeneity on the effectiveness of air sparging (Hinchee, 1994; Suthersan, 1997). The development of preferential pathways and short-circuiting can be a significant issue for in situ air sparging systems Predicted Relative Effectiveness for MTBE The most influential properties for determining the effectiveness of in situ air sparging for any contaminant are the contaminant s K oc, Henry s constant, solubility, and biodegradability (Johnson, 1998). As previously noted, contaminants are adsorbed onto soil within the saturated zone. During air sparging, contaminants transfer from the aqueous phase into the sparged air channels and are advected away. For strongly-adsorbed contaminants, partitioning into the aqueous phase is expected to be the rate-limiting step, resulting in longer remediation times and the potential for rebound following the cessation of sparging; however, MTBE has both a relatively low soil sorption coefficient and high solubility, which result in high dissolved-phase concentrations and low soil-phase concentrations. Consequently, MTBE adsorbed to the soil readily dissolves into the aqueous phase. While MTBE has a lower Henry s constant than BTEX compounds, it is still amenable to volatilization; therefore, air sparging will tend to volatilize MTBE from the dissolved phase and will not be limited by re-equilibration between adsorbed and dissolved MTBE. Similarly, upon cessation of sparging, MTBE concentrations will tend to rebound less than BTEX compounds due to MTBE s lower K oc. Table 4-4 illustrates the potential effectiveness of in situ air sparging for MTBE and BTEX compounds. Even though MTBE s Henry s constant is an order of magnitude lower than that of benzene, its solubility is an order of magnitude above that of benzene; thus, their equilibrium vapor-phase concentrations (although equilibrium is not typically achieved) in the sparge bubbles are expected to be of similar magnitude (Johnson, 1998). The higher the 43

64 Table 4-4 Estimates of BTEX and MTBE Equilibrium Vapor-Phase Concentrations in Sparge Bubbles Literature Approximate Predicted Equilibrium Theoretical Percent by Volume Solubility Vapor-Phase Compound Solubility (mg/l) in Gasoline (mg/l) Concentrate (ppmv) a Benzene 1, ,200 Toluene ,800 Ethylbenzene Xylenes to 1,300 MTBE 48, ,280 30,000 a ppmv = Parts per million by volume. equilibrium vapor-phase concentration of a given contaminant, the more effective in situ air sparging is for that contaminant In Situ Air Sparging Case Studies Several laboratory studies have evaluated removing MTBE from water using air sparging. To begin with, column studies using different soils showed that the rate of MTBE volatilization is very rapid in sandy soils (Mortensen et al., 2000). Results from another laboratory study showed that 85 percent of MTBE in water equilibrated with gasoline was successfully removed using air sparging (Bruce et al., 1998). MTBE recovery rates were shown to increase with increased air injection rates and water saturation levels. Results from field studies have revealed that if site hydrogeology is conducive to the successful use of in situ air sparging, this technology could be successful at removing MTBE from groundwater. For example, a paper by Bass (1996) presents 10 case studies that demonstrate the effectiveness of air sparging for remediating MTBE from groundwater. In seven of the 10 case studies, MTBE groundwater concentrations were reduced by 88 to 99 percent after 2 years of in situ air sparging. The placement of sparge wells and the rate and periodicity of sparging varied as a function of site-specific subsurface conditions. In lower permeability soils, where the likelihood of forming de-watered preferential air pathways is high, the sparging wells were pulsed to limit dewatering. Finally, minimal MTBE rebound was observed for most of these sites, as predicted by MTBE s low K oc In Situ Air Sparging Conclusions In situ air sparging is a well-established technology for BTEX remediation and relies on a combination of subsurface volatilization and enhanced biodegradation. As with most other in situ approaches, the effectiveness of air sparging is limited by site-specific subsurface conditions, such as low permeability soils. Short circuiting can be a significant issue with in situ air sparging systems. This technology is promising for MTBE remediation despite the apparent slow rate of MTBE biodegradation and low Henry s constant compared to BTEX 44

65 compounds. This is because MTBE is usually present primarily in the dissolved phase (i.e., not adsorbed to aquifer solids); therefore, if the hydrogeology of the site is amenable to air sparging for BTEX compounds, it is likely that it will also be effective for MTBE removal. Generally speaking, however, in situ air sparging is more effective for BTEX than for MTBE, mainly because of the increased BTEX biodegradation caused by the oxygenation of aquifers during air sparging. 4.5 In Situ Chemical Oxidation Processes Description The goal of chemically oxidizing contaminants is to convert contaminants into benign end products. The most effective oxidation processes are based on the production of free radicals. Free radicals are unstable, highly reactive, and generally short-lived chemical species with unstable electron states. For example, the hydroxyl radical, OH, is an uncharged species with one less electron than its much more stable form, OH-. The hydroxyl radical, therefore, has a strong tendency to extract that one electron from another chemical species, resulting in the oxidation of that species. The hydroxyl radical is a non-selective oxidant and can attack most organic compounds. The reaction of organic contaminants with OH is aggressive and rapid. Second order rate constants typically range between 10 9 to mole per second, allowing nearly instantaneous oxidation (Suthersan, 1997). Hydrogen peroxide (H 2 O 2 ), ozone (O 3 ), and potassium permanganate (KMnO 4 ) have all been used with varying success as oxidizing agents. The third strongest oxidant after fluorine and OH is ozone, which can either react with water to produce a hydroxyl radical or can react directly with the target compound. Hydrogen peroxide reacts with iron (II) or copper (II) to form Fenton s reagent, which subsequently reacts to form OH (Watts, 1998b). Table 4-5 summarizes chemical reactions that form hydroxyl radicals. Table 4-5 Production of Oxidizing Agents Reaction Oxidizing Agent Fe 2+ + H 2 O 2 Fe 3+ + OH + OH OH H 2 O 2 + O 3 3O 2 + 2OH OH 3H 2 O + O 3 6OH O 3 or OH With or without ferrous iron, hydrogen peroxide is the most commonly applied oxidant for the destruction of organic compounds, such as MTBE; however, the improper addition of hydrogen peroxide to soil and groundwater can result in highly exothermic reactions, which have caused at least two known explosions (Anderson, 1994). As the temperature of the free product rises due to reaction with an oxidant, volatile gasoline components could vaporize, creating a potentially dangerous situation if the gasoline components reach their explosion 45

66 limit. In general, in situ chemical oxidation is an effective destructive technology; however, the implementation of this technology poses some risks relative to the other technologies discussed. There are various commercially available and proprietary formulations of Fenton s reagent, which vary in aggressiveness and duration. In situ chemical oxidation strategies can be used in conjunction with in situ bioremediation. For example, as seen in Table 4-6, the reaction of hydrogen peroxide and ozone results in the production of oxygen, which will increase the dissolved oxygen content of the subsurface and, thus, stimulate aerobic biodegradation. Additionally, hydrogen peroxide/ozone injection can break down some of the more complex organics, creating molecules that are more easily degraded (Suthersan, 1997); however, due to the non-selectivity of oxidants, high concentrations of hydrogen peroxide and ozone will attack all organic matter, including microbial communities. Consequently, oxidants are often pulsed into the subsurface to achieve the benefits of increased oxygen and partial oxidation without sterilizing the local subsurface environment Effects of Contaminant and Site Characteristics For in situ chemical oxidation to be effective, the oxidizing agent must be able to contact then react with the chemicals of concern. In complex stratigraphies and low permeable zones, the effective delivery of the chemicals is limited (National Research Council, 1994). As previously noted, this limitation occurs with most remediation strategies. The effectiveness of chemical oxidation could also be limited by the presence of scavengers, which are chemical species, such as bicarbonate and carbonate ions, that readily react with oxidants, especially hydroxyl radicals (Anderson, 1994). Although the reaction rates of bicarbonate and carbonate ions with OH ( M -1 sec -1 and M -1 sec -1, respectively) are lower than those for organic compounds, the 1,000 concentrations of bicarbonates and carbonates are often more than 1,000 times greater than organic compounds. Consequently, for the target contaminant to be completely destroyed in the subsurface, concentrations of the oxidizing agent will need to be significantly greater (e.g., greater than 1,000 times) than the stoichiometric requirements suggested by the contaminant concentrations. An alternative approach to in situ chemical oxidation is not to completely destroy the organic contaminants, but to break them into more biodegradable molecules. In this case, less oxidizing additives are needed than is stoichiometrically required. Low reaction rates and the presence of other organic compounds may result in the production of intermediates or byproducts (Watts, 1998b). Since oxidizing agents will react with most organic compounds, most organic matter present in the vicinity of the application area will mineralize, which may reduce the rate of oxidation of the target contaminant (Suthersan, 1997). Fenton s chemistry is not expected to be effective in areas with high alkalinity (greater than 400 mg/l carbonate), high organic carbon content, or ph greater than 6.5 (Zogorski et al., 1997). 46

67 As mentioned, there is a danger of explosion when using in situ chemical oxidation under certain site conditions. Considering this, Fenton s reagent should not be used in the following situations (Anderson, 1994): Soil with high organic matter content or measurable free product (reactions will produce significant amounts of heat). Sites with the potential for uncontrolled vapor-phase migration, resulting in an unmanageable remediation process. Sites with USTs and pipelines near a source area that may be affected by the generation of a large amount of heat. Sites with possible uncontrolled ignition sources Predicted Relative Effectiveness for MTBE In general terms, the ability of a chemical to be oxidized can be thought of as its willingness to give up or donate electrons. MTBE has no double bonds, unlike BTEX compounds; thus, it is not as easily oxidized as BTEX compounds. Ex situ studies, however, have shown MTBE to oxidize under appropriate conditions (California MTBE Research Partnership, 1999). One concern related to the oxidation of MTBE is the production of intermediates, such as TBA, aldehydes, ketones, carboxylic acids, and acetone (Zogorski et al., 1997). These intermediates are more difficult to oxidize than MTBE; however, some of these intermediates, such as acetone, readily biodegraded. The formation of these intermediates illustrates the importance of appropriate performance monitoring during and after treatment In Situ Chemical Oxidation Case Studies The in situ chemical oxidation of MTBE has been successfully demonstrated in the field (United States Environmental Protection Agency, 1998a); however, while several vendors state that their technology is effective for remediating MTBE, in situ chemical oxidation has not been widely used. At one gasoline-impacted site, hydrogen peroxide was injected into the soil while vapors were removed using SVE during a pilot test. MTBE concentrations were reduced 80 to 98 percent in nearby monitoring wells in less than 12 hours; however, subsurface temperatures rose to about 140 F, resulting in a substantial amount of LNAPL volatilization, which could account for the MTBE disappearance. At another site, full-scale treatment of MTBEimpacted groundwater was performed by injecting hydrogen peroxide into the subsurface. Three injection events over 6 months caused substantial MTBE reductions, although groundwater sampling was limited. In another successful field-scale demonstration, hydrogen peroxide was injected at a site in New Jersey. MTBE concentrations dropped from 6,000 µg/l to below detection in less than a year (United States Environmental Protection Agency, 1998). The injection of Fenton s 47

68 reagent (hydrogen peroxide and ferrous sulfate) into groundwater at a third site in Texas reduced MTBE concentrations by 83 percent (Leethem, 2001) In Situ Chemical Oxidation Conclusions The complete destruction of a contaminant is an ideal end-point for a remediation scheme. In situ chemical oxidation relies on the use of ozone, hydrogen peroxide, potassium permanganate, or other oxidizing agents to react with the target contaminant or to form hydroxyl radicals that will attack MTBE. For organic compounds, the effectiveness of in situ chemical oxidation is limited by oxidant delivery in complex hydrogeological conditions, the effect of scavengers, and the safety of the process (i.e., violent reactions due to excess hydrogen peroxide or contaminant-infused vapor migration). In addition to the injection of ozone as an oxidant, technologies involving ozone sparging are currently being tested. Fine bubbles with a high surface-to-volume ratio are injected into the saturated zone to extract dissolved MTBE from contaminated groundwater. Ozone contained within the bubble and the thin film around the bubble is expected to react rapidly and destroy MTBE in the gas phase. One of the advantages of in situ chemical oxidation is the potential to destroy MTBE in a relatively short time frame. Technology limitations include effective oxidant delivery and the potential for byproduct formation. In situ chemical oxidation is currently considered an emerging technology for MTBE remediation due to the low number of reported successful field applications. More research is needed to identify variables that impact the effectiveness of in situ chemical oxidation for MTBE. 4.6 In Situ Bioremediation Description Bioremediation involves the use of microorganisms to either destroy or immobilize contaminants. In situ bioremediation of soil and groundwater contaminants has achieved a measure of success in both field tests and commercial-scale cleanups for a range of organic contaminants (National Research Council, 1993). At sites where the biodegradation of contaminants occurs without human intervention, the remediation process is referred to as intrinsic bioremediation, or bioattenuation. Bioattenuation is one mechanism of natural attenuation, which is later discussed in Section 4.7. When the conditions at a site require the implementation of a system to accelerate the rate of microbially mediated degradation reactions or to stimulate the activity of microorganisms by optimizing environmental conditions, the process is referred to as engineered bioremediation. Engineered bioremediation (also known as biostimulation ) typically involves the use of a system to supply oxygen or additional electron acceptors. In some cases, nutrients (such as phosphorus and nitrogen) and/or other growth-stimulating materials are also added (National Research Council, 1993). If the subsurface contaminant cannot be used by indigenous microbial communities as the sole source of carbon and energy, the addition of a primary substrate may 48

69 be needed. Finally, if the indigenous microbial community is not capable of contaminant degradation at an appreciable rate, the addition of a laboratory-enriched or genetically engineered microbial community at the site may be necessary. This type of engineered bioremediation is referred to as bioaugmentation Effects of Contaminant and Site Characteristics The local geology of the contaminated area plays a significant role in determining the success of in situ bioremediation. Site conditions, such as hydraulic conductivity and groundwater flow variations, are especially critical when implementing an engineered in situ bioremediation scheme since active manipulation of the subsurface environment is necessary. Field data have shown that the numbers of active microorganisms are generally higher in sandy transmissive soils than in clayey soils due to the increased flow of water containing the materials needed by microorganisms (Thomas et al., 1997). In addition, the contact between microorganisms and contaminants is greatest in high permeability zones (e.g., the hydraulic conductivity is greater than 10-4 cm/sec), resulting in increased contaminant biodegradation rates (Cookson, 1995). Environmental conditions, including temperature, ph, and moisture content, also affect the success of in situ bioremediation. Microbial metabolism typically accelerates with increasing temperature up to an optimum value after which cell activity sharply decreases. The optimum temperature for most subsurface microorganisms is between 20 and 40 C (Chapelle, 1992; LaGrega et al., 1994). Within 100 meters of the ground surface, soil temperatures are within 2 C of the mean annual surface temperature. This suggests that bioremediation is more likely to occur at appreciable rates in regions with temperate rather than arctic climates. In studies where the effect of temperature on the activity of MTBE-degrading microorganisms was evaluated, a 10 to 15 C increase in temperature resulted in an up to four-fold increase in the rate of MTBE degradation (Park and Cowan, 1997b; Steffan et al., 1997). Typically, however, temperature cannot be controlled in a cost-effective manner for full-scale field bioremediation applications. The optimum ph range for most microorganisms is between 6 and 8 (Chapelle, 1992; LaGrega et al., 1994). Soil ph values typically range between 5 and 9, depending on the buffering capacity of the subsurface environment (i.e., the amount of carbonate minerals present). When investigated, the activity of MTBE-degrading cultures was strongly identified as a function of ph with decreasing activities outside the neutral range (Eweis et al., 1997). Another factor that greatly affects contaminant biodegradation rates is the availability and subsequent recharge of electron acceptors once the original supply in the subsurface is depleted. Table 4-6 presents a list of byproducts from a number of microbially catalyzed oxidation-reduction reactions between hydrocarbons and electron acceptors. Oxygen produces the greatest amount of energy for microorganisms. Consequently, microorganisms preferentially use oxygen as the electron acceptor until its depletion. Microorganisms then 49

70 Table 4-6 Microbial Metabolism of Organic Matter Under Representative Aerobic and Anaerobic Conditions (Adapted from Suflita and Sewell, 1991) Process Electron Acceptor Metabolic Products Relative Potential Energy Aerobic Respiration O 2 CO 2, H 2 O High Denitrification NO 3 CO 2, N 2 Iron Reduction Fe 3+ CO 2, Fe 2+ Sulfate Reduction 2 SO 4 CO 2, H 2 S Methanogenesis CO 2 CO 2, CH 4 Low use nitrate followed by iron (III), maganese (IV), sulfate, and carbon dioxide. Groundwater recharge (e.g., infiltration of surface water) could increase oxygen concentrations, thereby causing microorganisms to shift from using a less energetically to a more energetically favorable electron acceptor; therefore, in zones with shallow contamination or in regions with significant groundwater recharge, environmental conditions are more likely amenable for successful bioremediation. Several studies have examined the effect of dissolved oxygen on the activity of MTBEdegraders. In the absence of oxygen, MTBE has been shown to be relatively resistant to microbial attack with some exceptions (Hurt et al., 1999; Yeh and Novak, 1995). In the presence of oxygen, decreasing microbial activity has been shown to be strongly correlated to reductions in dissolved oxygen concentrations. This phenomenon has been observed in laboratory studies (Park and Cowan, 1997b; Fortin and Deshusses, 1999a,b; Koenigsberg et al., 1999; Yang et al., 1998), pilot-scale bioreactors (Sun et al., 1996; Tang and Sun, 1997), and during field trials at gasoline-contaminated sites (Javanmardian and Glasser, 1997; Salanitro et al., 1999a,b) Predicted Relative Effectiveness for MTBE In general, branched alkyl ethers are relatively resistant to microbial degradation due to the difficulty associated with cleaving the ether bond (Mo et al., 1997) and the microbial attack of the branched tertiary or quaternary carbon structure (Squillace et al., 1997). Although early studies reported that MTBE is recalcitrant under both aerobic and anaerobic conditions (Fujiwara et al., 1984; Jensen and Arvin, 1990; Suflita and Mormile, 1993; Yeh, 1992; Yeh and Novak, 1995), more recent studies have revealed that several bacterial and fungal cultures from a number of environmental sources are capable of degrading MTBE either as a primary source of carbon and energy (Deeb et al., 2000b; Eweis et al., 1997; Fortin and Deshusses, 1999a,b; Hanson et al., 1999; Mo et al., 1997; Park and Cowan, 1997a,b; Salanitro et al., 1994) or cometabolically following growth on other substrates such as n- or iso-alkanes and aromatic compounds (Garnier et al., 1999; Hardison et al., 1997; Hyman et al., 1998; Koenigsberg et al., 1999; Steffan et al., 1997). Laboratory studies have reported both the partial degradation of MTBE to dead-end metabolic intermediates and its mineral- 50

71 ization to carbon dioxide. Most recently, the degradation of MTBE was shown to take place by indigenous microorganisms in laboratory microcosms and at field sites (Borden et al., 1997; Bradley et al., 1999; Church et al., 1997a,b, 1999; Salanitro et al., 1998, 1999a,b, 2000; Schirmer and Barker, 1998; Schirmer et al., 1998, 1999; Wilson et al., 2002). A detailed summary of some of the findings from the above-mentioned studies is available elsewhere (Deeb et al., 2000b; Stocking et al., 2000). The results from most laboratory and field studies suggest that bioremediation strategies involving direct metabolism, cometabolism, bioaugmentation, or some combination thereof can be used successfully at MTBE-impacted sites if conditions at the site are amenable to biological activity and oxygen/nutrient delivery In Situ Bioremediation Case Studies Currently, several field experiments are being conducted in an effort to evaluate the effectiveness of in situ bioremediation for the removal of MTBE from soil and groundwater (American Petroleum Institute, 1999; Chang, 1999; Envirogen, 1999; Salanitro et al., 1999a,b, 2000; Wilson et al., 2002). Preliminary results from a bioaugmentation study conducted at Port Hueneme, California, using a laboratory-enriched mixed culture revealed that oxygenamended biobarriers seeded with the culture were more effective than unseeded ones in degrading MTBE (Salanitro et al., 1999a,b, 2000). At this site, MTBE concentrations in groundwater decreased 90 to 99.9 percent with bioaugmentation. In an ongoing field trial at Port Hueneme, a pure bacterial culture and oxygen were injected with promising preliminary results (Chen et al., 2000). Results from pilot studies using oxygen injection at Vandenberg Air Force Base in California showed that bioremediation has a strong potential for success at MTBE-impacted sites (Wilson et al., 2002). In addition, a recent study has demonstrated the success of propane injection and bioaugmentation at an operating service station in New Jersey, where a 93-percent reduction in MTBE concentration was achieved within 2 months of system operation (Steffan et al., 2001) In Situ Bioremediation Conclusions A review of the available literature on the biodegradability of MTBE reveals that a number of cultures from diverse environments can either partially degrade or completely mineralize MTBE. MTBE is either utilized as a sole carbon and energy source or is degraded cometabolically by cultures grown on alkanes. Biological removal of MTBE has yet to be widely applied in the field, although at least three vendors are advertising laboratoryenriched MTBE-degrading microorganisms for use in both in situ and ex situ bioremediation applications (Envirogen, Environmental Resolutions, and Shell Global Solutions). The practical implementation of in situ bioremediation for MTBE is in the developing stages with more field studies needed to verify laboratory results; however, preliminary results from field studies using in situ bioaugmentation and/or oxygen injection at Port Hueneme, California, and Vandenberg Air Force Base, California, suggest that bioremediation has a strong potential for success. It is clear from these studies that the presence of oxygen is 51

72 critical for MTBE biodegradation; therefore, the addition of oxygen to MTBE-impacted aquifers could be beneficial, especially in the vicinity of gasoline release sources where rapid oxygen consumption is typically observed. In the case when indigenous microbial populations do not possess the ability to use MTBE as a primary substrate, a cometabolic substrate can be introduced to groundwater aquifers to support microbial growth and provide energy. Several compounds, including propane, iso-propanol, and n-butane, are documented cometabolic substrates for MTBE biodegradation (Hardison et al., 1997; Steffan et al., 1997). Some studies have suggested that even when indigenous MTBE-degrading cultures are present in subsurface environments, their concentrations may be too low to sustain MTBE metabolism (Salanitro et al., 1998); therefore, bioaugmentation with laboratoryenriched MTBE-degrading populations may be necessary to promote MTBE degradation at levels required to adequately mitigate contaminant migration. Based on the available studies, it is likely that a bioremediation strategy involving direct metabolism, cometabolism, bioaugmentation, or some combination thereof could be applied as a feasible and costeffective treatment method for MTBE contamination. 4.7 Natural Attenuation Description Natural attenuation encompasses several phenomena, which reduce, restrict, and/or eliminate contaminant migration in subsurface environments. The major physical, chemical, and biological processes that contribute to the natural attenuation of contaminants include the following (American Petroleum Institute, 1996): Adsorption to aquifer materials, leading to contaminant retardation. Dilution of contaminants via advection, dispersion, or diffusion. Degradation of contaminants either by abiotic means or by subsurface microorganisms under aerobic or anaerobic conditions. Contaminant loss via volatilization. The terms intrinsic remediation or remediation by natural attenuation are often used when natural attenuation is applied to remediate contaminated sites. The mechanisms responsible for the natural attenuation of petroleum hydrocarbons, especially BTEX compounds, are fairly well understood. As a result, the scientific protocols for applying natural attenuation at gasoline-impacted sites have been thoroughly delineated. Key evaluations of natural attenuation include considering if a plume is stable or shrinking and if natural attenuation can be effective at protecting off-site receptors Effects of Contaminant and Site Characteristics Geochemical conditions in the saturated zone play an important role in determining the success of natural attenuation. Because the major mechanism responsible for attenuating 52

73 petroleum hydrocarbons is most often biodegradation, subsurface conditions have to be amenable for microbial activity. In addition to favorable environmental conditions, such as temperature, ph, and moisture content, the presence of an adequate supply of electron acceptors and nutrients is required for bioattenuation rates to exceed contaminant migration rates Predicted Relative Effectiveness for MTBE Based on its unique chemical and physical properties, MTBE is neither readily retarded by aquifer materials nor easily volatilized from groundwater. Table 4-7 presents a summary of the predicted effectiveness of different natural attenuation mechanisms for MTBE relative to benzene. It is apparent that MTBE plumes are expected to migrate at faster rates than BTEX plumes, mostly because MTBE appears to be more resistant to microbial attack than BTEX compounds and because of MTBE s low retardation; thus, for a given release of MTBEblended gasoline, MTBE can be expected to move farther than other hydrocarbons and is more likely to impact larger areas of the subsurface. Because the degree of MTBE migration will be more extensive than that of other gasoline components, MTBE is more likely to encounter potential receptors, such as drinking water wells. In the absence of demonstrated microbial activity causing MTBE removal, dispersion is potentially the primary MTBE attenuating mechanism, and so the subsurface MTBE mass would not be reduced as fast as the mass of other hydrocarbons (Happel et al., 1998). The decreased attenuation potential of Table 4-7 Natural Attenuation Mechanisms and Their Effectiveness for MTBE Removal in the Subsurface Relative to Benzene Physical or Typical Typical Chemical Property MTBE Benzene Implications for Process of MTBE Value Value Natural Attenuation Dissolution Effective solubility <400 <30 MTBE concentrations higher (mg/l) than benzene concentrations in groundwater due to MTBE s higher solubility Volatilization Henry s law constant 30 a 305 a Volatilization of dissolved (from water) (atm/mol fraction) MTBE significantly slower than that of benzene Volatilization Vapor pressure 245 a 75.2 a Volatilization of MTBE (from (mm Hg) from free- or residual-phase non-aqueous at 20 C product faster than phase liquid) that of benzene Retardation Organic carbon 41 a 191 a Retardation of MTBE (by sorption) partition coefficient K oc less than that of benzene Biodegradation Aerobic Poor to fair Good Biodegradation of MTBE much slower than that Anaerobic Very poor Fair of benzene a From American Petroleum Institute (1996). 53

74 MTBE could also affect the width and shape of well-developed MTBE plumes. Since MTBE tends to be slowly attenuated, the downgradient or leading edge of MTBE plumes may be wider than that of BTEX plumes due to dispersion and diffusion, as was observed in the field by a number of researchers (Borden et al., 1997). Evaluations of the potential effectiveness of natural attenuation range from simplistic to more detailed; however, to demonstrate that natural attenuation is an acceptable remedial solution, detailed predictive groundwater modeling combined with long-term monitoring is typically required by regulatory agencies. A recent National Research Council report presents an indepth analysis of natural attenuation and describes the level of analysis likely needed to satisfy the concerns of all stakeholders when monitored natural attenuation is the recommended remedial strategy at a site (National Research Council, 2000) Natural Attenuation Case Studies A number of studies have evaluated the fate of MTBE in subsurface environments. While field results can be expected to differ between one site and another due to variations in site conditions, the outcomes from some of the studies presented here contradict each other more than expected (Table 4-8). These contradictions may be a result of MTBE s potentially variable attenuation characteristics with time or location within contaminated plumes. For example, if MTBE is cometabolically biodegraded in the presence of BTEX compounds, then MTBE biodegradation rates could be significant in the vicinity of the contaminant source, but would decrease in distance from the source as primary substrates are depleted. On the other hand, if BTEX compounds are preferentially degraded by soil microbial communities, the opposite would be observed. That is, MTBE biodegradation rates would be slowest in the vicinity of the source and fastest in the downgradient direction. Both possibilities require further laboratory and field investigations. Table 4-8 Observations from Representative MTBE Natural Attenuation Field Studies Study Field Site Location Borden et al., 1997; Sampson County, North Carolina Hurt et al., 1999; Elizabeth City, North Carolina Weaver et al., 1996, 1999; East Patchogue, New York Schirmer et al., 1998a,b; Schirmer and Barker, 1999; Borden Aquifer, Canada Field Observations Moderately fast decay rates near the source, but no decay downgradient. Low decay rates near the source, but significant mass losses downgradient. The estimated MTBE half-life was 0.1 years. No attenuation of a long plume (over 1 mile) was observed. Ninety-seven percent reduction in mass 8 years following a controlled release. Bioattenuation was assumed to be the dominant removal mechanism. 54

75 Just as the location within a plume can affect attenuation rates, so can the site s history of exposure to petroleum hydrocarbons. The natural attenuation of a secondary release at a site could potentially proceed at accelerated rates because the indigenous microbial populations at the site had prior exposure to petroleum hydrocarbons and have, thereby, developed the mechanisms to derive energy from hydrocarbon degradation reactions. This process is most often referred to as microbial acclimation or adaptation. In contrast, a past release at a site can sometimes reduce the natural attenuation rates of a secondary release. This can occur when the microbial degradation reactions responsible for attenuating a past release caused the depletion of oxygen, alternative electron acceptors, and nutrients at the site. Field research is needed to better understand the impact of a site s history on the fate and transport of MTBE in the subsurface Natural Attenuation Conclusions With the relatively recent discovery of MTBE in groundwater around the country, and with its observed persistence in subsurface environments at some sites, it is not yet clear whether natural attenuation is an acceptable option at the majority of MTBE-impacted sites. While it is clear that MTBE can biodegrade under a range of conditions, the biodegradation of MTBE is not a universal phenomenon at every site. Bioattenuation is a function of the petroleum release history at the site; thus, even if MTBE can be shown to biodegrade in aquifer samples, it is not clear whether biodegradation rates in the field are going to be rapid enough to mitigate plume migration and whether these rates can be sustained over time. Finally, the success of intrinsic biodegradation, as well as other attenuation mechanisms, is site-specific. There is some evidence to date that in certain hydrogeologic settings (flat gradients, groundwater flow rates of less than 0.1 foot per day), natural attenuation may be a feasible alternative for MTBE remediation. An estimate of the number of MTBE-impacted sites where monitored natural attenuation may be a viable remediation strategy is not available. Even if monitored natural attenuation is accepted, there are significant costs associated with monitoring and reporting requirements. Cost for these requirements typically range from $25,000 to $50,000 per year. Depending on the time value of money, these annual costs can represent a significant present value cost per site. 4.8 Overall Conclusions There are a number of technologies that have been and can be used to effectively remediate MTBE-impacted sites. This section evaluated the effectiveness of pump-and-treat or groundwater extraction followed by ex situ water treatment, SVE, MPE, in situ air sparging, in situ chemical oxidation, in situ bioremediation, and natural attenuation. Pump-and-treat, in situ air sparging, in situ chemical oxidation, in situ bioremediation, and natural attenuation are primarily applicable to the saturated zone while SVE is primarily applicable to the vadose zone. MPE is used to treat both the vadose and saturated zones. 55

76 The following conclusions can be drawn regarding the applications of the above remediation technologies to MTBE-impacted sites: The success of each of the above-discussed technologies depends largely on site hydrogeology, geochemistry, and the physical and chemical properties of the target contaminant. Due to the properties of MTBE, some of these remediation technologies are expected to be equally (if not more) effective for removing MTBE from the subsurface relative to BTEX compounds. In particular, pump-and-treat (for dissolved-phase MTBE), SVE (for separate-phase MTBE), MPE, and air sparging are expected to be relatively effective for MTBE. Approaches such as enhanced biodegradation and natural attenuation are not as effective for MTBE as for BTEX compounds. Pump-and-treat is expected to be successful in removing MTBE from groundwater due to the high solubility and low retardation of MTBE. The enhanced solubility of MTBE relative to BTEX suggests that MTBE subsurface concentrations can be reduced with fewer pore volumes of extracted groundwater. SVE is expected to be more effective in removing MTBE than BTEX compounds where site characteristics are favorable and when contaminants are present as residual or free-phase products in the vadose zone. The rapid application of SVE following a gasoline release is expected to be the most effective method for removing MTBE and BTEX compounds from the vadose zone since it reduces the overall time and cost of remediation. The combination of SVE and pump-and-treat, designated as MPE, allows for the remediation of both soil and groundwater at higher efficiencies than the separate application of each of these systems. Recent field studies suggest that MPE is effective in removing both MTBE and BTEX compounds from subsurface environments. As a result, MPE is increasingly being considered for use at MTBEimpacted sites. Because MTBE is primarily present in the dissolved phase in gasoline-contaminated aquifers, in situ air sparging is promising for MTBE remediation despite the apparent slow rate of MTBE biodegradation and its low Henry s constant. Recent field studies have shown that in situ air sparging effectively reduced MTBE concentrations over a period of 2 years at seven of 10 sites tested. Preliminary results from field studies using in situ bioaugmentation and/or oxygen delivery at Port Hueneme and Vandenberg Air Force Base suggest that bioremediation has a strong potential for success. Oxygen sources include air, pure oxygen, oxygen release compounds, hydrogen peroxide, and others. Based on the available studies, it is likely that an in situ bioremediation strategy involving direct metabolism, 56

77 cometabolism, bioaugmentation, or some combination thereof could be applied as a feasible and cost-effective treatment method for MTBE contamination. In situ chemical oxidation is currently considered an emerging technology for MTBE remediation due to the limited number of reported field applications. More research is needed to identify variables that impact the effectiveness of in situ chemical oxidation for MTBE. Natural attenuation as a remediation strategy may be less effective for MTBE relative to BTEX compounds due to MTBE s low retardation factor and slow rate of biodegradation, especially under anaerobic conditions. 57

78 58

79 5. Emerging Technologies, Techniques, and Process Enhancements Chapter 4 discussed a variety of remediation technologies that have been demonstrated to be partially or fully effective at MTBE-impacted sites. In any remediation effort, it is desirable to improve the efficiency of the process to accelerate the achievement of cleanup goals and/or to reduce costs. This chapter will discuss emerging technologies that have shown some promise, as well as techniques and process enhancements that can potentially be implemented with the remediation technologies discussed in Chapter 4 for improved efficiency. This chapter is not designed to provide an exhaustive review of emerging technologies, techniques, and process enhancements; rather, it provides an overview of the alternative options available if the reader wishes to consider non-traditional or non-established remediation technologies for an MTBE-impacted site. 5.1 Optimization of Pump-and-Treat/Groundwater Extraction Pulsed Pumping Pulsed pumping can sometimes be implemented to improve the efficiency of pump-andtreat/groundwater extraction operations (United States Environmental Protection Agency, 1996). Pulsed pumping is a means of lowering the average pumping rate of water from a contaminated aquifer to a rate closer to that of the slower release of contaminants from soils. In pulsed pumping, extraction pumps are not operated continuously, but are turned on and off at regular intervals. Pulsed pumping can increase contaminant mass removal where mass transfer limitations impede source reduction via the dissolution of contaminants into groundwater. During intervals when the extraction pumps are off, the dissolved contaminant concentration is allowed to increase as the contaminant diffuses from the source area and dissolves into the slower moving groundwater. When the extraction pumps are turned on, groundwater with a higher concentration of contaminants is removed, resulting in increased mass removal. The primary advantage of pulsed pumping is that a smaller volume of water is extracted for an equivalent amount of mass removed Adaptive Pumping In adaptive pumping, the well field is designed such that the operation of extraction and injection wells can be modified to reduce zones of stagnation (United States Environmental Protection Agency, 1996). One common practice is to assign wells for either injection or extraction. In this approach, wells may be alternately used for injection and extraction purposes. In addition, extraction wells can be periodically shut off, others turned on, and pumping rates adjusted to optimize remedial operations according to the results of plume monitoring activities. Previous studies have shown that this technique could significantly reduce the time required for site remediation. 59

80 5.2 Ex Situ Treatment Advanced Oxidation Processes (AOPs) AOPs destroy MTBE and other organic contaminants directly in the water through chemical oxidation. The removal of organic compounds from water by AOPs is primarily accomplished through the reaction of organic contaminants with highly reactive hydroxyl radicals (OH ) that can be produced through a variety of mechanisms. Compared to more established water treatment alternatives, such as air stripping and GAC, AOPs are generally considered to be an emerging technology. Some of the challenges with respect to the implementation of AOPs in water treatment are associated with the formation and fate of oxidation byproducts (e.g., TBA and tertiary butyl formate), non-selective radical oxidation, radical scavenging, and bromate formation (from ozone-based AOPs). Although it is possible to overcome these challenges, costs will increase as a result of greater energy usage, greater chemical dosage, and/or secondary treatment polishing steps. AOP technologies have been tested for MTBE destruction in both laboratory and pilot-scale studies. In addition, some full-scale applications have been reported. Recent studies using bench-scale reactors investigated MTBE and hydroxyl radical reaction kinetics in addition to MTBE oxidation pathways and byproduct formation (Acreo et al., 2001; Cater et al., 2000; Chang and Young, 2000; Kang et al., 1999; Safarzadeh-Amiri, 2001; Stefan et al., 2000). Common problematic byproducts include bromate (due to the reaction of ozone with bromide) and TBA, tertiary butyl formate, and acetone (Stefan et al., 2000). Some of these compounds are either more toxic than MTBE or more resistant to oxidation (Cater et al., 2000; Chang and Young, 2000). The California MTBE Research Partnership (1999) evaluated the following AOPs for the treatment of MTBE-contaminated water ex situ: Ozone/hydrogen peroxide (O 3 /H 2 O 2 ). Ozone/ultraviolet light radiation (O 3 /UV). Hydrogen peroxide/medium-pressure ultraviolet light radiation (H 2 O 2 /MP-UV). High energy electron beam irradiation (E-beam). Titanium dioxide-catalyzed ultraviolet light radiation (TiO 2 -catalyzed UV). Sonication/hydrodynamic cavitation. Fenton s reaction. The results of the study showed that the two most promising AOP technologies for MTBE removal from water are O 3 /H 2 O 2 and H 2 O 2 /MP-UV. Both of these processes are well understood and have been demonstrated at several bench- and field-scale sites to successfully remove MTBE from water. In addition to these two relatively well-established AOPs, E-beam 60

81 and ultrasonic cavitation are two emerging AOPs that were found to be promising based on their technical feasibility for removing MTBE from water. These technologies have already demonstrated success in disinfection and remediation applications. An American Water Works Association Research Foundation study is in progress to evaluate the feasibility of O 3 /H 2 O 2, UV/H 2 O 2, and E-beam for the removal of MTBE from potable water sources (American Water Works Association Research Foundation, 2001). Preliminary results from two bench-scale and three pilot studies are promising. A final report detailing the findings of this work was released in 2002 (Kavanaugh et al., 2003). Other studies have also demonstrated the effectiveness of AOPs. Chang and Young (2000) reported the destruction of MTBE (99.9 percent) in a bench-scale reactor involving the use of UV/H 2 O 2 as an oxidant. In conclusion, AOPs are believed to be most efficient for groundwater with MTBE concentrations ranging between 0.1 and 80 mg/l (Cater et al., 2000). While the use of AOPs is effective at treating high concentrations of MTBE, AOP systems can be expensive and can lead to the formation and accumulation of toxic or undesirable byproducts. More research is needed to evaluate the cost-effectiveness and performance of AOPs at field-scale level. Select AOPs are discussed below. Ozone/Hydrogen Peroxide (O 3 /H 2 O 2 ) When ozone is added to water, it participates in a complex chain of reactions that result in the formation of free radicals capable of MTBE destruction. Hydrogen peroxide can be combined with ozone to enhance the transformation of ozone to OH. A number of ozone/hydrogen peroxide systems has been used for the ex situ remediation of MTBE-impacted sites with influent MTBE concentrations of up to 660 mg/l and removal rates exceeding 99 percent (California MTBE Research Partnership, 1999). Hydrogen Peroxide/Medium-Pressure Ultraviolet Light Radiation (H 2 O 2 /MP-UV) Ultraviolet (UV) light radiation can destroy organic contaminants, including MTBE, through direct and indirect photolysis (Zepp, 1988). In direct photolysis, the absorption of UV light by MTBE places it in an electronically excited state, causing it to react with other compounds and eventually degrade. In contrast, indirect photolysis of MTBE is mediated by hydroxyl radicals that are produced when ozone or hydrogen peroxide are added to the source water either prior to or during UV irradiation. A variety of sources of UV light are available. Recently, MP-UV has received increasing attention because it has a greater potential for direct photolysis relative to low-pressure UV (LP-UV) and pulsed UV (P-UV); it radiates over a wider range of wavelengths (200 to 400 nanometers [nm]) than LP-UV lamps, which better facilitates the formation of hydroxyl radicals when hydrogen peroxide is present (hydrogen peroxide absorbs more in the higher wavelengths [250 to 300 nm]). In addition, 61

82 MP-UV lamps produce a greater UV output per lamp than LP-UV lamps; thus, MP-UV systems can be expected to use fewer lamps, take up less space, and require less maintenance. Energy costs, however, may be higher than with LP-UV systems. E-Beam Process High energy electron beam (E-beam) treatment refers to the use of ionizing radiation from an electron beam source to initiate chemical changes in aqueous contaminants. In contrast to other forms of radiation, such as infrared and UV, ionizing radiation from an E-beam is absorbed almost completely by the electron orbitals of the target compound, thus increasing the energy level of its orbital electrons. E-beam processes use the portion of the electromagnetic spectrum between 0.01 and 10 electron volt (ev) (Siddiqui et al., 1996a,b). Within to seconds, the E-beam irradiation of water results in the formation of electronically excited species, including ions and free radicals, along the path of the electrons. The combination of products that result from this reaction creates a unique environment where oxidizing and reducing reactions occur simultaneously (Allen, 1961). Oxidizing species (e.g., OH ), reducing species (aqueous electrons [e aq ], and hydrogen atoms [H ]) are the most reactive products of this reaction and control the rate of the E-beam process for MTBE destruction. Ultrasonic Cavitation In ultrasonic cavitation, contaminated water is irradiated with ultrasound waves, which pass through the medium in a series of alternate compression and expansion cycles. When the acoustic amplitude is large enough to stretch the molecules during its negative-pressure (rarefaction) cycle to a distance that is greater than the critical molecular distance to hold the liquid intact, microbubbles are created that then collapse in the subsequent compression cycle. This gives rise to extremes of temperature, which can trigger the thermal decomposition of MTBE in solution or thermal dissociation of water molecules to form extremely reactive radicals. The extreme conditions generated during cavitation decompose water to create both oxidizing (OH ) and reducing (H ) radical species (Skov et al., 1997; Kang and Hoffman, 1998). As in other AOPs, the primary mechanism for MTBE removal by cavitation is through reaction with hydroxyl radicals. Titanium Dioxide (TiO 2 )-Catalyzed UV When titanium dioxide (TiO 2 ), a solid metal catalyst, is illuminated by UV light (380 nm), valence band electrons are excited to the conduction band and electron vacancies (or holes) are created (Crittenden et al., 1996). These combinations of excited-state electron and holes are capable of initiating a wide range of chemical reactions; however, hydroxyl radical oxidation is the primary mechanism for organic contaminant destruction (Crittenden et al., 1996). The production of hydroxyl radicals can occur via several pathways, but as with many 62

83 of the other AOPs analyzed, is readily formed from hydrogen peroxide. Hydrogen peroxide is formed in three ways: Reduction of oxygen with conduction band electrons (Kormann et al., 1988). Oxidation of water by holes in the valence band (Kormann et al., 1988; Hong et al., 1987; Turchi and Ollis, 1990). Secondary reactions between oxidized organic matter (Kormann et al., 1988). Once hydrogen peroxide is formed, it can dissociate in the presence of UV radiation to form hydroxyl radicals (see H 2 O 2 /MP-UV) or react with other radicals (e.g., hydroperoxyl or superoxide radical) to form hydroxyl radicals. Hydroxyl radicals can also be formed from the direct reduction of TiO 2 -absorbed hydrogen peroxide by a conduction band electron (Al-Ekabi et al., 1989). Finally, hydroxyl radicals can be produced by the reaction of holes with a hydroxide (Hong et al., 1987; Turchi and Ollis, 1990; Sjogren, 1995) Synthetic Resin Sorbents Synthetic resin sorbents, like GAC, rely on the process of sorption to remove organic compounds from water. The primary advantages of resin sorbents over GAC are their on-site regenerability and their resistance to the competitive sorption of natural organic matter. Resin sorbents can be regenerated on-site through steam stripping or microwave irradiation. Results from laboratory and pilot-scale studies suggest that synthetic resins have the potential for success at MTBE-impacted sites. Davis and Powers (2000) identified two carbonaceous resins that exhibited three to five times greater sorption capacities than activated carbon for MTBE at a concentration of 1 mg/l. In batch and fixed-bed experiments, another synthetic resin was shown to remove both MTBE and TBA from water (Annesini et al., 2000). Limited pilot studies have also suggested that resins can efficiently remove MTBE from contaminated groundwater (California MTBE Research Partnership, 1999). The life expectancy and regenerability of the sorbent define the economics of using resins (Davis and Powers, 2000). Resin isotherm data for MTBE suggests that Ambersorb 563, a carbonaceous resin manufactured by Rohm and Haas (Philadelphia, Pennsylvania), is currently the resin industry s most promising candidate for removing MTBE from water (California MTBE Research Partnership, 1999). This resin can be regenerated using steam. Two independent studies found Ambersorb 563 to have a superior sorption capacity for MTBE compared to Filtrasorb 400, a coal-based GAC widely used by the water industry (California MTBE Research Partnership, 1999). Unfortunately, Rohm and Haas recently ceased production of Ambersorb 563, and the future of this material is unknown. 63

84 One of the advantages of using synthetic resins is the ease of regeneration relative to GAC. One of the limitations of resins is the experimentally observed reduction in MTBE sorption due to the presence of other gasoline constituents, such as BTEX compounds and TBA, in contaminated water. This is analogous to the effects of natural organic matter and BTEX on GAC sorption, as previously discussed. The use of synthetic resins is, therefore, costly relative to GAC, but resins can be designed to achieve a higher degree of selectivity and tend to have a longer lifetime than GAC. Limited data are available on the effects of background water quality toward the MTBE removal efficiency of resins. Available data suggest that the performance of Ambersorb resins is unaffected by ph (6.5 to 8.5), temperature (10 C versus 25 C), oxidants (hypochlorous acid, hydrogen peroxide, and ozone), the presence of natural organic matter, and the presence of TBA. Ambersorb resins have also been found to be resistant to biofouling; however, m-xylene, which can be considered representative of BTEX compounds, has been found to decrease MTBE sorption capacity when it is present at relatively high concentrations (43.2 mg/l) (Davis and Powers, 1999). Additional work is needed to characterize sorption properties of various resins and to evaluate the feasibility of field-scale applications. 5.3 In Situ Thermal Processes Thermal processes can be used to optimize SVE, MPE, and bioremediation operations. Increasing the temperature in the subsurface can improve the efficiency of the first two technologies by increasing the volatility of the contaminants and their desorption from soils. A higher temperature is advantageous for bioremediation because, as discussed in Chapter 4, microbial metabolism typically accelerates with increasing temperature up to optimum values ranging between 20 and 40 C (Chapelle, 1992; LaGrega et al., 1994). There are a variety of ways to apply heat to the subsurface. Three technologies six-phase heating, in situ radio frequency heating, and dynamic underground stripping are discussed in the following sections Six-Phase Heating Six-phase heating TM is an electrical technique used to resistively heat soil and create an in situ source of steam to strip contaminants, which are then captured using standard SVE technology (Heath and Garcia, 1999). Six-phase heating was developed by the Battelle Memorial Institute for the United States Department of Energy for enhancing the removal of volatile organic compounds from low-permeability soils. In six-phase heating, arrays of six vertical, angled, or horizontal electrodes are spaced around central vapor extraction wells. The design and placement of electrodes are optimized for each site based on the size and shape of the remediation area, site lithology, depth to groundwater, subsurface interval(s) and total depth of site impact, total organic carbon content, electrical resistivity of the soil at the site, and buried utilities and immediately adjacent surface structures (Heath and Garcia, 1999). A six-phase electrical potential is applied to the 64

85 electrode array, generating a voltage gradient throughout the zone of the array. By using a six-phase voltage pattern instead of a single-phase or three-phase voltage pattern, the voltage gradient between each electrode is kept constant, thus minimizing power density variations in the soil and creating a more uniform and larger heating pattern. As the electrical current generated by the voltage gradient passes through the soil, the resistance of the soil to the current flow causes the soil temperature to rise. The rise in temperature increases the volatility of volatile organic compounds. Furthermore, as the soils are heated to the boiling point of water, any soil moisture present turns to steam, thereby stripping the volatile organic compounds from the soil pore spaces. The vapor-phase volatile organic compounds and steam are collected by applying a vacuum to the central collection well (Ott, 1999). Six-phase heating has been applied successfully at over 13 sites, including a dense, nonaqueous phase liquid contaminated site in Skokie, Illinois (Heath and Garcia, 1999). In the absence of any published results of field applications for this technology at MTBE-impacted sites to date, it is difficult to speculate on whether the high costs typical of this technology at chlorinated solvent sites (mostly with dense, nonaqueous phase liquid) can be justified at sites with LNPAL and MTBE In Situ Radio Frequency Heating Radio frequency heating uses electromagnetic energy to heat soils in the same way that microwaves heat food. The process uses electrodes installed in soil borings. A device generates a radio frequency of the appropriate size to heat the particular soil based on the dielectric properties of the soil. The emitter or antenna generates heat in an area around the soil boring. The process is not dependent on soil water for operation, and temperatures are reported to reach 250 to 300 C. Excess soil water can actually limit the efficiency of the process Dynamic Underground Stripping Dynamic underground stripping is a remediation method that was developed by researchers at Lawrence Livermore National Laboratory and the College of Engineering at the University of California, Berkeley. Dynamic underground stripping removes volatile organic compounds below the water table by heating the subsurface above the boiling point of water, then removing both contaminant and water by vacuum extraction. Heating the subsurface not only vaporizes organic contaminants, but also enhances their removal by increasing their diffusion and desorption rates (Newmark and Aines, 1995). Dynamic underground stripping integrates three technologies: steam injection, electrical heating, and underground imaging (Newmark and Aines, 1995). Steam is pumped into injection wells, heating the contaminated soils to 100 C. Steam drives contaminated water toward the extraction wells, where it is pumped to the surface. When the steam front encounters contamination, volatile organic compounds are vaporized from the hot soil and are moved to the steam/groundwater interface, where they condense. Vacuum extraction after 65

86 full steaming of the contaminated zone continues to remove residual contaminants. The steam injection/vacuum extraction technique was developed at the University of California, Berkeley (Udell and Stewart, 1989; Udell and Stewart, 1990; Udell et al., 1991; Udell, 1994a). Descriptions of a steam injection system and its operational design are presented in Siegel (1994) and Udell (1994b). Electrical heating is done in conjunction with steam injection to heat less permeable areas. This technique heats clay and fine-grained sediments and causes water and contaminants trapped within the soil to vaporize and be forced into the steam-swept zones, where vacuum extraction removes them. Electrical heating is ideally suited for tight, clay-rich soil and/or near-surface contamination. Details on electrical heating construction and its operational design can be found in Siegel (1994). To monitor the progress of dynamic underground stripping operations, geophysical imaging methods are used to map the boundary between the heated areas and the cooler areas, which have not been remediated. Dynamic underground stripping has been successfully demonstrated at a number of sites, including a gasoline spill site at the Lawrence Livermore National Laboratory (Newmark and Aines, 1995). In the late 1990s, scientists at the Lawrence Livermore National Laboratory improved the efficiency of dynamic underground stripping by developing hydrous pyrolysis/oxidation. By introducing both heat and oxygen, this process was found to convert contaminants in groundwater to benign products such as carbon dioxide, chloride ions, and water. In 1997, this improved process was applied with great success at a site in Visalia, California, which was formerly used for treating utility poles with creosote and pentachlorophenol compounds. During the first 6 weeks of using hydrous pyrolysis/oxidation, contaminants were removed at a rate of 46,000 pounds per week compared to an average of 10 pounds per week using pump-and-treat (Walter, 1998). 5.4 In Situ Chemical Reduction The use of zero-valent metals as chemical reductants for the treatment of organic compounds is one of the latest innovations in remediation technologies (Gillham, 1996). Metals such as granular iron can be applied in the funnel and gate treatment system in which a porous wall of granular iron is constructed in the path of a contaminated groundwater plume. In principle, the iron wall is designed to react with contaminated water passing through it and reduces the contaminants to benign compounds, such as hydrocarbons, chlorides, and water. Other metals, such as zinc and tin, have also been found to be effective in reducing organic compounds, specifically chlorinated compounds (Boronina et al., 1995; Schlimm and Heitz, 1996). Palladium-coated iron particles have been demonstrated to improve reaction kinetics (Nyer and Vance, 2001). This technology is still in its developing stages and will need to be tested for its applicability to MTBE. Challenges associated with the implementation of zero-valent technology include (1) the production and accumulation of byproducts due to the low reactivity of some 66

87 compounds and (2) decreased metal reactivity over time, likely due to the formation of a surface passivation layer or to the precipitation of metal hydroxides (Fe[OH] 2 and Fe[OH] 3 ) and metal carbonates (FeCO 3 ) on the surface of the metal (Zhang et al., 1998). Potential improvements to this technology include the injection of nanoscale metal particles directly into contaminated aquifers and their attachment onto solid supports, such as activated carbon, zeolite, and silica, for ex situ treatment of contaminated water with metal (Zhang et al., 1998). 5.5 Phytoremediation Phytoremediation involves the use of plants or trees to either remove or degrade groundwater contaminants. Phytoremediation occurs by plant uptake of the contaminants, followed by translocation and accumulation into plant tissues or transpiration through leaves. In some cases, the breakdown of contaminants can take place either within plant tissues or in the rhizosphere (Rubin and Ramaswami, 2001). The transformation of MTBE inside plants is only possible once MTBE has been uptaken. The mechanisms responsible for MTBE metabolism by plants are still unclear and, as of yet, have not been studied in detail. In general, phytoremediation is most effective for contaminants with log K ow values in the range of 0.5 to 3.0 (Burken and Schnoor, 1998). Since MTBE has a log K ow within this range, phytoremediation is expected to be an effective technology at MTBE-impacted sites. In fact, several plant and tree species have been recently reported to uptake MTBE, including alfalfa plants (Zhang et al., 2001), poplar trees (Rubin and Ramaswami, 2001; Hong et al., 2001; Newman et al., 2000), and eucalyptus trees (Newman et al., 1998, 2000). In laboratory studies, poplar trees were shown to remove 30 percent of MTBE in test solutions at concentrations of 300 to 1,600 µg/l in a week (Rubin and Ramaswami, 2001). In this study, MTBE was largely untransformed during transport through the plant. In another study, the main mechanism for MTBE removal at high contaminant concentrations (10.2 mg/l) was also uptake and evapotranspiration from poplar leaves and stems. Between 37 and 67 percent of the MTBE was removed in a period of 10 days (Hong et al., 2001). In one case, significant groundwater uptake by deep-rooted poplar trees was shown to hydraulically contain an MTBE plume in Texas (Hong et al., 2001). The results of this study are promising and indicate that phytoremediation should be considered for further testing as an alternative remediation technology in shallow groundwater sites. In particular, Hong et al. (2001) suggest that deep-rooted trees may be able to serve as natural pumps that can supplement or replace pump-and-treat systems under certain site conditions (e.g., the depth to groundwater is less than 15 feet, the contamination in groundwater is restricted to depths of less than 10 feet). The specific limitations to the use of this technology have not yet been resolved. More research is needed to better design and implement engineering solutions involving phytoremediation at MTBE-impacted sites. 67

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89 6. Remediation Case Studies In this chapter, five case study sites that involve the remediation of subsurface MTBE are analyzed to better understand the field implementation of MTBE remediation technologies. Detailed summaries of the case study sites are presented in the Appendix. The purpose of this section is to present a synopsis of the primary lessons learned from these case studies. The lessons learned are organized by remediation technology and can be directly related to the technology evaluations presented in Chapter Site Summaries For this report, five case study sites were chosen for a detailed evaluation regarding the remediation of subsurface MTBE. These case study sites were chosen based on the following criteria: Sites have subsurface contamination by MTBE-blended gasoline. Sites have been remediated for at least 1 year by an active remediation system. Adequate data for groundwater trends and remediation system performance analyses were available. Design and operational information for the remediation systems were available. Capital and O&M costs for the system were available. Sites address a range of applicable remediation technologies. Although detailed remediation data for all of the sites were not available, general information is presented in those cases to give the reader a broader perspective on MTBE remediation. 6.2 Primary Lessons Learned This section presents a summary of the primary lessons learned from the remediation case studies presented in the Appendix. For each of the remediation technologies used at the case study sites, an evaluation of relevant data is presented to determine the primary remediation lessons that can be learned about each of the remediation and treatment technologies Pump-and-Treat/Groundwater Extraction As discussed in Chapter 4, pump-and-treat or groundwater extraction can be used for plume containment and/or remediation. Several of the case study sites used groundwater extraction, either as a singular remediation technology or in combination with other in situ technologies, such as SVE. 69

90 A review of initial MTBE concentrations in extracted groundwater can provide some insight regarding what concentrations in groundwater can be expected at an MTBE-impacted site. Data from extracted groundwater at several of the case studies were reviewed with the following results: Site 1 had a relatively low initial concentration of MTBE (130 µg/l). At this site, it is known that the spill source was gasoline that contained MTBE for octane boosting (about 3 to 8 percent MTBE by volume). Sites 3 and 4 had measured maximum observed MTBE concentrations prior to the remediation of 2,720 and 34,440 µg/l, respectively. These sites apparently were impacted by blended gasoline spills that contained MTBE at either 11 percent (RFG) or 15 percent (Oxyfuel). The MTBE concentration range associated with the octane-enhanced gasoline spill (Site 1) is consistent with reports from working professionals in areas that only use MTBE for octane boosting reasons (e.g., Florida). They have anecdotally reported that MTBE concentrations seen in groundwater are often not very high (less than 10 mg/l). In contrast, concentrations reported elsewhere commonly reach levels that are an order of magnitude higher (greater than 100 mg/l). This suggests that perhaps the MTBE plumes from lower percentage blends of gasoline are somewhat mass limited. Due to the small data set and the site-specific nature of this concentration data, it is inadvisable to use this information to predict the influent concentration ranges expected at an MTBE-blended gasoline spill. The reader should be aware that MTBE levels in extracted groundwater will depend upon a number of other variables, including time since the spill, recovery well location relative to the spill, degree of dispersion of MTBE mass, and the degree of mixing with groundwater. At any site, the best methods for quantifying the likely MTBE concentrations in extracted water are to: Sample water produced during a pumping test from the recovery well(s). Conduct a careful site investigation to accurately define the size of the plume and the likely concentrations in an extraction system. Pumping Performance As discussed in Chapter 4, MTBE s characteristics should allow groundwater extraction to be an effective method for removing dissolved-phase MTBE from the subsurface. The effective removal of MTBE from groundwater can be demonstrated by how the MTBE levels in monitoring wells and in extracted water decline in comparison to benzene levels over the same time frame. For most of the case study sites that used groundwater extraction, preferential removal of dissolved-phase MTBE was evident, even over the wide range of sitespecific conditions involved (e.g., initial concentrations, geology, pumping rate). 70

91 MTBE concentrations decreased quite rapidly at Site 3, which used both groundwater extraction and SVE. As presented in Table 6-1, MTBE levels in extracted groundwater were reduced by a factor of nearly seven (2,720 to 400 µg/l) during the first 1.8 years of operation. In contrast, benzene levels during the same time period declined by a factor of approximately three (3,970 to 1,270 µg/l). Table 6-1 Concentration Reductions with Groundwater Extraction and Treatment Systems Over Time Pumping Water System Influent System Influent Duration Extraction MTBE Benzene Site (Years) Rate, Q (gpm) Reduction Factor a Reduction Factor a to (With SVE) (With SVE) a Influent reduction factor is defined as the ratio of initial influent concentration to the treatment system (measured at pumping system startup) over the final influent concentration (measured at the end of operations or on the last sampling event for which there are data). Figure 6-1 is a plot of MTBE concentrations in extracted water versus time for Site 4, which also used SVE as a remediation technology. As shown, MTBE concentrations in extracted water declined from a high of 34,440 µg/l to just 272 µg/l in 3 years. This 127-fold decline for MTBE was far more than the 17-fold decrease of benzene over the same time period. Figure 6-1 also shows that very little rebound of MTBE levels was detected during the first 4 years of pumping, whereas rebound is commonly seen with many other hydrocarbon parameters. Approximately 5 years after pumping started, another gasoline spill apparently occurred (as indicated by the dramatic rise in MTBE levels); however, these relatively high Concentration (µg/l) 36,000 32,000 28,000 24,000 20,000 16,000 12,000 8,000 4,000 0 Benzene MTBE 06/ / / / / / / / / / / / / / /1996 Figure 6-1. Influent water concentrations for Site 4 pump-and-treat system (with SVE). 71

92 concentrations were reduced rather quickly by groundwater extraction during the following year. The fast decline of influent concentrations and lack of rebound are expected due to the very minimal adsorption of MTBE to the aquifer materials and the more rapid leaching of MTBE from the residual non-aqueous phase liquid. It is important to note that the hypothesis of preferential removal is specific to dissolvedphase MTBE. As with all gasoline components, the mass transfer of MTBE from residual phase to dissolved phase is limited by site-specific hydrogeological conditions. Although MTBE is expected to leach out of the residual phase more rapidly than other gasoline components, this leaching will be controlled by the degree to which subsurface water comes in contact with the residual phase (see Chapter 2) Treatment of Extracted Water Air Stripping For the case studies presented in the Appendix, air stripping was the most commonly used primary ex situ treatment method for MTBE-impacted waters. These case studies show that: Air stripping is a robust technology that can achieve high MTBE removal efficiencies. Off-gas treatment can cause air stripping to be more costly than competing technologies. Table 6-2 contains data on the air stripping systems used at Sites 1, 3, and 4. All of these sites used a single packed-tower air stripper. The data from these three sites are considered indicative of the type of performance that can be expected from air stripping for MTBE. Table 6-2 Air Stripper Performance at Three Case Study Sites Initial Average Average Average MTBE Air-to-Water MTBE Benzene Water Flow Influent Ratio Removal Removal Site Rate (gpm) (µg/l) (dimensionless) Rate (%) Rate (%) Comments >99.9 >99.9 Starting influents were quite low for both MTBE and BTEX 3 3 2,720 1, to >99.8 >99.8 None ,440 Unknown Unknown Unknown Air stripper only used for 2 months, no data available; switched back to GAC due to air stripper off-gas treatment problems 72

93 Air stripping is an effective technology as indicated by its application over a wide range of flow rates and influent MTBE concentrations (see Appendix). It is also reliable in the field, as operational run times at the case study sites were very high (greater than 90 percent). The reliability of air stripping is further proven by Case Study Site 3, where another technology (i.e., GAC) was abandoned due to operational problems (the remedial system was then switched to air stripping.) At this site, carbon adsorption was initially selected as the primary water treatment method; however, the treatment method was soon changed to the more reliable air stripping when MTBE breakthrough of GAC became problematic. As predicted by modeling (California MTBE Research Partnership, 1999), air-to-water ratios of greater than 150 are able to remove greater than 99 percent of MTBE in water, depending upon tower height and the size of the packing. Table 6-2 shows that air-to-water ratios ranging from 162 (Site 1) to 1,371 (Site 3) achieved MTBE removal rates ranging from 93 to greater than 99 percent. Table 6-2 also presents removal rates for benzene. For the air strippers at Sites 1 and 3, benzene removal rates were typically greater than 99 percent, while removal rates for MTBE ranged from 93 to greater than 99 percent. This difference in removal rates is consistent with the Henry s constants of the two compounds, which indicate that benzene can be stripped more easily than MTBE. Achieving greater than 99 percent removal of a contaminant is quite difficult, especially for air stripping in a single tower. If MTBE influent concentrations are high and/or if very low effluent treatment goals must be met (e.g., to meet drinking-water standards), multiple treatment units may be required. The results from Site 4 show that treating off-gas from air strippers can be problematic and costly. As presented in the Appendix, the designers initially selected liquid-phase GAC to treat extracted groundwater; however, as remediation progressed, it became clear that the GAC was not performing as planned (more details are given in the next subsection). As such, the designers switched to air stripping for water treatment; however, the treatment of the air stripper off-gas was so difficult and expensive that treatment was reverted back to liquidphase GAC (see Appendix). In summary, these case studies indicate that air stripping in a single tower is an effective technology for achieving high removal rates (93 to greater than 99 percent) of MTBE. Air-to-water ratios will likely have to exceed 150-to-1 to reach such high removal rates. It should be noted that the application of air stripping could be problematic and costly if offgas treatment is required. Air stripping may also require more than one tower in series if the effluent treatment goals are very low (such as to meet drinking-water standards) and greater than 99 percent removal is required. Granular Activated Carbon (GAC) GAC was the primary water treatment technology used at Sites 3 and 4. These applications of GAC were either unsuccessful or extremely costly. In general, these case studies show that 73

94 the effectiveness of GAC for MTBE removal is limited; however, under certain site conditions, GAC may still be the best choice for a water treatment system. In addition, these studies illustrate the critical nature of the frequent monitoring and maintenance of GAC systems for MTBE removal. Site 3 used three 600-pound vessels in-series to treat MTBE-impacted water at a flow rate of 3 gpm. MTBE concentrations in the treatment system influent ranged up to 2,720 µg/l. MTBE broke through the first two vessels within several weeks of startup. The system was then permanently switched to air stripping. It is unknown what type of GAC was used at this site. Site 4 used GAC as the primary water treatment method over a 6-year period. During the first 11 months of operation, the GAC treatment system was plagued by numerous breakthrough events, primarily by MTBE. Because of the poor performance of GAC, the site operators switched to air stripping for 2 months; however, the high cost and difficulty of treating the air stripper off-gas caused them to switch back to liquid-phase GAC treatment for the extracted water. The operators were reluctant to switch back because even with a low water flow rate (2 gpm), rapid GAC usage continued to be a problem. It is unknown what types of GAC were used at this site. The GAC system at Site 4 consisted of two 300-pound vessels in series. MTBE concentrations in the treatment system influent water ranged from 272 to 34,000 µg/l. During the 6 years of operation, both GAC vessels were changed an average of three times per year. MTBE desorption from the first carbon vessel was noted in 22 of the 73 monthly sampling events, which was more than four times as often as benzene desorption. Likewise, in a 6-year period, MTBE was desorbed 11 times by the second vessel in the series and benzene was desorbed only twice. As discussed by the California MTBE Research Partnership (1999), a MTBE desorption event can occur when the influent concentration decreases, allowing for adsorbed MTBE to dissolve back into the water stream flowing through the GAC. Desorption can also occur when other organic compounds (e.g., BTEX) increase in concentration, causing competitive adsorption effects to desorb MTBE molecules. If the GAC at Site 4 had been replaced every time that MTBE desorbed, then carbon change-outs would have occurred more than twice as often as their already frequent schedule. In summary, GAC can be used to remove MTBE from water; however, in field applications, MTBE breakthrough occurred quite rapidly. This can increase costs, complicate system operation, and potentially result in discharge violations. Because of the high potential for MTBE desorption and breakthrough, it is necessary to monitor influent and effluent conditions frequently and change-out GAC vessels in a timely manner. Of the two sites studied here that used GAC as a primary water-treatment technology, one (Site 3) was a failure (site operators switched to air stripping) while the other (Site 4) can be considered an operational success (in spite of frequent change-out requirements). Recent work has shown that coconut-shell GAC is more effective for MTBE removal than coal-based GAC (California MTBE Research Partnership, 1999). Unfortunately, it is unknown what types of GAC were used at the case study sites discussed above. 74

95 6.2.3 Soil Vapor Extraction (SVE) and Multi-Phase Extraction (MPE) SVE systems were used at four of the five case study sites. Site 2 used SVE as the sole remediation method. Sites 3 and 4 combined SVE with groundwater extraction (MPE). Site 5 combined SVE with several other subsurface remediation systems, including air sparging, free-product recovery, and a groundwater cutoff barrier. The presence of MTBE did not appear to affect the SVE system design (i.e., the blowers, piping, and vapor extraction wells were typical). None of the sites reported abandoning SVE because of MTBE-related complications, although several used SVE intermittently or at low flow rates to minimize off-gas treatment demands. As is typical, all the SVE systems appeared to remove significant contaminant mass from the vadose zone, and the amounts often far exceeded the quantities recovered as free or dissolved phase by the groundwater extraction systems (e.g., Site 3); however, the overall mass of recovered MTBE was not determined at any of the sites. The typical procedure for most SVE projects is to measure the extracted vapors with field instruments (e.g., photoionization detector or flame ionization detector) that do not quantify individual compounds; therefore, little can be concluded about overall MTBE recovery efficiencies using SVE. Since Site 2 used only SVE, groundwater concentrations can be evaluated to determine if dissolved-phase MTBE levels were impacted by SVE operations. Over 2.2 years of SVE operations, groundwater conditions at the site improved significantly. All 11 monitoring wells had distinct declines of dissolved-phase benzene and MTBE concentrations. One vapor extraction well had a benzene concentration decline from 980 to 0.86 µg/l (a 1,139-fold decrease) while MTBE concentrations declined from 3,500 to 220 µg/l (a 16-fold decrease). An adjacent monitoring well showed a benzene concentration decline from 670 to less than 0.5 µg/l (a 1,340-fold decrease) and an MTBE decline from 8,900 to 21 µg/l (a 423-fold decrease). As no other remedial systems were operating at Site 2 and no significant declining trends were noted before the SVE system was initiated, SVE and the improved in situ biodegradation are collectively credited with these significant reductions of BTEX and MTBE concentrations in groundwater. Also, the concentration declines in these wells are not due to the contaminant plume moving away, as all 11 wells showed contaminant declines, and the plume was encircled with monitoring wells. The benzene concentrations in groundwater declined at a faster rate than MTBE levels in all 11 wells monitored. This is likely a reflection of MTBE s greater resistance to volatilization from water (due to a lower Henry s constant) and minimal biodegradation. Considering MTBE s characteristics, the distinct decreases of dissolved-phase MTBE seen at Site 2 after 2.2 years of soil venting are somewhat surprising and not fully understood. Perhaps the SVE operation removed MTBE from the entrapped non-aqueous phase liquid, which prevented MTBE from recharging the groundwater. As a result, dissolved-phase MTBE levels in the groundwater were allowed to decline by natural attenuation. 75

96 SVE operations were also conducted at Sites 3, 4, and 5; however, since no concentration data are available for the soil gas or the extracted vapors, very little can be concluded about the effectiveness of SVE for MTBE removal at these sites. In addition, since these sites simultaneously used SVE and groundwater extraction, the impacts of the SVE system upon dissolved-phase MTBE concentrations cannot be isolated. It is evident that the two SVE systems operating at Site 5 removed a substantial mass of hydrocarbons, particularly from the vadose-zone soils above the free-product plume; however, based on laboratory analysis of the free product, MTBE had already been depleted from the free product (through natural processes) prior to initiating the SVE system. It is believed, therefore, that very little MTBE was recovered by the SVE system at Site 5. Groundwater quality data are available for a number of wells in the area treated by SVE at Site 5. MTBE concentration trends in these wells are inconsistent during the nearly 3-year period that venting occurred. MTBE levels in one well rose 84 percent over the 3 years, while MTBE levels in another well fell as much as 63 percent. The remaining wells have MTBE concentration changes somewhere between these two extremes. As a result, there is no clear trend of water-quality improvement in the wells installed near the SVE system. Site 5 demonstrates how applying SVE typically will not be beneficial for removing subsurface MTBE that has already been leached from the free-product and residual-phase gasoline in the vadose zone. At Site 4, MTBE was found to be predominant in the extracted vapors, which is consistent with theoretical predictions discussed in Chapter 4. High amounts of MTBE in the extracted off-gas resulted in rapid breakthrough of vapor-phase GAC used at the site. In summary, based on the four case study sites previously discussed, it appears that SVE can be an effective remediation method for MTBE. Data from Site 2 suggest that SVE mechanisms (i.e., volatilization from separate-phase product, aeration of the vadose zone) worked in combination with natural attenuation to rapidly decrease dissolved-phase MTBE concentrations in groundwater. In contrast, Site 5 demonstrates a typical scenario in which SVE was applied too late to recover MTBE, which apparently had already leached from separate-phase product (more quickly than other gasoline compounds). It is clear that the rapid implementation of SVE is critical so that separate-phase product in the vadose zone is subject to vapor extraction prior to MTBE leaching. Most of the sites reviewed here did not conduct compound-specific analyses of the extracted vapors and, as such, the performance of the SVE systems at removing MTBE cannot be scrutinized in detail. If the MTBE remediation capabilities of SVE systems are to be better understood, then operational procedures must be changed to better characterize vapor samples from the vadose zone and SVE off-gas. 76

97 6.2.4 Vapor Treatment Vapor treatment was used at four of the case study sites to treat the vapors from SVE systems and air strippers (Table 6-3). Site 3 combined the 550-cubic feet per minute (cfm) stripper off-gas stream with the 100-cfm air flow from the SVE system and treated the combined flow with a thermal oxidizer. Site 2 also used a thermal oxidizer to treat 650 to 700 cfm of off-gas produced by its remediation systems. For both of these oxidation systems, the reported destruction efficiencies were greater than 95 percent, though these were based upon influent/effluent readings taken with field detectors, such as flame ionization detector and photoionization detector instruments. No compound-specific analyses were conducted, so no specific conclusions can be made regarding MTBE destruction efficiencies. Table 6-3 Vapor Treatment Case Studies Source Average and Flow Destruction Rate or Removal Site (cfm) Method (%) Comments 2 SVE (700) Thermal oxidation All volatiles 95 Little information available 3 Stripper (550) Thermal oxidation All volatiles >99 Little information available SVE (100) for the combined 650-cfm flow 4 SVE (100) GAC, GAC GAC breakthrough on first and third days; then electric catox poor removal; Switched to catox for 15 months, Catox 98 then none required as concentrations were low 5 SVE Catalytic Catox unknown Little information available (unknown) oxidation At Site 4, vapor-phase carbon adsorption was initially used, but MTBE breakthrough of the GAC occurred on both the first and third days of operation. Vapor treatment was then switched to electric-fired catalytic oxidation (catox) for the first 15 months of SVE operation. During this period, the catox unit destroyed 96 to 100 percent of the influent hydrocarbon vapors. After 15 months of vapor extraction, the influent vapors dropped so low (8 parts per million by volume [ppmv] of hydrocarbons by photoionization detector) that secondary fuel costs soared. Initially, when soil gas influent concentrations were 135 ppmv volatile organic compounds, the cost for electricity (the secondary fuel) was only $40 per month. Eight months later, the influent concentrations had declined significantly (down to 14 ppmv of volatile organic compounds) and the secondary fuel costs had risen to $500 per month. Because of this, the SVE operations were temporarily halted and vapor treatment was switched back to GAC. With total volatile influent concentrations of zero to 21 ppmv, the GAC needed replacement three times in 1 year. Like Site 1, this site demonstrates the difficulty of treating high airflow, low-fuel content MTBE vapors with GAC. It also shows that high secondary fuel costs can result from treating low to moderate concentrations of MTBE with catalytic units. 77

98 In summary, the vapor treatment systems that operated at these four sites confirm performance trends discussed by the California MTBE Research Partnership (1999). GAC is very ineffective for high concentration off-gas streams, such as those typically encountered in the first few months of SVE operation. Vapor-phase GAC can be used effectively once vapor concentrations have decreased; however, at such low concentrations, consideration should be given to whether vapor treatment is needed at all. Thermal oxidation and catalytic oxidation units handled the total volatile loads encountered at the case study sites, but little data were available regarding specific performance on MTBE destruction. Also, little is known about the practicality of using thermal or catalytic oxidation for large flow (e.g., greater than 1,000 cfm) and low concentration air streams that might be encountered in the off-gas from larger air strippers. Due to the high air-to-water ratios needed to remove MTBE from water streams, it is likely that most air stripper off-gas streams will have low fuel contents. To maintain the thermal range necessary for the correct operation of thermox or catox units, secondary fuel is required. Initially, when contaminant levels are high, the secondary fuel requirements may be low; however, as remediation progresses, and contaminant levels decline, high secondary fuel demands can result. At Site 4, secondary fuel costs increased by a factor of 12 in just over 8 months. Based on these case studies, the best approach for off-gas treatment may be to plan on changing vapor treatment technologies one or more times over the few years that vapor treatment is required. Initially, when influent concentrations are high, thermal or catalytic oxidizer units may be the most reliable and cost-effective technologies. After several months, to perhaps a year or two, when influent levels are down, it may be more cost-effective to switch to GAC or even remove the vapor treatment altogether, if appropriate. While treating vapor flows containing MTBE can be problematic and/or costly, vapor-phase treatment may only be needed for a short time. Most SVE systems reach asymptotically low extraction concentrations within 1 to 4 years, and few SVE systems operate longer than this. As discussed earlier, while groundwater extraction systems can run for decades before all the hydrocarbons are removed, it seems likely that most MTBE will be extracted by pump-andtreat systems in a much shorter period of time, perhaps 5 to 10 years in some cases; thereafter, there may be no MTBE in the air stripper off-gas stream. Even if some low levels of MTBE remain in the water for longer periods of time, the very low concentrations may allow the airflow through the air stripper to be greatly reduced, which in turn greatly reduces vapor-phase treatment problems. In essence, while vapor-phase treatment of air stripper offgas or SVE discharge flows can be problematic when they contain MTBE, the problems may only exist for a few years, thereby limiting vapor treatment costs Air Sparging Systems Site 5 used a combination of remedial systems for MTBE and hydrocarbons, including SVE, free product skimming, and air sparging. Interpreting the performance of the individual 78

99 remedial systems at Site 5 is complicated by the simultaneous operation of these systems across the site; however, as analytical testing of the recovered free product showed that the product was depleted of MTBE, it is concluded that free-product skimming did very little to remove MTBE. As discussed in a previous section, the two SVE systems at Site 5 apparently did very little to remove MTBE from the subsurface; therefore, other than natural attenuation, the air sparging system appears to be the primary method of MTBE remediation at Site 5 (see Appendix). In April 1996, an air sparging performance test was performed at Site 5. This test provided evidence that sparging can physically remove MTBE from groundwater. As part of the 24-hour pilot test, MTBE vapor samples were collected in six passive vent outlets at two time intervals: 1 hour into the test and 12 hours into the test. These samples all contained MTBE vapors ranging from 3 to 76 ppmv. All six probes had higher readings at the first hour than in the twelfth hour, which is a typical pattern for sparging operations. Unfortunately, no baseline vapor samples were collected before the sparging began; however, soil gas sampling from 1 year before (May 1995) detected no MTBE in any of 10 soil gas probes located near the sparge lines (the detection limit is 1 ppmv). This provides some evidence that MTBE vapors were not present in the vadose zone under static conditions. In spite of the lack of baseline vapor data, sparging can still be clearly linked to MTBE removal by the vapor concentration patterns noted and by the substantial simultaneous offgassing measured in the six passive vents. Specifically, during the first and twelfth hours, significant upward air flows were measured in all six passive vents (11 to 38 standard cfm), thereby showing that the air flow in these probes and the MTBE vapors present in that air flow (3 to 76 ppmv of MTBE) were derived directly from sparging operations. Groundwater monitoring data from Site 5 provides further evidence of air sparging s ability to strip MTBE in situ. Six piezometers were closely monitored before and during the performance test. The dissolved-phase MTBE levels in these five piezometers decreased 31 to 82 percent during the 24-hour performance test (see the Site 5 description in the Appendix for details). Further evidence of in situ sparging activity is that dissolved oxygen levels in five of the six piezometers rose significantly, from a range before the test of 2.14 to 3.40 mg/l, up to a range of 3.39 to 8.87 mg/l after 24 hours of sparging. This dataset clearly demonstrates that air sparging removed MTBE via ex situ stripping during the performance test, as shown by the simultaneous occurrence of the following: Air being sparged into groundwater. Return air flow measured in passive vent wells. Elevated MTBE vapor concentrations in those same passive vent wells. Dissolved-phase MTBE levels decreasing in nearby piezometers. Dissolved oxygen levels increasing in nearby piezometers. 79

100 The long-term data regarding the performance of the air sparging system at this site are not as clear as those generated during the sparging test. The air sparging system at Site 5 is very large, with seven horizontal sparge lines, each approximately 150-feet long. Part of the sparge system had operated for about 2.5 years. Groundwater quality data from monitoring wells near the older sparge system do indicate decreased MTBE levels over the 2.5 years of operation. One well (MW-35), located just 10-feet downgradient of a sparge line, had an 86-percent decrease of MTBE levels after 2.5 years. Another nearby well (MW-34) also had a 92-percent decrease, but this well is 60-feet upgradient of the same sparge system. Since 60 feet is a long distance away for air sparging to impact an upgradient well, MW-34 is likely outside the influence of the sparge system; thus, the MTBE concentration decline in the upgradient control well (MW-34) is nearly the same as the treated well (MW-35) and so the sparge system cannot be directly credited with having a measurable long-term impact on MTBE levels. Many other nearby wells did not consistently monitor MTBE and so few other wells have data for additional comparison. The MTBE reductions noted in MW-34 and MW-35 must be due to advective/dispersive transport of the groundwater plume and/or natural biodegradation. Certainly, MTBE levels across the downgradient area generally dropped during the 2.5 years of air sparging. This lowering cannot be merely attributed to the advective/dispersive transport of the MTBE plume further downgradient. Several wells far downgradient have shown MTBE level declines and indicated that the plume may have even retracted back toward the source over the few years of sparging treatment. Based on steadily decreasing MTBE and BTEX concentrations that occurred before air sparging operations began, it appears that significant biodegradation occurs naturally at the site. Air sparging appears to have accelerated biodegradation (by increasing oxygen concentrations) and attenuation processes that were apparently already occurring at the site. In conclusion, the performance test at Site 5 clearly demonstrated that air sparging was physically volatilizing MTBE from groundwater. The test also demonstrated that dissolved oxygen was being added to groundwater (whether or not dissolved oxygen affected MTBE levels through enhanced biodegradation was not established). Since sparging s ability to volatilize MTBE was clearly demonstrated at Site 5, it is very likely that the full-scale sparge system has been at least partially responsible for the MTBE declines noted at the site after 2.5 years of sparging. Air sparging appears beneficial here, but MTBE and BTEX concentrations appear to be following declining trends that existed before the air sparging began; therefore, attributing MTBE reductions directly to the sparging operation cannot be done with confidence at this site. 80

101 7. Remediation Cost Estimates 7.1 Introduction Due to the site-specific nature of remediation operations and the numerous variables involved, it is challenging to quantify remediation costs associated with MTBE-impacted sites; however, some general inferences can be drawn. Key factors that could affect remediation costs include the following: The presence of other contaminants, such as BTEX. The range of concentrations of MTBE and the other contaminants. The location(s) of the contamination (vadose zone, capillary zone, and/or saturated zone). Hydrogeological parameters (e.g., hydraulic conductivity). Presence of on-site structures (e.g., building foundations, utilities, pavement). The size (length and total volume) of the affected area(s). The target cleanup concentrations. As noted in Chapter 4, remediation technologies used to remove BTEX from soil and groundwater, with the exception of in situ bioremediation and natural attenuation, have been demonstrated or predicted to be equally or more effective for removing MTBE. Consequently, for MTBE and BTEX plumes that are equivalent in size, MTBE is generally not expected to cause a significant increase in the lifecycle costs for remediation, assuming equivalent target cleanup levels. In those cases where plume dimensions are similar, but the mass of MTBE is significantly larger than the mass of BTEX, some increase in costs may occur due to the increase in ex situ treatment costs for extracted groundwater or soil vapor containing MTBE. The potential for MTBE plumes to migrate past BTEX plumes increases the scope of site characterization and remediation. An increase in site characterization costs is expected from the additional monitoring wells, sampling, and chemical analysis required. The lower the overlap of BTEX and MTBE plumes, the greater the volume of water that will require remediation. At the minimum, this will impact the costs of pumping. Bearing in mind that generalizing remediation costs for MTBE-impacted sites would be inappropriate due to the site-specific nature of remediation, this section was developed to highlight how the presence of MTBE can potentially impact the costs of remediating gasoline-impacted sites. In addition, it is intended to demonstrate how variables such as the time since the release of the MTBE-blended gasoline, the location of the contamination, and the size of the affected area can affect remediation costs using the technologies discussed in Chapter 4. The costs are discussed in the context of three hypothetical site scenarios, each characterized by a specific release history, extent of contamination, hydrogeological conditions, and contaminant level cleanup goals (20 µg/l for BTEX and MTBE). 81

102 7.2 Methodology For this exercise, three conceptual MTBE and BTEX contamination scenarios were developed. Each scenario is characterized by a specific release history, extent of contamination, hydrogeological conditions, and contaminant level goals. These scenarios and corresponding assumptions were presented to several vendors who provided cost estimates for various combinations of the technologies discussed in Chapter 4 that could practically be implemented at a site. The following technologies were evaluated for remediating MTBE and BTEX sites: SVE coupled with in situ air sparging. SVE coupled with ozonated air sparging. MPE coupled with hydrogen peroxide injection (for MTBE cleanup only). Enhanced bioremediation using Oxygen Release Compound (ORC). Pump-and-treat using GAC for ex situ water treatment. Pump-and-treat using biologically active fluidized bed reactors for water treatment. In situ chemical oxidation with hydrogen peroxide. In situ chemical oxidation with CleanOx Process. Costs for MTBE and BTEX remediation were calculated from capital and operating cost information provided by vendors. Additional costs were added for site work, piping, valves, and electrical work needed to install the system (if applicable), contractor profit, engineering, and contingency. For each outlined scenario, the vendors were asked to provide costs for BTEX and MTBE cleanup, as well as BTEX-only cleanup. Each vendor assumed that the BTEX and MTBE cleanup plume size was identical to the BTEX-only cleanup plume size for each scenario; therefore, the cost estimates are more accurate for sites with younger MTBE releases and older BTEX releases, where MTBE plume sizes are often equal to or less than BTEX plumes. Despite this limitation, the assumption of equal plumes sizes allowed a unit plume cost comparison. In addition, the vendors were asked to estimate the time required for complete remediation using their technology. 7.3 Contamination Scenarios Scenario A Young Shallow Release The first scenario, referred to as the Young Shallow Release, involves MTBE co-mingled with BTEX in shallow groundwater. Figure 7-1 shows a simplified schematic of the scenario with assumed dimensions. Several specific assumptions were provided to the vendors to explain the scenario and the details needed to establish a remediation strategy (Table 7-1). 82

103 20 Feet Free Product 15 Feet v = 3 Feet/Day BTEX and MTBE Figure 7-1. Scenario A LUST creating a dissolved plume 20 feet below the surface. Table 7-1 Plume Characterization of Scenario A Release History Release occurred from an underground gasoline storage tank. Detected within 1 month of its release by routine tank volume monitoring. Approximately 1,000 gallons of product are known to be released. Extent of Contamination Free product is present. Dissolved plume is moving away from the release point with MTBE and BTEX concentration levels of up to 200 mg/l. TPHd and TPHg are at high levels. Dissolved plume is within 200 feet of the release point and has an average plume width of 20 feet. Hydrogeologic Conditions Depth to groundwater is 20 feet. Shallow aquifer is comprised of fine to medium sand and is underlain by a substantial thickness of silty clay beginning at a depth of 35 feet. Total aquifer thickness is 15 feet. Hydraulic conductivity is 10 feet per day. Groundwater velocity is 3 feet per day. Horizontal hydraulic gradient is Vertical hydraulic gradient is negligible. Effective porosity is Fraction of organic carbon is TPHd = Total petroleum hydrocarbons (diesel). TPHg = Total petroleum hydrocarbons (gasoline). Based on discussions with the various vendors, ORC was determined to be economically impractical given the high volumes of ORC that would be stoichiometrically required to maintain an oxygenated subsurface environment for the given high concentration (200 mg/l) of BTEX and MTBE in the groundwater. Furthermore, ORC alone is not appropriate for soil remediation or free-product removal. Pump-and-treat options were also relatively expensive primarily due to the high capital costs of ex situ treatment equipment in this case, GAC and O&M costs for a longer remediation lifecycle. The least expensive options included: Air sparging coupled with SVE (in situ air sparging/sve) (with or without ozone in the air stream). MPE coupled with hydrogen peroxide. In situ chemical oxidation with hydrogen peroxide. 83

104 MPE had the shortest vendor-stated remediation period and, thus, was the most attractive technology due to its potential short remediation lifecycle and ability to attack contamination in both groundwater and soil. The cost of remediation using these eight technologies was stated by the vendors to be predominantly a function of the cross-sectional area of the plume either BTEX or MTBE. Under the assumption that the BTEX/MTBE plume size is identical to the BTEX-only plume, only the vendor estimate for ex situ treatment with GAC noted that the presence of MTBE would increase in situ remediation costs. There was also one vendor of MPE who believed that the presence of MTBE in the subsurface required the addition of hydrogen peroxide to the subsurface to achieve complete cleanup. The other vendors stated that MTBE would not increase remediation costs. Scenario A Conclusions Despite the fact that the vendors assumed that MTBE and BTEX plume sizes were identical, the vendor estimates from this scenario indicate that unit-remediation costs for MTBE were not more expensive than unit remediation costs for BTEX, using most remediation technologies. This suggests that for this scenario, the total cost of remediation will be proportional to the size of the plume and not the contaminants of concern. Of the eight technologies evaluated in this analysis, in situ chemical oxidation or in situ air sparging/sve appear to be the least expensive and most easily implementable remediation options for remediation at sites with plumes similar to the young shallow plume Scenario B Old Large Plume In Scenario B, referred to as the old large plume, MTBE is present in shallow groundwater downgradient from the leading edge of the BTEX plume, as shown in Figure 7-2. Specific assumptions were provided to the vendors to explain the scenario and the details needed to establish a remediation strategy and develop cost estimates (Table 7-2). High Water Level Low Water Level 20 Feet 15 Feet v = 3 Feet/Day BTEX MTBE Figure 7-2. Scenario B LUST with a MTBE plume ahead of a BTEX plume. 84

105 Table 7-2 Plume Characterization of Scenario B Release History Release occurred from an underground storage tank. Only detected during tank replacement. Potential release duration is 5 to 10 years and involves gasoline with lower levels of MTBE, 3 percent by volume. No volume loss was detected during routine tank monitoring; therefore, the leaks are assumed to be very slow. Volume of product is unknown, but potentially large due to the long period of leakage. Extent of Contamination No free product is present. Dissolved BTEX extends 250-feet downgradient from the source at maximum concentrations of 100 µg/l. Dissolved MTBE extends at least 1,500-feet downgradient from the BTEX plume at average concentrations of 1 mg/l and an average plume width of 175 feet. Hydrogeologic Conditions Depth to groundwater is 20 feet. Shallow aquifer is comprised of fine to medium sand and is underlain by a substantial thickness of silty clay beginning at a depth of 35 feet. Total aquifer thickness is 15 feet. Hydraulic conductivity is 10 feet per day. Groundwater velocity is 3 feet per day. Horizontal hydraulic gradient is Vertical hydraulic gradient is negligible. Effective porosity is Fraction of organic carbon is Following discussions with the various vendors, the costs for the two in situ chemical oxidation technologies were very high due to high initial capital and O&M costs required to remediate such a large portion of plume volume. Pump-and-treat options were also costly due to the estimated long period required for remediation. SVE coupled with air sparging using ozone was also a costly method due to the large number of sparge wells required; thus, air stripping/sve, MPE, and ORC appeared to be the most attractive alternatives from a cost perspective. As in the first scenario, except for the vendor who added oxidation to the MPE remediation strategy for only the MTBE scenario, ex situ treatment was the only remediation strategy that included a difference in the unit remediation costs for MTBE versus BTEX plumes; however, unlike the first scenario, the MTBE and BTEX release was significantly more expensive compared to the BTEX-only release because the MTBE plume was significantly longer and contaminated a significantly larger volume of water. Scenario B Conclusions In this older release scenario, it was assumed that the MTBE plume did not naturally attenuate at a similar distance from the source relative to the BTEX-only plume. The addition of MTBE to this scenario, therefore, resulted in significantly higher remediation costs relative to the BTEX-only plume; however, on a unit remediation basis, costs remained fairly similar for the two contamination scenarios. MTBE and BTEX concentrations were low (less than 1 mg/l) and contaminated a large volume of groundwater. For these conditions, cost estimates indicated that enhanced bioremediation was the most cost-effective strategy; 85

106 however, it is recognized that there is still uncertainty associated with MTBE bioremediation and, thus, this approach should be evaluated on a site-specific basis. Alternatively, air stripping/sve and MPE are both cost-effective remediation strategies Scenario C Large Vadose Zone Under Scenario C, MTBE is present in shallow groundwater downgradient from the leading edge of the BTEX plume and in deeper hydraulically connected aquifers (Figure 7-3). Specific assumptions provided to the vendors to explain the scenario and the details needed to establish a remediation strategy and develop cost estimates are tabulated in Table Feet 55 Feet BTEX v = 3 Feet/Day MTBE MTBE Figure 7-3. Scenario C LUST with contamination in shallow and deep aquifers. Despite relatively short time estimates for remediation predicted by all the vendors (less than 3 years), the in situ chemical oxidation technologies were significantly more expensive when compared to other options. This was due to high capital costs associated with the large number of injection wells and the high O&M costs for chemicals use. Due to the high hydraulic conductivity for this scenario and, therefore, high groundwater extraction rates, economies of scale made pump-and-treat relatively cost effective; however, due to the incremental treatment cost of MTBE relative to BTEX, pump-and-treat is not as cost effective as ORC, in situ air sparging/sve, or MPE to attain contaminant level goals. As with the two preceding scenarios, the vendors stated that MTBE did not increase the unit remediation costs of in situ air sparging/sve, ORC, or the hydrogen peroxide in situ chemical oxidation technology. One vendor stated that the presence of MTBE would result in an incremental increase associated with the use of MPE due to the addition of hydrogen peroxide to the subsurface, an addition which was not believed to be needed for the BTEXonly scenario. As in the previous scenario, this scenario involves MTBE-only contamination 86

107 Table 7-3 Plume Characterization of Scenario C Release History Release occurred from underground gasoline storage tanks. Only detected during tank replacement. Release duration is 5 to 10 years. Since no volume loss was detected during routine tank monitoring, the leak is assumed to be very slow. The volume of product lost is unknown, but potentially large due to the long period of leakage. Extent of Contamination No free product is present. Dissolved BTEX extends 250-feet downgradient from the source at maximum concentrations of 200 µg/l. Dissolved MTBE extends at least 1,500-feet downgradient from the BTEX plume with an average width of 150 feet and an average concentration of 100 µg/l MTBE. Although no BTEX appears to be present in the deeper hydraulically connected aquifer, MTBE concentrations are above detection levels. Hydrogeologic Conditions Depth to shallow groundwater is 80 feet. The shallow aquifer is comprised of fine to medium sand with inter-bedded silty clay. The shallow aquifer is underlain by silty clay with inter-bedded sand beginning at a depth of 135 feet, so that the total aquifer thickness is 55 feet. Underlying the silty clay zone is a regional drinking-water aquifer, which is heavily pumped, and is recharged, in part, by leakage through the overlying silty clay zone. Unused water supply wells that may have been improperly abandoned may be present in the area, which has resulted in the hydraulic connectivity between the two aquifers. The hydraulic conductivity of the shallow saturated zone is 30 feet per day. Groundwater velocity is 3 feet per day. Horizontal hydraulic gradient is feet per foot. Vertical hydraulic gradient is feet per foot. Effective porosity in the sand is Fraction of organic carbon in the sand is of groundwater, resulting in higher treatment costs for the MTBE and BTEX scenario relative to the BTEX-only scenario. Scenario C Conclusions Cost estimates for this scenario again indicated that in situ air sparging/sve was the most cost-effective remediation technology; however, this scenario also indicated that as the volume of contaminated water increases and the potential extraction flow rate increases, pump-and-treat becomes more cost effective due to economies of scale (e.g., capital equipment can be amortized over larger volume of treated groundwater); however, it is important to note that ex situ water treatment, required after groundwater is pumped out, does pose an incremental cost associated with the presence of MTBE at a LUFT site of approximately 30 to 60 percent (California MTBE Research Partnership, 1999). 87

108 7.4 Sensitivity Analysis The preceding three scenarios present, in a very general sense, a compilation of a significant amount of information; however, to understand the conclusions from the three scenarios fully, it is necessary to evaluate the effects of lowering the groundwater velocity and/or changing the treatment goal. For each of the scenarios, the vendors assumed a groundwater velocity of 3 feet per day and a treatment goal of 20 µg/l. Under most circumstances, the groundwater velocity will be less than 3 feet per day. Consequently, the vendors were asked to qualitatively discuss how the lower groundwater velocity would affect their cost estimates. The effects were found to vary depending on the technology as follows: A lower groundwater velocity facilitates air sparging and in situ chemical oxidation because the contaminated groundwater has a longer effective contact time with the reactive zone. A lower groundwater velocity increases pump-and-treat remediation time and, thus, the net present value cost of remediation. A lower groundwater velocity increases the costs for enhanced bioremediation due to the barrier approach used (i.e., a longer time is required for the contaminated pore volume to pass through the reactive wall). Similarly, lowering the treatment goal will have a varied effect on remediation technologies although, in each case, a lower goal results in an increase in remediation costs. For pumpand-treat and barrier technologies, a lower goal results in an exponential cost increase. The magnitude of this increase can be predicted using the batch flushing equation presented in Chapter 4. The cost for air sparging technologies will also increase, but not as dramatically as for pump-and-treat. Finally, in situ chemical oxidation strategies will also increase in cost, but the increase will be more linear. This is because a lower treatment goal will require a stoichiometric increase in oxidant. 7.5 Conclusions While each of the scenarios can be studied as a unique dataset, there are some useful insights that can be gained from studying the three scenarios together. In comparing the results from the three scenarios, in situ air sparging combined with SVE was the most cost-effective remediation technology. As the volume of contaminated groundwater increases, the cost-effectiveness of in situ air sparging/sve decreases; however, compared to the other technologies, it continues to be the lowest in cost. The application of ORC was cost-effective for plume management, assuming biodegradation is proven to be effective and reliable. There is still uncertainty associated with MTBE bioremediation and, thus, this approach should be evaluated on a site-specific basis; however, the application of ORC initially appears to be applicable in areas where low (less than 1 mg/l) initial MTBE and BTEX concentrations are present. When MTBE concentrations are high (greater than 1 mg/l), the in situ mixing 88

109 requirements of substrate or nutrient enhancements are substantially increased, resulting in higher costs. In situ chemical oxidation was cost effective for initial high MTBE concentrations; however, as the area or volume of contamination increased and concentrations become lower, the costs increased dramatically. In situ chemical oxidation, therefore, is only effective for the higher MTBE concentrations in source areas and will not be effective for large dissolved plumes. Finally, pump-and-treat has been shown to be less cost effective than other alternative remediation technologies because this technology requires a longer remediation time and because ex situ water treatment costs can be substantial. This cost evaluation verifies the prediction that MTBE remediation costs are likely of a similar magnitude compared to BTEX remediation costs when MTBE and BTEX plumes are equivalent in size. For each of the remediation technologies (excluding GAC and the vendor who combined oxidation with in situ air sparging/sve), costs were nearly the same independent of the presence of MTBE and assuming that both the MTBE and BTEX plumes are identical in size. This suggests that for the majority of sites surveyed, MTBE plumes have not yet migrated a significant distance past the BTEX plumes; otherwise, the presence of MTBE would result in a larger number of sites having higher costs. In summary, this section demonstrates that MTBE and BTEX plumes can be cost-effectively remediated with many technologies. In addition, as long as both the MTBE and BTEX plumes are equivalent in size, remediation costs are not expected to be significantly larger; however, if an MTBE plume has migrated past the associated BTEX plume, the site characterization and remediation costs will increase because of the greater area that needs to be characterized and remediated, as well as the greater volume of water that needs to be treated. 89

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111 8. Conclusions The overall objective of this report is to provide an in-depth evaluation of remediation strategies and technologies capable of addressing the UST cleanup challenges caused by MTBE. Chapter 2 reviewed the fate and transport of MTBE following a release of MTBEcontaining gasoline to the subsurface, with an emphasis on the relevance of these fate and transport characteristics on subsurface remediation strategies. Chapter 3 covered the development of remedial strategies at MTBE-impacted sites. Chapter 4 provided a summary of technologies showing the most promise to remediate MTBE-contaminated soil and groundwater. Chapter 5 provided a brief summary of emerging technologies that show promise of reducing the cost or duration of MTBE remediation. Case studies were presented in Chapter 6. Finally, Chapter 7 addressed the economic challenges associated with remediating sites under a range of release scenarios and site characteristics. The key conclusions drawn from the various chapters are as follows: MTBE s pure component solubility and volume fraction in gasoline are much higher than those of benzene. This will result in higher concentrations of MTBE in the source area relative to benzene. MTBE concentrations in groundwater have been observed at significantly lower levels than are theoretically possible, indicating that the complete dissolution of MTBE from the source is likely to require considerably more time than predicted based on equilibrium partitioning between the gasoline and water; thus, residual gasoline containing MTBE is likely to remain a long-term source of MTBE to groundwater, and more complete remediation of the residual product may be required than is the case for BTEX-only LUFT sites. MTBE plumes may migrate past BTEX plumes given MTBE s higher mobility and lower tendency to naturally biodegrade compared to BTEX compounds. MTBEimpacted sites, therefore, generally require more extensive site investigation. MTBE exhibits low sorption to aquifer materials and exists predominantly in the dissolved phase in aquifers. This property means that MTBE plumes move approximately at the rate of groundwater. It also means that groundwater flushing technologies should be effective in removing MTBE from hydraulically accessible soil matrices. In addition, air sparging is likely to be effective for remediating MTBE because of low adsorption to aquifer materials. The risk-based correction action approach for establishing cleanup standards is applicable at MTBE-impacted sites. There are significant challenges associated with selecting the appropriate combination of technologies to achieve optimum cleanup; however, tools are available to develop an optimum technical solution to the MTBE challenge. 91

112 Due to the properties of MTBE, available remediation technologies are expected to be equally, if not more, effective for removing MTBE from the subsurface relative to BTEX. In particular, pump-and-treat (for dissolved-phase MTBE), SVE (for separate-phase MTBE), MPE, and air sparging are expected to be relatively effective; however, other approaches such as enhanced biodegradation and natural attenuation are not likely to be as effective for MTBE relative to BTEX compounds. Pump-and-treat is expected to be successful in removing MTBE from groundwater due to MTBE s high solubility and low retardation. The enhanced solubility of MTBE relative to BTEX suggests that MTBE subsurface concentrations can be significantly reduced with fewer pore volumes of extracted groundwater. Where site characteristics are favorable and where contaminants are present as residual or free-phase products in the vadose zone, SVE is expected to be at least as effective in removing MTBE relative to BTEX compounds. The rapid application of SVE following a gasoline release is expected to be the most effective method for removing MTBE and BTEX compounds from the vadose zone since it reduces the overall time and cost of remediation. The combination of SVE and pump-and-treat, designated as MPE, allows for the remediation of both soil and groundwater at higher efficiencies than the separate application of each of these systems. Recent field studies suggest that MPE is effective in removing both MTBE and BTEX compounds from subsurface environments. As a result, MPE is increasingly being considered for use at MTBEimpacted sites. Because MTBE is primarily present in the dissolved phase in gasoline-contaminated aquifers, in situ air sparging is promising for MTBE remediation despite the apparent slow rate of MTBE biodegradation and its low Henry s constant. Recent field studies have shown that in situ air sparging effectively reduced MTBE concentrations over a period of 2 years at seven of 10 sites tested. Preliminary results suggest that bioremediation for MTBE-impacted sites have a strong potential for success. Based on the available studies, it is likely that an in situ bioremediation strategy involving direct metabolism, cometabolism, bioaugmentation, or some combination thereof could be applied as a feasible and cost-effective treatment method for MTBE contamination. In situ chemical oxidation is currently considered an emerging technology for MTBE remediation due to the lack of successful published field applications. More research is needed to identify variables that impact the effectiveness of in situ chemical oxidation for MTBE. 92

113 Natural attenuation as a remediation strategy may be less effective for MTBE relative to BTEX compounds due to MTBE s low retardation factor and slow rate of biodegradation, especially under anaerobic conditions. MTBE and BTEX plumes can be cost-effectively remediated using available technologies. In addition, as long as the MTBE and BTEX plumes are equivalent in size, remediation costs are not expected to be significantly larger for MTBE sites compared to BTEX-only sites; however, if an MTBE plume has migrated past the associated BTEX plume, the site characterization and remediation costs will increase because of the greater area that needs to be characterized and remediated and the greater volume of water that needs to be treated. 93

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115 References Acreo, J., S. Haderlein, T. Schmidt, M. Suter, and U. Von Guten (2001). MTBE oxidation by conventional ozonation and the combination ozone/hydrogen peroxide: Efficiency of the processes and bromate formation. Environmental Science and Technology, 35(21): 4,252-4,259. Acuna-Askar, K., A. J. Englande, C. Hu, and G. Jin (2000). Methyl tertiary-butyl ether (MTBE) biodegradation in batch and continuous upflow fixed-biofilm reactors. Water Science and Technology, 42(5-6): Al-Ekabi, H., N. Serpone, E. Pelizzetti, C. Minero, M.A. Fox, and R.B. Draper (1989). Kinetic studies in heterogeneous photocatalysis 2. TiO 2 -mediated degradation of 4-CP alone and in a 3-component mixture of 4-CP, 2,4-DCP and 2,4,5-TCP in air-equilibrated aqueous media. Langmuir, 5: Allen, A.O. (1961). The radiation chemistry of water and aqueous solutions. Van Nostrand- Reinhold, Princeton, New Jersey. American Petroleum Institute (1992). Pump & treat: The petroleum industry perspective. American Petroleum Institute Publication No. 4561, American Petroleum Institute, Washington D.C. American Petroleum Institute (1994). Transport and fate of dissolved methanol, methyl tertiary-butyl ether, and monoaromatic hydrocarbons in a shallow, sand aquifer. American Petroleum Institute Publication No. 4601, American Petroleum Institute, Washington D.C. American Petroleum Institute (1996). A guide to the assessment and remediation of underground petroleum releases, third edition. American Petroleum Institute Publication No. 1628, American Petroleum Institute, Washington D.C. American Petroleum Institute (1999). Personal communication with Tim Buscheck regarding American Petroleum Institute-funded controlled release of MTBE in Michigan led by Dr. Michael Barcelona. American Petroleum Institute (2000). Strategies for characterizing subsurface releases of gasoline containing MTBE. American Petroleum Institute Publication No. 4699, American Petroleum Institute, Washington D.C. American Society for Testing and Materials (1994). Provisional standard guide for risk-based corrective action. American Society for Testing Materials, West Conshohocken, Pennsylvania. 95

116 American Society for Testing and Materials (1998). Provisional standard guide for riskbased corrective action, PS American Society for Testing and Materials, West Conshohocken, Pennsylvania. American Water Works Association Research Foundation (2001). Progress Report 5: Removal of MTBE advanced oxidation processes. Malcolm Pirnie, Inc., Phoenix, Arizona. Anderson, W.C. (1994). Innovative site remediation technology: Chemical treatment. American Academy of Environmental Engineers, Annapolis, Maryland. Annesini, M., F. Gironi, and B. Monticelli (2000). Removal of oxygenated pollutants from wastewater by polymeric resins: Data on adsorption equilibrium and kinetics in fixed beds. Water Resources, 34(11): 2,989-2,996. Arulanantham, R., T. Buscheck, and M. Kavanaugh (1999). Assessment and management of MTBE-impacted sites. Results from an informal interview conducted during National Ground Water Association courses and companion California Regional Board courses. Barker, J.F., C.E. Hubbard, and L.A. Lemon (1990). The influence of methanol and MTBE on the fate and persistence of monoaromatic hydrocarbons in groundwater. Proceedings, Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection, and Restoration, National Ground Water Association, Westerville, Ohio, pp Bass, D.H. (1996). Performance of air sparging systems and review of case studies. Presented at Sparge-a-Rama, March 12, Bass, D.H., B. Riley, and T. Farrell (1994). Remediation of MTBE in groundwater. Proceedings, Superfund XV Conference, Volume one. Bass, D.H., and T.E. Sylvia (1992). Heated air stripping for the removal of MTBE from recovered groundwater. Petroleum hydrocarbons and organic chemicals in ground water. Ground Water Management, pp Borden, R.C., R.A. Daniel, L.E. LeBrun IV, and C.W. Davis (1997). Intrinsic biodegradation of MTBE and BTEX in a gasoline-contaminated aquifer. Water Resources Research, 33(5): 1,105-1,115. Boronina, T., K.J. Klabunde, and G. Sergeev (1995). Destruction of organhalides in water using metal particles: Carbon tetrachloride/water reactions with magnesium, tin, and zinc. Environmental Science and Technology, 29(6): 1,511. Bradley, P.M., J.E. Landmeyer, and F.H. Chapelle (1999). Aerobic mineralization of MTBE and tert-butyl alcohol by stream bed sediment microorganisms. Environmental Science and Technology, 33(11): 1,877-1,

117 Bradley, P., F. Chapelle, and J. Landmeyer (2001a). Methyl t-butyl ether mineralization in surface-water sediment microcosms under denitrifying conditions. Applied Environmental Microbiology, 67(4): 1,975-1,978. Bradley, P., F. Chapelle, and J. Landmeyer (2001b). Effect of redox conditions on MTBE biodegradation in surface water sediments. Environmental Science and Technology, 35(23): 4,643-4,647. Bruce, C., C. Gilbert, R. Johnson, and P. Johnson (1998). Methyl tert-butyl ether removal by in situ air sparging in physical model studies. Physical, chemical, and thermal technologies: Remediation of chlorinated and recalcitrant compounds, G. Wickramanayake and R. Hinchee, eds., Battelle Press, Columbus, Ohio, pp Burken, J., and J. Schnoor (1998). Predictive relationships for uptake of organic contaminants by hybrid poplar trees. Environmental Science and Technology, 32(21): 3,379-3,385. Buscheck, T.E., D.J. Gallagher, D.L. Kuehne, and C.R. Zuspan (1998). Occurrence and behavior of MTBE in groundwater. Proceedings, National Ground Water Association Southwest Focused Ground Water Conference: Discussing the Issue of MTBE and Perchlorate in Ground Water, National Ground Water Association, Westerville, Ohio. California MTBE Research Partnership (1999). Treatment technologies for removal of methyl tertiary-butyl ether (MTBE) from drinking water: Air stripping, advanced oxidation processes, granular activated carbon, synthetic resin sorbents. National Water Research Institute, Fountain Valley, California. California MTBE Research Partnership (2001). Treating MTBE-impacted drinking water using granular activated carbon. National Water Research Institute, Fountain Valley, California. Cater, S., M. Stefan, J. Bolton, and A. Safarzadeh-Amiri (2000). UV/H 2 O 2 treatment of methyl tert-butyl ether in contaminated waters. Environmental Science and Technology, 34(4): Chang, D. (1999). Personal communication regarding on-going bioaugmentation field experiment using microbes identified by Kate Scow. University of California, Davis, Department of Civil and Environmental Engineering, Davis, California (530) Chang, P., and T. Young (2000). Kinetics of methyl tert-butyl ether degradation and byproduct formation during UV/hydrogen peroxide water treatment. Water Research, 34(8): 2,233-2,240. Chapelle, F.H. (1992). Groundwater microbiology and biochemistry. Wiley and Sons, New York, New York. 97

118 Chen, C., H. Hohnbaum, A. Smith, K. Scow, D. Chang, and K. Jackson (2000). Prepared extended abstract. American Chemical Society, 40(1): Church, C.D., P.J. Tratnyek, J.F. Pankow, J.E. Landmeyer, A.L. Baehr, M.A. Thomas, and M. Schirmer (1999). Effects of environmental conditions on MTBE degradation in model column aquifers. Proceedings, Technical Meeting of United States Geological Survey, Toxic Substances Hydrology Program, Charleston, South Carolina, 7-12 March 1999, Water Resources Investigations Report C, Volume 3, U.S. Geological Survey, West Trenton, New Jersey, pp Church, C.D., L.M. Isabelle, J.F. Pankow, D.L. Rose, and P.G. Tratnyek (1997a). Method for determination of methyl tert-butyl ether and its degradation products in water. Environmental Science and Technology, 31(12): 3,723-3,726. Church, C., L. Isabelle, J. Pankow, P. Tratnyek, and D. Rose (1997b). Assessing the in situ degradation of methyl tert-butyl ether (MTBE) by product identification at the sub-ppm level using direct aqueous injection GC/MS. American Chemical Society, 37(1): Cookson, J. (1995). Bioremediation engineering: Design and application. R.R. Donnelley & Sons Company, San Francisco, California. Creek, D., and J. Davidson (1998). The performance and cost of MTBE remediation technologies. Proceedings, Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection, and Remediation, National Ground Water Association, Westerville, Ohio, pp Crittenden, J.C., D.R. Hokanson, D.W. Hand, T.N. Rogers, and E.J. Oman (1997) Environmental technologies design option tools (ETDOT) for the clean process advisory systems (CPAs) adsorption, aeration and physical properties software. National Center for Clean Industrial and Treatment Technologies, Michigan Technological University, Houghton, Michigan. Crittenden, J.C, D.R. Hokanson, D.W. Hand, T.N. Rogers, and E.J. Oman (1996). Environmental technologies design option tools (ETDOT) for the clean process advisory systems Adsorption, aeration and physical properties software. National Center for Clean Industrial and Treatment Technologies, Michigan Technical University, Houghton, Michigan. Crowley, J. (1998). Personal communication. Santa Clara Water District, San Jose, California (408)

119 Davidson, J.M., and R. Parsons (1996). Remediating MTBE with current and emerging technologies. Proceedings, Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection, and Remediation, National Ground Water Association, Westerville, Ohio, pp Davidson, J.M. (1995). Fate and transport of MTBE The latest data. Proceedings, Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection, and Remediation, Conference and Exposition, National Ground Water Association, Westerville, Ohio, pp Davis, S., and S. Powers (2000). Alternative sorbents for removing MTBE from gasolinecontaminated ground water. Journal of Environmental Engineering, 126(4): Deeb, R., S. Nishino, J. Spain, H. Hu, K. Scow, and L. Alvarez-Cohen (2000a). MTBE and benzene biodegradation by a bacterial isolate via two independent monooxygenase-initiated pathways. American Chemical Society, 40(1): Deeb, R., K. Scow, and L. Alvarez-Cohen (2000b). Aerobic MTBE biodegradation: An examination of past studies, current challenges and future research directions. Biodegradation, 11(2-3): Deeb, R., J. Hanson, K. Scow, and L. Alvarez-Cohen (2001a). Substrate interactions in BTEX and MTBE mixtures by an MTBE-degrading isolate. Environmental Science and Technology, 35(2): Deeb, R., S. Sue, D. Spiers, and M. Kavanaugh (2001b). Field studies to demonstrate the cost and performance of air stripping for the removal of MTBE from groundwater. Proceedings, Air and Water Management Association s 94th Annual Conference and Exhibition, Air and Waste Management Association, Pittsburgh, Pennsylvania. Dobbs, R.A., and J.H. Cohen (1980). Carbon adsorption isotherms for toxic organics. EPA-600/ , United States Environmental Protection Agency, Washington D.C. Dupasquier, D., S. Revah, and A. Auria (2002). Biofiltration of methyl tert-butyl ether vapors by cometabolism with pentane: Modeling and experimental approach. Environmental Science and Technology, 36(2): Einarson, M., and D. Mackay (2001). Predicting impacts of groundwater contamination. Environmental Science and Technology, 35(3): 66A-73A. Einarson, M.D. (1998). Natural contaminant pathways in shallow Santa Clara Valley sediments: Implications and opportunities for in situ remediation. Presentation at the Groundwater Resource Association Meeting, on May 6, 1998, in Oakland, California, Groundwater Resource Association of California, Sacramento, California. 99

120 Envirogen (1999). Personal communication with Rob Steffan regarding laboratory and preliminary field results on the use of propane to cometabolically degrade MTBE. Envirogen, Lawrenceville, New Jersey (609) Eweis, J.N., D.P. Chang, E.D. Schroeder, K.M. Scow, R.L. Morton, and R.C. Caballero (1997). Meeting the challenge of MTBE biodegradation. Presented at the 90th Annual Meeting and Exhibition of the Air and Waste Management Association, on June 8-13, 1997, in Toronto, Canada, Air and Waste Management Association, Pittsburgh, Pennsylvania. Finneran, K.T., and D.R. Lovley (2001). Anaerobic degradation of methyl tert-butyl ether (MTBE) and tert-butyl alcohol (TBA). Environmental Science and Technology, 35(9): 1,785-1,790 Fortin, N.Y., and M.A. Deshusses (1999a). Treatment of methyl tert-butyl ether vapors in biotrickling filters. 1. Rector startup, steady-state performance, and culture characteristics. Environmental Science and Technology, 33(17): 2,980-2,986. Fortin, N.Y., and M.A. Deshusses (1999b). Treatment of methyl tert-butyl ether vapors in biotrickling filters. 2. Analysis of the rate-limiting step and behavior under transient conditions. Environmental Science and Technology, 33(17): 2,987-2,991. Freeze, R.A., and J.A. Cherry (1979). Groundwater. Prentice-Hall, Inc., Englewood Cliffs, New Jersey. Freeze, R.A., and J.A. Cherry (1989). What has gone wrong? Ground Water, 27(4): Fujiwara, Y., T. Kinoshita, H. Sato, and I. Kojima (1984). Biodegradation and bioconcentration of alkylethers. Yugagaku, 33: Garnier, P., R. Auria, C. Auger, and S. Revah (1999). Cometabolic biodegradation of methyl t-butyl ether by Pseudomonas aeruginosa grown on pentane. Applied Microbiological Biotechnology, 51: Gillham, R.W. (1996). Recent advances in groundwater pollution control and remediation, M.M. Aral, ed., Kluwer Academic Publishers, Dordrecht, The Netherlands. Grasso, D. (1993). Hazardous waste site remediation: Source control. Lewis Publishers, Boca Raton, Florida. Griend, R., and M.C. Kavanaugh (1996). Evaluation of the effects of methyl tertiary-butyl ether on leaking underground fuel tank investigation and remediation programs. Presentation at the Western States Petroleum Association Workshop on MTBE Ground Water Issues, in Los Angeles, California, Western States Petroleum Association, Glendale, California. 100

121 Hanson, J.R., C.E. Ackerman, and K.M. Scow (1999). Biodegradation of methyl tert-butyl ether by a bacterial pure culture. Applied Environmental Microbiology, 65(11): 4,788-4,792. Happel, A.M., E.H. Beckenbach, and R.U. Halden (1998). An evaluation of MTBE impacts to California groundwater resources. Publication URCL-AR , Lawrence Livermore National Laboratory, Livermore, California. Hardison, L.K., S.S. Curry, L.M. Ciuffetti, and M.R. Hyman (1997). Metabolism of diethyl ether and cometabolism of methyl tert-butyl ether by a filamentous fungus, a Graphium sp. Applied Environmental Microbiology, 63(8): 3,059-3,067. Hatzinger, P., R. Steffan, and S. Drew (2001). New technologies for MTBE remediation: Biostimulation, bioaugmentation, and bioreactors. Contaminated Soil, Sediment and Water Journal, Association of Environmental Health and Sciences, pp Heath, W., and E. Garcia (1999). Downloaded in October 2000 (preparation date, March). Hinchee, R.E. (1994). Air sparging for site remediation. Lewis Publishers, Boca Raton, Florida. Hong, S., C.M. Duttweiler, and A.T. Lemley (1999). Analysis of methyl tert-butyl ether and its degradation products by direct aqueous injection onto gas chromatography with mass spectrometry or flame ionization detection systems. Journal of Chromatography, A 857(1): (12). Hong, M., W. Farmayan, I. Dortch, C. Chiang, S. McMillan, and J. Schnoor (2001). Phytoremediation of MTBE from a groundwater plume. Environmental Science and Technology, 35(6): 1,231-1,239. Hurt, K.L., J.T. Wilson, F.P. Beck, and J.S. Cho (1999). Anaerobic biodegradation of MTBE in a contaminated aquifer. Natural attenuation of chlorinated solvents, petroleum hydrocarbons, and other organic compounds, B.C. Alleman and A. Leeson, eds., Batelle Press, Columbus, Ohio. Hyman, M., P. Kwon, K. Williamson, and K. O Reilly (1998). Cometabolism of MTBE by alkane-utilizing microorganisms. Proceedings, First International Conference on Remediation of Chlorinated and Recalcitrant Compounds, G. B. Wickramanayake and R. E. Hinchee, eds., Battelle Press, Columbus, Ohio, pp Integrated Science and Technology, Inc. (IST) (1999). Comparative MTBE versus benzene plume behavior. BP Oil Company Florida facilities, 1349 Old Highway 41, Suite 225, Marietta, GA

122 Javanmardian, M., and H.A. Glasser (1997). In situ biodegradation of MTBE using biosparging. Extended Abstracts, 213th American Chemical Society National Meeting, American Chemical Society, Columbus, Ohio, pp Jensen, H.M., and E. Arvin (1990). Solubility and degradability of the gasoline additive MTBE, methyl tert-butyl-ether, and gasoline compounds in water. Contaminated Soil 90, F. Arendt, M. Hinsenveld, and W.J. Van Den Brink, eds., Kluwer Academic Publishers, Dordrecht, The Netherlands, pp Johnson, P. (1998). Assessment of contributions of volatilization and biodegradation to in situ air sparging performance. Environmental Science and Technology, 32(2): Johnson, P.C., and R.A. Ettinger (1997). Considerations for the design of in situ vapor extraction systems: Radius of influence vs. zone of remediation. Subsurface Restoration, C.H. Ward, J.A. Cherry, and M.R. Scalf, eds., Ann Arbor Press, Inc., Chelsea, Michigan. Johnson, P.C., C.C. Stanley, M.D. Kemblowski, J.D. Colthart, and D.L Byers (1990). A practical approach to the design, operating, and monitoring of in situ soil-venting systems. Groundwater Monitoring Review, 10(2): Kang, J., H. Hung, A. Lin, and M. Hoffman (1999). Sonolytic destruction of methyl tertbutyl ether by ultrasonic irradiation: The role of O 3, H 2 O 2, frequency, and power density. Environmental Science and Technology, 33(18): 3,199-3,205. Kang, J.W., and M.R. Hoffmann (1998). Kinetics and mechanism of the sonolytic destruction of MTBE by ultrasonic irradiation in the presence of ozone. Journal of Environmental Science and Technology, 32(20): 3,194-3,199. Kavanaugh, M., Z. Chowdhury, S. Kommineni, S. Liang, J. Min, J.P. Croue, N. Corin, G. Amy, E. Simon, W. Cooper, P. Tornatore, and M. Nickelson (2003). Removal of MTBE with advanced oxidation processes. American Water Works Research Foundation report, Denver, Colorado. Keller, A.A., O.C. Sandall, R.G. Rinker, M.M. Mitani, B.G. Bierwagen, and M.J. Michael. (1998). An integral cost-benefit analysis of gasoline formulation meeting California Phase II Reformulated Gasoline requirements, in Health and Environmental Assessment of MTBE, volume three, University of California Toxics Research and Teaching Program, University of California, Davis. Koenigsberg, S., C. Sandefur, W. Mahaffey, M. Deshusses, and N. Fortin (1999). Peroxygen mediated bioremediation of MTBE. Proceedings, Fifth International In Situ On-Site Bioremediation Symposium, B. C. Alleman, and A. Leeson, eds., Battelle Press, Columbus, Ohio, pp

123 Kormann, C., D.W. Bahnemann, and M.R. Hoffman (1988). Photocatalytic production of hydrogen peroxide and organic peroxides in aqueous suspensions of TiO2, ZnO, and desert sand. Environmental Science and Technology, 22: LaGrega, M.D., P.L. Buckingham, and J.C. Evans (1994). Hazardous waste management. McGraw-Hill, Inc., New York, New York. Landmeyer, J., F. Chapelle, H. Herlong, and P. Bradley (2001). Methyl tert-butyl ether biodegradation by indigenous aquifer microorganisms under natural and artificial oxic conditions. Environmental Science and Technology, 35(6): 1,118-1,126. Leethem, J. (2001). In situ chemical oxidation of MTBE and BTEX in soil and groundwater: A case study. Soil and Sediment Contamination: An International Journal, Spring 2001 Special Oxygenated Fuels Issue, pp Mace, R.E., and W.J. Choi (1998). The size and behavior of MTBE plumes in Texas. Proceedings, Conference on Petroleum Hydrocarbons and Organic Chemicals in Ground Water, National Ground Water Association, Westerville, Ohio. Mace, R.E., S. Fisher, D.M. Welch, S.P. Parra (1997). Extent, mass, and duration of hydrocarbon plumes from leaking petroleum storage tank sites in Texas. Bureau of Economic Geology, The University of Austin, Austin, Texas. Mackay, D. (1998). Fate and transport of MTBE in groundwater. Presentation at the Santa Clara Valley Water District s Leaking Underground Storage Tank Oversight Program, Santa Clara Valley Water District, San Jose, California. Malcolm Pirnie, Inc. (1998a). Fate and transport of MTBE in groundwater: Literature review. Malcolm Pirnie, Inc., Oakland, California. Malcolm Pirnie, Inc. (1998b). Technical memorandum: Modeling the volatilization of methyl tertiary-butyl ether (MTBE) from surface impoundments. Malcolm Pirnie, Inc., Oakland, California. McCulloch, B (1996). Manufacturer Literature. McCulloch Environmental Equipment Sales, Vacaville, California. McKinnon, R., and J. Dyksen (1984). Removing organics from groundwater through aeration plus GAC. Journal American Water Works Association, 76(5): Mo, K., C.O. Lora, A.E. Wanken, M. Javanmardian, X. Yang, and C.F. Kulpa (1997). Biodegradation of methyl t-butyl ether by pure bacterial cultures. Applied Microbial Biotechnology, 47:

124 Mortensen, A., K. Jensen, T. Sonnenborg, and E. Arvin (2000). Laboratory and numerical investigations of air sparging using MTBE as a tracer. Groundwater Monitoring and Remediation, 20(4): National Research Council (1993). In Situ bioremediation: When does it work? National Academy Press, Washington, D.C. National Research Council (1994). Alternatives for ground water cleanup. National Academy Press, Washington, D.C. National Research Council (2000). Natural attenuation for groundwater remediation. National Academy Press, Washington, D.C. Newman, L., S. Doty, K. Gery, P. Heilman, I. Muiznieks, T. Shang, S. Siemieniec, S. Strand, X. Wang, A. Wilson, and M. Gordon (1998). Phytoremediation of organic contaminants: A review of phytoremediation research at the University of Washington. Journal of Soil Contamination, 7(4): Newman, L., A. Wilson, L. Capron, S. Strand, P. Heilman, and M. Gordon (2000). Phytoremediation of methyl tert-butyl ether (MTBE). American Chemical Society, 40(1): Newmark R.L., and R.D. Aines (1995). Summary of the Gasoline Spill Demonstration- Dynamic Underground Stripping Project. UCRL-ID , Lawrence Livermore National Laboratory, Livermore, California. Nyer, E.K., and D.B. Vance (2001). Nano-scale iron for dehalogenation. Ground Water Monitoring and Remediation, 21(2): Ott, D.E. (1999). Downloaded in October Pankow, J.F. (1996). Calculated volatilization rates of fuel oxygenate compounds and other gasoline-related compounds from rivers and streams. Chemosphere, 33(5): Pankow, J.F., and J. Cherry (1996). Dense chlorinated solvents and other DNAPLs in groundwater. Waterloo Press, Portland, Oregon. Pankow, J. (1998). Dissolution of gasoline components. Presentation at the California MTBE Research Partnership s MTBE Source Protection Workshop in San Ramon, California. Park, K., and R.M. Cowan (1997a). Biodegradation of gasoline oxygenates. In Situ and On-Site Bioremediation: Volume 1, Battelle Press, Columbus, Ohio. 104

125 Park, K., and R.M. Cowan (1997b). Effects of oxygen and temperature on the biodegradation of MTBE. American Chemical Society, 37(1): Peargin, T.R. (1998). An empirical study of MTBE, benzene, and xylene groundwater remediation rates. Proceedings, 1998 Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection, and Remediation, National Ground Water Association, Westerville, Ohio. Pederson, T., and J. Curtis (1991). Soil vapor extraction technology. United States Environmental Protection Agency, Washington, D.C. Rice, D.W., R.D. Grose, J.C. Michaelsoen, B.P. Dooher, D.H. MacQueen, S.J. Cullen, W.E. Kastenberg, L.E. Everett, and M.A. Marino (1995.) California leaking underground storage tank (LUFT) historical case analyses. Lawrence Livermore National Laboratory, Livermore, California. Rixey, W.G., and S. Joshi (2000). Dissolution of MTBE from a residually trapped gasoline source - a summary of research results. American Petroleum Institute, Washington, D.C. Rubin, E., and A. Ramaswami (2001). The potential for phytoremediation of MTBE. Water Research, 35(5): 1,348-1,353. Safarzadeh-Amiri, A. (2001). O 3 /H 2 O 2 treatment of methyl tert-butyl ether (MTBE) in contaminated waters. Water Research, 35(15): 3,706-3,714. Salanitro, J.P., C.S. Chou, H.L. Wisniewski, and T.E. Vipond (1998). Perspectives on MTBE biodegradation and the potential for in situ aquifer bioremediation. Proceedings, Southwestern Regional Conference of the National Groundwater Association, National Ground Water Association, Westerville, Ohio. Salanitro, J.P., L.A. Diaz, M.P. Williams, and H.L. Wisniewski (1994). Isolation of a bacterial culture that degrades methyl t-butyl ether. Applied Environmental Microbiology, 60(7): 2,593-2,596. Salanitro, J., P. Johnson, G. Spinnler, P. Manner, H. Wisniewski, and C. Bruce (2000). Filedscale demonstration of enhanced MTBE bioremediation through aquifer bioaugmentation and oxygenation. Environmental Science and Technology, 34(19): 4,152-4,162. Salanitro, J., G. Spinnler, P. Maner, H. Wisniewski, and P. Johnson (1999a). Potential for MTBE bioremediation: In situ inoculation of specialized cultures. Proceedings, Conference on Petroleum Hydrocarbons and Organic Chemicals in Ground Water: Prevention, Detection and Remediation Conference, National Ground Water Association, Westerville, Ohio. 105

126 Salanitro, J.P., G.E. Spinnler, C.C. Neaville, P.M. Maner, S.M. Stearns, P.C. Johnson, and C. Bruce (1999b). Demonstration of the enhanced MTBE bioremediation (EMB) in situ process. Proceedings, In Situ and On-Site Bioremediation Conference, National Ground Water Association, Westerville, Ohio. Schirmer, M., and J.F. Barker (1998). A study of long-term MTBE attenuation in the Borden aquifer, Ontario, Canada. Ground Water Monitored Remediation, 18(2): Schirmer, M., J.F. Barker, and C.E. Hubbard (1998). Delineation and characterization of the Borden MTBE plume: An evaluation of 8 years of natural attenuation processes. American Petroleum Institute Publication 4668, American Petroleum Institute, Washington, D.C. Schirmer, M., B.J. Butler, J.F. Barker, C.D. Church, and K. Schirmer (1999). Evaluation of biodegradation and dispersion as natural attenuation processes of MTBE and benzene at the Borden field site. Physics and Chemistry of the Earth, Part B, 24(6): Schlimm, C., and E. Heitz (1996). Development of a wastewater treatment process reductive dehaloagenation of chlorinated hydrocarbons by metals. Environmental Progress, 15(1): 38. Schwarzenbach, R.P., P.M. Gschwend, and D.M. Imboden (1993). Environmental organic chemistry. John Wiley & Sons, Inc., New York, New York. Siddiqui, M., G.L. Amy, W.J. Cooper, C.N. Kurucz, T.D. Waite, and M.G. Nickelsen (1996a). Bromate removal by high energy electron beam process (HEEB). Journal American Water Works Association, 88(10): Siddiqui, M., G.L. Amy, and W.J. Cooper (1996b). Bromate ion removal by electric-arc discharge and high energy electron beam processes. Water disinfection and natural organic matter, R.A. Minear and G.L. Amy, eds., American Chemical Society, Symposium Series 649, Washington, D.C., pp Siegel, W.H. (1994). Design, construction, and operation of dynamic underground stripping facilities at Lawrence Livermore National Laboratory. Dynamic underground stripping project: LLNL gasoline spill demonstration report. Lawrence Livermore National Laboratory, Livermore, California, p Simmons, K., P. Kostecki, and E. Calabrese (1998). Ninth annual state by state soil cleanup standards. Soil and Groundwater Cleanup, pg Simmons, K., P. Kostecki, and E. Calabrese (1999). Ninth annual state by state groundwater cleanup standards. Soil and Groundwater Cleanup, pg 10-21,

127 Sjogren, J.C. (1995). Inactivation of phage MS-2 by titanium dioxide photocatalysis. Ph.D. Dissertation, Department of Chemical and Environmental Engineering, University of Arizona, Tucson. Skov, E.R., J.A. Pisani, and S.E. Beale (1997). Industrial wastewater treatment using hydroxyl radical oxidation by hydraulically induced cavitation. Proceedings of the American Institute of Civil Engineers Spring National Meeting in Houston, Texas, Session 86a, American Institute of Civil Engineers, New York, New York. Squillace P.J., J.F. Pankow, N.E. Korte, and J.S. Zogorski (1997). Review of the environmental behavior and fate of methyl tert-butyl ether. Environmental Toxicology and Chemistry, 16(9): 1,836-1,844. Stefan, M., J. Mack, and J. Bolto (2000). Degradation pathways during the treatment of methyl tert-butyl ether by the UV/H 2 O 2 process. Environmental Science and Technology, 34(4): Steffan, R.J., K. McClay, S. Vainberg, C.W. Condee, and D. Zhang (1997). Biodegradation of the gasoline oxygenates methyl tert-butyl ether, ethyl tert-butyl ether, and tert-amyl methyl ether by propane-oxidizing bacteria. Applied Environmental Microbiology, 63(11): 4,216-4,222. Steffan, R., P. Hatzinger, Y. Farhan, and S. Drew (2001). In situ and ex situ biodegradation of MTBE and TBA in contaminated groundwater. Proceedings, Petroleum Hydrocarbons and Organic Chemicals in Ground Water, National Ground Water Association, Westerville, Ohio. Stocking, A., R.A. Deeb, A. Flores, W. Stringfellow, J. Talley, R. Brownell, and M. Kavanaugh (2000). Bioremediation of MTBE: A review from a practical perspective. Biodegradation, 11(2-3): Suflita, J.M., and M.R. Mormile (1993). Anaerobic biodegradation of known and potential gasoline oxygenates in the terrestrial subsurface. Environmental Science and Technology, 27(6): Suflita, J.M., and G.W. Sewell (1991). Anaerobic biotransformation of contaminants in the subsurface. United States Environmental Protection Agency, Washington, D.C. Sun, P.T., J.P. Salanitro, and W.T. Tang (1996). Fate and biokinetics of methyl-t-butyl ether in activated sludge systems and its engineering significance. Proceedings, 51st Purdue Industrial Waste Conference Proceedings, Ann Arbor Press, Inc., Chelsea, Michigan, pp

128 Suthersan, S.S. (1997). Remediation engineering: Design concepts. Lewis Publishers, Boca Raton, Florida. Tang, W.T., and P.T. Sun (1997). Field evaluation of biological and non-biological treatment technologies to remove MTBE/oxygenates from petroleum product terminal wastewaters. Shell Development Company, Houston TX. American Petroleum Institute Publication 4655, American Petroleum Institute, Washington, D.C. Thomas, J.M., J.T. Wilson, and C.H. Ward. (1997). Microbial processes in the subsurface. Subsurface Restoration, C.H. Ward, J.A. Cherry, and M.R. Scalf, eds., Ann Arbor Press, Inc., Chelsea, Michigan. Travis, C.C., and C.B. Doty (1990). Can contaminated aquifers at Superfund sites be remediated? Environmental Science and Technology, 24(1): 1,464-1,466. Turchi, C.S., and D.F. Ollis (1990). Photocatalytic degradation of organic water contaminants: mechanisms involving hydroxyl radical attack. Journal Catalysis, 122: Udell K.S., N. Sitar, J.R. Hunt, and D.L. Stewart (1991). Process for in situ decontamination of subsurface soil and groundwater. United States Patent 5,018,576. Udell, K.S. (1994). Thermally enhanced removal of liquid hydrocarbon contaminants from soils and groundwater. Dynamic underground stripping project: LLNL gasoline spill demonstration report. Lawrence Livermore National Laboratory, Livermore, California, p. 5. Udell, K.S., and D.L. Stewart (1989). Mechanisms of in situ remediation of soil and groundwater contamination by combined steam injection and vacuum extraction. Proceedings, Symposium on Thermal Treatment of Radioactive and Hazardous Waste, AIChE Annual Meeting, November 6, San Francisco, California. Udell, K.S., and L.D. Stewart (1990). Combined Steam injection and vacuum extraction for aquifer cleanup. Proceedings, Conference of the International Association of Hydrogeologists, International Association of Hydrogeologists, Kenilworth, United Kingdom. United States Environmental Protection Agency (1989a). Evaluation of ground-water extraction remedies, volume one: Summary report. United States Environmental Protection Agency, Washington, D.C. United States Environmental Protection Agency (1989b). Evaluation of ground-water extraction remedies, volume two: Case studies. United States Environmental Protection Agency, Washington, D.C. 108

129 United States Environmental Protection Agency (1989c). Evaluation of ground-water extraction remedies, volume three: General site data, data base reports. United States Environmental Protection Agency, Washington, D.C. United States Environmental Protection Agency (1994). Methods for monitoring pump-andtreat performance. United States Environmental Protection Agency, Washington, D.C. United States Environmental Protection Agency (1995). Bioventing principles and practice. United States Environmental Protection Agency, Washington, D.C. United States Environmental Protection Agency (1996). Cleaning up the nation s waste sites: Markets and technology trends: 1996 edition. United States Environmental Protection Agency, Washington, D.C. United States Environmental Protection Agency (1997). Analysis of selected enhancements for soil vapor extraction. United States Environmental Protection Agency, Washington, D.C. United States Environmental Protection Agency (1998). Field applications of In Situ remediation technologies: Chemical oxidation. United States Environmental Protection Agency, Washington, D.C. United States Geologic Survey (1998). Simulating transport of volatile organic compounds in the unsaturated zone using the computer model R-UNSAT. Vainberg, S., A.P. Togna, P.M. Sutton, and R.J. Steffan (2002). Treatment of MTBEcontaminated water in a fluidized bed bioreactor. Journal of Environmental Engineering, in press. Walter, K. (1998). They all like it hot: Faster cleanup of contaminated soil and groundwater. Science and technology review. Lawrence Livermore National Laboratory, Livermore, California. Watts, R.J. (1998a). Hazardous wastes: Sources, pathways, receptors. John Wiley & Sons, Inc., New York, New York. Watts, R.J. (1998b). In situ application of modified Fenton s reagent spells success for BTEX, MTBE site cleanup. The Hazardous Waste Consultant, 16(3). Weaver, J.W, J.T. Haas, and J.T. Wilson (1996). Analysis of the gasoline spill at East Patchogue, New York. Proceedings, Non-Aqueous Phase Liquids (NAPLs) in Subsurface Environment: Assessment and Remediation, L. Reddi, ed. American Society of Civil Engineers, Washington, D.C., November 12-14, pp

130 Weaver, J.W., J.E. Haas, and C.B. Sosik (1999). Characteristics of Gasoline Releases in the Water Table Aquifer of Long Island. Proceedings, 1999 Petroleum Hydrocarbons and Organic Chemicals in Ground Water, API/NGWA, Houston, Texas, November 17-19, pp White, G., N. Russell, and E. Tidswell (1996). Bacterial scission of ether bonds. Microbiology Review, 60: Wilson, R., D. Mackay, and K. Scow (2002). In situ biodegradation supported by diffusive oxygen release. Environmental Science and Technology, 36(2): Yang, X., D. Tsao, M. Javanmardian, and H.A. Glasser (1998). Development of costeffective MTBE in situ treatment technologies. Presented at the ATV Vintermode om Grundvandsforurenin in Belje, Denmark. Yeh, C.K. (1992). Degradation of gasoline oxygenates in the subsurface. Ph.D. Dissertation, Virginia Polytechnic Institute and State University, Blacksburg, Virginia. Yeh, C.K., and J.T. Novak (1995). The effect of hydrogen peroxide on the degradation of methyl and ethyl tert-butyl ether in soils. Water Environment Research, 67(5): Zepp, R.G. (1988). Factors affecting the photochemical treatment of hazardous waste. Environmental Science and Technology, 22(3): 256. Zhang, W.X., C.B. Wang, and H.L. Lien (1998). Treatment of chlorinated organic contaminants with nanoscale bimetallic particles. Catalysis Today, 40: Zhang, Q., L. Davis, and L. Erickson (2001). Transport of methyl tert-butyl ether through alfalfa plants. Environmental Science and Technology, 35(4): Zheng, C., G.D. Bennett, and C.B. Andrews (1991). Analysis of groundwater remedial alternatives at a Superfund site. Ground Water, 29(6): Zheng, C., G.D. Bennett, and C.B. Andrews (1992). Reply to the preceding discussion by Robert D. McCaleb of Analysis of Groundwater Remedial Alternatives at a Superfund Site. Ground Water, 30(3): Zogorski, J., A. Morduchowitz, A. Baehr, B. Bauman, D. Conrad, R. Drew, N. Korte, W. Lapham, J. Pankow, and E. Washington (1997). Interagency Assessment of Oxygenated Fuels. Report coordinated by Office of Science and Technology Policy (OSTP), National Science and Technology Council, Executive Office of the President, Washington, DC. 110

131 Appendix: Summaries of Case Study Sites A. Site 1 Contamination History SOURCE MTBE-blended gasoline release from service station. Details of the release are unknown. SITE GEOLOGY Approximately 5 feet of medium sand overlying coquina limestone (i.e., coral and shell debris). Primary groundwater flow is in the limestone. The limestone has an extensive network of fractures and dissolution channels (i.e., secondary porosity), which allows for preferential flow and localized storage of gasoline contaminants. SPILL HISTORY Unknown, though it appears that at least two separate releases have occurred. The total spill volume and MTBE percentage are unknown; however, in Florida, MTBE is only used at lower percentages (3 to 8 percent by volume) as an octane enhancer. A release of MTBE-blended gasoline apparently occurred in the UST area during 1994, as evidenced by increased concentrations in nearby monitoring wells (MW-1 and MW-3). The plume is slow-moving, contained on-site, and fairly circular in shape due to aquifer characteristics (e.g., flat gradient). CONCENTRATIONS ENCOUNTERED Before December 1993, MTBE concentrations in the 15 monitoring wells ranged from non-detect to 960 µg/l. In March 1994 (3 months later), MTBE concentrations in the 15 monitoring wells ranged from non-detect to 37,000 µg/l. Remediation System OBJECTIVE Hydraulic containment of the plume and reduction of dissolved-phase concentrations. STANDARD Florida s standard for MTBE is 50 µg/l. REMEDIATION METHOD Groundwater pump-and-treat from three recovery wells. The pumping system operated at 30 gpm for 1 year, then 20 gpm for the next 2.5 years. The extracted water is treated by an air stripper. Pumping started December 1993 (the data presented is from December 1993 to June 1997). WATER TREATMENT METHOD Aeration by an air stripping tower (3-feet diameter, 25-feet tall). Air-to-water ratios ranged from 100-to-1 to 253-to-1. Treated water was discharged to an on-site infiltration gallery. 111

132 Remediation Performance PUMPING SYSTEM PERFORMANCE After 2.3 years of pumping, contaminant concentrations in the treatment system influent reached asymptotic conditions. In the first 2.3 years, benzene dropped steadily from 270 to 30 µg/l (a 9-fold decrease) in the system influent water. In the same period, the influent MTBE concentration dropped from 130 to 10 µg/l (a 13-fold decrease). After this, from March 1996 until June 1997, both MTBE and benzene levels have remained fairly stable, with an overall slight decline. As of June 1997, influent MTBE was 12 µg/l and influent benzene was 20 µg/l. MTBE concentrations in the groundwater declined significantly. In the first 3 years of pumping, MTBE in MW-6 declined from 960 to less than 1 µg/l (a 960-fold decline), while benzene declined from 770 to 1 µg/l (a 770-fold decline). At the same time, the MTBE in MW-7 declined from 37,000 to 77 µg/l (a 480-fold decline) while benzene declined from 14,000 to 310 µg/l (a 45-fold decline). TREATMENT SYSTEM PERFORMANCE Due primarily to fluctuating water extraction rates, the air stripper operated at air-to-water ratios ranging from 100-to-1 to 253-to-1 (average 162-to-1). At this range of air-to-water ratios, the air stripper consistently achieved greater than 99.9-percent MTBE removal over the 3.5 years of operation. The system also removed greater than 99.9 percent of BTEX components. GALLONS OF WATER TREATED Approximately 37,865,000 gallons over 3.5 years GALLONS OF MTBE REMOVED Approximately 1 gallon of MTBE recovered (37,865,000 gallons 24 µg/l [average MTBE concentration December 1993 to June 1997] = 0.9 gallons MTBE). Also, about 10 gallons of BTEX gasoline constituents were recovered over the same time period. Costs COSTS Table A-1 lists costs for the 30-gpm pump-and-treat system. Water treatment performed using air stripping. Table A-1 Costs for the 30-gpm Pump-and-Heat System System Total (3.5 Years of Operation) Capital and Installation $50,000 $50,000 Annual O&M $25,000 $87,500 Total $137,500 for 3.5 Years 112

133 COST PER 1,000 GALLONS OF WATER TREATED $3.63 per 1,000 gallons ($137,500/37,865,000 gallons = $3.63/1,000 gallons treated). Key Findings Groundwater pumping at this site was effective for on-site containment of MTBE and benzene plumes. The maximum MTBE concentration measured in site wells was 37,000 µg/l. The system continued operations after 3.5 years, as of June Pumping flow rates have varied from 20 to 30 gpm. For the first 2.3 years of pumping, MTBE concentrations of the stripping system influent decreased from 130 to 10 µg/l (10-fold decrease). For the same time period, benzene concentrations decreased from 270 to 30 (9-fold decrease). In site monitoring wells, MTBE concentrations decreased more quickly than benzene concentrations. Air stripping resulted in removal efficiency of greater than 99.9 percent for both MTBE and BTEX compounds using air-to-water ratio ranging from 100-to-1 to 253-to-1 (average 162-to-1); however, influent concentrations were low (maximum 130 µg/l). The presence of MTBE caused little cost impact to air stripping system. High air-towater ratios were used, but no off-gas treatment was required. B. Site 2 Contamination History SOURCE MTBE-blended gasoline release(s) from three 10,000-gallon USTs removed from service station in 1988 (relatively early MTBE use in California). Details of the release are unknown. SITE GEOLOGY Clay to 10 feet, underlain by poorly sorted sand with some silt, clay, and gravel to at least 48-feet deep. Depth to groundwater is 35 feet, and groundwater flows in the sand. SPILL HISTORY The initial gasoline contamination was discovered in MTBE contamination was discovered in March 1993 during the first MTBE sampling event. Based on elevated TPH concentrations, MTBE had likely been in subsurface earlier. The spill volume and MTBE percentage in released gasoline is unknown. CONCENTRATIONS ENCOUNTERED MTBE concentrations in groundwater ranged from non-detect to 17,000 µg/l prior to remediation. Free product up to 6-inches thick was repeatedly observed in one well. 113

134 Remediation System OBJECTIVE Remove separate-phase gasoline and MTBE from the subsurface (vadose zone) to minimize concentrations in groundwater. STANDARD California s Interim standard for MTBE in drinking water is 35 µg/l. REMEDIATION METHOD SVE via five extraction wells. Total Q = 430 to 997 cfm; average = 700 cfm. SVE operation started May 1994 and continued until July 1996 (2.2 years), when operation was ceased. VAPOR TREATMENT METHOD Extracted vapors treated by thermal oxidation, with designed destruction efficiency of greater than 95 percent. Remediation Performance SVE SYSTEM PERFORMANCE During 2.2 years of operation, the SVE system removed a significant quantity of gasoline from the vadose zone. In this time, the SVE system operated 82 percent of the time (645 days of a possible 788) and approximately 5,850 gallons of gasoline (as measured by TPHg) were extracted. Extracted vapors showed a distinct and steady decline in TPHg concentrations, which dropped from 2,220 ppmv to 12.1 ppmv (a 181-fold decrease). MTBE has a higher vapor pressure than the BTEX components and, therefore, was likely preferentially removed from the vadose zone by the SVE operations: however, no measurements of the specific composition of the extracted vapors were made to confirm this. Individual parameters like benzene and MTBE were not quantified in the influent or effluent air streams. Over 2.2 years of SVE operation, groundwater conditions at the site improved significantly. One SVE well had a benzene concentration decline from 980 to 0.86 µg/l (a 1,139-fold decrease). In the same well, over the same time period, MTBE concentrations declined from 3,500 to 220 µg/l (a 16-fold decrease). An adjacent monitoring well showed a benzene concentration decline from 670 to less than 0.5 µg/l (a 1,340-fold decrease) and an MTBE decline from 8,900 to 21 µg/l (a 423-fold decrease). All 11 wells monitored had distinct benzene and MTBE dissolved-phase concentrations declines. Thereafter, MTBE and BTEX levels stayed near non-detect and the SVE treatment was stopped 6 months later. SVE operations appear to have reduced BTEX and MTBE concentrations in groundwater. These decreases in groundwater concentrations can be attributed, at least in part, to enhanced in situ biodegradation. The benzene concentrations in groundwater declined more than the MTBE levels in all 11 wells monitored. This is likely to be a reflection of MTBE s greater resistance to volatilization from water (due to a lower Henry s constant) and relatively slow biodegradation. 114

135 VAPOR TREATMENT SYSTEM PERFORMANCE Over 2.2 years of operation, the thermal oxidizer destroyed an average of 95.8 percent of the recovered volatile organic compounds. GALLONS OF WATER TREATED Not applicable. Groundwater treatment occurred in situ. GALLONS OF MTBE REMOVED Unknown as the influent vapor composition was not measured. Costs COSTS Table A-2 lists costs for this moderate-sized SVE system. Table A-2 Costs for the Moderate-Sized SVE System System Total (2.2 Years of Operation) Capital and Installation $100,000 $100,000 Annual O&M $18,000 $40,000 Total $140,000 for 2.2 Years COST PER 1,000 GALLONS OF WATER TREATED - Not applicable as no water was pumped. Key Findings The SVE system at this site removed a large quantity of gasoline from the vadose zone and appears to have significantly decreased MTBE and BTEX concentrations in groundwater. Because of the relatively high vapor pressure of MTBE, it is likely that the SVE system preferentially removed MTBE from the vadose zone; however, extracted vapors were not analyzed for specific compounds. It appears that SVE operations caused significant reductions in dissolved-phase MTBE and BTEX concentrations in groundwater. These reductions are presumed to be a result of volatilization from the aqueous phase and enhanced biodegradation due to SVE operations. Other natural attenuation mechanisms may also be responsible for the reductions. Concentrations of BTEX components decreased more quickly than MTBE; this is likely to be a reflection of MTBE s greater resistance to volatilization 115

136 from water (due to a lower Henry s constant) and relatively slow biodegradation. It is unclear exactly how and why MTBE concentrations in groundwater decreased. The costs for remediating this site were not heavily impacted by the presence of MTBE. The effectiveness and duration of the SVE/vapor treatment operations were not significantly different than if only BTEX treatment was required. MTBE and BTEX reached non-detectable levels in groundwater at approximately the same time. C. Site 3 Contamination History SOURCE MTBE-blended gasoline release from a rusted, single-walled steel UST (275 gallons). SITE GEOLOGY Five to 15 feet of sandy soils over fractured schist bedrock. Groundwater flow is in the fractured bedrock. There are 70 private water supply wells within a one-half mile radius. SPILL HISTORY The release was discovered November 1993 when gasoline odors were noticed in a supply well 125-feet downgradient from the UST. The spill volume and MTBE percentage are unknown. CONCENTRATIONS ENCOUNTERED Initial sampling of the impacted supply well showed 1,295 µg/l of MTBE, thereafter rising to 4,610 µg/l before remediation began. MTBE distribution is irregular due to fracture flow and to many operating supply wells, but the plume shape was generally oval. Of the 28 MTBE-impacted supply wells, only four exceeded 40 µg/l, so the high-concentration MTBE area is small. The one highly impacted well (maximum MTBE at 4,610 µg/l) appears to be at the intersection of major fracture systems, suggesting it is located along the predominant contamination pathway. Remediation System OBJECTIVE Stop plume migration and reduce dissolved-phase concentrations in the impacted wells. STANDARD Vermont s health advisory for MTBE in drinking water is 40 µg/l. REMEDIATION METHOD Groundwater pump-and-treat from highly impacted water supply well (average Q = 3 gpm). The extracted water was initially treated by GAC and then by aeration. Shallow unsaturated soils and bedrock near the source were treated with SVE (maximum Q = 100 cfm). The vapors from the air stripper and SVE system were treated by regenerative thermal incineration. Groundwater pumping started August 1994 and continued until June SVE started November 1994 and continued until June WATER TREATMENT METHOD Pumped water initially treated by three 600-pound GAC vessels in series. MTBE broke through the first two vessels in just a few weeks, so the primary water treatment was switched to aeration. The air stripper used an air-to-water ratio of 1,371-to-1 (550 cfm blower, average water Q = 3 gpm). Treated water was discharged to the local publicly owned treatment works. 116

137 VAPOR TREATMENT METHOD A single thermal oxidation unit was used for treating the air stripper off-gas and SVE off-gas. Little additional information regarding vapor treatment is available. Remediation Performance PUMPING SYSTEM PERFORMANCE During 1.8 years of operation, MTBE levels in the pumped groundwater influent were reduced by a factor of 6.8 (2,720 to 400 µg/l). During the same period, the benzene declined by a factor of 3.1 (3,970 to 1,270 µg/l). As such, it appears that MTBE was preferentially removed from the groundwater, as compared to benzene. TREATMENT SYSTEM PERFORMANCE The 1,371-to-1 air-to-water ratio used here resulted in 99.8 to greater than 99.9 percent removal of benzene over 1.8 years of pumping to date, while MTBE removal was 93 to greater than 99.9 percent. The regenerative thermal incinerator reportedly had a hydrocarbon destruction efficiency of 99.4 percent. GALLONS OF WATER TREATED Approximately 2,200,000 gallons. GALLONS OF MTBE REMOVED Unknown. As of January 1996 (the last date of complete data available), the pump-and-treat system had removed 160 gallons of gasoline and the SVE system had removed 1,060 gallons of gasoline. SVE removed 87 percent of gasoline and the pump-and-treat system removed 13 percent of the total 1,220 gallons known to have been recovered. Presumably, additional gasoline was destroyed in situ by enhanced biodegradation due to oxygen brought to the contaminated region by the SVE system operations. Costs COSTS - Costs for this small-scale combined SVE and 3-gpm pump-and-treat system are listed in Table A-3. Table A-3 Costs for the Small-Scale Combined SVE and 3-gpm Pump-and-Treat System Pump-and-Treat SVE Total Capital and Installation $30,425 $136,100 a $166,525 Annual O&M $15,225 $12, 500 b $46,100 Total $57,825 $154,800 $212,625 for 1.8 Years a Includes the full $100,000 cost of thermal vapor treatment unit. b Includes $10,000 per year for electrical cost. 117

138 COST PER 1,000 GALLONS OF WATER TREATED $26.28 per 1,000 gallons ($57,825/2,200,000 gallons = $26.28/1,000 gallons treated). This does not include a portion of the vapor treatment costs for addressing the air stripper off-gas; including this would raise the per 1,000-gallon cost even higher. The SVE system was more costly on a per-day basis than pump-and-treat because the SVE capital costs (including vapor treatment unit) were higher, SVE operating costs were higher (due to secondary fuel consumption), and the SVE run time was shorter (less operation days to amortize costs). Specifically, SVE cost $389 per day to install and operate ($154,800/398 days) while pump-and-treat cost $116 per day to install and operate ($57,825/495 days). However, because SVE accounted for 90 percent of the gasoline recovered, the SVE was more cost effective than pump-and-treat. Specifically, SVE costs $146 per gallon of gasoline recovered ($154,800/1,060 gallons) while pump-and-treat costs $361 per gallon of gasoline recovered ($57,825/160 gallons). The SVE system proved technically and cost effective at this site (this is in spite of the fact that these calculations unfairly attribute all vapor treatment costs to the SVE). Key Findings D. Site 4 The pumping system at this site performed well for MTBE removal from groundwater. The dissolved-phase MTBE concentrations decreased twice as quickly as those of benzene. This difference is attributed partly to less retardation of MTBE in the aquifer and partly to the operations of the SVE system. The SVE system performed well for removing separate-phase gasoline from the vadose zone. Although no data are available, it is likely that MTBE was preferentially removed from the separate-phase product due to its relatively higher vapor pressure. The air stripper operated at an air-to-water ratio of 1,371-to-1, which allowed for 93 to greater than 99.9 percent removal of MTBE. The thermal incineration system used for to treat off-gas from the SVE system and from the air stripper was adequate for the destruction of MTBE. This site demonstrates that a combination of technologies can effectively remediate an MTBE-impacted site. In addition, a quick response to the gasoline releases at this site allowed for use of the more cost-effective SVE, which effectively limited the spread of MTBE into the groundwater system. Contamination History SOURCE MTBE-blended gasoline release from a UST at a service station. Details of the release are unknown. 118

139 SITE GEOLOGY Ten to 18 feet of medium-coarse sand and gravel (with some cobbles and boulders) overlying bedrock. Groundwater flow is in the sand and gravel unit. SPILL HISTORY Details unknown. Gasoline contamination was discovered in 1989; MTBE contamination was discovered in 1990 when first analyzed. An additional small MTBE-blended gasoline release apparently occurred in mid The spill volume and MTBE percentage in released gasoline are unknown. CONCENTRATIONS ENCOUNTERED At the start of remediation in June 1990, MTBE concentrations in groundwater ranged from non-detect to 3,800 µg/l in the nine monitoring wells sampled. No time-series MTBE data exist for conditions prior to remediation. Remediation System OBJECTIVE Stop the plume migration and reduce dissolved-phase concentrations. STANDARD Massachusetts cleanup levels for MTBE are site-specific; none was established for this site. The water effluent discharge standard was 100 µg/l. REMEDIATION METHOD Groundwater pump-and-treat from one recovery well (average Q = 2 gpm). The extracted water was treated by GAC and discharged under a National Pollutant Discharge Elimination System permit. Groundwater pumping started June 1990 and continued until August Shallow unsaturated soils near the source was treated with SVE from three wells (maximum Q = 100 cfm). A full-scale SVE system started January 1993 and continued until August 1996 (3.8 years), when operations ceased. WATER TREATMENT METHOD Extracted water was treated by two 300-pound GAC vessels in series. After 11 months of liquid-phase GAC treatment, water treatment was changed to air stripping. After 2 months, air stripping operations ceased due to difficulties in cost-effectively removing MTBE from the air stripper off-gas. The water treatment was then switched back to liquid-phase GAC. VAPOR TREATMENT METHOD Vapor-phase adsorption by GAC was planned for use, but GAC breakthrough occurred on the first and third days of operation. Vapor treatment was then switched to electric-fired catalytic oxidation for the first 15 months (January 1993 to March 1994) of SVE operations. During this period, the catox unit destroyed 96 to 100 percent of the influent hydrocarbon vapors. By March 1994, influent vapors dropped so low (8 ppmv of hydrocarbons by a photoionization detector) that secondary fuel costs (i.e., electricity) soared. Because of this, the SVE operations were temporarily halted between March 1994 and May Vapor treatment switched to vapor-phase GAC after May With influent concentrations of non-detect to 21 ppmv, GAC needed replacement three times in 1 year. The SVE system was no longer needed after August 1996 because the extracted vapors were non-detect for hydrocarbons. 119

140 Remediation Performance PUMPING SYSTEM PERFORMANCE During the first 3 years of operation (June 1990 to July 1993), benzene levels in the extracted groundwater (i.e., the treatment system influent) declined by a factor of 17 (11,000 to 636 µg/l). During the same 3-year period, MTBE levels declined by a factor of 127 (34,440 to 272 µg/l). As such, it appears that MTBE was preferentially removed from the groundwater, as compared to benzene. In the same time frame, a monitoring well just downgradient from the recovery well showed a benzene decline by a factor of 32 (2,500 to 78 µg/l), while MTBE declined by a factor of 35 (7,100 to 198 µg/l). From July 1993 until early 1995, MTBE and BTEX concentrations stayed low and continued their asymptotic decline. Then another gasoline spill apparently occurred. Benzene increased slightly over the next few months (from a steady range of 15 to 38 µg/l before the spill to a high of 362 µg/l after the spill), but MTBE increased dramatically (from a steadily declining 34 µg/l before the spill to a high of 17,300 µg/l after the spill). MTBE levels in a monitoring well downgradient from the tank pit rose from 79 to 59,200 µg/l after the spill. Because the pump-and-treat system and the SVE system were immediately remediating this new spill, the MTBE levels quickly declined, reaching 803 µg/l by August 1996 (3 years later). Again, after this second spill incident, it appears that MTBE was preferentially removed from the groundwater, as compared to benzene. In the 6.2 years of operation, the pump-and-treat system removed 11 gallons (66 pounds) of hydrocarbons. SVE SYSTEM PERFORMANCE The SVE system started in January 1992, but operations ceased almost immediately as vapor-phase carbon breakthrough occurred on the first day and again on the third day. Once all the mobile separate-phase product had been pumped/bailed from the subsurface (300 gallons total), SVE operation began again in January After January 1993, the SVE system operated for 3.6 years (1,309 days) at 100 cfm and removed 155 gallons (approximately 932 pounds) of hydrocarbons. SVE influent concentrations declined from an initial high of 138 to 0.0 ppmv (as measured with a photoionization detector) over those 3.5 years. Since vapor composition analyses were not conducted throughout the project, specific component recovery (e.g., MTBE) can not be determined. SVE operations ceased in August 1996 as no more hydrocarbons had been recovered by the system in several months. WATER TREATMENT SYSTEM PERFORMANCE Even with the low water flow of 1 to 2 gpm, the 300-pound GAC vessels had to be replaced often, primarily due to MTBE breakthrough. Over 6.2 years of operation, benzene only broke through the first vessel five times and through the second vessel twice. MTBE, however, broke through the first vessel 22 times and the second vessel 11 times. In summary, MTBE s presence required nearly five times as much liquid-phase GAC to be used and this necessitated more frequent GAC change-outs. The drastically increased frequency of change-outs also mandated more frequent analysis of the influent, midfluent 120

141 (between the two GAC vessels), and effluent, so that MTBE breakthrough could be caught early and discharge violations avoided. VAPOR TREATMENT SYSTEM PERFORMANCE The catox unit destroyed 96 to 100 percent of the influent hydrocarbon vapors. By March 1994, influent vapors dropped so low (8 ppmv of hydrocarbons by a photoionization detector) that secondary fuel requirements rose dramatically. The cost for electricity (the secondary fuel) was only $40 per month in April 1993, when soil gas influent concentrations were 135 ppmv of volatile organic compounds. After 8 months, when the influent concentrations declined significantly (down to 14 ppmv of volatile organic compounds), the secondary fuel electricity cost had risen to $500 per month. The SVE system was not operational March 1994 to May 1994 while the vapor treatment method was switched to vapor-phase GAC. In the first year of operation, with flow rates of 100 cfm and influent concentrations averaging 5 ppmv of volatile organic compounds (maximum = 21 ppmv), the 300-pound vapor-phase GAC vessels needed replacement three times. Later, the influent went non-detect for several months; hence, the SVE operation was ceased (in August 1996). GALLONS OF WATER TREATED 3,500,000 gallons in 6.2 years. GALLONS OF MTBE REMOVED Unknown. Through the operation of both the SVE and pump-and-treat, 165 gallons of hydrocarbons (including MTBE) were recovered in 6.2 years. Presumably, more gasoline was destroyed in situ by enhanced biodegradation due to SVE operations. Costs COSTS - Costs for this small-scale combined SVE and pump-and-treat system are listed in Table A-4. Table A-4 Costs for the Small-Scale Combined SVE and Pump-and-Treat System Pump-and-Treat Total Capital and Installation $430,000 $430,000 Annual O&M $33, 000 $194,000 Total $624,000 for 6.2 Years COST PER 1,000 GALLONS OF WATER TREATED - This cost can not be determined as pump-and-treat costs can not be isolated from SVE costs, given the available data. 121

142 Key Findings Data from this site indicate that pumping caused the preferential removal of dissolved-phase MTBE in groundwater in comparison to benzene. The performance of liquid-phase carbon treatment at this site was negatively impacted by the presence of MTBE in the influent water. MTBE broke through the carbon system five times more frequently than benzene, dramatically increasing carbon change-out requirements. In addition, the presence of MTBE caused more frequent samplings of influent, midfluent, and effluent. SVE appears to have been effective for removing volatile hydrocarbons at this site; however, no conclusions can be made regarding MTBE removal efficiency because composition analyses of extracted vapors were not performed. Vapor treatment was effective using catalytic oxidation (i.e., 96- to 100-percent destruction efficiency) though secondary fuel costs were high, particularly as influent concentrations declined with time. The performance of vapor-phase carbon was very poor due to the presence of MTBE, which broke through the treatment vessels almost immediately (i.e., on the first and third day of operations). When influent SVE vapors were moderately high, vapor treatment by GAC was ineffective. MTBE broke through on the first day and again 2 days later. When SVE vapors had dropped to low influent levels (maximum = 21 ppmv), the small GAC vessels needed replacement on a reasonable basis (three times in 1 year), thus showing that vapor-phase treatment by GAC can be effective with low influent concentrations. E. Site 5 Contamination History SOURCE Petroleum hydrocarbon release(s), including MTBE-blended gasoline spilled at a bulk terminal and truck load-out facility. SITE GEOLOGY Fine to coarse sand, with minor amounts of silt and clay. The depth to groundwater ranges from 7- to 15-feet below grade. There are numerous private water wells (used primarily for irrigation) immediately downgradient of the site; the transport of contaminants may have been accelerated because some of these wells operated after the release. SPILL HISTORY The release of petroleum hydrocarbons was discovered in 1992 during a soil gas survey of the site. The presence of MTBE was first reported in 1993 when groundwater samples were collected and analyzed for various parameters, including MTBE. Blended gasoline were stored at and moved through the terminal from 1988 to No other details regarding documented petroleum releases were available. 122

143 CONCENTRATIONS ENCOUNTERED During June/July 1993, 44 private water wells and 24 monitoring wells were sampled. MTBE concentrations ranged from non-detect to 4,000 µg/l. Later sampling events showed that dissolved-phase MTBE concentrations in the 24 monitoring wells ranged from non-detect to 140,000 µg/l. An extensive free-product plume (longer than 900 feet) is also present at the site. Remediation System OBJECTIVES There are four remediation systems being used at the site, two SVE systems, a product cut-off barrier, and an air sparging system. The systems each have separate objectives, as described below: SVE systems Remove free-phase product and residual hydrocarbons from subsurface soils in two different areas. Product cut-off barrier Block the migration of free product and remove product from the subsurface. Air sparging system Remove dissolved-phase contaminants from groundwater. STANDARD The state s standard for MTBE is not defined at this time. Remediation Methods SVE Systems Two systems have been installed to address different areas of contamination. System A is located near the primary source area and System B is located downgradient of the site. System A consists of five extraction wells, transfer piping, in-line water knockouts, SVE extraction skid with regenerative blower, and a vapor control unit (flare). The operation of System A began in August System B consists of 18 extraction wells, two valve boxes, SVE extraction skid with blower, and a vapor control unit (catalytic oxidizer). Four of the wells originally used for System B were taken offline in March 1996 for use in the air sparging system. The operation of System B began in November In both of the systems, flow rate, vacuum, temperature, and vapor concentrations (BTEX and total volatile hydrocarbons) were measured for each of the operating SVE wells. Specific information regarding system flow rates and well completion details was not available. Product Cutoff Barrier A subsurface product barrier and skimmer system were installed in July 1994 to prevent the migration of free-phase floating product on the shallow water table. The system consists of impermeable barrier panels, product collection wells installed in the gravel backfill upgradient of the barrier, pneumatically-driven skimmer pumps, an air compressor and supply lines, and piezometers installed in the backfill. The impermeable barrier panels partially penetrate the saturated thickness of the aquifer, allowing the flow of groundwater 123

144 below the panels. Floating product is removed from the water-table surface on the upgradient side of the barrier using skimmer pumps. Free-product skimming continued from startup in 1994 until October 1995 when there was no longer any measurable free product in any of the skimmer wells or the piezometers along the barrier. Six months later, free product was also not detected in any of the skimmer wells or piezometers since system shutdown in October Air Sparging System The air sparging system consists of three horizontal in situ sparging wells, four horizontal engineered trench sparger wells, riser pipes and well headers, well clean-out and gate valve manways, a compressor, and an aftercooler. The three horizontal wells have been in operation since February These wells, which each span approximately 150 feet, are located 18 to 21 feet below ground surface, at the base of the aquifer. Air injection into each well was typically between 100 and 140 standard cfm. The four trench wells began operating in March The trench wells are installed at the base of the aquifer, with depths ranging from 20 to 24 feet below ground surface. The horizontal well lengths varied from 160 to 210 feet. The well trenches were backfilled with gravel to increase air distribution. Passive vent pipes and piezometers were installed above the water table in the air sparging trench backfill. Process data including air flow rates, temperature, and pressure have been collected routinely since startup of the sparging system. Groundwater is monitored to assess concentrations of dissolved BTEX and total volatile hydrocarbons in the vicinity of the sparging system. WATER TREATMENT METHOD None required. VAPOR TREATMENT METHOD A flare unit was used for vapor control. No further details are available. Remediation Performance AIR SPARGING SYSTEM PERFORMANCE Regarding the 2 years of air sparging, although MTBE concentrations have decreased significantly in some wells since the start of air sparging, there is no consistent trend. For example, the MTBE concentration in MW-35, which is about 15-feet downgradient of the older sparging system, decreased to 86 percent from a high of 73,000 µg/l from approximately August 1994 to 10,000 µg/l in March 1996; however, the MTBE concentration in MW-34, which is about 30-feet upgradient of the older sparging system (and MW-35), decreased similarly (92 percent) from a high of 63,000 µg/l from approximately July 1994 to 5000 µg/l in March These data indicate that the distinct MTBE concentration decrease in downgradient MW-35 cannot be readily correlated to effects of the air sparging systems. In addition, no distinct trend can be noted for MTBE concentrations in wells immediately downgradient of the new sparging system. Concentrations in three of these wells varied from a decrease of 74 percent to an increase of 18 percent based on the last 6 months of monitoring data. In summary, data from the site 124

145 monitoring wells does not clearly indicate that air sparging is effective for MTBE removal from groundwater. To help establish the effectiveness of the air sparging system, a performance test was conducted in conjunction with expanding the sparging system in April/May During the 24-hour test, MTBE concentrations in groundwater and vapor were measured to determine the impact of the sparging system. The air flow rate into the sparging wells varied between 55 to 116 standard cfm, equivalent to an air-to-water ratio of 100-to-300. In five of six wells/piezometers, MTBE concentrations in groundwater decreased over the 24-hour test period. Table A-5 lists the approximate initial concentrations and the concentrations measured at the end of the 24-hour test period. Table A-5 Approximate Initial Concentrations and Concentrations Measured at the End of a 24-Hour Air Sparging System Performance Test Initial MTBE Final MTBE Percent Piezometer (µg/l) (µg/l) Removal P-9 8,100 4, P-10 3,800 5,200 37* P-11 3,900 2, P-12 1, P As indicated above*, one piezometer (P-10) experienced a concentration increase of approximately 37 percent. The project scientists concluded that this well became clogged during testing, thereby invalidating the results. Based on results from the other piezometers, MTBE concentrations in the sparging trench decreased from 31 to 82 percent over the 24-hour test. These percentage decreases are similar to those found for total volatile hydrocarbons during the same test. Benzene concentrations at the start of the test were quite low, causing percent reduction comparisons to be misleading. In addition to the concentration decreases measured in the groundwater, MTBE stripping was evidenced during the sparging performance test by detectable MTBE concentrations in the vapor extracted from trench gravels during the sparge testing; however, no background vapor measurements made prior to the start of the test were included in the report. These data show that in situ stripping of MTBE did occur due to the air sparging system; however, the limited amount of data over such a short period of time preclude strong conclusions being made regarding the long-term effectiveness of air sparging for MTBE removal. 125

146 SVE SYSTEM PERFORMANCE MTBE concentrations in the wells near the source area SVE system (System A) have been fairly constant since the initiation of the SVE system. During the period from June/July 1993 to March 1996, concentrations in these wells have varied from an increase of 84 percent in MW-15 to a decrease of 63 percent in MW-22. Based on these data, it appears that SVE is not effective for removing dissolved-phase MTBE from water at this site. Effectiveness of the SVE system is expected to be poor for dissolved-phase MTBE due to the low Henry s Law constant of MTBE and the fact that the MTBE at this site appears to primarily be in the dissolved phase. PERFORMANCE OF PRODUCT CUTOFF BARRIER By April 1996, over 4,800 gallons of free product were recovered from behind the cutoff barrier; however, based on data collected from the site, it did not appear that there were significant concentrations of MTBE in the free product that existed onsite. Three upgradient wells with free product were sampled in August 1995 to test for the presence of MTBE. Although MTBE was not detected in the samples from any of these three wells, the detection limits were very high (i.e., 1,000,000 to 2,000,000 µg/l). Free-phase product analyses were also carried out on samples from five wells in December None of these samples detected oxygenated additives, including MTBE (detection limit = 10,000 µg/l). Because it appears that the free-product plume does not contain MTBE, the performance of the cutoff barrier is not directly relevant to MTBE remediation. TREATMENT SYSTEM PERFORMANCE No effluent treatment required. GALLONS OF WATER TREATED Not applicable as the treatment occurred in situ. Costs Cost data for installation and O&M of these remediation systems are not available. Key Findings Some 3 to 7 years after the MTBE-blended gasoline spill occurred, the dissolvedphase MTBE is far downgradient of the remediation systems and has been transported far beyond the downgradient extent of the BTEX compounds. Operation of the SVE systems at the site does not appear to have impacted dissolvedphase MTBE (i.e., SVE was applied too late to be effective for MTBE). Performance testing of the air sparging system indicated that MTBE concentrations in the sparge trench were reduced by 31 to 82 percent over the 24-hour test, which was reportedly run at an air-to-water ratio of 100 to 300. Monitoring well data from wells near the sparging system suggest that sparging reduced MTBE concentrations, although the correlations are not fully clear. 126

147 The downgradient extent of this large MTBE plume appears to have stabilized and even retracted somewhat from 1993 to This may be due to: a) natural dilution due to advective/dispersive spreading of the plume; and/or, b) MTBE mass recharge reduction in the core of the plume due to operating SVE and air sparging systems. 127

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