Column test-based optimization of the permeable reactive barrier (PRB) technique for remediating groundwater contaminated by landfill leachates.

Journal of Contaminant Hydrology 168 (2014) 1–16

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Column test-based optimization of the permeable reactive barrier (PRB) technique for remediating groundwater contaminated by landfill leachates Dan Zhou a,b, Yan Li a,b,⁎, Yinbo Zhang c, Chang Zhang d, Xiongfei Li e, Zhiliang Chen c, Junyi Huang a,b, Xia Li f, Giancarlo Flores g, Masashi Kamon h a Guangdong Provincial Key Laboratory of Marine Resources and Coastal Engineering, School of Marine Sciences, Sun Yat-sen, University, 135 Xin'gang RD.W., Guangzhou 510275, PR China b Key Laboratory for Aquatic Product Safety of Ministry of Education, School of Marine Sciences, Sun Yat-sen, University, 135 Xin'gang RD.W., Guangzhou 510275, PR China c South China Institute of Environmental Science, Ministry of Environmental Protection, No. 7 West Street, Yuancun, Guangzhou 510655, PR China d Shandong Bonaray Analysis Instrument Technology Co., Ltd, Building A5, High and New Technology Industrial Development Zone, Jining 272000, PR China e Guangdong Provincial Environmental Technology Center, 28 Modiesha Avenue, Xingang Dong Road, Guangzhou 510308, PR China f Nanhai Environmental Technology Center of Foshan City, Environmental Protection Building, 4 New RD. 3S., Guicheng, Foshan 528200, PR China g Graduate School of Engineering, Kyoto University, Yoshida-Honmachi, Kyoto 606-8501, Japan h National College of Technology, 355 Chokushicho, Takamatsu-shi, Kagawa 761-8058, Japan

a r t i c l e

i n f o

Article history: Received 19 May 2014 Received in revised form 27 August 2014 Accepted 3 September 2014 Available online 16 September 2014 Keywords: Permeable reactive barrier (PRB) Groundwater Contamination Remediation Component ratio Thickness Longevity

a b s t r a c t We investigated the optimum composition of permeable reactive barrier (PRB) materials for remediating groundwater heavily contaminated by landfill leachate, in column tests using various mixtures of zero-valent iron (ZVI), zeolite (Zeo) and activated carbon (AC) with 0.01–0.25, 3.0–5.0 and 0.7–1.0 mm grain sizes, respectively. The main contributors to the removal of organic/ inorganic contaminants were ZVI and AC, and the optimum weight ratio of the three PRB materials for removing the contaminants and maintaining adequate hydraulic conductivity was found to be 5:1:4. Average reductions in chemical oxygen demand (COD) and contents of total nitrogen (TN), ammonium, Ni, Pb and 16 polycyclic aromatic hydrocarbons (PAHs) from test samples using this mixture were 55.8%, 70.8%, 89.2%, 70.7%, 92.7% and 94.2%, respectively. We also developed a systematic method for estimating the minimum required thickness and longevity of the PRB materials. A ≥309.6 cm layer with the optimum composition is needed for satisfactory longevity, defined here as meeting the Grade III criteria (the Chinese National Bureau of Standards: GB/T14848/93) for in situ treatment of the sampled groundwater for ≥10 years. © 2014 Published by Elsevier B.V.

1. Introduction For landfills already in operation, replacement of failed bottom layers with new impervious lining systems is a difficult task. The PRB method utilizes a passive permeable treatment wall filled with reactive materials, which is installed in the path of a contaminated groundwater plume. When passing through the wall, contaminants in the groundwater are ⁎ Corresponding author. Tel.: +86 20 39332201; fax: +86 20 39332159. E-mail address: [email protected] (Y. Li).

http://dx.doi.org/10.1016/j.jconhyd.2014.09.003 0169-7722/© 2014 Published by Elsevier B.V.

removed by degradation, precipitation and sorption processes due to physical, chemical, biochemical or integrated interactions between the contaminants and reactive materials (Thiruvenkatachari et al., 2008). The reactive materials packed inside the PRB wall are selected depending on the main contaminants present in the contaminated groundwater. Commonly used reactive materials include zero-valent iron (ZVI), activated carbon (AC) (Rijnaarts et al., 1997), zeolite (Zeo) (Kovalick and Kingscott, 1995) and peat (Kao and Yang, 2000). The most widely utilized of the three, in both laboratory studies and field applications, is ZVI (Gavaskar et al., 1998;

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D. Zhou et al. / Journal of Contaminant Hydrology 168 (2014) 1–16

Köber et al., 2002): more than 60% of the PRBs installed worldwide were reportedly iron-based by 2004 (ITRC, 2005). ZVI has a high reduction potential, −440 mV (Simon and Meggyes, 2000), and thus acts primarily as a reductant in most systems, transferring electrons to the contaminants, e.g., halogenated hydrocarbons and some inorganic ions (Thiruvenkatachari et al., 2008). Thus, it can activate various useful reaction mechanisms for contaminant removal (e.g., redox reactions, precipitation, and sorption) (Calabrò et al., 2012; Xenidis et al., 2002). ZVI has been proven to be a highly efficient material for removing heavy metals (Mn, Ni, Pb, Cu, Zn, etc.) (Blowes et al., 2000), petroleum hydrocarbons (Guerin et al., 2002), nitrates (Gandhi et al., 2002), and halogenated hydrocarbons. However, use of ZVI alone as the reactive medium has limitations with regard to the long-term hydraulic properties and removal efficiency due to the deactivation, corrosion, and clogging of the barrier pores (Li et al., 2006; Liang et al., 2005; Ruhl and Jekel, 2012). In particular, anaerobic iron corrosion increases the pH inside iron PRBs and promotes precipitation of secondary minerals (Carniato et al., 2012), which can have detrimental effects on the longevity of the PRB. However, use of granular mixtures of ZVI with other reactive materials in various weight or volume ratios may help to solve these problems, and reduce the amount of relatively expensive ZVI required. Activated carbons (ACs) are carbonaceous materials that have chemically heterogeneous surfaces. Their composition and structure depend on the starting material and production methods. However, their surfaces typically have high densities of phenolic and carboxylic groups (inter alia) and they have high capacities for removing a wide range of contaminants, both organic — e.g., phenols, BTEX, PCE and TCE (Bone, 2012) and COD — and inorganic (Dong et al., 2009; Thiruvenkatachari et al., 2008). Heavy metals may also be removed by AC (DiNardo et al., 2010; Köber et al., 2002; Nakagawa et al., 2003; Panturu et al., 2009; Scherer et al., 2000; USEPA, 1998). Thus, ACs, mostly in granular form (GAC), were commonly used materials in the early stages of PRB technology. Sorption by AC is strongly influenced by the solution pH. High pH causes ionization of the carboxylic and hydroxylic groups on AC surfaces (Peng et al., 2003), which increases interactions of water molecules with the surfaces and thus decreases adsorption of particularly hydrophobic contaminants. Liu and Pinto (1997) also reported a decrease in phenol adsorption when the pH was reduced from 6.3 to 3.07. In addition, the efficacy of AC may be affected by groundwater constituents, e.g. Cornelissen et al. (2005) found that natural organic matter may compete with contaminants for the binding sites and consequently reduce their sorption rates. Thus, as noted by Obiri-Nyarko et al. (2014), careful characterization of the aquifer and management would be required if AC were to be considered as a barrier material. Like ZVI, it clearly has limitations as a single material, but it could potentially be a valuable complement. Zeolites contain water molecules, alkali and alkaline earth metals, and are therefore also potentially useful as PRB reactive substrates owing to their high capacities for ion-exchange (200–400 meq/100 g), adsorption (specific surface areas up to 145 m2/g), catalyzing beneficial reactions, and molecular sieving (ITRC, 2011; Roehl et al., 2005). Their surfaces also have diverse pore structures, allowing selective adsorption of contaminants. The high ion-exchange capacities of zeolites are

attributed to their permanent negative charges, which develop from isomorphic substitution. These charges are not pH dependent and are usually balanced by alkali and alkaline earth metals such as Na+, Ca2+, K+ and Mg2+ (Peric' et al., 2004). Natural zeolites generally have relatively large particle sizes, making them suitable for use as reactive media. However, their low organic carbon content limits their sorption of organic compounds. Examples of natural zeolites include analcime, chabzite, clinoptilolite, heulandite, natrolite, philipsite, mordenite and stilbite (Coombs et al., 1997). Clinoptilolite has been extensively used for removing (mainly) cationic contaminants such as Pb, Cu and Cd (Peric' et al., 2004). However, zeolite/or surfactant-modified zeolite has capacities to remove a wide range of contaminants including heavy − 3− metals, NH+ 4 , NO3 , PO4 , radionuclides, PCE and BTEX, with efficiencies ranging between 80% and 100% (Bowman et al., 1995; Katz et al., 2006; Li et al., 1999; Park et al., 2002; Peric' et al., 2004; Ranck et al., 2005; Vidic and Pohland, 1996). Numerous investigations have shown that leachates from many landfills in both developing and developed countries often contain high concentrations of various contaminants including: dissolved organic matter (e.g., CH4, volatile fatty acids, humic, and fulvic compounds); inorganic macrocomponents (e.g., Ca2+, 2− − − Mg2+, NH+ 4 , Cl , SO4 , HCO3 ); heavy metals (e.g., Cd, Cr, Cu, Ni, Pb, Zn, Cu); and xenobiotic organics including halogenated hydrocarbons, aromatic hydrocarbons, phenols, chlorinated aliphatics, etc. (Christensen et al., 1994, 2001; Dong et al., 2009; Onay and Pohland, 1998; Renou et al., 2008; San and Onay, 2001; Zhou et al., 2006). Due to a lack of waste classification and the habits of citizens in South China, municipal solid waste in landfills of the region typically comprises more than 40% kitchen waste, 25% paper, plastic, metal and glass waste, 14% garden and textile waste, and 20% construction waste (Lin et al., 2010). Hence, landfill leachates in this region generally have more complex composition than leachates in developed countries and the most abundant contaminants in groundwater that they contaminate are ammonium, heavy metals, and organic compounds (Ni et al., 2004). PRB technology for in situ remediation of groundwater treatment has been widely studied or applied since it was invented. According to Bone (2012), 624 studies on PRBs were published between 1999 and 2009, of which approximately 40% was laboratory-based investigations and 32% was field studies. Contaminants that have been treated with PRBs to date include halogenated aliphatic hydrocarbons, metals, metalloids, radionuclides, pesticides, petroleum hydrocarbons, and nutrients emanating from agricultural systems (Blowes et al., 1998; Bone, 2012; Conca et al., 2002; Köber et al., 2002; Luo, 2007; USEPA, 2002). However, few studies to date have investigated the application of PRBs for remediating groundwater containing as diverse and high concentrations of contaminants as leachates from landfills in South China. Contamination profiles and numerous other relevant factors are site-specific (Henry et al., 2008; Ott, 2000; USEPA, 1998). Thus, the design of a PRB generally involves a series of steps that include: a preliminary technical and economic assessment; characterization of the aquifer geology, geochemistry and contaminant levels at the site; selection of appropriate reactive materials (types and ratios); treatability studies (batch and column tests); engineering design (including the location, orientation, dimensions and longevity of the


PRB); choice of construction method; formulation of a monitoring plan; and detailed financial analysis (Gavaskar et al., 2000; Obiri-Nyarko et al., 2014). Therefore, the first requirement for a robust PRB design is a good understanding of the site and aquifer characteristics, including the site geology, aquifer hydrogeology (Erto et al., 2011; Klammler and Hatfield, 2008; McMahon et al., 1999; Powell et al., 1998), geochemistry (Puls, 2006) and microbial activity. Detailed knowledge of the contaminated plume is also clearly essential (Erto et al., 2011; Powell et al., 1998), including the spatio-temporal distribution of the contaminants, directions and rates of groundwater flow, and preferential flow paths to avoid bypassing or overflowing of the contaminants (Federal Remediation Technologies Roundtable [FRTR], 2002; Henry et al., 2008; ITRC, 2011). The choice of reactive materials is generally influenced by: the type of contaminants to be removed (organic and/or inorganic), their concentrations, and the mechanisms needed for their removal (e.g. biodegradation, sorption or precipitation) (McGovern et al., 2002); the hydrogeological and biogeochemical conditions of the aquifer; the environmental/health impacts; the materials' mechanical stability (capacity to maintain hydraulic conductivity and reactivity over time), and their availability and cost. Single materials were frequently applied in early stages of PRB technology. However, combinations of materials, which may be biotic, abiotic or both, are frequently applied nowadays because they have several advantages. Inter alia, they can improve permeability, reduce costs, increase the number of mechanisms available for single or multi-contaminant removal, enhance and accelerate removal rates, and (thus) substantially improve the long-term performance of barriers. Numerous studies have confirmed the potential benefits of using mixtures of materials, including the following. Ma and Wu (2008) used two abiotic materials, zero-valent zinc (Zn0) and ZVI, for degrading trichloroethylene (TCE) and found that the mixture degraded it three times faster than ZVI alone. Saberi (2012) found that mixtures of Ni/Fe nanoparticles improved Pb2+ removal. Moraci and Calabrò (2010) noted that a mixture of iron and pumice provided effective Cu and Ni removal with long-term hydraulic conductivity. Mixtures of two biotic materials, compost and mulch, were also reportedly effective for degrading chlorinated solvents (Henry et al., 2003; Öztürk et al., 2012), and better than either material individually (AFCEE, 2008; Ahmad et al., 2007). Mulch is more stable than compost, due to its high lignin content, and thus serves as a long-term source of organic C, while compost readily decomposes and releases nutrients that can be easily accessed by microbes that participate in the anaerobic degradation of the chlorinated solvents. The potential utility of a combination of biotic and abiotic materials such as ZVI and organic substrates for the degradation of chlorinated solvents and immobilization of inorganic contaminants has also been reported (Mueller et al., 2004). In addition, a batch study by Pratt et al. (1997) showed that altering the composition of reactive mixtures, e.g., adding calcite, quartz sand, or pyrite, in reactors may alter both the precipitates formed and the morphology of ZVI surfaces. Furthermore, combinations of reactive materials can also improve the permeability of barriers (ITRC, 2005). Clearly, appropriate combinations can improve the efficiency of treating various contaminants as they provide different removal mechanisms and can enhance permeability. Thus, by

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using a suitable combination of ZVI, AC, zeolite, peat and/or other materials it should be possible to remediate many types of contaminated groundwater. However, in some cases combinations of materials may have no effects, or even antagonistic (inhibitory) rather than synergistic (stimulatory) effects on PRBs' contaminant removal and overall performance. For example, Köber et al. (2002) noted that a combination of GAC and Fe0 did not improve the removal of TCE and MCB due to the reduction in adsorption capacity of the GAC by the Fe0. Scherer et al. (2000) observed that coupling metals can have a catalytic or synergistic effect, but may also result in the formation of oxide layers that may impair the barrier's performance. The effect of combining materials on the overall performance of the PRBs depends on numerous factors, including the ratio of materials in the mixture. Therefore, the reactivity and long-term interactive effects of the materials should be considered when selecting material combinations for PRBs, in addition to all the factors mentioned above (Obiri-Nyarko et al., 2014). Furthermore, it is important to ensure that the hydraulic permeability of the PRB, including screens and reactive media, is at least twice that of the aquifer, to avoid problems due to gradual reductions in permeability via precipitation of iron oxides/hydroxides, carbonates and or other metal precipitates in the treatment media or filter layers (USEPA, 1998). Thus, the type and ratio of the reactive materials in the mixture are major determinants of the likelihood of achieving the planned contaminant removal efficiencies and hydraulic conductivities. In groundwater contaminated by landfill leachate, inorganic 2− − − − − species (e.g., Ca2+, Mg2+, NH+ 4 , Cl , SO4 , NO3 , NO2 , HCO3 ) and heavy metals (e.g., Cd, Cr, Cu, Ni, Pb, Zn, Cu) can be removed by sorption, ion-exchange, reduction and/or precipitation processes (Thiruvenkatachari et al., 2008), which ZVI and zeolite can offer. Some xenobiotic organic compounds can also be removed by sorption and reductive–oxidative degradation processes (Thiruvenkatachari et al., 2008) that ZVI, AV and zeolite can offer. However, persistent, chemically stable organic pollutants (which may be highly toxic) must be largely removed by sorption, retardation and biodegradation (Yang et al., 2010), rather than ion-exchange and reductive–oxidative degradation. AC has high sorption capacity, thus mixtures of ZVI, AC and zeolite may provide high, cost-effective and synergistic capacities for removing both organic and inorganic contaminants in groundwater contaminated by landfill leachate. In addition to all the factors mentioned above, the rates of contaminant removal are strongly influenced by the selected materials' grain size. Small particles have high specific surface areas, which promote interaction with contaminants. Thus, ZVI has been used mostly in the form of chips, jet blasting media, iron foams and pellets, particulates and powders, or as Fe-filler material for concrete (Obiri-Nyarko et al., 2014). However, if they are too small ZVI particles can be easily pulverized after oxidation, which decreases the hydraulic conductivity and thus the longevity of reactive barriers. Thus, 0.074 to 2.4 mm ZVI grains have been frequently used as PRB materials (Ruhl and Jekel, 2012), and most often 0.25 to 2 mm grains with specific surface areas of 0.5 to 1.5 m2/g (ITRC, 2011; Mackenzie et al., 1999; Powell et al., 1998; Tratnyek et al., 2003). The adsorption capacities of both AC and zeolite also decrease with reductions in their specific surface areas (and thus increases in their

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particle sizes), so they are generally used as reactive PRB materials in granular form. However the molecular sieving capacity of zeolite depends more on the sizes of the vestibules in the molecular lattices than the surface area (reference). Furthermore, the cost is a major factor to consider when selecting reactive materials, thus commercial availability and convenience are also important. Considering costs, specific surface areas and hydraulic conductivity, we postulated that a mixture of fine ZVI (0.01–0.25 grains, mostly close to 0.25 mm), coarse zeolite (3.0–5.0 mm, to optimize porosity) and intermediate AC (0.7–1.0 mm grains) may provide high, cost-effective removal efficiencies and satisfactory hydraulic conductivity. Target removal efficiencies are generally determined by local or national groundwater quality standards. The criterion applied here is that concentrations of focal contaminants after treatment should meet the Grade III standard of groundwater quality issued by the Chinese National Bureau of Standards (GB/T14848/93). Both laboratory batch and column studies have been frequently used to obtain information about various geochemical and microbiological phenomena in PRB treatment systems. Batch tests require little time, but column experiments provide dynamic flow conditions that closely approximate those expected in a PRB system in field deployments (Gavaskar et al., 2000; Henderson and Demond, 2007; ObiriNyarko et al., 2014). Thus, tests with columns composed of mixtures of ZVI, zeolite and AC in various ratios should provide the best indications of the optimal ratios for cost-effectively removing contaminants while maintaining satisfactory hydraulic conductivity. After reactive materials have been selected, the dimensions, location and orientation of the barrier must be defined. Two important, mutually dependent parameters are the capture zone and residence time (Obiri-Nyarko et al., 2014). The capture zone refers to the width of the barrier required to intercept the entire contamination plume. The residence time is defined as either the contact time between the contaminated groundwater and reactive material required to achieve treatment goals (FRTR, 2002; Gillham et al., 2010; Ott, 2000; Puls, 2006), or the time that contaminated groundwater takes to pass through the reactive materials in the PRB (Calabrò et al., 2012; Li and Benson, 2010). The design of a PRB must ensure that the residence time, defined in the latter manner, is sufficient to treat the contaminant(s) of interest. With given types of contaminants and concentrations in the groundwater, the residence time is mainly determined by the groundwater velocity and the thickness of the reactive materials in the PRB. Another key parameter is the longevity of the barrier, defined as the time that a PRB continues to treat contaminants at designed levels (Henderson and Demond, 2007; ITRC, 2011; Robertson et al., 2000). Remediating contaminated groundwater using PRB is an advanced technology that is generally 50% cheaper than the Pump & Treat technique, as no routine maintenance and surface buildings are required (Schad and Grathwohl, 1998). It is also being exploited in numerous field applications, particularly in the USA and Canada (Calabrò et al., 2012), which have provided extensive operational information. In principle, when designing a PRB just sufficient amounts of reactive materials should be used to reduce contaminant concentrations to target values. However, when dissolved contaminants come into contact with the PRB, numerous

reactions occur that form precipitates and gradually reduce the barrier's removal efficiency, porosity, permeability and (hence) longevity (Furukawa et al., 2002; Mackenzie et al., 1999; Moon et al., 2008; Phillips et al., 2000). A complication is that little is known about the long-term performance of the reactive materials used in PRBs, even ZVI, the most commonly used material used to date (Geranio, 2007; Obiri-Nyarko et al., 2014). Numerous authors have reported that its performance declines with time, due to the accumulation of secondary mineral precipitates and production of gas, mainly CO2 and N2 (Kamolpornwijit et al., 2003; Morrison, 2003; Vikesland et al., 2003). These processes gradually impair barrier hydraulics due to pore filling (Eykholt et al., 1999; Liang et al., 2005; Mackenzie et al., 1999), and reduce the material's reactivity by decreasing its reactive surfaces (Jeen et al., 2008; Kamolpornwijit et al., 2003; Li et al., 2006; Liang et al., 2000; O'Hannesin and Gillham, 1998; Vogan et al., 1999; Wilkin and Puls, 2003). Fouling can also reduce PRBs' porosity and hence hydraulic conductivity (Henderson and Demond, 2007). Furthermore, variations in flow velocities due to geological heterogeneity and groundwater geochemistry can exacerbate the effects of fouling (Blowes et al., 2000; Li et al., 2005). Generally, high flow rates and high concentrations of mineral-forming ions enhance mineral precipitation. These reductions in reactivity, porosity, and hydraulic conductivity of the ZVI cause reorientation of flow paths, changes in flow rates, seepage velocity, and residence times, and deterioration of treatment efficiency (Jin et al., 2009; Kamolpornwijit et al., 2003; Vikesland et al., 2003; Zolla et al., 2009). Laboratory column tests and field studies of PRBs containing ZVI have shown that the pH rises quickly near the entrance face and subsequently levels off in the range of 9–10 (Wilkin et al., 2003, 2009; Furukawa et al., 2002). Li et al. (2005, 2006) reviewed the types and quantities of secondary minerals formed in PRBs, and found that the most common minerals are iron oxides (magnetite, ferrous sulfide, hematite), iron (oxy)-hydroxides (ferrous hydroxide, ferric hydroxide, green rust, goethite, lepidocrocite), carbonates (calcite, aragonite, siderite) and marcasite. These materials are reportedly responsible for annual reductions in porosity and hydraulic conductivity ranging from 0.0007 to 0.03, and from 14.2% to 66.7%, respectively. However, calcium carbonates and siderite are typically found near PRBs' entrance faces, whereas magnetite, ferrous hydroxide, green rust and iron oxyhydroxides form throughout a PRB (Furukawa et al., 2002; Mackenzie et al., 1999; Phillips et al., 2003; Sarr, 2001; Wilkin et al., 2005). In addition, porosity reductions are typically greatest near the entrance face, for instance Wilkin et al. (2003) found that the porosity of an iron medium in a PRB decreased by 0.032 within 25 mm of the entrance face, and b0.00002 at 80 mm, after 8 years operation. Thus, numerous processes have complex, interacting effects on the longevity of a PRB including the heterogeneous formation of mineral precipitates and gases, with accompanying spatio-temporal variations in reductions in reactivity within it. The optimal thickness of reactive materials in a PRB is a trade-off between maximizing effectiveness and minimizing construction costs. Of course, reductions and variations in the thickness reduce the barrier's longevity (ITRC, 2011), but for an optimal balance between cost, removal efficiency and longevity a barrier should have optimal types and ratios of reactive


materials and minimum dimensions (Erto et al., 2011; Henry et al., 2008; Kacimov et al., 2011). The minimum thickness in this context is defined as the thinnest layer of reactive materials that can reduce concentrations of contaminants at the outlet to target levels. This parameter can be roughly calculated from estimates of flow velocities, residence times and removal efficiencies per unit length of the reactive materials obtained from column test data (Obiri-Nyarko et al., 2014). However, the minimum thickness may not be the most cost-effective due to various unpredictable changes that may affect the barrier's longevity, despite the availability of several detailed design methodologies (Ahmad et al., 2007; Fallico et al., 2010; Gavaskar et al., 2000; ITRC, 2005, 2011; Muegge, 2008; ObiriNyarko et al., 2014; USEPA, 1998). After life-cycle comparisons of PRBs and pump-and-treat systems for groundwater remediation, Higgins and Olson (2009) concluded that ZVI-type PRBs must last at least 10 years to be environmentally superior to pump-and-treat systems in all impact categories. At a given site, when the optimum ratio and minimum thickness of reactive materials have been obtained from laboratory batch or column studies (Muegge, 2008), the longevity is considered to be approximately proportional to the thickness of the PRB reactive materials. The optimum thickness of reactive materials can thus be determined from the viable longevity (N10 years) and estimates of the minimum thickness. In the study presented here we performed laboratory column tests to investigate the optimal composition of permeable reactive materials to remove contaminants (from groundwater heavily polluted with leachate from a landfill in South China) with high efficiency and acceptable hydraulic conductivity. We also developed a systematic method to estimate the minimum thickness of PRB materials required to meet the current quality standards when treating such contaminated groundwater for 10 years.

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size range of 0.7–1.0 mm and consisted of coconut carbon, ash (≤5%) and iodine (950–1000 mg/g). Plexiglas columns (length 90.0 cm, inner diameter 10.0 cm) were used to determine the optimal thickness and composition of PRB materials to meet the current local groundwater standards (Fig. 1). Since fine soil particles may prevent samples from flowing into a column, a compacted 5 cm layer of quartz (0.2 to 1.0 mm grains) was placed at the bottom of each column to filter out solid particles in the leachate and allow it to flow equably into the column. An 80 cm layer of the uniformly mixed reactive materials was then placed on top of the quartz layer. Mixtures with eight weight ratios of ZVI, zeolite and AC were used in the tests (designated mixtures A–H; Table 1). In each case the mixture was packed into the column and consolidated by gently tapping the outside of the column with a rubber hammer. Hollow tubes with an inner diameter of 0.5 cm were inserted into the center of each column at 5.00 cm intervals to enable sampling of the pore water passing through the reactive materials. Small holes were bored into the tubes to allow the pore water to enter from all flow directions when sampling. Using natural contaminated groundwater as the inflow, the optimal design parameters of the PRB, such as the thickness and composition of the reactive materials, were determined directly based on the minimum length from the bottom of the column to the sampling port at which the concentrations of the contaminants in the pore water satisfied the groundwater quality standard. Before each experiment, all columns were flushed with 2.00 L ultrapure water, then fully saturated with de-aerated

2. Materials and methods 2.1. Experimental setup and sampling Leachate flowing in a small ditch from a closed landfill in Guangdong, China, was sampled during both dry and wet seasons, then stored in 1 L brown glass bottles at +4 °C until analyses or tests. The leachate was comprehensively analyzed to identify all contaminants whose concentrations exceeded the Chinese National Standard (GB/T14848/93). Only contaminants exceeding the standard were considered further in developing the PRB technique. For reasons detailed in the Introduction, commerciallyavailable ZVI, AC and zeolite with suitable, complementary grain sizes were selected as the permeable reactive materials to test. The ZVI reactant (Fengyao Chemical Industry , Liyang, China) contained Fe0 (N 98%) with impurities of manganese (0.35%) and traces of silicon, carbon, sulfur and phosphorus. Its grain size ranged from 0.01 to 0.25 mm, which is smaller than the range of 0.075–2.4 mm reported by Eljamal et al. (2011), resulting in a larger superficial area between the leachate and iron particles. The zeolite reactant (Haisheng Water Purification Plant Material, Gong Yi, China) contained aluminosilicate (N95%) and had a grain size range of 3.0–5.0 mm. The AC (Guangzhou Jietan Trade Co., Ltd, Guangzhou, China) component had a grain

Fig. 1. Apparatus used to determine the optimal thickness and composition of reactive materials in the PRB.

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Table 1 Proportions of reactive materials and main characteristics of the PRB. Mixture

ZVI:Zeo:ACa

Weight (g)

Volume (cm3)

Porosity (%)

RT (hours)b

A B C D E F G H

10:10:80 10:45:45 10:80:10 30:10:60 30:35:35 30:60:10 50:10:40 50:40:10

4500 5853 6400 5500 7335 9094 7774 11,794

6473 6490 6500 6380 6473 6306 6444 6361

45.64 43.02 39.70 42.60 39.44 41.80 43.20 41.30

98.47 93.07 86.00 90.60 85.10 87.87 92.70 87.60

a

ZVI refers to zero-valent iron, Zeo to zeolite and AC to activated carbon. RT refers to the residence time (Calabrò et al., 2012), i.e., the time taken for the leachate to pass through the reactive materials from the inlet to the outlet. b

ultrapure water. The experiment was then commenced by pumping undiluted leachate sampled during the dry season (the most heavily polluted samples) through each column, simultaneously, from a 50 L plastic barrel via eight fine polyethylene tubes at a constant flow rate of 0.5 mL/min using peristaltic pumps (Longer BT100). The room temperature was kept at 20 °C during the column tests. The leachate was supplied to the plastic barrel in turn, 10 L every 48 h, and 138.24 L leachate was consumed totally within 576 h. As mentioned above, the leachate was stored at + 4 °C all the time except the 48 h passing from the barrel through the eight columns. Therefore, the changes in the initial concentration of the leachate were minor within 576 h. Tubes placed at 0 (inlet), 20.0, 50.0 and 80.0 cm (outlet) from the bottom of the mixture layer were selected as the sampling ports (Fig. 1). During the experiment, syringes were used to collect samples from all four ports for inorganic analysis (30 mL samples) with 50-mL high density polyethylene bottles, and COD analysis (50 mL samples) with 100-mL serum bottles. In addition, 50 mL samples were collected in 100 mL serum bottles by syringe from the inlet and outlet ports to analyze 16 PAHs prioritized by the US EPA (reference). Samples were collected on days 0, 3, 7, 12, 18 and 24 (of pumping the leachate), then stored at +4 °C until analysis.

surrogates (naphthanene-d8, acenaphthene-d10, phenanthrene-d10, benzo[a] anthracene-d12, perylene-d12) were added for recovery calculation. The 16 dissolved PAHs in the samples were extracted by solid-phase extraction using a HCC18 SPE cartridge (6.0 mL, 500.0 mg, CNW, Germany). Prior to fractionation, the cartridges were conditioned by washing with 10.0 mL of methanol followed by 10.0 mL of deionized water. All samples were passed through the cartridges under vacuum at a flow rate of 6.00 mL/min. After drying, the cartridges were eluted consecutively with 15.0 mL ethyl acetate followed by 15 mL dichloromethane. The effluent was collected in 50 mL tubes, dried by anhydrous sodium sulfate, then concentrated to 1.00 mL under a nitrogen stream and transferred to 1.50 mL vials for GC–MS analysis. An Agilent 7890A GC system with a 5975C mass selective detector was used for the quantitative analysis of PAHs. The GC was equipped with an Agilent DB-5 MS capillary column (30.0 m × 0.25 mm × 0.25 μm), using helium as carrier gas at a constant flow rate of 1.0 mL/min. The injector temperature was held at 280 °C. The GC oven temperature was maintained at 70 °C for 1 min, then increased to 300 °C at a rate of 6 °C/min, and finally held at 300 °C for a further 10 min. 2.3. Optimization of the composition, thickness and longevity of the PRB materials A certain mixture of PRB reaction materials was considered to be effective if concentrations of all the contaminants in the treated effluent were lower than the limits issued by the government for groundwater remediation projects. The total amount of a given contaminant (Mrmv, mg) removed during the experiment in time T (s) (as shown in Fig. 2) was calculated using the integration tools within ORIGIN 8.0: N X C i Δt i Mrmv ¼ Q C 0 T−

! ð1Þ

i¼1

2.2. Analytical methods COD was measured using the potassium dichromate method (Eaton and Franson, 2005), and pH was measured with a pH meter (Sartorius PB-10). Nitrite, nitrate and ammonium were determined by ion chromatography (IC) using a Dionex 600 IC system equipped with a CS12A cation column and AS14 anion column (Dionex Company, USA). Eleven metal elements (Ba, Be, Cd, Cr, Cu, Mn, Mo, Ni, Pb, Zn, and Co) and two non-metal elements (As, Se) were determined by Inductively Coupled Plasma Emission Spectrometer-Atom Emission Spectrometry (ICP-AES), using a Thermo Jarrell-Ash IRIS/h instrument. To analyze the 16 priority PAHs, a blank control of ultrapure water was prepared and m-terphenyl was added to the blank and all samples as an internal standard. PAH standard mixture solutions were also prepared by diluting a stock standard mixture (10.0 mg/L) to obtain six concentrations (0.050, 0.100, 0.500, 1.000, 5.000 and 10.000 μg/L), and five deuterated

Fig. 2. Sketch showing expected change with time in concentration of a contaminant in the effluent during PRB treatment: C0 refers to the initial concentration in the leachate before treatment; CL, the concentration in the effluent after passing through the initial reactive materials; C, the concentration in the effluent. The gray area enclosed by the C0 line, C line and Y axis indicates the mass removed by the PRB materials.


where Q is the groundwater flux (cm3/s), C0 is the initial concentration of the contaminant in the groundwater (mg/L), N is the number of subintervals during the interval T, Ci is the concentration at the ith subinterval (mg/L), and ti is the time length of the ith subinterval (s). Using data from the column tests, the total removal efficiency of each PRB mixture for a given contaminant (MR, mg/cm), and the removal efficiency per unit reaction time (MRT, mg/cm3/s) were calculated as: M R ¼ Mrmv =L0 S

ð2Þ

M RT ¼ MR =T

ð3Þ

where, L0 is the length from the inlet to the sampling port (cm), S is the cross-sectional area of the column (78.5 cm2 in our tests), and T is the time taken for the concentration of the contaminant to rise from the lowest concentration that the treatment provided to a stable level (e.g., the initial concentration). The minimum thickness of the PRB materials (LMIN, cm), required to ensure that contaminant concentrations in the effluent complied with the current standards was estimated as: LMIN ≥vðC–C BL Þ=MRT

ð4Þ

where v is the velocity of groundwater passing through the PRB materials (cm/s), C is the concentration of the focal contaminant in the contaminated groundwater (mg/L) at the site where the PRB project was planned to be constructed, assumed to be 0.02 times the concentration in the dry season sample (Cui et al., 2011), and CBL is the smaller of CB and CL, where CB is the official groundwater quality limit (mg/L), and CL is the lowest concentration remaining in the groundwater after treatment with the PRB materials (mg/L), as defined in Fig. 2. The hydraulic conductivity (K) of the natural uncontaminated red soils in the studied landfill site is reportedly 0.00072 to 3.216 m/d (Zhong et al., 2009), and the hydraulic gradient (I) is about 0.05 in hilly areas. Thus, the actual groundwater velocity calculated from Darcy's law is 3.60 × 10−5 m/d to 0.16 m/d, and was assumed to be 0.10 m/d. The longevity of the PRB materials depends mainly on their thickness, the removal efficiency and contaminant flux in the groundwater. To estimate the longevity (TL, years) of the reactive medium for a given contaminant in an actual subsurface situation, the following equation was used: T L ¼ LMR =vðC–C BL Þ

ð5Þ

where, L is the thickness of PRB reactive materials, which should be larger than LMIN for a given contaminant. 2.4. Hydraulic conductivity of the PRB materials To evaluate the permeable capacity of each mixture, the constant head permeability method (Head and Keeton, 2008) was applied to determine the change in hydraulic conductivity during the experiment. The hydraulic conductivity (K) was calculated assuming Darcy's law as follows: K ¼ QL=ΔhS

ð6Þ

7

where K is the hydraulic conductivity (cm/s), L is the length (cm) over which the head pressure drop occurs; Δh is the difference in head pressure (cm), and S is the cross-sectional area of the column (cm2). 3. Results and discussion 3.1. Main contaminants in the leachate The comprehensive analysis of contaminants showed that − COD and concentrations of NH+ 4 , NO2 , Mn, Ni, Se, and Be exceeded Grade III Chinese national limits (GB/T14848/93) in both the dry and wet season samples, whereas levels of Mo, Pb, Zn, Cd, Cu, and Co only exceeded the limits during the dry season (Table 2). Organics in landfill leachate are known to vary with the type and age of the landfill. However, in landfills receiving waste from cities in South China the main organic contaminants are volatile fatty acids, humic and fulvic compounds, and halogenated hydrocarbons (Zhou et al., 2006). To save time in the following column tests we focused the organic analyses on the contaminants that greatly exceeded the official limits, are highly toxic or representative of significant classes of substances. More specifically, we chose: COD as a representative measure of organic contaminants; + − − NH+ 4 and total nitrogen (TN, including NH4 , NO2 and NO3 ) as representatives of inorganic contaminants; Ni (which exceeded the limit 14.6-fold, more than any other metal) and Pb (which exceeded the limit 5.8-fold, more than any other toxic metal, i.e., Hg, Cr, Pb, Cd, and As) as representative metallic contaminants; and the 16 EPA priority PAHs known to be highly carcinogenic to humans. We examined changes in concentrations of all of these contaminants during treatment with each of the PRB mixtures. 3.2. Optimization of the PRB mixtures 3.2.1. Removal efficiency of organic contaminants 3.2.1.1. COD. COD concentrations sampled at the 20 cm, 50 cm and outlet ports (i.e., after the leachate had passed through 20 cm, 50 cm and 80 cm of the reactive materials, respectively) are illustrated in Fig. 3. Generally, COD concentrations were lowest at the 80 cm sampling port throughout the treatment, and remained relatively constant up to around 72 h, as the removal efficiency remained close to maximal. They then be gradually increased before approaching a steady state at slightly less than the initial concentration at around 288 h, suggesting that removal efficiencies along all three path lengths gradually decreased and approached zero (Fig. 3a), or a low constant value (Fig. 3b and c) at this time. Furthermore, at times less than 72 h the measured COD concentrations were highest at the 20 cm sampling port (Fig. 3a), and lowest at the 80 cm sampling port (Fig. 3c). The mixtures that resulted in the lowest concentrations in the effluents at the 20 cm sampling port were G and H, indicating that increasing the proportion of ZVI increased COD oxidation rates. At the 50 cm sampling port, the lowest concentrations were obtained with mixtures A and B at T b 72 h, suggesting that the reduction of COD concentration resulted mainly from adsorption by Zeolite and AC, since the oxidation–reduction of COD did not occur completely in such a short time. At the 80 cm

8


Fig. 3. Time courses of COD concentrations in the effluents from PRB mixtures of indicated compositions and thicknesses: initial concentration hereafter refers to the concentration in ground water sampled during the dry season.

sampling port, effluent concentrations were similar in tests with all of the eight mixtures during the first 72 h, suggesting that their removal efficiencies were very similar during early stages, provided that the path length was sufficiently long. At the 20 cm port (Fig. 3a), COD concentrations gradually decreased in tests with mixtures A–F until the 168th hour, in contrast to the patterns observed at the 50 and 80 cm ports (Fig. 3b and c), although they were almost as high as the initial concentration at the inlet during the first 72 h. However, the time courses of the COD concentration at 20 cm with mixtures G and H were similar to those observed at the 50 cm and 80 cm sampling ports. These results indicate that the penetration time provided by the 20 cm path length was too short for AC and zeolite to absorb the contaminants in the leachate, but sufficient for the oxidation–reduction reaction between ZVI and COD, hence reactions with the ZVI component dominated at the start of COD removal. The ZVI contents in mixtures A–F were too low for effective removal of the COD. However, as the operational time increased, the ion-exchange, adsorbing, and catalytic capacities of zeolite (ITRC, 2011), and the adsorption capacity of AC (Bone, 2012) began to dominate after 72 h. From T = 288 h until the end of the column test COD remained at a high and relatively constant level, suggesting that the zeolite component had little effect on COD removal due to the extremely high concentrations of dissolved contaminants in the influent, which rapidly reduced the cationexchange capacity and effective surface area. At the 50 cm and 80 cm sampling ports (Fig. 3b and c), the lowest concentrations were obtained with mixtures G and H at T N 288 h, suggesting that a high proportion of ZVI could reduce the time needed for COD removal. Ion exchange reactions between zeolite and metal ions also presumably reduced the adsorption capability of the molecular sieve, resulting in COD desorption from the PRB materials in late stages. Only the organic compounds carrying oxidizing groups were removed by ZVI, hence the removal capacity remained stable after the AC component had become saturated. As shown in Table 3, in tests with mixtures A–F there were minor differences in COD concentration at 576 h both between mixtures and between samples collected from the 20, 50 and 80 sampling ports. These findings further indicate that removal efficiencies of mixtures A–F decreased to similar very low levels due to the consumption of the reactive materials and decline of reactive surfaces of the material by precipitation of minerals (Jeen et al., 2008; Li et al., 2006) provided that the reaction time

was sufficiently long. In contrast, mixtures G and H retained a little more removal capacity due to their higher proportions of ZVI. Overall, the mixtures with the highest percentage of AC (A and B) resulted in the lowest initial COD concentrations in the effluent (Fig. 3b and c) during the first 72 h. However, mixtures G and H provided the lowest concentrations after longer operation times, and mixture G (with 50% ZVI, 10% zeolite and 40% AC) provided the best overall COD removal. 3.2.1.2. 16 priority PAHs. In tests with all eight mixtures, the concentrations of the 16 priority PAHs defined by the US EPA (ATSDR, 2005) measured at the 80 cm sampling port decreased from the initial concentration of 0.10 mg/L to less than 0.01 mg/L (Fig. 4). Therefore, all mixtures showed high performance in PAH removal. As illustrated in Table 3, compared with the initial concentration, the differences in concentrations of the PAHs sampled from the eight tested PRB mixtures after 576 h were minor. These findings indicate that their removal was

Table 2 Chemical characteristics of the landfill leachate. Components

COD pH NH+ 4 NO− 2 − NO3 As Ba Be Cd ΣCr Cu Mn Mo Ni Pb Se Zn Co 16 PAHs

Concentration (mg/L) Quality standarda

Dry season

Wet season

– 6.50 ~ 8.52 0.20 0.02 20.00 0.05 1.00 0.0002 0.01 – 1.00 0.10 0.10 0.05 0.05 0.01 1.00 0.05 –

5436.08 7.90 1488.70 b10.00 b10.00 0.04 0.30 b0.006 0.035 1.55 2.72 1.32 0.14 0.73 0.29 0.10 2.80 0.08 0.10

1607.1 8.61 701.10 b10.00 18.70 0.025 0.27 b0.006 0.0025 0.44 0.094 0.43 0.042 0.25 0.037 0.10 0.46 0.02 b

– No values issued. a Refers to Grade III of the quality standard of groundwater issued by the Chinese National Bureau of Standards (GB/T14848/93). b Not detected.


9

Table 3 Initial concentrations at the inlet and concentrations at the three outlet ports at the end of the column tests with each type of PRB material. COD (mg/L) I A B C D E F G H

(1)

5436.08

20

50

(3)

80

(4)

4483.01 4681.15 4392.19 4359.17 4268.35 4433.47 4367.42 4483.01

4474.75 4466.50 4268.35 4672.90 4144.51 4111.49 4235.33 4301.38

4466.50 4639.87 4284.86 4045.44 4499.52 4367.42 3789.50 3673.92

I

20

50

80

0.73

0.40 0.46 0.43 0.42 0.56 0.40 0.36 0.34

0.39 0.45 0.41 0.35 0.37 0.33 0.32 0.35

0.41 0.50 0.37 0.39 0.37 0.45 0.35 0.28

Ni (mg/L)

A B C D E F G H

NH+ 4 (mg/L)

TN (mg/L) (2)

K (m/d)

I

20

50

80

I

20

50

80

I

80

1773.30

1518.72 1435.29 1542.25 1505.88 1505.88 1505.88 1343.32 1544.39

1512.30 1508.02 1614.97 1465.24 1424.60 1512.30 1283.42 1499.47

1518.72 1488.70 1576.47 1422.46 1409.63 1435.29 1105.88 1219.25

1488.70

1187.50 1087.50 1157.50 1027.50 906.50 1027.50 958.50 876.50

1057.50 1107.50 1127.50 1047.50 988.50 1027.50 892.50 981.50

986.50 929.50 878.50 806.50 852.50 1037.50 614.50 842.50

36.69 55.03 55.03 27.52 18.34 55.03 15.72 13.76

18.34 27.52 9.17 36.69 27.52 10.01 9.17 5.50

I

20

50

80

I

20

50

80

I

80

0.29

0.05 0.05 0.10 0.04 0.04 0.05 0.05 0.04

0.05 0.09 0.13 0.05 0.05 0.08 0.03 0.06

0.05 0.11 0.11 0.07 0.07 0.08 0.04 0.05

7.90

8.47 8.80 8.92 8.76 9.11 9.09 9.16 9.27

8.14 8.27 8.79 8.87 8.89 8.99 9.34 9.44

8.15 8.29 8.55 8.90 9.07 9.24 9.17 9.45

0.10

0.0053 0.0035 0.0069 0.0045 0.0052 0.0040 0.0032 0.0048

Pb (mg/L)

pH

PAHs (mg/L)

(1) refers to the initial concentrations at the inlet of the column, (2), (3) and (4), the concentrations at the 20 cm, 50 cm, and 80 cm sampling ports at the end of the column tests with each type of PRB material, respectively.

mainly due to the adsorption and/or molecular sieving capacities of the zeolite and AC, while reactions between PAHs and ZVI made minor contributions due to the high chemical stability of PAHs. 3.2.2. Removal efficiency of inorganic contaminants Nitrogen is another contaminant found at high concentrations in landfill leachates, in the form of inorganic nitrogen (nitrate, nitrite and ammonium), organic nitrogen (proteins, amino acids, nucleic acids, chitin, murein etc.) and diverse degradation products (Reynolds and Richards, 1996). The initial concentrations of total nitrogen and ammonium in the leachate were 1770 mg/L and 1490 mg/L, respectively. Changes in the concentrations of total nitrogen and ammonium in the effluent

with testing time are shown in Figs. 5 and 6, respectively. At all three sampling ports used, the concentrations of total nitrogen and ammonium in the effluent were initially constant up to T = 72 h, and then increased gradually with time, before reaching a steady state close to the initial concentration at T N 288 h. This suggests that removal efficiencies of all the mixtures remained close to maximal during the first 72 h, then gradually declined as the reactive materials were consumed and eventually approached zero (Figs. 5a and 6a) or a low constant value (Figs. 5b and 6b, 5c and 6c) after 288 h. Because of the shorter path length, concentrations of total nitrogen and ammonium in the effluent at the 20 cm port were initially considerably higher than those at the 50 cm port and the outlet. The removal of leachate contaminants is likely to involve oxidation–reduction reactions between ZVI and the contaminants, and sorption–desorption of the contaminants to zeolite and AC. The reductant ZVI can transfer electrons to nitrate, generating nitrogen, ammonium and nitrite, via the following reactions (Dong et al., 2002, 2003): −

2þ

2NO3 þ 5Fe þ 6H2 O→5Fe − NO3 − NO3 þ NH4

Fig. 4. Time courses of the total concentration of the 16 priority PAHs measured at the 80 cm sampling port in effluents from PRB mixtures of indicated compositions and thicknesses.

2þ

þ 4Fe þ 7H2 O→4Fe 2þ

þ Fe þ H2 O→Fe −

þ

−

þ N2 þ 12OH

þ

þ NH4

− NO2

þ OH ⇌NH3 ↑ þ H2 O:

−

ð7Þ

þ 10OH

ð8Þ

−

ð9Þ

þ 2OH

ð10Þ

As shown in Eqs. (7)–(10), increases in OH− concentration and NH+ 4 generation shift the equilibrium of the reversible reaction (Eq. (10)) to the right. Thus, NH3 is generated and escapes into air. Furthermore, a cell may be formed when ZVI particles contact AC particles, creating a negative ZVI pole and positive carbon pole, promoting R-NO2 + Fe0 + H+ → R-NH2 + Fe2 ++ H2O reactions followed by ionization of the R-NH2 to NH+ 4 . All these reactions reduce the TN

10


Fig. 5. Time courses of total N concentration in effluents from PRB mixtures of indicated compositions and thicknesses.

concentration, but ammonium was still present at relatively high concentrations. Under anoxic conditions, both anaerobic denitrification and anaerobic ammonium oxidation (anammox) processes promote removal of nitrogenous contaminants from the leachate. Denitrifying bacteria can utilize H2 derived from ZVI donating electrons to water (Dong et al., 2009) to convert nitrate to nitrogen gas, and the overall anammox process can be described by the following equation (Strous et al., 1998): þ

−

−

þ

NH4 þ 1:32NO2 þ 0:066HCO3 þ 0:13H →1:02N2 − þ 0:26NO3 þ 0:066CH2 O0:5 N0:15 þ 2:03H2 O:

ð11Þ

Furthermore, zeolite and AC have high sorption capacities, i.e., high ion-exchange and adsorption capacities. Thus, the reaction chain represented by Eqs. (7)–(9) may be propagated by these physiochemical processes. For instance, nitrate may first be deoxidized by ZVI to form intermediate products, e.g., ammonium, nitrite and nitrogen gas, which then undergo adsorption and ion-exchange reactions with the zeolite and AC components, which in turn drive the oxidation–reduction reactions with ZVI. Accordingly, concentrations of nitrogenous contaminants in the solution decreased, presumably coupled with continuous corrosion of ZVI and gradual reductions in the adsorption capacities of zeolite and AC. However, the ammonium concentration in the effluent may temporarily increase + above its initial level due to the conversion of NO− 3 to NH4 , as shown in Eq. (8) and Fig. 6a.

At the 20 cm port, TN concentrations in tests with mixtures A–F and H were almost as high as the initial concentration at the inlet during the first 72 h (Fig. 5a). The time course of the TN concentration at this port was only similar to those observed at the 50 cm and 80 cm sampling ports in tests with mixture G. Thus, the TN removal efficiency was only high when the reactive mixture contained both high ZVI and high AC contents, indicating that the reactions of these constituents with the nitrogenous contaminants dominated during the first 72 h. Within this period all the mixtures efficiently removed NH+ 4 (Fig. 6a), indicating that all three PRB components (ZVI, zeolite, and AC) efficiently reacted with contaminating NH+ 4 through the reaction chain from Eqs. (7)–(10). Furthermore, as shown in Figs. 5 and 6, mixture G generally provided the lowest TN concentrations throughout the experiment and at all path lengths, suggesting that ZVI and nitrate rapidly reacted. The mixtures with high ZVI contents also generated more intermediate products. The main reaction was the one between ZVI and NO− 3 as shown in Eq. (8), and the main intermediate product was deduced to be NH+ 4 , which could be efficiently adsorbed by AC. As shown in Table 3, in tests with mixtures A–F there were minor differences in TN concentration at 576 h both between mixtures and between samples collected from the 20, 50 and 80 sampling ports (as previously observed for COD concentrations). These findings further indicate that removal efficiencies of mixtures A–F decreased to similar very low levels due to the consumption of the reactive materials and decline of reactive surfaces of the material by precipitation of minerals (Jeen et al.,

Fig. 6. Time courses of ammonium concentration in effluents from PRB mixtures of indicated compositions and thicknesses (Grade III standard of ammonium (GB/T14848/93) ≤ 0.2 mg/L).


2008; Li et al., 2006) provided that the reaction time was sufficiently long. In contrast, mixtures G and H retained a little more removal capacity due to their higher proportions of ZVI. at the 576th hour However, the concentration of NH+ 4 decreased with the path length through which the leachate passed, for every mixture except F and with increases in the proportion of ZVI among the mixtures at each sampling port (Table 3). These findings indicate that ion exchange/sorption of intermediate products (e.g., iron ions) by zeolite promoted the + production of NH+ 4 , thereby reducing the NH4 removal efficiency, according to Eq. (8). Of the eight mixtures studied, mixture H contained the same percentage of ZVI as G, but a higher zeolite and lower AC content, while C had the highest zeolite content. Both H and C exhibited low performances for the removal of nitrogenous contaminants, suggesting that the zeolite component had a much lower removal capacity for the intermediate products generated in the reaction chain than AC. Mixtures containing low proportions of ZVI also showed low removal capacities for nitrogenous contaminants, although they contained more zeolite and AC than mixture G. Overall, the main contributors to the removal efficiency of nitrogenous contaminants were ZVI and AC, and mixture G exhibited the highest efficiency for the removal of TN and ammonium. 3.2.3. Heavy metal removal efficiency Changes in total Ni and Pb concentrations at the three sampling ports as a function of testing time are shown in Figs. 7 and 8, respectively. Like the trends observed for COD and nitrogenous contaminants, concentrations of both these contaminants gradually increased after the first 72 h. At the end of the experiment, the concentrations of Ni and Pb became stable at 0.34 ~ 0.56 mg/L and 0.038 ~ 0.098 mg/L, respectively, at the 20 cm port, 0.33 ~ 0.45 mg/L and 0.032 ~ 0.13, respectively, at the 50 cm port, and 0.28 ~ 0.51 mg/L and 0.037 ~ 0.075 mg/L, respectively, at the 80 cm port, corresponding to maximum removal efficiencies of 63.1% and 86.2% for Ni and Pb, respectively. It is clear in Figs. 7 and 8 that mixture G removed Ni and Pb most efficiently (since the concentrations were lowest in effluents from this mixture, and corroborative changes were observed throughout the whole treatment process), most likely due to the high percentages of ZVI and granular AC. However, differences in Ni and Pb concentrations at 576 h were small

11

both between mixtures and between samples collected from the 20, 50 and 80 sampling ports (Table 3), indicating that removal efficiencies of all the reactive materials decreased to low, stable levels due to consumption of the reactive materials and the decline of reactive surfaces of the material caused by secondary mineral precipitates (Jeen et al., 2008; Li et al., 2006), provided that the reaction time was sufficiently long. 3.3. pH A high ZVI content should promote the removal of organic and inorganic contaminants from the leachate. However, the use of high amounts of ZVI also inevitably causes some environmental problems, the main one being increased pH of the effluent. The pH of effluents from mixture G approached 9.5 (Fig. 9), consistent with possible ZVI corrosion. The oxidation of ZVI consumes H+ and acidic functional groups in the organics, e.g., \OH and \COOH, resulting in a reduction in H+ concentration and corresponding increase in OH− concentration in the effluent, thereby increasing the pH (Thiruvenkatachari et al., 2008). The differences in pH at 576 h were small among samples from the 20 cm, 50 cm and 80 cm sampling ports (Table 3), but the pH increased with increases in the proportion of ZVI, indicating that the pH remained almost constant after sufficient reaction time. 3.4. Hydraulic conductivity Changes in hydraulic conductivity (K) as a function of testing time in columns with each of the eight mixtures are shown in Fig. 10. Clearly, the hydraulic conductivity decreased with increasing PRB treatment time. Such reductions may be due to precipitation of minerals (e.g., CaCO3, FeCO3, and Fe(OH)2) (Li et al., 2006) or formation of gas bubbles (e.g., CO2 and N2) (Kamolpornwijit et al., 2003) driven by reactions between the leachate and PRB materials, which are likely to have unpredictable effects on the contaminant removal performance and PRB longevity (Wantanaphong et al., 2006). The value of K ranged from 13.75 m/d to 55.03 m/d at the start of the experiment and 5.50 m/d to 36.69 m/d after 550 h. In our tests, the grain size of ZVI was relatively small compared with those of zeolite and AC. Although the measured porosities of the eight columns (Table 1) were similar, columns

Fig. 7. Time courses of total Ni concentration in effluents from PRB mixtures of indicated compositions and thicknesses (Grade III standard of Ni (GB/T14848/ 93) ≤ 0.05 mg/L).

12


Fig. 8. Time courses of total Pb concentration in effluents from PRB mixtures of indicated compositions and thicknesses (Grade III standard of Pb (GB/T14848/ 93) ≤ 0.05 mg/L).

with a higher proportion of ZVI, such as G and H, probably contained a larger number of small pores, resulting in smaller hydraulic conductivities. The actual K values of the natural uncontaminated hilly red soils in the landfill reportedly range from 0.00072 m/d to 3.216 m/d (Zhong et al., 2009). Therefore, the K values of the PRB materials would be sufficiently large for unhindered groundwater flow in the subsurface if PRB techniques were applied at the focal landfill. However, the conductivity of each mixture was much smaller than its initial value at 576 h (Table 3), and two of the largest differences were for mixtures C and F, suggesting that high proportions of zeolite tend to increase fouling of the reactive materials. 3.5. Target thickness and longevity of the PRB materials In tests with mixture G the concentrations of contaminants at the outlet increased with time, and even exceeded the official limits in late stages. However, during the first 72 h NH+ 4 concentrations at the 20, 50, and 80 cm sampling ports were 68.4 mg/L, 4.0 mg/L and undetectable, respectively, Ni concentrations were 0.39, 0.16 and 0.01 mg/L, respectively and Pb concentrations were 0.16, 0.01 and 0.01 mg/L, respectively. All the concentrations at the 80 cm outlet complied with the corresponding limits in the Standard (0.20, 0.05 and 0.05 mg/L, respectively), indicating that mixture G had the ability to reduce contamination to target concentrations. It should be noted that lab-scale column tests were performed to

determine the removal capacity of each mixture of reactive materials, i.e., the removal efficiency per unit volume of reactive materials and reaction time, which could be used to estimate the minimum thickness and longevity of a PRB for a real field application. Unsurprisingly, the contaminant removal efficiency clearly increased with increases in the path length of the mixtures, as illustrated for mixture G (identified as the most effective PRB material) in Fig. 11. Removal efficiencies of COD, Ni, Pb, TN and NH+ 4 , increased sharply with increases in the path length up to 288 h, suggesting that effective reactions occurred between the mixture and all these contaminants until this time. The removal efficiencies subsequently decreased and remained at a low level, suggesting that the materials were almost exhausted along the whole path and that the PRB device required renewal. The pH decreased from around 9.0 to 6.5 before the 168th hour, then increased and remained stable and high (at around 9.0) along the whole path length. However, during the first 168 h the oxidation–reduction reactions of the mixture with the contaminants were not sufficiently quick to dominate within the first 20 cm of the PRB mixture. The predominant reaction was therefore adsorption of NH+ 4 by AC and zeolite, which resulted in an increase in pH, according to Eq. (10). However, with increases in the path length the oxidation–reduction reaction became predominant, and increasing amounts of ferric ion were produced, resulting in the precipitation of white Fe(OH)2 by Fe2+ + OH− → Fe(OH)2↓, and brown Fe(OH)3 by Fe3+ + OH− → Fe(OH)3↓. These processes consumed large

Fig. 9. Time courses of pH in effluents from PRB mixtures of indicated compositions and thicknesses (Grade III standard of pH (GB/T14848/93) = 6.50–8.52).


Fig. 10. Time courses of hydraulic conductivity of PRB mixtures of indicated compositions.

quantities of OH−, and thus caused the fall in pH with increases in path length. After 168 h the reactions between ZVI and the contaminants became weak, since the oxidation–reduction reactions were much faster than the adsorption, while AC and zeolite still had effective adsorption capacity. Thus, NH+ 4 and organic contaminants in the leachate were adsorbed by AC and zeolite, resulting in the generation of OH− (Eq. (10)). Generally, weakly acidic functional groups (e.g., \OH and \COOH) can release H+ by ionization. However, the reduction of organic contaminants caused the decline of these groups due to the adsorption of AC and zeolite, resulting in the generation of OH− by the hydrolysis of these weakly acidic groups. All these processes resulted in the pH increasing with increases in the path length.

13

When the path length and reaction time were sufficient, Fe2+ produced by the reaction between ZVI and H+ combined with the organic contaminants in the leachate, which caused the release of H+, and thus the decrease in pH. However, this reaction readily approaches equilibrium and shifts to the opposite direction. Therefore, changes in pH with the path length after 288 h were small, and the pH was high in tests with all of the mixtures in the late stages (Fig. 9). All these findings suggest that the leachate tended to be alkaline after this PRB treatment. The theoretical longevity of mixture G was estimated and is shown in Table 4. Redox reactions between ZVI and heavy metals may occur very rapidly, and adsorption of heavy metal ions to AC may approach equilibrium within about 60 min, according to Cao et al. (2011). The estimated average velocity of the groundwater in the focal landfill was approximately 0.1 m/d. The time (0.2 days) required for the groundwater to pass through the minimum thickness of reactive material was sufficient to allow completion of the redox reactions, precipitation, and sorption between the PRB materials and contaminants in the groundwater. The main determinant of the viability of the PRB method for groundwater remediation is the minimum longevity of the PRB materials with respect to the five main contaminants. For the focal landfill here, the calculated minimum thickness of the optimum PRB mixture is 2.04 cm, and the thickness required to meet the minimum longevity of 10 years required for granular iron PRBs to be more cost-effective than pump-and-treat systems (Higgins and Olson, 2009) is 309.6 cm (Table 4). These results demonstrate that the PRB method is a practicable approach in terms of both the longevity and treatment efficiency. There are two general types of PRB designs for field applications: funnel-and-gate and continuous gate (Powell et al., 1998; Thiruvenkatachari et al., 2008). Most continuous

Fig. 11. Changes in contaminant removal efficiency with increases in path length in mixture G.

14


Table 4 Calculated longevities for indicated thicknesses of mixture G (ZVI:Zeo:AC in a 5:1:4 ratio).

COD NH+ 4 TN Ni Pb Subtotal a b c

Concentrationa (mg/L)

Standardb (mg/L)

LMINc (cm)

Longevities for different thicknesses (years) 20 cm

50 cm

80 cm

100 cm

200 cm

309.60 cm

108.72 29.77 35.47 0.015 0.006

34.94 0.00 2.14 0.0058 0.0025

1.57 1.25 2.04 0.85 0.62 2.04

0.84 1.05 0.65 1.54 2.12 0.65

2.11 2.64 1.61 3.86 5.32 1.61

3.37 4.22 2.58 6.18 8.51 2.58

4.22 5.28 3.23 7.72 10.64 3.23

8.43 10.55 6.46 15.44 21.27 6.46

13.05 16.33 10.00 23.90 32.93 10.00

Refers to the contaminant concentration in the groundwater. The smaller of either the issued standard or specification of groundwater quality or the lowest concentration that remains after treatment with the PRB materials. Minimum thickness of the PRB materials required to satisfy current standards.

gate PRBs are constructed using the trenching method, and only 30–90 cm thick (Naftz et al., 2002). Thus, the funnel-andgate approach should be used for a field PRB construction with a greater thickness of reactive materials and longer operational life. When a PRB approaches the end of its life due to exhaustion of reactive materials, and fouling of the pores due to secondary mineral precipitation and/or excessive decline of removal efficiency the reactive materials can be excavated and new materials installed into the PRB frame. The optimal dimensions, location and orientation of a PRB are generally site-specific due to the uniqueness of site characteristics (Henry et al., 2008; Ott, 2000; USEPA, 1998). However, the weight ratio of ZVI/zeolite/AC identified as optimal for reactive materials in this study should also be suitable for remediating groundwater contaminated by similar landfill leachates. 4. Conclusions We investigated the optimum composition of PRB reactive materials for the remediation of groundwater heavily contaminated by a landfill leachate and proposed a systematic method for estimating the minimum thickness and longevity toward five types of contaminants. The main contributors to the removal efficiency of organic and inorganic contaminants in the leachate were ZVI and AC, and the most effective weight ratio of a ZVI/zeolite/AC mixture was found to be 5:1:4, based on the measured reactivity and hydraulic conductivity. The average reductions in chemical oxygen demand (COD), and contents of total nitrogen (TN), ammonium, Ni, Pb and 16 polycyclic aromatic hydrocarbons (PAHs) were 55.8%, 70.8%, 89.2%, 70.7%, 92.7% and 94.2%, respectively. Thus, the PRB method most efficiently removed Pb and the 16 PAHs. The minimum thickness of PRB materials required to satisfy current Chinese standards for the quality of groundwater (GB standard GB/T14848/93) was estimated to be 2.04 cm. The results also show that a 309.6 cm thick layer of the optimal PRB mixture should meet the standard for at least 10 years. Acknowledgments The authors gratefully acknowledge Dr. John Blackwell (Sees-editing Ltd.) for his editing of this paper. This work was supported by the Chinese National Science Foundation (grant nos. 41072182; 21177162), Guangdong Major Science and Technology Project (grant no. 2012A030700008), Guangzhou Science and Technology Project (grant no. 2010Z1-E101), and the industry support project by Shandong Province: the

industrialization and application research of hermetic highflux intelligent electromagnetic digestion system.

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Laboratory column study for evaluating a multimedia permeable reactive barrier for the remediation of ammonium contaminated groundwater.

Cr(VI)-contaminated groundwater remediation with simulated permeable reactive barrier (PRB) filled with natural pyrite as reactive material: Environmental factors and effectiveness.

Permeable Adsorptive Barrier (PAB) for the remediation of groundwater simultaneously contaminated by some chlorinated organic compounds.

Remediation of MSW landfill leachate by permeable reactive barrier with vegetation.

Geochemical modelling for predicting the long-term performance of zeolite-PRB to treat lead contaminated groundwater.

Comparison of permeable reactive barrier, funnel and gate, nonpumped wells, and low-capacity wells for groundwater remediation.

First evidence of fish genotoxicity induced by heavy metals from landfill leachates: the advantage of using the RAPD-PCR technique.

Removal of ammonium-nitrogen from groundwater using a fully passive permeable reactive barrier with oxygen-releasing compound and clinoptilolite.

An overview of permeable reactive barriers for in situ sustainable groundwater remediation.

Use of the landfill water pollution index (LWPI) for groundwater quality assessment near the landfill sites.

Perfluoroalkyl acids in municipal landfill leachates from China: Occurrence, fate during leachate treatment and potential impact on groundwater.

Field survey of enteric viruses in solid waste landfill leachates.

Cement kiln dust (CKD)-filter sand permeable reactive barrier for the removal of Cu(II) and Zn(II) from simulated acidic groundwater.

Detection of enteric viruses in solid waste landfill leachates.

Optimization-based multicriteria decision analysis for identification of desired petroleum-contaminated groundwater remediation strategies.

Hydroxyl radical (OH) scavenging in young and mature landfill leachates.

Analysis of microbial community structure and composition in leachates from a young landfill by 454 pyrosequencing.

A bench-scale assessment of ozone pre-treatments for landfill leachates.

Hydraulic performance of a permeable reactive barrier at Casey Station, Antarctica.

activated carbon permeable reactive barrier.

Arsenic removal from contaminated groundwater by membrane-integrated hybrid plant: optimization and control using Visual Basic platform.

Chlorobenzene degradation by bacteria isolated from contaminated groundwater.

Pyrite-based autotrophic denitrification for remediation of nitrate contaminated groundwater.

Strategies of management for the whole treatment of leachates generated in a landfill and in a composting plant.