Effects of Temperature and Acidic Pre-Treatment on Fenton-Driven

Jan 29, 2009 - Research Center, P.O. Box 1198, Ada, Oklahoma 74820, and. U.S. Environmental Protection Agency, Office of Research and. Development ...
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Environ. Sci. Technol. 2009, 43, 1493–1499

Effects of Temperature and Acidic Pre-Treatment on Fenton-Driven Oxidation of MTBE-Spent Granular Activated Carbon E U N S U N G K A N †,§ A N D S C O T T G . H U L I N G * ,‡ National Research Council, Robert S. Kerr Environmental Research Center, P.O. Box 1198, Ada, Oklahoma 74820, and U.S. Environmental Protection Agency, Office of Research and Development, National Risk Management Research Laboratory, Robert S. Kerr Environmental Research Center, P.O. Box 1198, Ada, Oklahoma 74820

Received August 25, 2008. Revised manuscript received December 17, 2008. Accepted December 19, 2008.

The effects of temperature and acidic pretreatment on Fentondriven chemical oxidation of methyl tert-butyl ether (MTBE)spent granular activated carbon (GAC) were investigated. Limiting factors in MTBE removal in GAC include the heterogeneous distributionofamendedFe,andslowintraparticlediffusivetransport of MTBE and hydrogen peroxide (H2O2) into the “reactive zone”. Acid pretreatment of GAC before Fe amendment altered the surface chemistry of the GAC, lowered the pH point of zero charge, and resulted in greater penetration and more uniform distribution of Fe in GAC. This led to a condition where Fe, MTBE, and H2O2 coexisted over a larger volume of the GAC contributing to greater MTBE oxidation and removal. H2O2 reaction and MTBE removal in GAC increased with temperature. Modeling H2O2 transport and reaction in GAC indicated that H2O2 penetration was inversely proportional with temperature and tortuosity, and occurred over a larger fraction of the total volume of small GAC particles (0.3 mm diameter) relative to large particles (1.2 mm diameter). Acidic pretreatment of GAC, Fe-amendment, elevated reaction temperature, and use of small GAC particles are operational parameters that improve Fenton-driven oxidation of MTBE in GAC.

Introduction There may be as many as 250,000 releases of methyl-tertbutyl ether (MTBE) associated with leaking underground fuel tanks in the United States (1). MTBE is mobile and persistent in the environment due to its high water solubility, low Henry’s constant, and relative biorecalcitrance under most environmental conditons (2, 3). These properties also contribute to the difficulty and expense in treating MTBEcontaminated water. Granular activated carbon (GAC) is one method used in water treatment to remove MTBE. Selection of this technology relative to other technologies (i.e., air stripping, advanced oxidation) is dependent on site-specific * Corresponding author phone: (580) 436-8610; fax: (580) 4368614; e-mail:[email protected]. † National Research Council. ‡ U.S. Environmental Protection Agency. § Current affiliation: National Research Council. Department of Chemical and Petroleum Engineering, United Arab Emirates University, Al-Ain, United Arab Emirates. 10.1021/es802360f CCC: $40.75

Published on Web 01/29/2009

 2009 American Chemical Society

factors and includes the MTBE concentration in water, flow rate, removal efficiency requirements, and water quality parameters. Once the GAC is spent with MTBE, it is regenerated and placed back in service (i.e., reused), replaced with virgin GAC, or disposed. In nearly all cases involving GAC regeneration, the spent GAC is thermally regenerated either on-site or transported to a thermal regeneration facility and regenerated off-site. Fenton-driven treatment of the GAC is a treatment option under development to regenerate the GAC on-site and in situ. Additionally, Fenton-driven treatment of the spent-GAC could be performed prior to disposal to reduce the mass of adsorbate and potential for leaching after disposal. Fenton-driven regeneration of MTBE-spent granular activated carbon (GAC) involves a pretreatment step in which iron (Fe) is amended to the GAC. Subsequently, MTBE is adsorbed and concentrated onto the carbon surfaces and MTBE oxidation and GAC regeneration are achieved by amending H2O2. The reaction between H2O2 and the Fe in the GAC results in the formation of hydroxyl radicals ( · OH) that oxidize MTBE. The reaction rate constant between · OH and MTBE is high (1.6 × 109 L mol-1 s-1) (4) indicating the vulnerability of MTBE to radical attack in Fenton systems. Tertiary butanol (TBA) and acetone are the primary reaction byproducts from MTBE oxidation (5, 6). Subsequent oxidation of these intermediates involves the formation of a variety of carboxylic acids which also undergo oxidation by · OH, ultimately forming CO2. Formation of the highly reactive, nonselective · OH in proximity to high concentrations of target organics near carbon surfaces favors high transformation efficiency and overcomes slow reaction kinetics involving Fenton oxidation in dilute aqueous systems (5). During oxidation MTBE mass transfer and mass transport from the GAC (solid) to the bulk fluid involves the following steps: (1) desorption from solid to liquid phase, (2) diffusive transport within the pores involving pore and surface diffusion, (3) diffusive transport through a quiescent film surrounding the particle, and (4) advective transport into the bulk solution (Figure 1) (7, 8). The rate-limiting mechanism may be simplified to pore diffusion limitation if contaminant desorption is fast relative to pore diffusion (9). H2O2 diffusion from the bulk solution into the GAC pores reacts predominantly with the Fe immobilized in the GAC (10). Consequently, · OH are generated in the aqueous phase near the GAC surface. Since · OH are extremely short-lived, oxidation of MTBE and other contaminants occurs near the GAC surface where the · OH originate. The relative amount of MTBE, or other contaminants, transformed in either the aqueous or solid (sorbed) phases has not been quantified. Enhanced mass transfer of contaminants from the GAC surface to solution results from the steep concentration gradients established in the GAC pores during oxidation. Based on this conceptual model, the effect of temperature was investigated as a means to potentially improve MTBE oxidation and GAC regeneration. The initial reaction of H2O2 (Rxn 1) has a higher activation energy relative to other reaction steps (11) and the catalyzed H2O2 reaction is exothermic often resulting in an increase in solution temperature. Fe3++H2O2 f Fe2+ + · HO2+H+

(Rxn 1)

An increase in the solution temperature attributed either to the H2O2 exothermic reaction or externally applied heat, will enhance MTBE desorption, intraparticle and thin film diffusion of MTBE and H2O2, and the H2O2 reaction rate. VOL. 43, NO. 5, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. Schematic of the proposed mechanisms of (1) intraparticle MTBE mass transfer (desorption), (2) MTBE diffusive mass transport (pore + surface), intraparticle diffusive transport of H2O2, (3) MTBE diffusive transport outward through the quiescent film, H2O2 diffusive transport from the bulk solution through the quiescent film into the GAC pores, and (4) MTBE and H2O2 mixing in bulk solution. Further, H2O2 reacts with immobilized Fe (Fenton-like reaction) resulting in the formation of · OH and oxidation of MTBE by · OH. Intraparticle surface and pore diffusive transport involves a tortuous pathway quantified by the tortuosity factor (τF) ) actual transport distance (∑i)1 di)/direct path (DP). Fenton-driven regeneration of GAC involves the simultaneous occurrence of these multiple mechanisms. Previous research has not addressed the cumulative thermal effects of Fenton-driven contaminant destruction and GAC treatment. An increase in reaction temperature during the homogeneous Fenton reaction increased the H2O2 reaction rate and the rate of contaminant transformation (11-14). Specifically, sequential 10 °C increases in solution temperature during the homogeneous Fenton oxidation of 2,4-dichlorophenoxyacetic acid enhanced the contamination removal reaction rates by a factor of 1.6-2.5 (avg. 2.2) (11). Management of H2O2 concentration, application frequency, or external heat applied to the treatment unit could be used to regulate the temperature of the solution recirculated through the GAC treatment unit to optimize treatment performance. Acidic Fe solutions amended to GAC resulted in the immobilization of Fe predominantly within a narrow interval (i.e., approximately 18-20 µm) on the periphery of the GAC particle (10). Mechanisms of Fe immobilization in GAC include ion exchange, adsorption, surface complexation, and precipitation (15, 16). The pH at point of zero charge (pHpzc) is the pH at which positive and negative surface charges are equal and the GAC surface has a net charge of zero. Electrostatic repulsion of Fe2+ or Fe3+ in solution by positively charged surfaces of the GAC may hinder penetration and deposition of Fe deep in the GAC particle. Acidic pretreatment of GAC alters the surface chemistry by increasing the carboxylic and lactonic acidic surface oxide functional groups (17), increases the cation exchange capacity and Fe sorption (18), lowers the pHPZC (17, 19) and consequently, can increase the dispersion or quantity of metal catalysts within the GAC (17, 20-22). Similarly, we propose that acidic pretreatment of the GAC will lower the pHPZC of the GAC and may reduce repulsive forces between the cations (Fe2+, Fe3+) in solution and positively charged surface sites on the GAC. Application of the Fe solution at a pH above the pHPZC of the GAC will lower the electrostatic repulsion and allow greater Fe penetration and dispersion into the GAC. By extending the Fe-rich zone further into the GAC and increasing the volume of the GAC particle favorable for the Fenton mechanism, it is proposed that intraparticle desorption and diffusion 1494

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limitations will decline and contribute to an increase in overall MTBE removal (i.e., transformation) efficiency. The objectives of this study were to systematically investigate the effects of acidic pretreatment of GAC on Fe immobilization and distribution, in conjunction with the role of temperature on MTBE desorption and diffusion, H2O2 diffusion and reaction, and consequential effects on MTBE oxidation and GAC regeneration.

Methods and Materials The GAC (URV, 8 × 30 mesh, Calgon Carbon Corp., Pittsburgh, PA) was derived from bituminous coal and activated in a manner to minimize H2O2 reactivity (23). The GAC was rinsed with deionized (DI) water, dried in an oven at 105 °C, and stored in a desiccator until used. The surface area and pore volume of the GAC was 1290 m2/g and 0.64 mL/g, respectively (10). pH Point of Zero Charge. The pH at point of zero charge (pHPZC) of the GAC was determined using the pH drift method (24). Deionized (DI) water (50 mL) amended with NaCl (0.01 M) was placed in 100 mL amber vials and sparged with N2 (∼200 mL/min; 10-15 min) to eliminate CO2 and to stabilize pH. The pH was adjusted (pH 2-11) in a series of vials by adding either HCl or NaOH while purging the headspace with N2. GAC (0.15 g) was added and the vial was capped immediately. The final pH (pHFINAL) was measured in each of the vials after 48 h and plotted versus the initial pH (pHINITIAL). The pHPZC was determined graphically at the intersection of pHFINAL and the line pHFINAL) pHINITIAL. Acid Pretreatment, Fe Amendment, MTBE Adsorption, and Oxidative Treatment. GAC was prepared under three different conditions. A stock solution was prepared immediately before the Fe solution was amended to GAC by dissolving ferrous chloride (FeCl2 · 4H2O) into DI water (50 mg/L as Fe). Untreated, Fe-amended GAC (5 g/reactor) was sequentially (×3) suspended in a solution of Fe2+ (200 mL, 50 mg/L, pH 5-5.5). Acid-treated, Fe-amended GAC was amended with nitric acid (pH 3.0, 4 days) followed by the Fe amendment procedures described above. The GAC was

rinsed with DI water to eliminate the chloride residual in the suspension. Fe-unamended GAC received as is, was used as a control to assess the roles of Fe amendment and acid pretreatment. The MTBE adsorption step was performed by amending an MTBE solution (0.12 L, 1.8 mg/L) to water-saturated GAC (5 g of GAC, 15 mL of deionized water; pH 3.0-3.5) in 125 mL side-armed Erlenmeyer flasks wrapped in aluminum foil and parafilmed. The sidearm was fitted with a Teflon tube containing GAC that captured volatile emissions but allowed O2(g) and CO2(g) to be vented from the reactor. The postsorption MTBE solution was sampled after equilibrium (>3 d) in replicate and analyzed. Differences between initial and final concentrations were used to calculate the mass of MTBE adsorbed to the GAC. Postoxidation MTBE in GAC was measured by extracting subsamples (1 g) in 10 mL of methanol and analyzing the extract for MTBE. Individual reactors were used to assess the role of temperature at 25, 35, 45, and 55 °C. Constant temperature was maintained in a temperature-controlled shaking water bath (Cole Parmer, Vernon Hills, IL). Fenton oxidation of the MTBE-spent GAC was performed with six, sequential applications of H2O2 (1.9 mL of 30%) (pH of 3.0-3.5) on separate days. · OH scavenging by high concentrations of H2O2 is a probable source of overall oxidation inefficiency (10) and multiple applications of H2O2 were applied to minimize spikes in concentration and · OH scavenging. H2O2 concentrations were measured with time under complete mix conditions ([H2O2]INITIAL ∼9.8 g/L). MTBE removal efficiency in the GAC was evaluated by measuring the MTBE in the GAC after Fenton oxidation. The GAC in the sidearm traps was analyzed for MTBE. Volatile MTBE losses were minor (1-4%) and values were used to correct the estimates of MTBE removal. The MTBE desorption + diffusion rate from MTBE-spent GAC (25-55 °C) was evaluated using the fill and draw method. This involved postsorption applications of DI water (55 mL) to MTBE-spent GAC (5 g) every 15 min and measuring MTBE in solution. The rate of desorption + diffusion was calculated as the mass of MTBE desorbed from the GAC divided by the mass of GAC and time period of desorption. In all cases tested, the mass of MTBE on the GAC was not significantly depleted (97-98% remaining) and aqueous concentrations were significantly below the equilibrium MTBE concentration when the aqueous sample was collected. Analytical Methods. Analytical methods used in the measurement of GAC metals, aqueous phase and GAC MTBE, H2O2, and iron concentrations are available in the Supporting Information. Scanning electron microscopy/energy dispersive X-ray spectrometry (SEM/EDS) imaging and microanalysis of GAC particles was performed to assess the depth-dependent elemental composition of GAC particles. Detailed procedures are presented in the Supporting Information and have been reported elsewhere (10).

Results and Discussion Accumulation of Fe on the periphery of the untreated GAC was significant but the Fe content declined rapidly over a short transport distance (Figure 2). Low levels of Fe measured in the periphery (0-20 µm) of the acid-treated GAC suggests that the Fe penetrated further into the GAC. It is noted that Fe is not measured by the SEM-EDS beyond ∼20 µm into the GAC. The pHPZC of the acid-treated, GAC (pHPZC ) 4.2) was reduced from the background pHPZC (pHPZC ) 5.5) and was less than the pH of the Fe solution amended to the GAC (pH 5-5.5). Under this condition, negative charges on the GAC surface attract and adsorb cations in solution (i.e., Fe2+, Fe3+) and favor greater dispersion of the metal catalysts in the GAC (20).

FIGURE 2. Atomic chemical composition of Fe in acid-treated and untreated Fe-amended GAC. Measurements were by SEM/ EDS and the depth-dependency of Fe was derived by varying the accelerating voltage (5, 10, 20, 30 keV). Two particles were selected from test reactors and two sites were randomly selected on each particle (n ) 4). The surface area interrogated by EDS (47,000 µm2 per site) represented approximately 0.3-4.2% of the external surface area of 8 × 30 GAC particles (i.e., assuming spherical GAC particles). The radius of the GAC particles (300-1200 µm) represents GAC retained by 8 × 30 sieving. Residual Fe in solution with acid-treated GAC ([Fe]AQUEOUS ) 10.8 mg/L) (3-4 days contact time) was greater than the untreated GAC (2.7 mg/L). Correspondingly, the average mass of Fe immobilized in the GAC was greater using the untreated Fe-loading method (5.6 g/kg (n ) 4); 95% confidence interval (CI) 5.4-5.9 g/kg) than in the acid pretreated GAC (4.7 g/kg (n ) 4); 95% CI 4.6-4.8 g/kg). This apparent anomaly is partially attributed to the high degree of Fe sorption, precipitation, and immobilization in the periphery of the untreated GAC despite an anticipated increase in Fe penetration into the acid pretreated GAC (Figure 2). Using selective chemical extraction procedures, it has been determined that Fe amended to GAC was predominantly poorly ordered, amorphous Fe resulting from precipitation (25). The amorphous Fe in conjunction with the high stability of the Fe(III)-OsFe(II) interactions (26) provides a mechanism by which ferrous Fe is immobilized and that other reactions (i.e., precipitation, complexation) and immobilization mechanisms are compounded in the Fe-rich periphery of the GAC. Variability in the initial [MTBE]GAC (42.2-42.4 mg/kg) was insignificant among the Fe-unamended, untreated Fe-loaded, and acid pretreated GAC. MTBE oxidation and removal increased as a result of acid-treatment and Fe-amendment, and a temperature-dependent increase in H2O2 reaction and MTBE removal was measured under all conditions tested (Figure 3a, b). Enhanced MTBE oxidation and removal in the acid-treated, Fe-amended GAC was attributed to greater VOL. 43, NO. 5, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 3. (a) Pseudo first-order H2O2 reaction rate constants (kH2O2) (r2 > 0.97) for acid-treated and untreated Fe-amended, and Fe-unamended GAC at 25, 35, 45, and 55 °C. Under all conditions tested, kH2O2 increased with acid pretreatment, Fe-amendment, and increasing temperature. (b) Temperature-dependent MTBE removal in the GAC. The initial MTBE loading (42.2-42.4 mg MTBE/kg GAC) was similar in all cases. The average mass of Fe immobilized in the untreated GAC was 5.6 g/kg (n ) 4) (95% confidence interval (CI) 5.4-5.9 g/kg) and 4.7 g/kg (n ) 4) (95% CI 4.6-4.8 g/kg) in the acid pretreated GAC. MTBE removal increased with Fe-amendment, acid pretreatment, and temperature. penetration of Fe into the GAC. Tertiary butanol (TBA) and acetone, the main reaction byproducts from MTBE oxidation, undergo significant transformation in Fenton-driven regeneration of MTBE-spent GAC (5). Negligible accumulation of these reaction products was measured in the GAC and is consistent with a previous study involving similar GAC and oxidative conditions (5). Results are consistent with a conceptual model where the Fe distribution and size of the reaction zone within the GAC particle increased due to acid pretreatment. Two conceptual models are used to illustrate differences between MTBE and H2O2 transport and reaction within the GAC particle (Figure 4a, b). In the untreated, Fe-amended GAC particle, Fe is predominantly immobilized on the periphery of the GAC particle (Figure 4a). Under this condition, H2O2 transport into the GAC particle is impeded by the high rate of reaction in the Fe-rich zone. Exterior of the GAC particle, MTBE oxidation in solution is limited due to low concentrations of Fe and MTBE. Interior of the GAC particle, past the Fe-rich zone, H2O2 reaction and MTBE oxidation are also limited due to low concentrations of Fe. Consequently, only in the narrow Fe-rich zone where H2O2, Fe, and MTBE coexist, are conditions favorable for MTBE oxidation. Under this condition, MTBE oxidation is dependent on desorption and intraparticle (outward) diffusion. In Fe-unamended GAC, the background Fe distribution is probably the most uniform but H2O2 reaction is the lowest (Figure 3a). Slow H2O2 reaction contributes to higher H2O2 persistence in the treatment system and greater diffusive transport distances into the GAC particle. However, MTBE oxidation is most limited (Figure 3b) due to low concentrations of background Fe. In the acid-treated, Fe-amended GAC, results suggest that Fe penetrated further into the GAC and Fe immobilization was more uniformly distributed than in the untreated Feamended GAC (Figure 2). It is assumed in this conceptual model that Fe distribution is uniform and extends to the center of the particle (Figure 4b). Under this condition, the reaction of H2O2 is not limited to the Fe-rich periphery of the 1496

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GAC particle. Rather, H2O2 may penetrate further into the GAC and react over a larger volume of the GAC particle. Consequently, the “reactive zone” is larger and the MTBE transport distance (i.e., from the GAC interior) required to reach this zone is shorter. The combination of these factors is attributed to the increase in MTBE oxidation (Figure 3b). Contrasting the temperature-dependent relative increase in calculated values of MTBE diffusion, measured values of MTBE desorption + diffusion, and Fenton-driven MTBE removal provides insight into MTBE mass transfer, mass transport, and transformation in GAC (Figure 5). The diffusivity of MTBE in water was calculated using the Wilke-Chang equation (eq 1) (27) and varied by nearly a factor of 2 over the temperature range (25-55 °C) used in these experiments (Table 1). DC,W ) T × 7.4 × 10-8(ΦWMW)1⁄2⁄(µW × VC0.6)

(1)

where DC,W ) diffusivity of chemical C in water (cm /s at 1 atm.), µW ) viscosity of water (centipoise), T ) absolute temperature (K), MW ) molecular weight of water (18 g/mol), VC ) molar volume of chemical C at normal boiling point (128 and 27.2 cm3/g-mole for MTBE and H2O2, respectively), and ΦW ) association parameter for water (2.26, dimensionless) (27). MTBE desorption + diffusion values, measured using the fill and draw procedures, increased with rising temperatures. A similar increasing trend in calculated MTBE diffusion values, and in measured MTBE desorption + diffusion values, with temperature suggests that diffusion is the limiting step to MTBE transport from GAC (Figure 5). This is consistent with others (9) who reported contaminant desorption and diffusion are important steps in the mass transport of contaminants in GAC, but mass transport is predominantly controlled by diffusive transport. MTBE removal did not exhibit a slope similar to that of MTBE desorption + diffusion, or MTBE diffusion (Figure 5). This suggests a mechanism other than MTBE transport limited MTBE transformation and removal from the GAC. 2

FIGURE 4. (a) Untreated, Fe-amended GAC. Fe immobilization occurred within a short transport distance into GAC particles resulting in a narrow interval over which Fenton-driven oxidation of MTBE was favorable. (b) Acid-treated, Fe-amended GAC. Surface chemistry changes in GAC from acid pretreatment resulted in greater penetration of Fe into the GAC. Consequently, the reactive zone in which MTBE oxidation is favorable is extended over a larger volume of the GAC particle. A critical analysis of the temperature-dependence of H2O2 fate and transport in GAC provides additional insight. H2O2 diffusion in macroporous solids is functionally dependent on the tortuosity in the GAC (eq 2). DH2O2,GAC ) DH2O2,W ⁄ τf

(2)

where DH2O2,GAC ) diffusivity of H2O2 in water saturated GAC (cm2/s), DH2O2,W ) diffusivity of H2O2 in water (Table 1) (cm2/ s), and τf ) tortuosity (dimensionless). Tortuosity (τf) is the ratio of the actual transport distance to the shortest (direct) pathway distance (Figure 1). A tortuosity factor of 1 indicates that the actual transport pathway is equal to the shortest transport pathway and tortuosity is negligible. A τf value of 10 for H2O2 in activated carbon (28) was assumed to represent the tortuosity conditions in our study. A mathematical model involving H2O2 diffusion and reaction in GAC was developed to assess the transport distance of H2O2 in GAC (eq 3). A numerical differential equation solver (Berkeley Madonna, version 8.3, Berkeley Madonna Inc., CA) was used to solve eq 3.

(

)

2 ∂C ∂2C ∂C ) DH2O2,GAC + 2 - kH2O2C ∂t r ∂r ∂r

(3)

where C ) concentration of H2O2 (mg/L), r ) radius of GAC particle (µm), and kH2O2 ) pseudofirst order H2O2 reaction rate constant (s-1). The GAC particle was assumed to have a spherical shape and uniform distribution of Fe (i.e., acid-treated, Fe-amended GAC). H2O2 reaction rates in the acid-treated, Fe-amended

reactors were simulated in GAC particles at different temperature (25, 55 °C), particle size (radius (r) ) 300, 1200 µm), and tortuosity factor (1-10). H2O2 reaction was modeled using pseudo first-order kinetics consistent with laboratory results (Figure 3a). Simulations involved segmented GAC particles in ten (10) differential volumes (spherical shells) with equal radii (re ) r/10) from the GAC center (i.e., re ) 30 and 120 µm for r ) 300 and 1200 µm, respectively). The thickness of an outer layer extending beyond the carbon particle was sized in a manner to ensure the H2O2 concentration was similar to the bulk phase concentrations measured in the study. Therefore, the H2O2 concentration on the outside edge of the first layer of the carbon particle corresponded to measured H2O2 concentrations in the reactor. Simulations for three separate differential volumes (spherical shells) of the GAC particle were critically analyzed (i.e., the outermost, midpoint, and innermost differential volumes). Modeling results indicate that H2O2 significantly penetrated the outermost differential volume at both 25 and 55 °C, however, deeper penetration into the GAC particle was significantly limited (i.e., midpoint and innermost spherical shells) (Figure 6). H2O2 concentrations were higher and penetration was greater at 25 than 55 °C, but concentrations were less than 10% of the initial concentration. H2O2 concentrations in GAC were inversely proportional with tortuosity. Specifically, at lower tortuosity, a smaller fraction of H2O2 reacts in the periphery of the GAC particle allowing H2O2 to penetrate deeper into the GAC (refer to Supporting Information Figure SI-1). For example, assuming a tortuosity of 5 (T ) 55 °C), H2O2 reaches the midpoint and innermost VOL. 43, NO. 5, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 5. Relative increases in the temperature-dependent estimates of MTBE diffusion, and measured values of MTBE desorption + diffusion, and MTBE removal (acid-treated, Fe-amended GAC).

TABLE 1. Temperature Dependency of MTBE Diffusivity in Water temper- viscosity ature of water (°C) (centipoise) 25 35 45 55

0.90 0.73 0.60 0.51

MTBE overall H2O2 diffusivity increase diffusivity in water in MTBE in water (× 10-5 cm2/s)a diffusivityb (× 10-5 cm2/s)a 0.85 1.10 1.36 1.66

1 1.29 (1.29) 1.60 (1.24) 1.95 (1.22)

2.2 2.8 3.4 4.2

a Diffusivity calculated from eq 1. b Relative increase in MTBE diffusivity from 25 °C in parenthesis.

spherical shells at 10-15% of the initial H2O2 concentration. Results also indicate that H2O2 penetrates a larger fraction of the total volume of small GAC particles (300 µm) more rapidly relative to large GAC particles (1200 µm) (refer to Figure SI-2). Overall, modeling results suggest that H2O2 penetration and MTBE oxidation is inversely correlated with the size of GAC particles. The dimensionless Thiele-modulus (Φ) for porous, spherical particles (28, 29) can be used to critically analyze the relative roles of intraparticle H2O2 diffusion and reaction in GAC (eq 4). 1

Φ ) [dGAC ⁄ 6] × [(kH2O2 × MGAC ⁄ Vpores) ⁄ (CGAC × DH2O2,GAC)] 2

(4) where dGAC ) diameter of GAC particle (µm) (600 - 2400 µm), MGAC ) mass of GAC (g) (5 g), Vpores ) volume of GAC 1498

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FIGURE 6. H2O2 penetration into GAC particles during oxidation (25, 55 °C, [H2O2]INITIAL ∼ 9500 mg/L, tortuosity factor (τf) ) 10). The radius of the GAC particle is 1200 µm and simulations are presented for the innermost, midpoint, and outermost differential volumes of an idealized spherical GAC particle. pores (mL) (2.2 mL/g GAC), and CGAC ) concentration of GAC in slurry (0.1 g/ mL). Given the temperature-dependent values for kH2O2 (Figure 3a) and tortuosity-corrected values of DH2O2, GAC (Table 1, eq 2), Φ ) 0.6 and 0.9 for small GAC particles (dGAC ) 300 µm) at 25 and 55 °C, respectively, and Φ ) 2.5 and 3.4 for large GAC particles (dGAC ) 1200 µm) at 25 and 55 °C, respectively. The effectiveness factor (η) is the ratio of the actual H2O2 reaction and the ideal reaction (assuming no diffusional limitations) and is a measure of the reduction in the H2O2 reaction rate attributed to H2O2 diffusion limitations (29). For example, Φ , 1, η ) 1 indicates no pore diffusion limitations and no reduction in reaction; Φ ) 1, η ∼ 0.76 indicates some limitation; and Φ . 1, η ∼ 1/Φ indicates strong pore diffusion limitations. In our study, it is estimated that some H2O2 diffusion limitation occurred in small GAC particles (dGAC ) 300 µm; 25-55 °C) where Φ ∼ 1 and η ∼ 0.76. In large GAC particles (dGAC ) 1200 µm; 55 °C) where Φ ) 2.5 - 3.4, η was estimated to be 0.29-0.4 and indicated significant H2O2 pore diffusion limitations and consequently, a reduction in H2O2 reaction in the GAC. These results are consistent with model results which indicated H2O2 diffusion transport limitations in large GAC particles (dGAC ) 1200 µm) (refer to Figure SI-2). Enhanced MTBE oxidation and removal in GAC measured at higher solution temperature was predominantly attributed to an increase in MTBE and H2O2 diffusion. Optimal treatment conditions for Fenton-driven regeneration of MTBE-spent GAC include acid pretreatment of small GAC particles, amendment of Fe in a solution at a pH > pHPZC, and Fentondriven regeneration at elevated temperatures.

Acknowledgments We acknowledge Drs. R. G. Arnold, W. P. Ela, and R. E. Sierka (University of Arizona), and M. Blankenship, T. Pardue, and Dr. B. Pivetz (Shaw Environmental Inc., Ada, OK) for their

assistance. The U.S. Environmental Protection Agency, through its Office of Research and Development, funded and managed the research described here. It has not been subjected to Agency review and therefore does not necessarily reflect the views of the Agency, and no official endorsement should be inferred.

Supporting Information Available Additional details of analytical methods used in this research and numerical modeling results of H2O2 transport (diffusion) and reaction in a GAC particle. This material is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Johnson, R.; Pankow, J.; Bender, D.; Price, C.; Zogorski, J. MTBE - To What Extent Will Past Releases Contaminate Community Water Supply Wells? Environ. Sci. Technol. 2000, 34 (9), 210A– 217A. (2) Keller, A. A.; Sandall, O. C.; Rinker, R. G.; Mitani, M. M.; Bierwagen, B. An evaluation of physicochemical treatment technologies for water contaminated with MTBE. Ground Water Monit. Rev. 2000, (114), 114–126. (3) Sutherland, J.; Adams, C.; Kekobad, J. Treatment of MTBE by air stripping, carbon adsorption, and advanced oxidation: technical and economical comparison for five ground waters. Water Res. 2004, 193–205. (4) Buxton, G. V.; Greenstock, C.; Hellman, W. P.; Ross, A. B. Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms, and hydroxyl radicals ( · OH/ · O-) in aqueous solution. J. Phys. Chem. Ref. Data. 1988, 17 (2), 513–886. (5) Huling, S. G.; Jones, P. K.; Ela, W. P.; Arnold, R. G. Fentondriven chemical regeneration of MTBE-spent GAC. Water Res. 2005, 39 (10), 2145–2153. (6) Stefan, M. I.; Mack, J.; Bolton, J. R. Degradation pathways during the treatment of methyl tert-butyl ether by the UV/ H2O2 process. Environ. Sci. Technol. 2000, 34 (4), 650–658. (7) De Las Casas, C.; Bishop, K.; Bercik, L.; Johnson, M.; Potzler, M.; Ela, W.; Saez, A. E.; Huling, S. G.; Arnold, R. G. In-Place Regeneration of GAC using Fenton’s Reagents; Clark, C. , Lindner, A. , Eds.; ACS symposium series 940; American Chemical Society: Washington DC, 2006; pp 43-65. (8) Crittenden, J. C.; Hand, D. W.; Arora, H.; Lykins, B. W. Design considerations for GAC treatment of organic chemicals. AWWA 1987, 79, 74–82. (9) Crittenden, J.; Hutzler, N.; Geyer, D.; Oravitz, J.; Friedman, G. Transport of organic compounds with saturated groundwater flow: Model development and parameter sensitivity. Water Resour. Res. 1986, 22 (3), 271–284. (10) Huling, S. G.; Jones, P. K.; Lee, T. R. Iron optimization for Fentondriven oxidation of MTBE-spent granular activated carbon. Environ. Sci. Technol. 2007, 41, 4090–4096. (11) Lee, Y.; Lee, C.; Yoon, J. High temperature dependence of 2,4dichlorophenoxyacetic acid degradation by Fe3+/H2O2 system. Chemosphere 2003, 51, 963–971.

(12) Lunar, L.; Sicilia, D.; Rubio, S.; Perez-Bendito, D.; Nickel, U. Degradation of photographic developers by Fenton’s reagent; Condition optimization and kinetics for metol oxidation. Water Res. 2000, 34 (6), 1791–1802. (13) Lee, C.; Yoon, J. Temperature dependence of hydroxyl radical formation in the hν/Fe3+/H2O2 and Fe3+/H2O2 system. Chemosphere 2004, 56, 923–934. (14) Lopez, A.; Mascolo, G.; Detomaso, A.; Lovecchio, G.; Villani, G. Temperature activated degradation (mineralization) of 4-chloro3-methyl phenol by Fenton’s reagent. Chemosphere 2005, 59, 397–403. (15) Hegenberger, E.; Wu, N. L.; Phillips, J. Evidence of strong interaction between iron particles and an activated carbon support. J. Phys. Chem. 1987, 91, 5067–5071. (16) Pakula, M.; Biniak, S.; Swiatkowski, A. Chemical and electrochemical studies of interactions between iron (III) ions and an activated carbon surface. Langmuir 1998, 14, 3082–3089. (17) Zhu, Z. H.; Radovic, L. R.; Lu, G. Q. Effects of acid treatments of carbon on N2O and NO reduction by carbon supported copper catalysts. Carbon 2000, 38, 451–464. (18) Kinoshita, K. Carbon. Electrochemical and Physicochemical Properties; John Wiley and Sons Inc.: New York, 1988; 475 pp. (19) Noh, J. S.; Schwarz, J. A. Effect of HNO3 treatment on the surface acidity of activated carbons. Carbon 1990, 28, 675–682. (20) Tseng, H. H.; Wey, M. Y. Effects of acid treatments of activated carbon on its physiochemical structure as a support for copper oxide in DeSO2 reaction catalysts. Chemosphere 2006, 62, 756– 766. (21) Chen, W.; Parette, R.; Zou, J.; Cannon, F. S.; Dempsey, B. A. Arsenic removal by iron-modified activated carbon. Water Res. 2007, 41, 1851–1858. (22) Mondal, P.; Balomajumder, C.; Mohanty, B. A laboratory study for the treatment of arsenic, iron, and manganese bearing ground water using Fe 3+ impregnated activated carbon: effects of shaking time, pH and temperature. J. Hazard Mater. 2007, 144 (1-2), 420–426. (23) Hayden, R. Calgon Carbon Corporation, Pittsburgh, PA, 2001, Personal Communication. (24) Lopez-Ramona, M. V.; Stoecklib, F.; Moreno-Castillaa, C.; Carrasco-Marina, F. On the characterization of acidic and basic surface sites on carbons by various techniques. Carbon 1999, 37, 1215–1221. (25) Huling, S. G.; Jones, P. K.; Ela, W. P.; Arnold, R. G. Repeated reductive and oxidative treatments on granular activated carbon. J. Environ. Eng. 2005, 131 (2), 287–297. (26) Roden, E. E.; Zachara, J. M. Microbial reduction of crystalline iron(III) oxides: influence of oxide surface area and potential for cell growth. Environ. Sci. Technol. 1996, 30 (5), 1618–1628. (27) Welty, J. R., Wicks, C. E., Wilson, R. E., Rorrer G. L., Eds. Fundamentals of Momentum, Heat, and Mass Transfer; John Wiley and Sons, Inc.: New York, NY, 2000. (28) Georgi, A.; Kopinke, F. Interaction of adsorption and catalytic reaction in water decontamination processes. Part I. oxidation of organic contaminants with hydrogen peroxide catalyzed by activated carbon. Appl. Catal., B 2005, 58, 9–18. (29) Schmidt, L. D. The Engineering of Chemical Reactions; Oxford University Press: New York, 1998; 536 pp.

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