Phase Transfer of Palladized Nanoscale Zerovalent Iron for

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Phase Transfer of Palladized Nanoscale Zerovalent Iron for Environmental Remediation of Trichloroethene Sourjya Bhattacharjee† and Subhasis Ghoshal*,† †

Department of Civil Engineering, McGill University, Montreal, Quebec H3A 0C3, Canada S Supporting Information *

ABSTRACT: Palladium-doped nanoscale zerovalent iron (Pd-NZVI) has been shown to degrade environmental contaminants such as trichloroethene (TCE) to benign endproducts through aqueous phase reactions. In this study we show that rhamnolipid (biosurfactant)-coated Pd-NZVI (RLPd-NZVI) when reacted with TCE in a 1-butanol organic phase with limited amounts of water results in 50% more TCE mass degradation per unit mass of Pd-NZVI, with a 4-fold faster degradation rate (kobs of 0.413 day−1 in butanol organic phase versus 0.099 day−1 in aqueous phase). RL-Pd-NZVI is preferentially suspended in water in biphasic organic liquidwater systems because of its hydrophilic nature. We demonstrate herein for the first time that their rapid phase transfer to a butanol/TCE organic phase can be achieved by adding NaCl and creating water-in-oil emulsions in the organic phase. The significant enhancement in reactivity is caused by a higher electron release (3e− per mole of Fe0) from Pd-NZVI in the butanol organic phase compared to the same reaction with TCE in the aqueous phase (2e− per mole of Fe0). XPS characterization studies of Pd-NZVI show Fe0 oxidation to Fe(III) oxides for Pd-NZVI reacted with TCE in the butanol organic phase compared to Fe(II) oxides in the aqueous phase, which accounted for differences in the TCE reactivity extents and rates observed in the two phases.

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INTRODUCTION

contamination, because it contains the major mass fraction of the contaminant. An impediment to achieving direct degradation of TCE in the NAPL using NZVI or Pd-NZVI is the hydrophilic nature of the nanoparticles which prevents their migration into the NAPL phase. In a study by Saleh et al.,6 NZVI particles were coated with a novel triblock copolymer (poly(methacrylic a ci d )- blo ck -p ol y (m e t hy l m e t hacr y lat e)- blo ck- p o l y (styrenesulfonate)) consisting of an anchoring component as well as a hydrophobic and a hydrophilic component, to target the TCE NAPL-water interface; but the ability of these triblock copolymer coated NZVI particles to degrade TCE NAPL was not demonstrated, which is an important criterion for remediation performance. Only one study to date has evaluated the ability of NZVI to degrade TCE directly in the NAPL7 wherein NZVI was directly emplaced into a TCE NAPL which was amended with varying amounts of water. The rate of TCE degradation was observed to be strongly dependent upon the quantity of water present within the NAPL. No studies have been reported on degradation of TCE NAPL by Pd-NZVI. Efforts to create higher contact between NZVI and TCE motivated the implementation of emulsified NZVI (ENZVI)

Nanoscale zerovalent iron (NZVI) can degrade trichloroethene (TCE), a toxic but common environmental pollutant in groundwater, through reductive dechlorination reactions to ethane and other innocuous byproducts.1 Pd0 deposited on the NZVI surface (Pd-NZVI) can enhance the rate of TCE degradation by acting as a hydrogenation catalyst and/or by shuttling electrons to TCE via the formation of a galvanic couple with Fe0.2 However, one of the major challenges in addressing TCE contamination in groundwater stems from its tendency to migrate deep into the aquifers because it is denser than water (density 1.46 g/mL) where the TCE nonaqueous phase liquid (NAPL) acts as a long-term source of contamination because of slow dissolution of TCE in the groundwater. In typical in situ applications of Pd-NZVI, the nanoparticles would be coated with surface modifiers such as polyelectrolytes and surfactants in an aqueous solution and injected into the aquifer. The surface modifier coating is essential in order to improve Pd-NZVI subsurface transportability by reducing particle aggregation in the aqueous media and deposition on groundwater media.3−5 However, this remediation approach achieves degradation of only the dissolved aqueous fraction of TCE at any point in time, and the degradation of the TCE in the NAPL is limited by the rates of dissolution. It would be more desirable to target and degrade the TCE NAPL, the source-zone for the groundwater © XXXX American Chemical Society

Received: April 3, 2016 Revised: June 17, 2016 Accepted: July 5, 2016

A

DOI: 10.1021/acs.est.6b01646 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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and monorhamnolipid with M.W. 504 g mol−1) was purchased from Jeneil Biosurfactant Co. (Saukville, WI). Gas standards of ethane, ethylene, methane (99% purity), and 1-, cis-, transbutene (1000 ppm in N2) were obtained from Scotty Specialty Gases. Chloroethenes (vinyl chloride and cis-1,2- and trans-1,2dichloroethene) and hexenes (cis-3- and trans-3- ≥ 95%) were obtained from Sigma-Aldrich. Methanol and 1-butanol (99% purity) were purchased from Fisher Scientific. Water used in experiments was Millipore double deionized water. NZVI Synthesis. Bare NZVI particles were synthesized using a procedure described previously.18 Briefly, 0.019 M NaBH4 was added dropwise at 3 mL/min using a syringe pump to a continuously mixed aqueous solution of 0.07 M FeSO4· 7H2O which was prepared in 30% methanol, followed by a mixing time of 1 h. The resulting NZVI suspension was washed three times with methanol and dried under nitrogen and stored in sealed vials in an anaerobic glovebox (Coy Laboratories) containing high purity N2/H2 (95%:5%). RL-Pd-NZVI Synthesis. 40 mg of dried NZVI was added in a 60 mL vial containing 18.8 mL of H2O and sonicated for 10 min to disperse the nanoparticles. Next RL was coated onto NZVI by addition of 0.2 mL of 10 g/L TOC RL stock to the NZVI suspension and mixing on a table top shaker at 300 rpm at 25 ± 1 °C for 20 h (RL-NZVI). Thereafter, a 0.4 mL ethanolic solution of 1 g/L of Pd-acetate was added to the RLNZVI and sonicated for 10 min to synthesize RL-Pd-NZVI (Pd(O2CCH3)2 = 1 wt % of NZVI). Phase Transfer Protocol. 0.1 mL of pure TCE was then added to the RL-Pd-NZVI suspension which resulted in the formation of an immiscible oil phase at the bottom of the vial. Thereafter 4 mL of butanol was added resulting in a clear separate phase at the top of the aqueous solution. All of the components were then subjected to sonication (37 kHz frequency, FisherBrand 11203 Ultrasonicator) for 10 min, which created a grayish-black suspension. Finally, 1.5 mL of 6 M NaCl was added to commence the phase separation and transfer of nanoparticles into the organic phase which occurred within 10 min. The sonication of the TCE/RL-Pd-NZVI/water mixture had to be carried out in the presence of butanol and was followed by addition of NaCl to facilitate efficient phase transfer. For no phase transfer systems, 5.5 mL of H2O was added to the vial after the RL-Pd-NZVI preparation step followed by 0.1 mL of pure TCE. Prior to use, all aqueous and nonaqueous components used in the synthesis protocols and phase transfer and no phase transfer systems were deoxygenated by bubbling nitrogen. Reactivity Studies. TCE degradation experiments were carried out in 60 mL vials capped with crimp-sealed butyl rubber septa, and samples were prepared in the anaerobic glovebox. No phase transfer systems (termed “SYSTEM A”) were mixed at 250 rpm on a shaker, while phase transferred systems (termed “SYSTEM B”) were kept quiescent. Degradation products were quantified through headspace measurements of reaction vials in GC (FID and MS). Further details are provided in the SI. Due to the presence of TCE NAPL in the test systems and the use of stoichiometrically limited amounts of Pd-NZVI compared to TCE, headspace TCE concentrations in SYSTEMS A and B measured over the duration of TCE degradation experiments were constant. This was because the overall degradation of TCE NAPL by Pd-NZVI resulted in relatively small decreases compared to the TCE NAPL mass in

for the in situ treatment of TCE source zones.8−10 In this approach, aqueous suspensions of NZVI were mixed with oil to emulsify NZVI in oil−water droplets, and TCE was partitioned at high concentrations into the oil phase in close proximity to NZVI. The authors observed a 9% increase in TCE degradation with ENZVI compared to systems containing nonemulsified NZVI, with stoichiometrically excess Fe0.10 Although lab scale studies showed improved performance of ENZVI over NZVI for TCE removal due to enhanced partitioning and degradation of TCE,10 field scale application of ENZVI highlighted challenges with its subsurface delivery due to its viscous nature (∼2000 cP).9 Phase transfer processes which enable the translocation of nanoparticles between aqueous and nonaqueous phases can be immensely advantageous in various applications of nanoparticles because they can facilitate chemical reactions between reactants which are originally dispersed in different phases. For instance, Li et al.11 studied the hydrogenation of olefins in an aqueous/1-butanol biphasic system using rhodium nanoparticles. The nanoparticles were stabilized with a thermoregulated ligand Ph2P(CH2CH2O)nCH3 (n = 16) that enabled them to switch from the aqueous to the butanol phase and vice versa by changing the temperature. The nanoparticles phase transferred into butanol were catalytically very active. Generally to enable such phase transfer, specific ligands are employed often coupled with modifications in the physicochemical properties (e.g., ionic strength, temperature) of the organic or aqueous phases.12−15 However, these processes are difficult to engineer due to challenges associated with the selection or synthesis of an efficient ligand and preservation of nanoparticle functionality in the transferred phase.14,15 Although not employed for nanoparticles, the concept of phase transfer of reactive agents has been investigated previously to study the oxidative dechlorination of TCE NAPL using MnO4− ions.16,17 Seol and Schwartz16 employed phase transfer agents such as tetraethylammonium, tetrabutylammonium, and pentyltriphenylphosphonium cations, to transfer MnO4− directly into the TCE NAPL. The authors monitored Cl− ion generation as an indicator for TCE breakdown and observed up to 50% increase in Cl− ions (over a reaction time of 60 min) with the use of phase transfer agents, compared to Cl− generation without the use of phase transfer agents, suggesting that oxidative decomposition of TCE in the NAPL was promoted. In our study, we have employed a tertiary organic phase which is miscible with TCE NAPL, to enable a biosurfactantcoated Pd-NZVI to interact with TCE in a NAPL through a simple phase transfer process. We investigated the TCE degradation rates and extent, end products generated, and changes in nanoparticle morphology and surface chemistry in order to evaluate the overall applicability potential of such a nanoparticle phase transfer process for environmental remediation. We observed a 50% increase in the amount of TCE NAPL degraded by phase transferred Pd-NZVI in the organic phase compared to that degraded in the aqueous phase with major fraction of end products (90%) being nontoxic ethene, ethane, and butenes.

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MATERIALS AND METHODS Chemicals and Gases. Ferrous sulfate heptahydrate (99%), sodium borohydride (≥98.5%), and palladium acetate (99%) were purchased from Sigma-Aldrich. Rhamnolipid (RL) JBR215 (mixture of dirhamnolipid with M.W. 650 g mol−1 B

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Figure 1. (a) TEM image of NZVI and (b) chemical structures of mono- and di rhamnolipid.40

Figure 2. Phase transfer process: (a) TCE is found settled as a pure NAPL phase at the bottom of the vial, while RL-Pd-NZVI is in the aqueous phase. (b) Formation of an emulsion after sonication. (c) Phase transferred RL-Pd-NZVI in 1-butanol phase after NaCl addition.

(37%), and the liberated H2 gas was measured using a GCTCD. X-ray photoelectron spectroscopy (XPS) and transmission electron microscopy (TEM) were performed for nanoparticles before and after reaction using a VG Escalab 3MKII instrument and a Tecnai G2F20 S/TEM, respectively. Further details on instrument and sample preparation are included in the SI. Optical microscopy images were obtained by placing a drop of the butanol phase water-in-oil emulsion on a slide and observing it under an Olympus BX51 microscope. Phase Characterization. The aqueous and organic phases after the phase transfer process were characterized on a mass basis, in the absence of Pd-NZVI to avoid rapid degradation losses of TCE and in the absence of any headspace to avoid partitioning losses of volatile components. The difference in the mass of the vial was noted after the addition of each component. After the phase separation was completed, the aqueous phase was carefully removed and placed in a 60 mL vial. Based on the air/water partitioning of TCE, butanol, and ethanol, their concentrations in the aqueous phase were determined through headspace measurements in GC-FID. The salt concentration in the aqueous phase was determined using a conductivity meter (Fisher Scientific Traceable Conductivity Meter). Water content in the organic phase was measured using a Karl Fischer coulometric titrator (Mettler Toledo C30 Compact Karl Fischer Coulometer). A detailed schematic and stepwise explanation is provided in Figure S1.

the systems. Thus, tracking the TCE disappearance with time was not feasible. Therefore, TCE degradation products were quantified at each intermediate time point, and a carbon mass balance approach was used to obtain the corresponding TCE degraded. Calibration standards were prepared by adding known quantities of the gas standards in the reactors set up exactly like SYSTEMS A and B but without the Pd-NZVI. Analytical Methods. The TOC content of RL was determined using a TOC analyzer (Shimadzu Corp.), and 1 g/L TOC corresponds to 1.7 g/L mass concentration of RL. The mass of RL adsorbed to the NZVI surface was estimated by measuring the difference between the unadsorbed RL in solution after equilibration (i.e., 20 h mixing of NZVI and RL) and the total RL dose. The NZVI was separated by centrifugation (6500g, 20 min) and then retained in a vial by the use of a super magnet (K&J Magnetics Inc.), while the supernatant was decanted and analyzed in the TOC analyzer. Analysis revealed complete sorption of RL to NZVI. The mass of Pd deposited on NZVI particles was measured using an ICP-OES (Thermo ICap Duo 6500). The RL-PdNZVI was separated from solution using centrifugation followed by magnetic separation, and then the nanoparticles and supernatant were separately acid digested in aqua regia (3:1 HCl:HNO3). Pd deposited on NZVI was 0.5% w/w NZVI. Fe0 content was measured using acid digestion as mentioned elsewhere.1 The NZVI particles were acid digested in HCl C

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Figure 3. Composition by weight percent for (a) organic (butanol/TCE) phase exculding Pd-NZVI and (b) aqueous phase excluding Pd-NZVI. (c) Optical microscopy of the organic phase at 10× magnification. (d) Optical microscopy of organic phase at 60× magnification.

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RESULTS AND DISCUSSION Nanoparticle Characterization. NZVI used in this study, shown in the TEM image in Figure 1a, consisted of spherical particles in the 20−100 nm size range arranged in chainlike aggregates with a measured BET surface area of 25 m2/g. RL has been shown to improve the transport of Pd-NZVI in a range of subsurface media at relatively low concentrations3,4 and has a chemical structure shown in Figure 1b. RL-Pd-NZVI was in the similar size range as NZVI but appeared more oxidized as discussed in a later section. Phase Transfer and Characterization of Phase Components. The phase transfer of RL-Pd-NZVI was facilitated using butanol and NaCl as follows. An aqueous suspension of RL-Pd-NZVI was sonicated in the presence of TCE NAPL and butanol (Figure 2a) which resulted in the formation of stable oil in water microemulsions of TCE (Figure 2b). Subsequent addition of NaCl decreased the ionic interactions of RL with water due to charge screening (reduction of Debye length). This resulted in the Winsor I− III−II transition19−21 which led to the formation of a continuous organic phase consisting of water in oil microemulsion and an excess aqueous phase. RL was bound to PdNZVI through carboxylate functional groups22 and caused the simultaneous transfer of the nanoparticles into the organic

phase upon NaCl addition (Figure 2c). The phase transferred RL-Pd-NZVI was extremely stable in the organic phase, and destabilization of the nanoparticles has not been observed (even after completion of degradation reactions) in systems kept quiescent for over 1.5 years. Butanol was used to facilitate the phase transfer based on its history of implementation for TCE remediation by TCE mobilization or by bioremediation.23,24 Butanol can be delivered in the subsurface effectively25 and has advantages of being miscible with TCE NAPL and having low solubility in water. Emulsification of TCE NAPL can be achieved in the subsurface by the shear forces generated by fluid flow through porous media.25−27 The composition of the aqueous and organic phases after the phase transfer process was characterized and is shown in Figure 3a. Ethanol was added during the Pd-NZVI synthesis procedure and therefore was also incorporated in the component analysis. As shown in Figure 3a, the organic phase was primarily made up of butanol (79% by wt.) and consisted of 5% TCE by weight. Some fraction of NaCl was also present within the organic phase. A weight basis breakup of the components is provided in Table S1. On a weight basis, 96% of the initial TCE mass added to the vial migrated into the organic phase (Figure S2). RL was assumed to be completely in the organic phase due D

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Figure 4. (a) Schematic of SYSTEM A and SYSTEM B used to assess TCE degradation. In SYSTEM A, RL-Pd-NZVI reacts with TCE in the aqueous phase, while in SYSTEM B, RL-Pd-NZVI reacts with TCE in organic phase. (b) TCE degradation profile with time for SYSTEM A and SYSTEM B fitted with pseudo-first-order rate law. Error bars represent standard deviation from measurements of triplicate systems. Inset equations represent pseudo-first-order TCE degradation rate constants for SYSTEM A (kobs‑A) and for SYSTEM B (kobs‑B).

TCE + x·e− + y·H+ → products + z·Cl−

to its role in emulsification. Additionally, RL (due to its dark brown color) was visually observed to accumulate in the organic phase. The organic phase was also imaged under a microscope. In Figure 3c (image taken under 10× magnification), the formation of emulsion droplets with an average size of 2 μm can be clearly seen. All droplets have dark edges due to the presence of Pd-NZVI. A 60× magnification allowed us to observe the location of RL-Pd-NZVI (Figure 3d). As seen in the figure, the droplet is surrounded by dark particles which can be attributed to aggregated RL-Pd-NZVI particles. TCE Degradation in the Organic Phase. The key feature that must be preserved when RL-Pd-NZVI is transferred into the organic phase is its reactive functionality for TCE NAPL degradation. To this end, we evaluated the TCE degradation rate and extent by RL-Pd-NZVI in the aqueous phase and compared it with the phase transferred system. The two batch reactors are schematically shown in Figure 4a and are explained below: 1) SYSTEM A (aqueous phase reactions): SYSTEM A consists of an aqueous suspension of RL-Pd-NZVI saturated with dissolved TCE (8.4 mM TCE), in equilibrium with a TCE NAPL. This represents a system where Pd-NZVI reactions occur solely in the aqueous phase. (2) SYSTEM B (organic phase reactions): SYSTEM B consists of a TCE NAPL which is completely dissolved in the organic phase into which the RL-Pd-NZVI is phase transferred. In SYSTEM A and SYSTEM B, rapid/instantaneous equilibration of TCE between the various phases occurred and RL-Pd-NZVI had access to stoichiometrically excess amounts of TCE throughout the duration of the degradation experiments (Figure S3). Therefore, the TCE degradation rate in both systems can be attributed solely to the reaction kinetics at the RL-Pd-NZVI surface rather than any mass transfer limitations due to slow dissolution rates from the NAPL. Degradation of TCE by RL-Pd-NZVI can be represented using the following reactions:1 Fe0 → Fe2 + + 2e−

(1)

Fe 2 + → Fe3 + + e−

(2)

(3)

Eq 1 shows that Fe0 oxidation yields 2 electrons resulting in generation of Fe2+ species which may further oxidize to Fe3+ yielding an additional electron. In eq 3, x represents the stoichiometric amount of electrons required for TCE dechlorination. The profiles for TCE degradation by RL-Pd-NZVI in the aqueous phase (SYSTEM A) and in the organic phase (SYSTEM B) are shown in Figure 4b along with the reaction rate constants. The TCE degradation profile was best fitted with a pseudofirst-order rate equation. The integrated rate equation is shown in eq 4 M t = Me + (M 0 − Me)e−kobst

(4)

where Mt is the mass of TCE in the reactor at any time t, M0 is the initial mass of TCE in the reactor, Me is the mass of TCE in the reactor at the end of the degradation reaction, and kobs is the observed pseudo-first-order TCE degradation rate constant (day−1). This study employed stoichiometrically limited amounts of Pd-NZVI compared to TCE, and the plateaus observed for SYSTEM A and SYSTEM B in the amount of TCE degraded (Figure 4b) are due to the attainment of stoichiometric end points by Pd-NZVI rather than an incomplete degradation (explained below). Therefore, eq 4 incorporates the stoichiometric end points in the form of Me in the rate calculations. As is evident from Figure 4b, the TCE degradation rate and extent were significantly improved in the organic phase reaction (SYSTEM B) as compared to the aqueous phase reaction (SYSTEM A) with a 4-fold faster dechlorination rate and degradation of nearly 50% higher TCE mass (232 ± 5 μmoles in SYSTEM B compared to 156 ± 4 μmoles in SYSTEM A). The RL-Pd-NZVI particles employed in both systems were identical. Previously we reported that the presence of free RL in solution can affect the TCE degradation rate and extent by coating the exposed Pd sites.22 However, in this study, as mentioned earlier, no free RL existed in solution which could potentially affect the TCE degradation rates or extents. E

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Figure 5. TEM images of (a) unreacted RL-Pd-NZVI, (b) RL-Pd-NZVI in SYSTEM A after reaction with TCE, and (c) RL-Pd-NZVI in SYSTEM B after reaction with TCE.

Figure 6. Fe 2p3/2 XPS spectra of (a) unreacted RL-Pd-NZVI, (b) RL-Pd-NZVI in SYSTEM A after reaction with TCE, and (c) RL-Pd-NZVI in SYSTEM B after reaction with TCE.

Cl (Table S2). High resolution scan for the Fe 2p3/2 XPS spectra for unreacted RL-Pd-NZVI particles is shown in Figure 6a. The deconvoluted peaks at 711.3 and 714.4 eV revealed that the surface of unreacted RL-Pd-NZVI primarily consisted of Fe(III) oxides and hydroxides (Fe2O3 and FeOOH)30,31 with a contribution of zerovalent iron (Fe0) seen at 707.2 eV. This is in agreement with the typical structure of NZVI consisting of a Fe0 core and a shell of iron oxides and hydroxides.18 After undergoing reaction with TCE in the organic phase, particle surfaces in SYSTEM B (Figure 6c) did not show a considerable difference in the oxidation states except the disappearance of Fe0. However, for nanoparticles exposed to TCE in the aqueous phase (SYSTEM A), a new peak corresponding to Fe(II) oxide (FeO) appeared around 709.6 eV30(Figure 6b). The relative intensity of Fe3+ peaks was also lower. Overall this suggests that iron nanoparticles in aqueous phase reactions (SYSTEM A) primarily transformed into oxides in the +2 oxidation state during reaction with TCE, while those in organic phase reactions (SYSTEM B) oxidized to the +3 state only. This was in qualitative agreement with the different morphological characteristics of iron oxides seen in TEM images (Figure 5b and 5c). The implications of these differences in the surface chemistry of Pd-NZVI nanoparticles in the aqueous and organic phases are observed most clearly when comparing the extents of degradation achieved in those phases. TCE degraded in the organic phase was 232 μmoles, while that in the aqueous phase was 156 μmoles (Figure 4b). Given that Pd-NZVI was stoichiometrically limited relative to TCE, the difference in the amounts degraded can only be explained by differences in the number of electrons available. The higher moles of TCE degraded per mole of Fe0 in the organic phase are afforded by

Thus, in order to probe the differences arising in the reactivity characteristics, we recovered the nanoparticles from SYSTEM A and SYSTEM B at the end of their reactive lifetime and conducted TEM and XPS analysis to gain insight into the changes brought about in the particle morphology and surface chemistry. Figure 5a shows that before reaction, RL-Pd-NZVI consisted of spherical particles between 20 and 100 nm arranged as chains with Pd deposits on the surface (Figure S4). The nanoparticles appear partially oxidized (confirmed by EDS analyses shown in Figure S5) which can be attributed to the reaction of NZVI with water during the overnight mixing process with RL. After having undergone reaction with TCE in the aqueous phase, RL-Pd-NZVI in SYSTEM A does not appear as distinct particles but rather as small needlelike clusters (Figure 5b). EDS of those particles (Figure S6) suggest the formation of iron oxides, which as seen in the TEM image are packed closely together. In contrast, RL-Pd-NZVI particles extracted after reaction with TCE in the organic phase in SYSTEM B (Figure 5c) have a coarse structure, and a hollowed out core can be observed for some particles. EDS analysis of the hollowed out particles showed a strong peak for Pd (Figure S7). This is consistent with previous reports where reaction of Pd-NZVI with water resulted in the outward diffusion of Fe ions creating a hollowed out structure, while Pd0 migrated progressively inward.28 Acicular particles are also observed in Figure 5c which are typically the morphology of lepidocrocite.29 To further identify the chemical identities of the oxides formed, XPS analysis was carried out on the particles. Low resolution survey scans for unreacted RL-Pd-NZVI showed the presence of Fe, O, and Pd, while those for RL-PdNZVI in SYSTEMS A and B yielded peaks for Fe, O, Pd, and F

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Environmental Science & Technology Table 1. Electron Balance in SYSTEMS A and B SYSTEM

initial μmoles of unreacted Fe0a

A

572

B

initial electrons available

total electrons used for TCE degradationb

μmoles of Fe0 remaining at end of degradation reactionc

electron balance

987 (±26)

5

87 ± 2%

1563 (±33)

3

91 ± 2%

1144 (2e− per μmole Fe0) 1716 (3e− per μmole Fe0)

a Calculated from hydrogen liberated after acid digestion of unreacted Pd-NZVI. bCalculated from stoichiometry (refer to Tables S3 and S4 in the SI). cCalculated from hydrogen liberated after acid digestion of reacted Pd-NZVI.

caused oxide growth close to the nanoparticle surface (in effect progressively burying the Pd0 sites underneath), while in SYSTEM B a change in RL configuration resulted in Fe(III) oxides predominantly precipitating away from the nanoparticle surface as seen in Figure 5c. Pd0 sites on the NZVI surface are the reactive sites for TCE degradation, and therefore it can be envisioned that Fe(II) oxides in SYSTEM A retarded the TCE access to the Pd0 sites thereby causing slower degradation rates and likely lower degradation extents by inhibiting the conversion of Fe(II) to Fe(III). We observed a pseudo-firstorder degradation rate constant kobs of 0.099 day−1 for RL-PdNZVI in the aqueous phase reactions (SYSTEM A), while in the organic phase reactions (SYSTEM B) the RL-Pd-NZVI exhibited a higher kobs of 0.413 day−1 due to the relatively easier access of TCE to Pd0. The passivation of Pd-NZVI in aqueous phase reactions has not been reported in earlier studies with adsorbed polyelectrolytes or surfactant layers on Pd-NZVI,22,37,38 most likely because it was carried out under conditions where Pd-NZVI was stoichiometrically in excess of TCE. Under stoichiometrically excess conditions of Pd-NZVI, there would be an adequate supply of electrons to degrade TCE rapidly, and therefore the extent of oxide growth near the nanoparticle surfaces caused by adsorbed polyelectrolytes or surfactants may not be sufficient enough to cause a passivation effect and adversely affect the degradation rate. Degradation End Products. Generation of toxic end products is undesirable in the environmental remediation of TCE. Therefore, a detailed end product characterization for degradation of TCE in the organic phase and in the aqueous phase was carried out. Figure 7 shows that the major end products (80%) in both systems at the end of the reactive lifetime are nontoxic ethene and ethane. Butenes, which are usually the coupling products of

the release of 3 electrons from the nanoparticles compared to the release of 2 electrons in the aqueous phase. To verify whether the lower extent of TCE degradation by RL-Pd-NZVI in the aqueous phase reactions (SYSTEM A) was due to lack of Fe0 oxidation, the particles were extracted and acid digested after the TCE degradation reaction. The liberated H2 was used to estimate the moles of Fe0 remaining.1 As seen in Table 1, less than 1% of the initial Fe0 remained unused in both systems. Moreover, an electron balance using a 2e− conversion scheme for SYSTEM A (eq 1) and a 3e− conversion scheme for SYSTEM B (eqs 1 and 2) yielded approximately 90% balance. Studies investigating the degradation of chlorinated contaminants by green rust have previously reported the rapid degradation of carbon tetrachloride and tetrachloroethylene facilitated by the conversion of Fe(II) oxides to Fe(III) in the presence of zerovalent noble metals such as Ag0 and Pt0.32−34 The differences observed in the iron oxidation may be a result of a combination of the following factors. In the aqueous phase in SYSTEM A, RL-Pd-NZVI was in contact with significantly larger amounts of water (1.38 mol) compared to RL-Pd-NZVI in the organic phase in SYSTEM B (0.013 mol). Although hydrogen evolution could not be detected in SYSTEM A and SYSTEM B during the TCE degradation experiments, separate experiments conducted with 1,2-DCA (a compound resistant to degradation by Pd-NZVI) revealed that higher amounts of hydrogen were evolved in SYSTEM A (aqueous phase reactions) compared to SYSTEM B (organic phase reactions). Further details of the experiments are provided with Figure S8, in the SI. These results suggest that degradation of TCE via the atomic hydrogen species generated at the Pd site (H2 → 2H•) through the reduction of water (Fe0 + 2H+ → Fe2+ + H2) in the aqueous phase (SYSTEM A) was more dominant than in SYSTEM B and may have led to a preferential generation of more Fe(II) oxides. Another contributing factor could be related to the adsorbed layer configuration of RL on Pd-NZVI. Greenlee et al.35 observed that surface modifiers adsorbed on NZVI can significantly influence the oxidation mechanisms and the growth of the iron oxide-hydroxide phase. Tratnyek et al.36 conducted aging experiments of NZVI in water, with and without the presence of natural organic matter (NOM), and observed that the formation of green rust precipitates decreased in the presence of NOM and also noted a slower rate of Fe0 content decrease. They attributed these effects to the inhibition of reaction between NZVI and H2O due to the coating of NZVI with unreactive organic matter. It is likely that the configuration of the adsorbed RL layer in the aqueous phase (SYSTEM A) and organic phase (SYSTEM B) was dissimilar due to the butanol-water interface in the latter system, which affected the oxidation processes differently and resulted in the growth of different oxides. In SYSTEM A the closely arranged clusters of Fe(II) oxides seen in Figure 5b could be a result of a more compact configuration of RL on the NZVI surface which

Figure 7. End products generated at the end of TCE degradation in SYSTEM A and SYSTEM B. G

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Environmental Science & Technology acetylene and ethene,39 make up 10% of the remaining minor products. Other minor products are provided in Table S5 and were less than 1% of the total degradation products generated. Certain differences are however observed in SYSTEM B compared to SYSTEM A. In SYSTEM B we observe that ethane and dichloroethenes (1,1-DCE and cis-1,2-DCE) constitute higher amounts of the end products compared to SYSTEM A. For instance the formation of slightly higher amounts of DCEs in SYSTEM B (9%) is observed compared to SYSTEM A (5%). The differences arising in the end products could be due to a shift in the dechlorination pathway of certain reaction intermediates within the organic phase. As discussed above and shown in Figure S8, TCE degradation via reactions with H• were more likely in SYSTEM A. Further studies are required to gain an exact mechanistic understanding of these processes.

Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS The research was funded by the Natural Sciences and Engineering Research Council of Canada (Grant Nos. 203158-11 and 413978-12). S.B. was supported by McGill Engineering Doctoral Awards. We thank Josianne Lefebvre (Polytechnique Montreal) and David Liu (McGill) for assistance with XPS and TEM-EDS analyses, respectively.

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IMPLICATIONS Direct elimination of the TCE as NAPL can significantly speed up the remediation of contaminated aquifers. Biphasic treatment systems which can emplace reactive NZVI from the aqueous to the nonaqueous phase may serve as an effective strategy to achieve this goal. Results from batch TCE degradation studies demonstrated that the phase transfer of RL-Pd-NZVI into the nonaqueous organic phase resulted in the degradation of 50% more TCE NAPL per unit mass of Fe0 (232 μmoles TCE degraded in organic phase against 156 μmoles in aqueous phase for the same 572 μmoles of Fe0, i.e. 0.4 mol TCE degraded/mol Fe0 in organic phase against 0.27 mol TCE degraded/mol Fe0 in aqueous phase) at a 4-fold faster rate along with generation of nontoxic products. This was afforded by the rapid release of 3 electrons per mole Fe0 in the organic phase (facilitated by Pd0 and by iron oxide precipitation away from the Pd-NZVI surface) compared to 2 electrons per mole Fe0 in the aqueous phase. However, further studies are needed to assess the phase transfer performance in saturated porous media and to investigate the feasibility of scale-up of such systems. This would include optimizing subsurface delivery methods of the various constituents to achieve efficient phase transfer and addressing concerns with the mobilization of TCE NAPL as well as evaluating the reusability and recovery of the salt injected to address concerns arising from the possibility of increased groundwater salinity. However, the implementation of the phase transfer systems should be easily achieved for treatment of waste TCE NAPL in reactors.

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.6b01646. Methods for GC, TEM, and XPS analysis; schematic and description of phase characterization; EDS of TEM images presented in Figure 5; relative atomic abundances of oxides identified for RL-Pd-NZVI before and after reaction; and a quantified list of end products generated at end of TCE degradation (PDF)

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DOI: 10.1021/acs.est.6b01646 Environ. Sci. Technol. XXXX, XXX, XXX−XXX

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DOI: 10.1021/acs.est.6b01646 Environ. Sci. Technol. XXXX, XXX, XXX−XXX