Activation of Peroxydisulfate by Ferrite Materials for Phenol Degradation

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Activation of Peroxydisulfate by Ferrite Materials for Phenol Degradation Yue Li, Roya Baghi, Jan Filip, Syful Islam, Louisa Jane Hope-Weeks, and Weile Yan ACS Sustainable Chem. Eng., Just Accepted Manuscript • DOI: 10.1021/ acssuschemeng.8b05257 • Publication Date (Web): 29 Mar 2019 Downloaded from http://pubs.acs.org on March 30, 2019

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ACS Sustainable Chemistry & Engineering

Activation of Peroxydisulfate by Ferrite Materials for Phenol Degradation

Yue Li1, Roya Baghi2, Jan Filip3, Syful Islam1, Louisa Hope-weeks2, Weile Yan1 * 1Department

of Civil, Environmental and Construction Engineering, Texas Tech University,

Texas, United States 2Department

3Regional

of Chemistry and Biochemistry, Texas Tech University, Texas, United States

Centre of Advanced Technologies and Material, Palacký University, Olomouc, Czech

Republic

*Corresponding author. Tel: (806) 834 3478; Fax: (806) 742-3449 Email address: [email protected]

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Abstract Persulfate such as peroxydisulfate (PDS) is among the most widely applied oxidants for breaking down organic contaminants in water. The oxidation power arises from conversion of persulfate to sulfate radical or other reactive oxidants. Ferrite materials are good candidates for catalytic activation of persulfate owing to its ability to incorporate a variety of transition metals in the structure, stability against aqueous dissolution, and magnetic susceptibility allowing catalyst separation and reuse. In this study, ferrite spinels incorporating zinc, nickel, cobalt, or copper were synthesized with an epoxide-driven sol-gel method and were annealed at 350 °C and 700 °C, respectively. The particles were evaluated for activating PDS using phenol as a model organic contaminant. Cu-ferrite annealed at the low temperature (350 oC) was identified to be the most active ferrite for PDS activation. This solid consists of predominantly CuFe2O4, while at the higher annealing temperature, decomposition of CuFe2O4 to Fe2O3 and CuO and significant increase in particle size resulted in severe loss of PDS activation ability. Remarkable increases in phenol oxidation rate was observed above pH 9.0 and were attributed to PDS activation by phenoxide. The presence of methanol, bicarbonate or chloride ion (1 – 5 mM) significantly slowed down phenol oxidation, whereas the addition of tert-butyl alcohol did not affect the degradation rate, indicating the dominant oxidant is sulfate radical. Comparison of Cu-ferrite against reference metal oxides suggests that the catalytic performance of Cu(II) sites in the ferrite phase is comparable to those in the highly active but leachable CuO, and Cu-ferrite demonstrated good reusability during repeated phenol oxidation experiments. Keywords: persulfate, peroxydisulfate, ferrite, Cu-ferrite, CuFe2O4, phenol, trichloroethene, TCE


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Introduction Persulfate anions, including peroxymonosulfate (PMS, SO52-) and peroxydisulfate (PDS, S2O82-), have gained increasing interest in research and applications as oxidants for degrading organic contaminants in remediation systems such as in situ chemical oxidation (ISCO).1-5 Compared to PMS, PDS is more widely used in field deployments due to its price competitiveness and its higher aqueous stability (half-life in water at 25 oC on the order of 100 days).3, 6 Although PDS has a relatively high oxidation potential (E° = 2.0 V),7 direct oxidation of contaminants by PDS is slow, and the oxidation efficiency is greatly enhanced when catalytic activation of persulfate releases sulfate radical (SO4•-, E° = 2.5 – 3.1 V depending on pH 8) or other reactive oxidants.9 Comparison is often made between persulfate and hydrogen peroxide (H2O2)-based ISCO processes. Hydroxyl radical (HO•), produced via peroxide activation by base or transition metals10, is a highly reactive (E° = 1.8 – 2.7 V depending on pH) but non-selective oxidant.11-12 HO• oxidizes organic molecules by hydrogen abstraction or addition, while SO4•- is more prone to direct electron transfer reactions,13-14 making SO4•- more reactive towards aromatic molecules with electron-donating substituents.12, 15-16 In addition, the reactions between peroxide and aquifer minerals trigger decomposition of H2O2 to water and oxygen,17-18 whereas persulfate is considerably more stable in groundwater matrices.3, 19 In environmental applications, these properties translate to desirable outcomes including greater subsurface mobility, longer contact time between the oxidizing agent and target containments, and selectivity for certain contaminants among background constituents in the underground environment. As direct reactions between PDS and contaminants do not proceed efficiently at an ambient temperature, a variety of methods for PDS activation including the use of heat,14, 20 alkaline chemicals,21 carbon-based substrates (including soil organic matter),8, 22-24 or metal 3 ACS Paragon Plus Environment


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catalysts 9, 25 have been documented. In alkaline water, SO4•- reacts with OH- to form HO•,13 and the two radicals have been observed as the main reactive oxidants in most prior studies, although non-radical oxidation pathways involving surface-mediated electron transfer between persulfate and the contaminants have been proposed in recent studies as well.6, 22 Among metal-based activators, iron oxides (e.g., magnetite) have been evaluated previously because of their natural abundance in the aquifer media.26-27 The catalytic cycle of iron, however, relies on efficient reduction of Fe(III) to Fe(II) by persulfate, which is considerably slow in comparison to other steps in the radical chain reactions.3 Other transition metals such as copper(II)28 and silver(I)9 were found to be efficient in catalyzing the decomposition of PDS. Several studies have observed CuO to possess superior activity for PDS and PMS activation and the metal is considered more environmentally benign than other active metals such as cobalt.6, 29 However, significant leaching of Cu ion from CuO into the aqueous phase precludes the particles from repeated use in practical applications. Ferrites belong to a class of spinels and are oxides of divalent metals and ferric iron (Fe3+) in the cubic crystal system with its nominal formula being MeFe2O4 and commonly referred to as Me-ferrite, where Me denotes a divalent metal.30 The most common ferrite is magnetite (Fe3O4), but the ferrous ion can be replaced by other divalent metals to form various mixed metal ferrites.30 A handful of synthesis methods have been documented for the making of ferrite materials for environmental applications, including co-precipitation,31-32 sol-gel,33-35 and hydrothermal36 methods. Careful control of the hydrolysis and calcination conditions during the synthesis procedures can give rise to ferrite materials with well controlled particle size and high surface areas. The catalytic function of ferrites can be optimized by incorporating different transition metals into the spinel structure. Meanwhile, the magnetic properties of ferrites provide 4 ACS Paragon Plus Environment

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the benefit of facile separation of the catalysts to minimize effluent treatment and enable catalyst reuse. Recent works on Cu-ferrite (i.e., CuFe2O4) suggest good efficiency for the activation of PMS, and the solid catalysts are amenable to magnetic separation with very low copper leaching.35, 37-38 Fewer studies have examined PDS activation by ferrite, and the results were inconsistent. Guan et al., reported no apparent activity with Cu-ferrite towards atrazine,35 while other studies observed efficient degradation of contaminants with carbon nanotube-supported or standalone Cu-ferrite.39-40 These findings suggest catalyst performance is strongly influenced by material synthesis parameters and the aqueous reaction conditions. In the present study, a series of spinel ferrites were produced using an epoxide-enabled sol-gel method with divalent metals including Zn(II), Ni(II), Co(II), or Cu(II). Different from previous studies where organic ligands such as citrate were used in the sol-gel process to stabilize colloidal complexes, which necessitates ligand removal through a thermal gasification step,35, 37 the epoxide addition method permits mild and controllable quenching of protons from hydrated metal ions, forming gels of uniformly sized colloids connected in a continuous network.41 Following the gel formation, supercritical solvent extraction retains the interconnected architecture and forms aerogels with high porosity and specific surface area. Annealing the gel at different temperatures allows us to specifically examine the effect of thermal treatment on the catalytic behavior, as it has been reported that spinel has a tendency to partially decompose and the process is temperature and metal-dependent.30 The objectives of this study were to assess the efficiency of ferrites produced via epoxide-addition method for the catalytic oxidation of phenol by PDS, to examine the effects of water chemistry parameters, and to identify the active sites on the catalyst and the reactive oxidants involved in phenol oxidation using aqueous batch experiments and material characterization approaches. 5 ACS Paragon Plus Environment


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Experimental Section Synthesis of Ferrites The list of materials and reagents used is given in the Supporting Information. Synthesis of the ferrites was performed under ambient conditions following a sol-gel process reported previously.30 Ferrite in which Fe(II) was replaced with divalent salt of zinc, cobalt, nickel, or copper was denoted as Zn-, Co-, Ni-, or Cu-ferrite, respectively. For all ferrites, 0.33 mmol FeCl3.6H2O was added to a silanized glass vial and hydrated with 0.5 mmole of deionized water. 0.17 mmol of the relevant divalent metal chloride salt was added to FeCl3 solution above, followed by the addition of 4.7 mmol of propylene oxide to the mixed solution of divalent metal and ferric ions. After vortexing for 15 seconds, the gels were set aside and allowed to gel undisturbed. The copper-based gel was synthesized based on the method reported previously.41 Copper bromide was employed in lieu of the chloride salt due to slow and incomplete sol-gel transition with the chloride precursor.41 3.3 mmol of copper bromide was dissolved in 5 ml dimethylformamide to make a dark green solution. Following that, 13 mmol deionized water and 30 mmol of epichlorohydrin were added to the copper solution, which resulted in the formation of Cu-based gel. As a reference material, iron oxide (Fe2O3) was prepared using a similar sol-gel process by adding 17 mmol of propylene oxide to a solution of 1.6 mmol of FeCl3 in 3 mL ethanol.42 All gels (ferrites, Cu or Fe oxides) were aged in closed vials without disturbance for one day, immersed in acetone bath for a week with acetone exchanged daily, and then transferred to a SPIDRY critical point dryer to change out the solvent in the gel with liquid CO2 at 11°C and ca. 850 psi over a period of 3-4 days. Once the solvent was removed, the gels were brought beyond the critical point of CO2 at 38°C and 1100 psi before the drying chamber was depressurized to ambient pressure slowly. The as-prepared aerogels were annealed at 350°C or 6 ACS Paragon Plus Environment

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700°C for 6 h at a temperature ramping rate of 1 °C/min. Details on material characterization methods are available in the SI. Batch experiments The reactivity of ferrites for PDS activation was examined through batch experiments in 40-mL borosilicate glass vials at ambient temperature (20 ± 1 °C). All aqueous solutions used in the experiments were prepared with distilled-deionized water (> 18 MΩ•cm, DDI). Ferrite (0.75 g/L) was added to un-buffered phenol solutions at a concentration ranging from 0.11 to 0.21 mM. The final pH of these unbuffered experiments was in the range of 4.3 to 5.9 (Table S1). When investigating the effect of pH, diluted HCl or NaOH was used to adjust the pH of the suspension containing phenol and ferrites to pH 4.5-9.5 before adding PDS. For experiments investigating the effect of background electrolytes, a similar procedure was followed except that bicarbonate or chloride stock solution was added prior to pH adjustment. An experiment was started by the addition of a small aliquot of fresh PDS stock solution to the reaction solution, resulting in 1 mM PDS in the reactor. Periodically, 2-ml samples were taken at predetermined intervals and were filtered through 0.2-μm syringe filters (Millipore SLLG025SS). Samples were immediately spiked with concentrated sodium nitrate solution and methanol to reach a concentration of 10 mM and 470 mM, respectively, to quench sulfate radical,43 hydroxyl radical, and other reactive oxidants.44 The effect of was further assessed using trichloroethene (TCE) as a probe contaminant. The reactor used in TCE degradation experiments was identical to that of phenol experiments except that the vial was capped with a PTFE-lined MininertTM valve. pH adjustment was performed after the addition of catalyst and PDS. The vial was immediately capped and spiked with TCE stock solution (saturated TCE solution in DDI) to start the reaction. Periodically, 50 μL of head space gas (10 mL headspace was maintained above 30 mL of 7 ACS Paragon Plus Environment


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aqueous solution) was sampled using a gas-tight syringe and was analyzed immediately using gas chromatography (GC). Analytical methods Concentration of phenol was determined by a high performance liquid chromatography instrument (HPLC, Agilent 1100) equipped with a GP-C18 column (Sepax Technologies) and a UV-Vis DAD detector. Mixture of methanol and DDI at a volume ratio of 40:60 was used as the eluent at 1 mL/min. Phenol quantitation was made based on UV absorbance at 270 nm with reference to a 5-point calibration curve. The analysis of PDS was performed following the spectrophotometric method reported in the literature.45 Specifically, 1 mL of sample was amended with 1.7-mL color development solution (1 M potassium iodide in 100 mM sodium bicarbonate) fresh prepared before sample analysis. The mixture was manually shaken vigorously for 1 minute and allowed to sit for 15 minutes before spectrophotometric measurement of the absorbance at 400 nm. TCE analysis was performed using a GC system reported in our earlier studies.46-47

Results and Discussion Activity of Ferrite Series Four types of ferrite were prepared at an annealing temperature of 350 oC and assessed for their reactivity in activating PDS to oxidize phenol. As shown in Figure 1a, the control experiments containing only phenol or ferrites indicate that there was negligible adsorption of phenol onto ferrite surface and direct oxidation of phenol by ferrite did not occur within the experiment time frame. Furthermore, PDS alone was unable to transform phenol. Among the four substituted 8 ACS Paragon Plus Environment

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ferrites, Zn-ferrite did not lead to noticeable phenol oxidation. Other ferrites (Co-, Ni-, and Cuferrites) brought about continued loss of phenol in the presence of PDS. To confirm that the activation of PDS was enabled by ferrite solids as opposed to metals leached into the solution, a parallel experiment was conducted in which Cu-ferrite was filtered out at 20 min and the remaining aqueous phase was allowed to continue mixing. Insignificant phenol degradation after Cu-ferrite removal (Figure 1b) demonstrates convincingly that the activation of PDS was mediated by the Cu-ferrite particles, not dissolved metal ions, which is consistent with earlier reports.35, 37 Among the three types of ferrites, Cu-ferrite demonstrates the highest activity in PDS activation, followed by Ni-ferrite and Co-ferrite in order of decreasing activity. For this reason, Cu-ferrite was used for further investigations. Effect of Annealing Temperature Cu-ferrites annealed at 350 °C and 700 °C, respectively, were evaluated for phenol degradation. The results indicate that there is a vast difference in the activity of Cu-ferrite annealed at the two temperatures (Figure 1b). Cu-ferrite annealed at 350 °C gave rise to over 95% of phenol oxidation within 90 min, while the solids annealed at 700 °C was ineffective in eliciting the oxidation process. To understand the material properties of Cu-ferrites produced at the two annealing temperatures, the solids were characterized by TEM, XRD analysis, and Mossbauer spectroscopy. The material annealed at 350 °C appears as uniformly sized spheroidal particles of approximately 5-10 nm (Figure. 2a). In contrast, Cu-ferrite annealed at 700 °C displayed more well-defined crystal facets and the crystallites span a large range of sizes from 50 nm to several hundreds of nanometers (Figure 2b). Sintering of particles to form clusters of particles can be discerned as well. The XRD spectrum of the Cu-ferrite aerogel annealed at 350 °C, as depicted in Figure 2c, 9 ACS Paragon Plus Environment


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indicates the presence of CuFe2O4, CuO and hydrated copper chloride (Cu7Cl4(OH)10•H2O). Copper chloride was from the synthesis precursors (ferric chloride and cupric bromide) and its persistence reflects relatively slow hydration of Cu-chloride complexes.41 The diffraction peaks of the low temperature annealed Cu-ferrite are broad, which points to the nanocrystalline nature of the solid formation. The same set of mineral phases are present in the aerogel annealed at 700 °C,

however, much sharper peaks observed with the higher temperature solids are indicative of

long-range ordering. This observation agrees with the substantial increase in particle size and crystal facets noted in the TEM images. While CuO constitutes a minor phase (12 wt.%) in solids annealed at 350 oC, it accounts for a significant portion (31 wt.%) of the crystalline material at 700 oC based on XRD spectra fitting with component phases. Additionally, hematite (α- Fe2O3) was clearly present in the 700 oC ferrite sample. The emergence of hematite and CuO as major phases implies that CuFe2O4 had decomposed to the respective single metal oxides during high temperature annealing. Moreover, increasing the annealing temperature from 350 oC to 700 oC caused a sharp decrease in the specific surface area of Cu-ferrite from 170 m2/g to 15 m2/g and a loss of over 95% of the pore volume (Table S2). These changes are in accordance with the morphologies observed of the two solids under TEM. The significant loss of the activity of ferrite aerogel upon high temperature annealing is therefore attributed to the decrease in solid surface area and the occurrence of phase separation during the high temperature treatment. In the XRD analysis, phase separation and formation of Fe2O3 and CuO were clearly noted in solids annealed at the higher temperature. Whether individual metal oxides exist in the material prepared with low-temperature annealing cannot be confirmed with XRD data alone due to their poorly crystalline character. In light of this, scanning-transmission-electron-microscopy with energy-dispersive X-ray spectroscopy (STEM-XEDS) images were acquired for the Cu-ferrite 10 ACS Paragon Plus Environment

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annealed at 350 oC. STEM-XEDS analysis permits mapping of individual elements at a sub-nm resolution, thereby revealing the spatial distribution of Cu and Fe and distinguishing CuFe2O4 from Fe2O3 or CuO. In Figure 3, the high angle annular dark field (HAADF) image of a cluster of Cu-ferrite was first obtained (Figure 3a), followed by a raster scan by a finely focused electron beam across the image area, generating an X-ray spectrum carrying chemical compositional information at each pixel.48 This leads to the construction of elemental maps as shown in Figure 3 b-d. The spatial extents and the relative brightness of O, Cu, and Fe maps correspond with each other very well, suggesting the three elements are co-locating. The analysis was performed over six random locations. Other than co-localized Cu and Cl hot spots representing unconsumed copper chloride residues as detected by XRD (Figure S1), no distinctive Fe or Cu-rich regions was identified. Collectively, the XRD and STEM-XEDS analysis results confirm the dominant presence of Cu-ferrite in the solids annealed at a lower temperature. Additionally, Mӧssbauer spectroscopic characterization was employed to analyze the presence of Fe(II) in the fresh prepared Cu-ferrite, as Fe(II) reacts rapidly with PDS giving rise to sulfate radical.49 The room temperature Mӧssbauer spectrum of Cu-ferrite is given in Figure 2e. The relatively small shift in spectral center line ( 9 is activated by phenoxide or hydroperoxide ions, Figure 4b overlays the experimentally measured phenol oxidation rate constant onto the speciation change of phenol and hydrogen peroxide as a function of pH. The fairly good correspondence between the phenol removal rate and the onset 13 ACS Paragon Plus Environment


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of phenoxide formation suggests that phenoxide contributes to the activation of persulfate in the present system. To eliminate the confounding role of phenoxide activation, the pH effect was further investigated using a non-protonic aliphatic contaminant, trichloroethene (TCE). Figure 5 shows TCE oxidation in the presence of PDS and Cu-ferrite at various initial pH. Compared to phenol, the oxidation rate of TCE is considerably slower. The degradation rate shows a weak dependence on the initial pH in the range of 6 – 9.5. Liang et al. reported that PDS conversion to sulfate radical is independent of pH within a circumneutral range (5 – 9) and alkaline activation is not the dominant mechanism until at pH ~ 12.14 The TCE data is in agreement with this observation, suggesting effective PDS activation is attainable at typical groundwater pH. Effect of background ions Many studies have observed strong effects of (bi)carbonate or chloride ions on the degradation of organic contaminants by PDS. These studies have seen carbonate casting varying effect on probe contaminant oxidation, ranging from inhibitive35, 7, indifferent,53 to rate-enhancing effects29, 35, 53. This diverging behavior reflects the possible existence of different activation pathways of PDS impacted by the nature of the catalyst and solution chemistry. It is noted that the introduction of carbonate introduces alkalinity and alters the solution pH. To decouple the carbonate effect from that of pH, we adjusted pH of the solution containing the background ion before the addition of PDS so that the initial pH was consistently at ~ 5.6. The results are shown in Figure 6a. As shown, the presence of 1 mM chloride ion or bicarbonate ion adversely impacted the phenol degradation rate by suppressing the reaction rate by approximately 80%. No further decrease in phenol oxidation rate was observed when chloride or carbonate concentration was increased to 5 mM.


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Bicarbonate ion reacts with sulfate radical with a second-order rate constant in the range of 2 – 9 ×106 M-1‧s-1 at near neutral pH,43,54 while the reaction between sulfate radical and phenol is much faster at ~ 109 M-1‧s-1.55 The rate constant of bicarbonate scavenging hydroxyl radical (9 × 106 M1‧s-1)56

was also several magnitudes lower than that of phenol reacting with hydroxyl radical (~

109 - 1010) M-1‧s-1.57 Given the large difference in rate constants, the inhibitive impact of carbonate cannot be fully accounted for by the radical quenching effect. Other possible explanations may include HCO3- forming unreactive surface complexes or the complexes preferentially consuming the surface-born reactive oxidants.15 Compared to carbonate, chloride reacts more rapidly with sulfate and hydroxyl radical, with the rate constants on the order of 108 109 M-1‧s-1 in acidic to neutral pH43, 44. As such, the slowing down of phenol degradation may be mainly due to scavenging of reactive radicals. It is worth noting that the reaction rates did not further decrease with increasing concentrations of bicarbonate or chloride ions from 1 mM to 5 mM. The remnant activity may be explained by the presence of surface-bound radical or a nonradical pathway, which were not captured by the background ions. This is further investigated through radical scavenging experiments as follows. Identifying Reactive Oxidants and Catalytic Active Species In studies on activation of persulfate chemicals by earth-abundant transition metals, copper oxides or iron-copper mixed oxides were found to possess a superior catalytic ability. Two oxidation mechanisms, including radical 29, 35, 37 and non-radical pathways6 have been brought up in the literature. To investigate the reactive oxidant involved in phenol oxidation in the present system, methanol or tert-butyl alcohol (tBA) was applied as a radical quenching agent. The rate constants of HO• with methanol and tBA are on the same order of magnitude (9.7 x 108 M-1‧S-1 and 3.8 – 7.6 x 108 M-1‧S-1, respectively),25, 44 while SO4•- reacts considerably slower with tBA (4 15 ACS Paragon Plus Environment


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– 9.1 x 105 M-1‧S-1)25 than with methanol (0.2 - 2.5×107 M-1‧S-1)43. Therefore, methanol is a nonselective quencher for both hydroxyl (HO•) and sulfate (SO4•-) radicals, whereas tBA consumes hydroxyl radical preferentially.35 As depicted in Figure 6b, phenol removal was significantly impaired by introducing 400 mM methanol into the solution. However, the presence of equivalent concentration of tBA had a negligible effect on phenol oxidation, indicating minimal participation of HO• in the reaction. Therefore, the dominant oxidant in this system is sulfate radical. On the other hand, neither radical scavengers nor background ions were able to completely suppress phenol oxidation. A small but consistent degree of phenol oxidation in the presence of methanol or background ions suggests the presence of a more specific, possibly nonradical, oxidation mechanism in our system, although it is contribution is much lower than the dominant sulfate radical pathway. The consumption of persulfate during a phenol oxidation experiment was monitored (Figure S2). The data shows that PDS consumption increased with the amount of phenol degraded. The average ratio of PDS consumption over phenol degradation was approx. 4 during the course of the experiment. Compared to the more reactive but less specific hydroxyl radical (e.g., the reported oxidation utilization efficiency during phenol oxidation by Fe(II)/Fe(III) catalyzed H2O2 conversion is well below 5% at a circumneutral pH 18, 58), sulfate radical exhibits significantly higher efficiency of contaminant oxidation per mole of oxidant consumed. The nearly constant stoichiometric ratio observed throughout the experiment also implies that a high oxidant utilization efficiency can be attained even at a very low phenol concentration. These attributes are advantageous for PDS application in the field. To identify the active metal species in Cu-ferrite and to compare the catalytic performance of Cu-ferrite against individual metal oxides, we prepared mesoporous CuO and Fe2O3 aerogels 16 ACS Paragon Plus Environment

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following the same epoxide-addition sol-gel synthesis method except that only Cu(II) or Fe(III) precursor was added to make the respective oxide. The materials were annealed at 350 oC and 700 oC and subsequently tested for their ability to oxidize phenol in the presence of PDS. The results evidently show that Fe2O3 is inefficient at generating reactive oxidants (Figure S3). While CuO annealed at 350 oC catalyzes rapid PDS conversion, CuO annealed at 700 °C is barely active (Figure S3). TEM images reveal that the high temperature treatment caused CuO to undergo significant crystal growth and sintering (Figure S3), which may account for the diminished catalytic activity. For CuO annealed at 350 oC, oxidant formation is clearly mediated on the solid surface, as a parallel experiment with CuO particles removed by filtration at 10 min saw complete loss of activity. However, a direct comparison of the apparent phenol oxidation rate constants (kobs) between Cu-ferrite and CuO is not permissible due to the differences in surface area and the quantity of Cu sites on the solid surface. To correct for this effect, the surface area-normalized rate constant (kSA) and surface Cu sites-normalized rate constant (kcu) are computed as follows: (1)

(2)

Where As is the surface area of the respective solids as listed in Table 2,

is the atomic

fraction of Cu in the solid matrix as measured by STEM-XEDS (provided in Table 2), m is the mass loading of solids, and V is the reactor volume. In Eq. (2), the relative surface densities of Cu sites of Cu-ferrite and CuO were estimated using the specific surface area and the atomic percentage of Cu in each solid, respectively. The results are shown in Figure 7. Notably, the



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performance of Cu-ferrite is on par with that of CuO when the rates were compared on the basis of kCu. This finding supports the premise that the activity of the solid catalysts stems primarily from the Cu sites, and incorporating Cu(II) in a spinel structure does not compromise its activity relative to the well studied CuO catalysts. Catalyst Stability and Amenability to Reuse Cu-based catalysts have been extensively evaluated in wet oxidation processes. However, significant leaching of copper ion in acidic and neutral solutions implies that the catalyst performance would rapidly deteriorate with repeated use, and the dissolved metal ions in the effluent warrant additional treatment steps. For this reason, incorporation of Cu(II) in a ferrite structure to form a more stable heterogeneous catalyst for aqueous phase applications has been explored in several studies including this work. Previous studies have showed a high activity for PMS activation, with contaminant half-life time (t1/2) within 10 min at a catalyst loading of 0.1 g/L for a variety of contaminants.6, 35, 38 Activation of PDS by Cu-ferrite, however, has generated inconsistent findings in the literature, with studies reporting negligible oxidation of atrazine35 to efficient degradation of diethyl phthalate.39 The discrepancy may arise from a multitude of factors, but the characteristics of catalysts are believed to play a critical role. In this study, we have shown that low-temperature annealing is strongly preferred as CuFe2O4 is unstable and decomposes at an elevated temperature. Conventional sol-gel methods employ organic ligands to stabilize metal complexes, which entails thermal combustion to remove the templating agents.35, 39

In contrast, the epoxy addition sol-gel method explored here employs a mild and well

controlled hydrolysis condition that would favor the formation of ferrite over discrete ferric or cupric oxides. The sensitivity of catalysts to synthesis routes should therefore be captured in consideration in developing Cu-ferrite materials for practical applications. 18 ACS Paragon Plus Environment

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As with other ferrite materials, the catalysts synthesized here are amenable to magnetic separation for potential reuse. In this study, we harvested Cu-ferrite after oxidation reactions using a hand-held magnet and the particles were rinsed and placed in a new phenol solution to test catalyst reusability. The recycled particles attained similar rates of phenol oxidation over 3 cycles (Figure S5). Furthermore, analysis of the aqueous solution in which Cu-ferrite was immersed for up to 45 h shows that less than 3% of Cu in the solid phase was released (Table S4), suggesting good feasibility for catalyst reuse or deployment in continuous treatment processes.

Acknowledgement This study is partly supported by the funding from the U.S. National Science Foundation (CHE1308726). The authors thank Dr. Juliusz Warzywoda for assistance in BET analysis.

Supporting Information Information about experiment initial and final pH, particle porosimetric properties, additional STEM-XEDS analysis data of Cu-ferrite, and catalytic performance of recycled Cu-ferrite are available in the Supporting Information. This material is available free of charge via the Internet at http://pubs.acs.org.



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Table 1. Phenol degradation rate constants of selected experiments

Type of Solids

Specific Surface Area by B.E.T method (m2/g)

Atomic % of Cu by STEMEDX

Cu-ferrite at 350 °C

170

19 +/6.5%

CuO at 350 °C

96

62 +/6.0%

Initial pH

Apparent rate constant (min-1)

Normalized rate constant by surface Cu sites (min-1‧m-2) a

4.5

2.5×10-2

2.6×10-2

5.6

2.2×10-2

2.3×10-2

7.1

3.6×10-2

3.8×10-2

9.0

3.9×10-2

4.0×10-2

9.4

8.6×10-2

8.9×10-2

5.6

2.4×10-2

2.5×10-2

a

Calculated from apparent rate normalized by catalyst surface concentration and atomic percentage of Cu in the solid matrix.



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(a)

(b)

Figure 1. a) Degradation of phenol by PDS in the presence of various substituted ferrites. The dashed line indicates adsorption of phenol by the ferrites. b) Degradation of phenol by PDS in the presence of Cu-ferrite annealed at 350 °C and 700 °C, respectively. In the experiment with open circle (○), Cu-ferrite solids were filtered out at 20 min, and concentration of phenol in the remaining aqueous phase was continually monitored. [Phenol]o = 0.21 mM, [PDS]o = 1 mM, Me-ferrite loading = 0.75 g/L.


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Figure 2. TEM images of Cu-ferrite annealed at 350 °C (a) and at 700 °C (b). X-ray diffraction spectra of Cu-ferrite annealed at 350 °C (c) and at 700 °C (d). (e) Mossbauer spectra of Cuferrite and CuO (both annealed at 350 oC).



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Figure 3. Elemental map of Cu-ferrite annealed at 350 °C by Scanning Transmission Electron Microscopy-Energy-dispersive X-ray Spectroscopy (STEM-EDX). (a) High-angle annular dark field image, (b) – (d) O, Cu, and Fe maps of the particles in (a). Scale bar represents 20 nm.


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Figure 4. (a) Effect of pH on the degradation of phenol in PDS and Cu-ferrite system. (b) Trends of apparent phenol degradation rate constants and percentage ionization of phenol and hydrogen peroxide as a function of pH. (c) Effect of pH on the degradation of trichloroethene (TCE). [Phenol]o = 0.21 mM, [TCE]o = 0.14 mM, [PDS]o = 1 mM, Cu-ferrite loading = 0.75 g/L.



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Figure 5. Effect of pH on the degradation of trichloroethene (TCE) in PDS and Cu-ferrite system. [TCE]o = 0.14 mM, [PDS]o = 4 mM, Cu-ferrite loading = 0.2 g/L. Inset shows the apparent 1storder rate constant as a function of pH.


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Figure 6. Effect of background ions (a) and radical quenchers (b) on phenol degradation in the Cu-ferrite/PDS system. In (b), 430 mM methanol or 400 mM tert-butyl alcohol (tBA) was introduced into the solution before the addition of PDS. The pH of all experiments was approximately 5.6. [Phenol]o = 0.21 mM, [PDS]o = 1 mM, Cu-ferrite loading = 0.75 g/L.



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Figure 7. Pseudo first-order rate constants for phenol degradation in PDS/Cu-ferrite and PDS/CuO systems. Rate constants were normalized by total surface area in the systems and by surface Cu sites respectively. [Phenol]o = 0.21 mM, [PDS]o = 1 mM, and solids loading was 0.75 g/L and 0.41 g/L for Cu-ferrite and CuO, respectively. The difference is solids loading was to provide equivalent number of Cu sites of each material.


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TOC Graphic

SYNOPSIS Cu-ferrite aerogel prepared using an epoxy driven sol-gel method enabled efficient catalytic conversion of peroxydisulfate for aqueous organic contaminant oxidation.


Activation of Peroxydisulfate by Ferrite Materials for Phenol Degradation

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