Selenium Removal from Sulfate-Containing Groundwater Using

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Selenium Removal from Sulfate-Containing Groundwater Using Granular Layered Double Hydroxide Materials Man Li, Lisa M Farmen, and Candace K. Chan Ind. Eng. Chem. Res., Just Accepted Manuscript • DOI: 10.1021/acs.iecr.6b04461 • Publication Date (Web): 13 Feb 2017 Downloaded from http://pubs.acs.org on February 18, 2017

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Industrial & Engineering Chemistry Research

Selenium Removal from Sulfate-Containing Groundwater Using Granular Layered Double Hydroxide Materials Man Li,1 Lisa M. Farmen,2 and Candace K. Chan1* 1. Materials Science and Engineering, School for Engineering of Matter, Transport and Energy, Arizona State University, Tempe, AZ 85287; *[email protected] 2. Crystal Clear Technologies, Inc., 2828 SW Corbett, Suite 145, Portland, OR 97201

KEYWORDS: Selenate removal; power plant water; column test, layered double hydroxides

ACS Paragon Plus Environment

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ABSTRACT

Due to recent changes in regulatory discharge limits, selenium removal from electric utility wastewaters is becoming an important issue. In this work, Mg-Al-CO3 layered double hydroxide (LDH) in granular form was evaluated for treatment of selenium-containing groundwater in small scale column tests. The LDH material showed good capacity for removing selenate from groundwaters that contained very high levels of sulfate and total dissolved solids. Column tests were investigated using the granular LDH to treat groundwater containing trace levels of selenium < 2 ppb. Removal of sulfate using chemical pre-treatment of the groundwater resulted in about 3X higher selenium loading onto the granular LDH. The structural changes in the media after exhaustion and analysis of the effluent water from the column test were also studied to better understand the species removed by the LDH. These results show that the LDH is a promising sorbent for removing selenium from wastewaters with high levels of sulfate and background species.


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1. Introduction Although an essential trace element, selenium can enter the aquatic environment through contamination from agricultural drainage1 and industrial (e.g. oil refining, mining) wastewaters and can be toxic to humans and wildlife at concentrations as low as 2 – 5 µg Se/L2,3. The U.S. Environmental Protection Agency’s maximum allowable contaminant level for selenium in drinking water is currently 50 µg/L4, but regulatory limits for wastewaters from the steam electric power industry have been recently (as of November 2015) lowered, which may present new treatment challenges. For example, the maximum point source discharge limits for selenium in wastewaters from flue gas desulfurization (FGD) is only 12 µg/L for any one day5. However, since selenium is found as a constituent of coal at the level of several mg per kg6,7, FGD wastewaters from coal-fired power plants can contain high levels of selenium on the order of 1 – 10 mg/L8,9. Selenium at trace levels can also be an issue for non-coal-fired power plants if these seleniumcontaining waters are used as make-up waters for the cooling water. A large amount of water is required to extract heat from the power plant condenser on the low pressure side of the steam turbine and dissipate it in the wet cooling towers through evaporation10. When groundwater containing naturally occurring trace levels of contaminants is used as make-up water for the cooling water, the level of ions become concentrated in the cooling tower blowdown. As a result, the development of treatment methods for groundwaters with trace levels of selenium are also important. The various chemical, biological, and physical treatment strategies for selenium removal from water have been summarized in several recent reviews9,11. The use of physical methods such as sorption is attractive since it is a simple and low-cost process. However, sorption of selenium


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onto solid-phase extraction materials is complicated by the fact that selenium is typically found as the oxoanions biselenite (HSeO3-, Se(IV)) or selenate (SeO42-, Se(VI)) in natural waters12,13. This means that sorbents must contain positively charged surfaces in order to effectively adsorb the selenium species. Layered double hydroxides (LDHs) are attractive sorbents and ion-exchange materials for xII anion removal14,15,16. LDHs can be represented as [ (OH)2]x+[ ⁄ ] • mH2O, where M

is a divalent metal, MIII is a trivalent metal, An- is an exchangeable anion with a valence of n, and the x value is in the range of 0 – 0.3317. LDHs are anionic clays containing positively charged layers of brucite-like octahedral hydroxide sheets separated by compensating anions and water molecules that can be exchanged with other anions18,19,20,21. LDHs also display a unique structural feature known as the “memory” or “reformation” effect 22,23. Prior studies have shown that calcination of crystalline LDHs can cause decarbonation, dehydroxylation, and loss of the layered structure to form a periclase structure; subsequent exposure of the calcined LDH into water results in rehydration and recovery (reformation) of the initial layered structure from the non-layered, periclase form24 (Fig. 1). During rehydration, anions present in the water can also be intercalated or incorporated in the reconstructed LDH15, providing another mechanism for anion removal from water. LDHs have greater affinities for multi-charged anions with high charge density17,18,25, making them particularly attractive for the removal of the oxoanionic forms of chromium, arsenic, and selenium over other monovalent anions that might be also present in the water (e.g. chloride). However, this feature can make the sorption capacities of LDHs susceptible to interference from carbonate and sulfate anions.


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Figure 1. Crystal structure of layered double hydroxide (LDH) and the periclase structure formed after calcination. The layered structure is reconstructed upon exposure of the periclase structure to water. Although there have been many reports investigating the fundamental sorption properties of LDHs in batch tests, there have been limited studies evaluating the implementation of LDHs in packed bed columns, particularly for selenium removal. Recent work by Chubar et al. demonstrated that Mg-Al-CO3 LDH (i.e., MII = Mg2+, MIII = Al3+, A = CO32-) was effective for removing selenium from spiked de-ionized (DI) water in a packed bed26. The LDH was prepared with a novel alkoxide-free sol-gel synthesis that resulted in interlayer carbonate ions that could be easily exchanged with selenate, Se(VI)27. The study found that 4300 bed volumes (BVs) of water containing 50-60 µg/L Se(VI) could be treated using the LDH. However, the water matrix only contained 0.01 M NaCl as background electrolyte. Introduction of sulfate at a concentration 74 times higher than the Se(VI) level decreased the Se(VI) removal efficiency by 2.6 times due to simultaneous removal of the competing sulfate anions. These results show that although LDHs are effective sorbents for removing Se(VI), sulfate present in the water can interfere with binding sites and decrease the adsorption capacities, particularly in water matrices with high sulfate levels but only trace levels of selenium. There are also limited studies of the performance of


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LDHs as sorbents in more complex water matrices, where multiple anions and cations might be present. Here, commercially available Mg-Al-CO3 LDH in granular form was evaluated for the removal of selenium from groundwater samples containing naturally occurring levels of selenium < 2 ppb and sulfate around 50 ppm, a difference in concentration of 20,000. The asreceived granules displayed structural features associated with the calcined, dehydrated form of LDH. Fixed bed, small scale column tests were conducted to study the selenium removal capacities in dynamic conditions. Pre-treatment of the water with BaCl2 to remove sulfate was found to be effective for improving the Se(VI) adsorption capacity of the LDH. Detailed characterization of the exhausted media using X-ray diffraction (XRD), Fourier-Transform infrared spectroscopy (FTIR) and surface area measurements, in addition to analysis of the effluent water samples, were performed in order to study the materials properties of the media after use, the species removed, and the potential removal mechanisms. 2. Methods Detailed descriptions of the materials, reagents, water samples, and column test methods are provided in the Supplementary Information (Fig. S1-S2). Granular LDH was obtained from Sasol Germany GmbH (PURALOX MG 63 HT – Granulate). A powdered LDH was obtained from Sigma-Aldrich to use as a reference sample for materials characterization. Synthetic water samples were prepared from ultrapure de-ionized (DI) water spiked with selenate or selenite. Groundwater samples were obtained from one of the wells used as makeup water for the cooling towers at Salt River Project’s Santan Generating Station in Gilbert, AZ.


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3. Results and discussion 3.1. Physical properties of LDH The physical properties of the LDH were characterized using XRD, SEM, and BET analysis. The as-received LDH displayed a pattern consisting of two broad reflections (Fig. 2a) that are attributed to a magnesium/aluminum oxide solid solution with cubic periclase structure (Fig. 1)24,28.

Figure 2. XRD patterns of granular LDH (a) as-obtained sample; (b) after immersion in DI water; exhausted media from column test conducted in (c) groundwater, (d) groundwater pretreated with BaCl2; (e) powdered LDH with layered structure and reference pattern of Mg0.667Al0.333(OH)2(CO3)0.167(H2O)0.5 from PDF 01-089-0460 (bottom). Previous studies showed that calcination of Mg-Al-CO3 LDHs results in dehydroxylation between 70 – 190 oC and decarbonation above 360 oC 24,29, leaving the non-layered periclase materials, as depicted in equation (1) 30.


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∆

Mg1-xAlx(OH)2(CO3)x/2 → Mg1-xAlxO1+ x/2 + x/2 CO2 + H2O

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(1)

This suggests that the interlayer water and carbonate anions were removed by the manufacturer through calcination. SEM imaging (Fig. S3) showed that the material consisted of micron-sized granules formed from aggregated particles. The specific surface area obtained from the BET measurement was ca. 91 m2/g. Sonication of the material resulted in nanosheet-like particles about 200 nm in size (Fig. S3, inset). FTIR spectroscopy of the LDH (Figure 3a) showed that the interlayer hydration deformation bands (δH2O) typically found at around 1635 cm-1 27,31 were absent in the sample, confirming the absence of water molecules inside the interlayer space. Interlayer carbonate anions were identified by the band at 1365 cm-1 16,32.

Figure 3. FTIR spectra of granular LDH (a) as-obtained, (b) after reconstruction and rehydration in DI water, (c) exhausted media from column test conducted in BaCl2 pre-treated well water.


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Immersing the granular LDH into DI water confirmed the ability of the material to regenerate the layered structure, which would also be important for its capacity for oxoanion removal. As shown in Figure 2b, the XRD pattern showed the reflections associated with the layered rhombohedral crystal structure of LDHs 15,33,34. However, the reflections for the reconstructed granular LDH samples were broad, suggesting a low crystallinity. The reference XRD pattern for powdered LDH with layered structure showed much sharper reflections in comparison (Fig. 2e). FTIR analysis of the granular LDH reconstructed in DI water revealed bands at 450 cm-1 from the lattice vibrations of Mg,Al-oxide octahedral sheets, at 550 cm-1, 670 cm-1, and 772 cm-1 from Mg/Al-OH translations, and at 940 cm-1 from Al-OH deformation 27,31,32. 3.2. Selenium adsorption properties Analysis of the obtained groundwater showed that the total selenium concentration was 1.75 ppb Se and the amount of sulfate was 36.6 ppm, or more than 20,000 times higher. The determination of selenium speciation at trace levels requires advanced analytical techniques 35,36 and was not conducted for the groundwater samples. However, previous studies have shown that LDHs are effective for removing both oxoanionic forms of selenium 14,32. Our jar test data also confirmed the efficacy of the granular LDH for removal of both Se(VI) and Se(VI) from spiked DI water solutions (Fig. S4). Since the as-obtained media was comprised of the dehydrated, periclase-form of the LDH, jar test studies were performed to elucidate the Se(VI) removal mechanism in DI water containing trace levels or high concentrations of Se(VI). For the layered LDH, it is expected that the Se(VI) will bind on the particle surfaces and can also be removed via anion-exchange with the interlayer anions and water molecules. However, when the as-obtained media (with periclase structure) is immersed into a solution containing Se(VI), the Se(VI) can also become incorporated into the


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material as the layered structure is reconstructed. At trace levels (initial [Se] = 3.43 ppb), the jar test results showed that the periclase-form of the LDH required 120 min for 100% removal of the Se(VI) (Fig. 4a). In contrast, the Se(VI) could be removed within 2 min of exposure to the layered form of LDH (prepared by reconstructing the as-obtained LDH in DI water, Supporting Information). This implies that the Se(VI) removal was predominately through fast surface adsorption on the layered LDH, while the reconstruction of the layered structure from the periclase-form slows the removal of Se(VI). When the Se(VI) level was increased (initial [Se] = 72.9 ppm), the Se(VI) removal was also slower for the periclase-form over the layered one (Fig. 4b). Due to the high Se(VI) levels in this case, there are insufficient surface sites for Se(VI) adsorption available and anion-exchange in the LDH interlayer space is required, which is a slower process. However, it appears that at longer exposure times (e.g., 22 h), the periclase-form can remove more Se(VI) than the layered form (Fig. S5). This is likely due to the combination of Se(VI) being removed by interlayer anion-exchange in addition to its incorporation into the reconstructed layered structure, rather than just anion-exchange alone.


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Figure 4. Jar test results using 1 g/L LDH in DI water spiked with (a) 3.43 ppb, (b) 72.9 ppm Se(VI). The as-obtained LDH had the non-layered, periclase structure. The layered LDH was obtained by reconstructing the as-obtained LDH in DI water.

To understand the role of sulfate in the groundwater on the selenium removal characteristics of the as-obtained granular LDH, jar tests were performed in water solutions intentionally spiked with large amounts of Se(VI), namely 100 ppm. The sorbent dosage was 1 g/L and the exposure time was 22 h to decrease the likelihood of kinetic limitations for adsorption. As shown in Table 1, the LDH could remove 74% of the Se(VI), which corresponded to an adsorption capacity of 89.5 mg/g. Other inorganic sorbents based on aluminum oxides37, iron oxides38,39,40, manganese oxides41,42, chitosan/clay composites43, and silica44 have reported Se(VI) capacities below 20 mg/g. For Mg-Al LDHs, a study on materials prepared using alkoxide sol-gel, alkoxide-free solgel, and hydrothermal precipitation methods reported adsorption capacities of 4, 45, and 18 mg/g, respectively, at pH 7 32. From these results, we can see that the granular LDH displays a


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superior adsorption capacity compared to other metal oxide and LDH-based sorbents for selenate. Table 1. Jar test results using 1 g/L as-obtained LDH and 22 h exposure time in solutions spiked with 100 ppm Se(VI)

Water

% Se removed

Se loading (mg/g)

DI water + 100 ppm Se(VI)

74%

89.5

Groundwater + 100 ppm Se(VI)

33%

41.1

Groundwater + BaCl2 + 100 ppm Se(VI)

65%

72.8

As shown in Table 1, when the LDH was exposed to the groundwater spiked with 100 ppm Se(VI), the adsorption capacity decreased by more than half and only 33% of the Se(VI) was removed. The much lower removal efficacy suggests the presence of competing species in the water. Therefore, removing the sulfate from the water using a chemical pre-treatment step was investigated. Barium chloride, BaCl2, can precipitate out sulfate as BaSO4 and can also coprecipitate selenate at the same time45. Performing the jar test in groundwater pre-treated with BaCl2 to remove the sulfate showed that the amount of Se(VI) that could be removed was almost the same as what was observed in DI water, with loadings of 72.8 mg/g observed. These results confirm that sulfate removal from the water can be effective for improving the Se(VI) adsorption capacity of LDH sorbents.


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To investigate the selenium removal characteristics of the granular LDH in dynamic situations, column tests were performed using an empty-bed contact time (EBCT) of 30 minutes to ensure that sufficiently long exposure times were used. Figure 5a shows the breakthrough curve for the column test performed using the as-obtained groundwater (initial [Se] = 1.75 ppb). As shown in Figure 5a, the breakthrough volume was Vb = 2.5 L. The exhaustion volume, Ve, was equal to 3 L of water (327 bed volumes, BVs) processed. The mass of the dried LDH-granular was determined to be 7.7 g. Based on the breakthrough data, the adsorption ability of the sorbent was calculated as Qe = 0.65 µg Se/g LDH (Table 2). This low adsorption capacity is consistent with the interference of sulfate with the selenium binding sites on the LDH, and also the much lower level of [Se] in the water compared to the amount of sulfate (from [S]). The [S] during the column test was also monitored and the breakthrough curve is shown in Figure 5a. The Vb for [S] was similar to the one observed for [Se], but with a loading of 18 mg S/g LDH, which suggests that the Se binding sites were saturated by sulfate.


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Figure 5. Breakthrough curve of column test with granular LDH as bed material to remove selenium, EBCT = 30 min; (a) groundwater with initial [Se] = 1.75 ppb and [S] = 36.6 ppm; (b) groundwater pre-treated with BaCl2 and initial [Se] = 1.3 ppb (1.64 ppb before BaCl2 pretreatment).


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Table 2. RSSCT results for LDH-granular in well water (W1); BV is the bed volume; is the capacity of loading selenate at breakthrough; is the exhaustion capacity; AER is the adsorbent exhaustion rate.

Water

BV (cm3)

Sorbent Mass (g)

(L)

(µg/g)

(L)

(µg/g)

AER (g/L)

V (L) by 1 g

Groundwater

9.18

7.7

2.50

272

0.57

3.00

327

0.65

2.57

0.39

Groundwater with BaCl2 pretreatment

8.74

7.9

9.75

1115

1.6

15.55

1779

1.9

0.51

1.97

Next, column tests were performed using the groundwater pre-treated with BaCl2. The initial selenium concentration in the water decreased slightly from 1.64 ppb to 1.3 ppb after BaCl2 was added, suggesting that some of the selenium was removed by co-precipitation. The initial [S] in the water was 41.5 ppm; after BaCl2 addition it decreased to 1.9 ppm. The [S] in the water in the sample of the effluent after 135 BVs was 0.9 ppm, suggesting that there may have been some sulfate adsorption on the LDH media. The column test results using the pre-treated water are shown in Figure 5b. Compared with the column test performed using water without BaCl2 pretreatment, breakthrough occurred much later, with the Vb = 9.7 L and Ve = 15.55 L (1779 BVs) (Table 2), which is much higher than without BaCl2 pre-treatment. Based on the breakthrough data, the exhaustion capacity was Qe = 1.90 µg Se/g. This indicates that BaCl2 pre-treatment can help decrease the amount of sulfate in the water, leading to a roughly 3X increase in the amount of Se(VI) removed by granular LDH. The adsorption exhaustion rate (AER) also decreased from 2.57 to 0.51 g/L. The [S] concentration did not increase at the Vb observed for Se, suggesting that


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sulfate adsorption was not as large of a contributor to the exhaustion of the Se(VI) binding sites on the LDH. Nonetheless, the exhaustion capacity of Qe = 1.90 µg Se/g obtained in the BaCl2treated water is still much lower than what was observed by Chubar et. al (Qb = 481.1 µg/g) when performing the column test in background electrolyte containing only NaCl26. To identify other potentially interfering species in our groundwater samples, ICP-MS analysis was performed on the effluent water at breakthrough from the column test conducted using the BaCl2-treated groundwater. Figure 6 shows the influent and effluent concentrations for various metal species analyzed. Other than removing selenium, the LDH was also effective for completely removing chromium and uranium (initial concentrations < 10 ppb) and removing some of the magnesium, calcium, and strontium (initial concentrations between 10 – 100 ppm) present in the groundwater. On the other hand, slight increases in arsenic and aluminum concentrations were observed.

Figure 6. Metals concentration in BaCl2 pre-treated groundwater with blue and red text indicating the influent and effluent concentrations, respectively. The removal percentage is also shown.


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Since LDHs have already been shown to be effective for adsorption of chromium and arsenic oxoanions through a similar mechanism as for selenium33,46,47,48,49, the changes in these concentrations can be explained by the simultaneous removal of these species on the LDH. The slight increase in the [As] could be from desorption of the arsenic oxoanions due to exhaustion of the LDH, but this could not be confirmed without having the corresponding breakthrough curve for [As]. Regarding the changes in the cation concentrations in the effluent, studies have also shown that certain metal cations can interact with anions adsorbed on the LDHs and become removed 50, despite LDHs being anionic clays with positively charged surfaces. Specifically, uranium cationic species were shown to be removed on Mg-Al LDHs through the formation of complexes with the carbonates adsorbed on the LDHs 51,52,53,54, although previous studies focused on much higher uranium levels (80 ppb – 750 ppm) than observed here. Previous studies also showed that calcined LDH, which releases hydroxide anions during rehydration and reconstruction of the layered structure according to equation (2), can remove Mg2+ and Ca2+ from solution through its reaction with these OH- species and precipitation of Mg(OH)2 and Ca(OH)2 30,55. Mg1-xAlxO1+ x/2 + x/n An- + (1+ x/2) H2O → Mg1-xAlx(OH)2Ax/n + xOH-

(2)

Since the pH of the effluent at breakthrough was measured to be 9.3, only a little higher than the influent pH of 8.3, it is possible that the divalent cations in the groundwater could act to buffer the increase in pH resulting from the reaction in equation (2) by forming the hydroxide precipitates. This mechanism could explain the lower levels of Ca2+, Mg2+, and Sr2+ in the effluent. Since Mg-Al LDHs typically do not undergo dissolution unless below pH 5 56, the slight increase in [Al] observed in the effluent is not yet understood. Previous studies have observed Al3+ leaching from LDHs even at pH 7 though the formation of complexes with organic


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compounds57, but further detailed study of the composition of the groundwater samples would be needed in order to determine if this were occurring. 3.3. Characterization of LDH after selenate removal tests The exhausted granular LDH from the column test performed with BaCl2 treated groundwater was further characterized using BET, XRD, and FTIR. The BET surface area for the LDH decreased to 41 m2/g after the column test, confirming the surface adsorption processes. The XRD analysis showed the exhausted LDH adopted the rhombohedral crystal structure (Fig. 2cd). This indicates that the water molecules together with the anions were incorporated back into the LDH to form the crystalline layered structure15,33. However, the XRD patterns for the LDH recovered from the column test using groundwater pre-treated with BaCl2 (Fig. 2d) looked similar as the one obtained for the LDH used in the groundwater without pre-treatment (Fig. 2c). Both of these XRD patterns also looked similar, although with slightly larger peak widths, to the one obtained from the LDH reconstructed in DI water (Fig. 2b). The reflections for all of the reconstructed granular LDH samples were broad, suggesting a low crystallinity. These results indicate that XRD is not able to distinguish between the different intercalated anion species in the reconstructed granular LDH. FTIR characterization of the exhausted granular LDH from the column test performed with BaCl2 treated groundwater also confirmed the recovery of the LDH crystal structure and also provide evidence of oxoanion adsorption. As shown in Figure 3c, the LDH bands associated with interlayer carbonate and Mg,Al-OH bonds could be observed, in addition to a new band at 861 cm-1 that was not seen in the spectrum from the LDH reconstructed in DI water (Fig. 3b). Previous studies observed a band at 849-862 cm-1 attributed to the ν(Se-O) vibrations of Se(VI)O-Mg,Al-LDH complexes27, so this is the likely origin of this feature. However, the adsorption


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of arsenate and chromate on Mg-Al-LDH was observed at 852 cm-1 58 and 887 cm-1 28, respectively. Since this is in the same range, it is possible that the observed band could have contributions from the adsorption of the arsenic and chromium oxoanions as well. Additionally, the exhausted LDH-granular spectrum exhibited features between 1000 – 1300 cm-1, which are not observed in the LDH reconstructed in DI water. These could be attributed to interlayer surface adsorbed sulfate26,59 since the BaCl2 may not have completely removed all the sulfate in the water. Conclusions The as-obtained granular LDH exhibited good efficacy for the removal of selenium from water. Due to the non-layered, periclase structure of the media, reconstruction of the layered structure occurs when the media is immersed into water. For trace levels of selenate, surface adsorption of selenate can occur quickly on the layered form of LDH, while at higher levels the ion-exchange plays a larger role due to the insufficient amount of surface sites. For the nonlayered periclase form, the reconstruction of the layered structure is a slow process, but ultimately improves the capacity for Se(VI) removal. The LDH was effective in jar tests for removing selenium from groundwater that contained very high levels of competing sulfate anions, with removal capacities > 41 mg/g if the selenate and sulfate concentrations were roughly the same (100 ppm vs. ca. 40 ppm, respectively). Lowering the sulfate levels through addition of BaCl2 resulted in a higher selenate adsorption capacity > 72 mg/g. Column tests were performed using the groundwater containing [Se] < 2 ppb Se and 20,000 times higher [S] with and without BaCl2 pre-treatment. Using the BaCl2 pre-treatment resulted in about three times higher loading of selenium with an adsorption capacity of 1.9 µg Se/g, indicating that sulfate removal will help to improve the sorption capacities of the LDH. Other methods for sulfate


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removal may also be effective and more appropriate for large-scale demonstrations in the field. Characterization of the effluent water and exhausted media confirmed that selenium was adsorbed onto the LDH. Other constituents in the water, such as chromium, arsenic, uranium, strontium, magnesium, and calcium were also removed, although not all of these species may have necessarily adsorbed onto the LDH. These results show that although the selenium concentration in the groundwater was very low compared to the other background ions, the granular LDH was an effective sorbent for its removal. While the observed adsorption capacities in the column tests were low, the results are promising considering the trace levels of selenium in the groundwater. Supporting Information. Detailed experimental methods; photograph of the column test setup; method used to estimate exhaustion capacity; SEM image of granular LDH; jar test results of LDH removal of 1 ppm selenite and selenate from DI water; jar test results of LDH removal of trace and high levels of selenate over long exposure times. Corresponding Author *Candace K. Chan, [email protected] Author Contributions The manuscript was written through contributions of all authors. All authors have given approval to the final version of the manuscript. Funding Sources


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The work was funded through the Arizona State University/Salt River Project Joint Research Program, under the project “Selenium Removal from SRP Waters with Nano-layered Solid Phase Extraction Materials”. Partial support from the National Science Foundation Nanosystems Engineering Research Center for Nanotechnology Enabled Water Treatment (ERC-1449500) is also acknowledged. ACKNOWLEDGMENT We would like to thank D. Pelley, F. Fuller, and B. Schuster from Salt River Project for their support of this research and providing water samples. We are grateful to P. Westerhoff, A. Bowen, D. Hanigan, K. Hristovski, and J. Markovski for helpful discussions. We also acknowledge the use of facilities within the LeRoy Eyring Center for Solid State Science and Goldwater Environmental Laboratory at ASU. REFERENCES (1)

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Selenium Removal from Sulfate-Containing Groundwater Using

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