RedOx-controlled sorption of iodine anions by hydrotalcite composites

Abstract

The radioactive contaminant iodine-129 (I-129) is one of the top risk drivers at radiological waste disposal and contaminated groundwater sites where nuclear material fabrication or reprocessing has occurred. However, currently there are very few options available to treat I-129 in the groundwater, partially related to the complex biogeochemical behavior of iodine in the subsurface and occurrence of I-129 in the multiple chemical forms. We hypothesize that layered hydrotalcite materials containing redox active transition metal ions offer a potential solution, benefiting from the simultaneous adsorption of iodate (IO₃⁻), and iodide (I⁻) anions, which exhibit different electronic and structural properties and therefore may require dissimilar hosts. To test this hypothesis, Cr³⁺-based materials were selected based on the rationale that Cr³⁺ readily reduces IO₃⁻ in solution. It was combined with either redox-active Co²⁺ or redox-inactive Ni²⁺ so that two model materials were prepared by hydrothermal synthesis including Co²⁺–Cr³⁺ or Ni²⁺–Cr³⁺ (M–Cr). Obtained M–Cr materials composed of Co²⁺–Cr³⁺ or Ni²⁺–Cr³⁺ layered hydrotalcite and small fractions of Co₃O₄ spinel or Ni(OH)₂ theophrastite phases were structurally characterized before and after uptake of periodate (IO₄⁻), IO₃⁻, and I⁻ anions. It was found that the IO₃⁻ uptake is driven by its chemical reduction to I₂ and I⁻. Interestingly, in the Co²⁺–Cr³⁺ hydrotalcite, Co²⁺ and not Cr³⁺ serves as a reductant while in the Ni²⁺–Cr³⁺ hydrotalcite Cr³⁺ is responsible for the reduction of IO₃⁻. A different uptake mechanism was identified for the IO₄⁻ anion. The Co²⁺–Cr³⁺ hydrotalcite phase efficiently uptakes IO₄⁻ by a diffusion-limited ion exchange mechanism and is not accompanied by the redox process, while Cr³⁺ in the Ni²⁺–Cr³⁺ hydrotalcite reduces IO₄⁻ to IO₃⁻, I₂ and I⁻. Iodide exhibited high affinity only to the Co–Cr material. The Co–Cr material performed remarkably well for the removal of IO₃⁻, I⁻ and total iodine from the groundwater collected from the US DOE Hanford site, WA, USA, outperforming non-redox active hydrotalcites (e.g., Mg²⁺–Al³⁺) reported previously. This work demonstrates that redox-controlled sorption can be a highly effective method for the treatment of anions based on elements with mobile oxidation states. Further, multiple anions of interest could be simultaneously removed through a combination of approaches.

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Introduction

Iodine is a naturally occurring element that is important for human health and is vital for thyroid function. On the other hand, absorption and concentration of radioactive iodine increases risk of thyroid cancer. Significant amounts of radioactive iodine have been released to the environment due to the activities associated with production of nuclear weapons and from accidents at nuclear power plants. Among iodine radioisotopes, by-products of uranium and plutonium fission I-131 and I-129 are particularly hazardous. Iodine-131 with a half-life of about 8 days, is a major public health concern after nuclear accidents (such as at Chernobyl or Fukushima Daiichi nuclear power plants).¹ Isotope I-129 has a very long half-life of about 15.7 million years, and its levels build up in the environment over time. Iodine-129 is a risk-contributing contaminant of environmental and health concern at U.S. Department of Energy nuclear waste storage sites such as the Savannah River and Hanford Sites.² At the Hanford Site located in Washington State, management of I-129 contamination in the 200 Area vadose zone is an essential component for the Hanford Site cleanup program and protection of the Columbia River. One potential management approach is removal of radioiodine from groundwater for safe long-term storage. To date this remains a challenging task.^3–5 In subsurface aquatic environments, iodine exhibits complex biogeochemical behavior and occurs in multiple inorganic and organic chemical forms and oxidation states varying from −1 to +7.^6,7 In the Hanford groundwater, iodine exists predominantly as iodate (IO₃⁻), which accounts for up to about 70% of total iodine with iodide (I⁻) and organoiodine being minor components.⁴ This implies that to be effective, separation methods should satisfy two requirements including high affinity for the dissimilar forms of iodine and sufficient selectivity for different iodine species to ensure their removal from the complex solutions containing excess of common environmental anions (carbonate, chloride, sulphate, other) found in groundwater.

For several decades considerable research effort has been directed toward identifying synthetic materials for removing or attenuating transport of iodine in the groundwater.^8–10 Clay-like hydrotalcite materials commonly known as layered double hydroxides (LDHs) are attractive candidates as they have been extensively studied for the removal of harmful anions from contaminated water.¹¹ The LDH materials can be represented by the general formulae [M_1−x²⁺M_x³⁺(OH)₂]^x+(Aⁿ⁻)_x/nyH₂O, where M²⁺ is a divalent and M³⁺ is a trivalent cation; the value of x is equal to the molar ratio of M³⁺/(M²⁺ + M³⁺), whereas Aⁿ⁻ is the interlayer anion. These materials usually uptake anions through ion exchange. Variation of the identities of M²⁺, M^m+, and Aⁿ⁻ give rise to a large class of isostructural materials with physicochemical properties that can be tailored to achieve high selectivity and capacity for the uptake of an anion of interest¹² and also be regenerated for reuse.¹³ Application of hydrotalcites for the selective sorption of various anions of iodine has been recently reviewed.^8,9 Among tested hydrotalcites, Mg²⁺–Al³⁺ materials showed promise for iodine removal and have been preferentially investigated. For instance, iodide has been immobilized using Mg²⁺–Al³⁺ LDH by exchanging the parent Cl⁻ anion with I⁻.¹⁴ In another study, calcined Mg²⁺–Al³⁺–CO₃²⁻ material was used to remove iodide from contaminated ground water.¹⁵ Hydrotalcite-type compounds consisting of brucite-like positively charged layers with the similar Mg²⁺–Al³⁺–CO₃²⁻–NO₃⁻ chemical composition showed promise for IO₃⁻ uptake.^16,17 The main drawback of the Mg²⁺–Al³⁺ LDH materials is insufficient selectivity toward iodide and iodate anions. Several studies demonstrate that iodine removal efficiency of uncalcined Mg²⁺–Al³⁺ materials is limited by their high affinity toward carbonate^12,13 present in groundwater, raising concerns regarding their applicability for environmental separations.^8,9 Presence of other anions also affects iodine removal. Kang et al. reported that Mg²⁺–Al³⁺ hydrotalcite exhibits high efficiency of iodide uptake in competing anion free solutions.¹⁸ However, when this and other hydrotalcites were tested in dilute sodium salt (nitrate, nitrite, aluminate, hydroxide) solutions the iodide uptake was significantly reduced.^19,20 A similar effect was observed for iodide adsorption using Zn-bearing hydrotalcites in sodium chloride solutions.²¹

One potential approach to increase LDH selectivity toward iodine adsorption is to introduce soft polarizable metal ions into the LDH structure which have tendency to strongly bind soft bases such as iodine anions.²² Indeed, Hg and Ag-bearing sulfide minerals demonstrate high affinity for iodine anions as reported in the review by Mattigold et al. (and references therein),¹⁰ however toxicity limits their utilization for groundwater treatment. We hypothesize that the presence of the borderline soft redox-active metal ions in the LDH structure will have a weaker but similar effect and may improve their selectivity toward adsorption of the iodine anions. Another objective of this work is to examine if the application of LDH material containing redox-active transition metals is beneficial for selective uptake of the iodine anions due to their potential conversion to the oxidation state and chemical form most complimentary with the electronic and structural environment of the LDH material. Introduction of a secondary redox uptake mechanism in addition to anion exchange can be particularly useful in designing hydrotalcite materials for the simultaneous adsorption of iodate and iodide anions, which exhibit different electronic and structural properties and therefore may require dissimilar hosts. In fact, this can pave the way for a multi-modal approach to removal of multiple anions, where some of the anions can be removed by a simple ion-exchange mechanism, while the others can be removed through more complicated chemical or redox pathways. As proof-of-concept, we are examining borderline soft, redox-active transition metals such as Co²⁺ and Cr³⁺ for the selective uptake of anions of iodine, and their subsequent redox change. Indeed, our previous exploratory work has indicated that among the large array of LDH compounds synthesized and tested for their affinity toward iodate and iodide, Cr³⁺-containing hydrotalcite materials exhibited effective IO₃⁻ uptake from the Hanford groundwater at high sorption capacity.²³ These scoping results prompted us to elucidate and compare uptake mechanisms for iodine anions using two materials based on Co²⁺–Cr³⁺ and Ni²⁺–Cr³⁺ (referred in the text as Co–Cr and Ni–Cr or jointly as M–Cr) as model redox-active compounds containing a combination of the borderline soft Co²⁺ and Ni²⁺ divalent transition metal ions. In addition, Co²⁺ can be easily oxidized to Co³⁺ while Ni²⁺ is expected to retain a +2 oxidation state; this difference allows elucidation of the role of the model redox mobile Cr³⁺ in the uptake of the iodine anions. It is understood that Cr-based materials are unlikely candidates for the practical applications due to the toxic properties of Cr(VI). Rather it serves as a convenient redox-active model allowing benchmarking of redox-driven ion exchange mechanisms and design of environmentally friendly advanced hydrotalcite materials to manage challenging redox mobile subsurface contaminants that include not only iodine, but As, Hg, Tc, and others as well. In this work, the mechanism and sorption behavior of M–Cr toward three different anionic species of iodine including IO₄⁻, IO₃⁻ and I⁻ was explored. Even though periodate is anticipated to be the only minor iodine species in the subsurface, its investigation is essential to understand redox behaviour of iodine upon incorporation in the hydrotalcite structure.

Experimental

Materials

All chemicals (reagent grade) were purchased from Sigma Aldrich or Baker & Adamson Chemicals and used without further purification. Distilled deionized (DI) water was used for the preparation of the aqueous solutions. Iodine uptake from the groundwater was measured using 25 mM solutions of IO₄⁻, IO₃⁻ or I⁻ prepared using the well 299-W19-36 water collected from the US DOE Hanford site, WA, USA. The major constituents of the groundwater are listed in Table S1.†

Material synthesis

While coprecipitation methods have been widely used for the synthesis of LDH composites of the form M–Cr, this work employed hydrothermal methods, which we find result in higher crystallinity. The modified synthetic procedure for the M–Cr (M = Co or Ni) hydrotalcites was adapted from hydrothermal syntheses reported elsewhere.²⁴ In a general procedure, the divalent transition metal and Cr³⁺ nitrate or chloride salts were dissolved in ∼50 mL deionized water at a 4 : 1 molar ratio. The solution was stirred for about an hour followed by the addition of 0.5 M NaOH solution under stirring until a pH of 12 was achieved. The mixture was transferred to a Teflon-lined autoclave and kept at 110 ± 5 °C for 72 hours to yield the aggregate. The obtained solid was gravity-filtered, washed with excess water, and dried. To determine the metal composition of the obtained products, massed amounts of the materials were dissolved in concentrated HNO₃, diluted with water as appropriate and analyzed by Inductively Coupled Plasma-Optical Electron Spectroscopy. The elemental inorganic carbon analysis was conducted by Atlantic Microlab, Inc. (Norcross, GA).

Kinetics of iodine uptake

Uptake of IO₄⁻ or IO₃⁻ anions from DI water and Hanford groundwater collected from the well 299-W19-36 was monitored and quantified by Raman spectroscopy. Raman measurements were performed with an InPhotonics RS2000 high-resolution Raman spectrometer equipped with a thermoelectrically cooled charged coupled device (CCD) detector operating at −55 °C; a 670 nm 150 mW diode laser as the excitation source; and focused fiber optic probe RamanProbe™ operated in a 180° back reflection mode. An integration time of 10 s was used for each acquisition, and 10–20 acquisitions were taken and averaged for each sample. Typical sample preparation involved rigorous mixing 100 mg of M–Cr material with 1 mL of 0.025 M KIO₃ or KIO₄ solution in DI or groundwater in a glass vial. The vial was then centrifuged at 4000 rpm for about 10 minutes until the sorbent settled at the bottom. The liquid phase was continuously monitored for 3 hours. After 3 hours, the solid LDH material was re-suspended in the solution and allowed to stir overnight. Raman measurements were resumed at 24 hours post initial contact. The spectra were baseline-corrected and normalized to the water band. The Raman intensity of the IO₃⁻ band at 803 cm⁻¹ (ref. 25) or IO₄⁻ band at 793 cm⁻¹.²⁶ in the initial 0.025 M aqueous solutions were compared to the intensities of the corresponding bands at each time point and used to quantify the iodine uptake by the materials.

Characterization techniques

Understanding the structural and morphological features of the LDH materials before and after iodine exposure was achieved by X-ray diffraction (XRD), scanning electron microscopy (SEM)/electron dispersive spectrometry (EDS), and transmission electron microscopy (TEM) studies. Changes of the metal center chemical environment upon iodine uptake as well as iodine speciation in the LDH matrix was done through X-ray photoelectron spectroscopy (XPS) and in situ TEM studies. To prepare iodine-loaded LDH samples, ∼100 mg of the LDH material was contacted with 5 mL of the 0.5 M I⁻, IO₃⁻, or IO₄⁻ DI water solution under agitation at room temperature. In 24 hours the liquid phase was removed from the sample after centrifugation, and the loaded composite was rinsed with DI water and air-dried.

XRD patterns of the samples were recorded on a Philips X'pert Multi-Purpose Diffractometer (MPD) (PANAlytical, Almelo, The Netherlands) equipped with a fixed Cu anode operating at 45 kV and 40 mA. XRD patterns were collected in the 5–100° 2θ-range with 0.04° steps at a rate of 5 s per step. Phase identification was performed using JADE 9.5.1 from Materials Data Inc., and the 2012 PDF4+ database from ICSD. The lattice parameters and volume-averaged crystallite sizes were calculated from whole-pattern fitting TOPAS v5 (Bruker AXS GmbH, Germany).

XPS data were recorded on a Phi5000 Versa Probe system equipped with a monochromatic Al Kα X-ray source (1486.7 eV) and a hemispherical analyzer. Powder samples were mounted using double-sided carbon conductive tape attached to a stainless steel sample holder. The instrument was calibrated for Cu (2p_3/2) at 932.6 (±0.1 eV) with the FWHM (full with at half maximum) of 0.98 eV. The surface charge was eliminated by charge neutralizer and correction of data referencing the 284.5 eV C 1s peak. The percentages of individual elements detected were determined from the relative composition analysis of the peak areas of the bands on the basis of the relative peak areas and their corresponding sensitivity factors to provide relative compositions. XPS peak fitting was done using the software XPSPEAK41 with Shirley type background and 20% GL (Gaussian–Lorentzian ratio). The satellite peaks were fitted with parameters given in literature.^27,28

SEM analysis was performed using the FEI Quanta 3DFEG Dual Beam microscope operated at 10–20 kV. The samples were prepared by two independent methods; in the first method, the sample particles were dispersed onto carbon tape and coated with ∼5 nm of carbon to minimize charge effects. In the second method, samples were mounted on the tape and polished using typical metallographic techniques to avoid colloidal silica for polishing. Compositional analysis was performed with Oxford 80 mm² SDD electron dispersive spectrometry (EDS) detector. For quantitative EDS analysis, calculated K factors provided by INCA software were used. No correction for absorption within the specimen was performed. Both secondary electron images (SE) and backscatter electron images (BSE) were recorded.

TEM analysis was performed using the FEI Titan 80-300 operated at 300 kV. The information limit of the microscope was below 0.1 nm. Imaging of the focused ion beam (FIB)-prepared samples was done under low electron dose conditions to minimize beam-induced damage. All images were digitally recorded using a CCD camera and analyzed using Gatan digital micrograph. Compositional analysis was performed with EDAX Si(Li) EDS detector.

Batch contact testing

The M–Cr materials were tested in the systematic batch contact experiments using Hanford groundwater collected from the well 299-W19-36. The total iodine content in the groundwater was determined by ICP-MS to be 8.6 ± 0.9 μg L⁻¹. In the batch contact tests to determine M–Cr sorption efficiency for total iodine, unmodified groundwater was used. To determine sorption efficiency of M–Cr for the individual iodine anions, concentration of IO₃⁻ or I⁻ in the groundwater was adjusted with either KIO₃ or NaI to 15 to 130 000 μg L⁻¹. Periodate is not expected to be present in the Hanford groundwater,⁴ and was not tested in the batch contact studies. For each batch contact sample, 70 ± 5 mg of material was contacted with 15 mL of the test groundwater solution. The samples were placed in the 20 mL glass scintillation vials, agitated for 24 hours at room temperature, and centrifuged. The contact time was selected based on the sorption kinetics measurements demonstrating that sorption is complete in 24 h. The initial pre-contact and post-contact solutions were subjected to the ICP-MS analysis to determine iodine concentration. To quantify the efficiency of iodine removal, the distribution coefficient, K_d (mL g⁻¹), was calculated by using eqn (1),where C_i is the initial concentration of iodine in the groundwater, C_g is the equilibrium concentration of iodine in the post-contact groundwater, V_GW is the volume of iodine solution in groundwater in milliliters, and W_C is the weight of the dry M–Cr material in grams.

Iodine analysis was done using ICP-MS. The instrument was calibrated using high-purity National Institute of Standards and Technology (NIST)-traceable calibration standards to generate calibration curves and verify continuing calibration during the analytical run. All standard preparations and sample dilutions were done using a 0.5% Spectrasol solution CFA-C to prevent build-up/memory of iodine in the ICP-MS introduction system.

Results and discussion

M–Cr synthesis and characterization

In the material synthesis, the 4-fold molar excess of the divalent transition metal relative to the trivalent Cr³⁺ was used based on the previous observations that LDH materials with high M²⁺ : M³⁺ stoichiometric ratio facilitated IO₃⁻ (ref. 16 and 29) and I⁻ (ref. 28 and 29) uptake. For the [M_1−x²⁺M_x³⁺(OH)₂]^x+(Aⁿ⁻)_x/nyH₂O hydrotalcite, the over layer charge is determined by the M²⁺ : M³⁺ ratio,³⁰ and materials with low charge density resulting from low content of M³⁺ ions favor soft charge-diffuse iodine anions. Based on the ICP-OES analysis, the M : Cr molar ratio in the obtained materials was determined to be 2.8 ± 0.1 for Co–Cr and 4.4 ± 0.1 for Ni–Cr composites, respectively. Elemental inorganic carbon analysis suggests presence of the carbonate in both materials, so that the CO₃²⁻ : Cr molar ratio is 0.36 ± 0.02 and 0.174 ± 0.005 for Co–Cr and Ni–Cr products, respectively; atmospheric CO₂ dissolved in the alkaline synthetic mixtures being the source of carbonate.

IR spectra of the obtained solid materials exhibited a moderately strong band at ∼1350 cm⁻¹ as well as bands in the 820–850 cm⁻¹ region which are assigned to the asymmetric stretching and the out-of-plane deformation of free interlamellar carbonate anions with D_3h symmetry in the composites.³¹ Additional bands observed at 1480–1490, 1388 (shoulder), and 1067 cm⁻¹ are also attributed to intercalated carbonate ions, but in the lower C_2v symmetry,^32,33 suggesting binding of these carbonates by the hydrotalcites. A broad band centered at ca. 3440 cm⁻¹ is due to the ν_O–H stretching vibration of hydrogen-bonded geminal OH groups and water molecules, and that at 1636 cm⁻¹ corresponds to the bending of water.³⁴ Other IR bands appeared at the 500–700 cm⁻¹ spectral region, and were assigned to the vibrations of the metal–oxygen bonds.³⁵ Interestingly no presence of nitrate anion in the prepared materials was observed by IR spectroscopy suggesting preferential incorporation of the carbonate anion available due to the aerated conditions of the synthesis.

XRD studies of the materials were conducted to identify the structure of the composites. The XRD patterns of the Co–Cr and Ni–Cr materials (Fig. 1A and B) are characterized by the presence of two distinct phases, one phase being significantly more crystalline than the other, so that M–Cr have composite structures. The diffractograms of the highly crystalline phases in the Co–Cr and Ni–Cr materials are consistent with the respective structures of the cubic spinel Co₃O₄ (JCPDS no. 09-0418)³⁶ and theophrastite, the anhydrous form of Ni(OH)₂ (JCPDS no. 59-0463).³⁶ Based on the relative intensities, it should be noted that while a significant fraction of the cubic spinel Co₃O₄ is present in the Co–Cr composite, the Ni(OH)₂ phase constitutes only a tiny portion of the Ni–Cr composite. The second phase in both composite materials exhibits a nearly identical broad and highly polycrystalline diffraction pattern that is strikingly similar to the previously reported hydrotalcite composite of the formula Mg₆Al₂CO₃(OH)₁₆·4H₂O (JCPDS no. 14-0191)³⁶ characterized by a layered structure with intercalated CO₃²⁻ anions. The diffraction pattern of this polycrystalline phase also matches that of previously observed hydrotalcite-like Co–Cr and Ni–Cr carbonates obtained by the coprecipitation method.^32,33,37 The XRD profiles of the polycrystalline phases obtained in this work are sharper and more intense than ones obtained by the simple co-precipitation method as illustrated in above reports by del Arco et al. and Kooli et al.,^32,33,37 which is attributed to the improved crystallinity of the hydrotalcite products resulting from the prolonged hydrothermal syntheses. In unison with literature, the XRD results indicate that the hydrotalcite phases of the prepared M–Cr composites possess highly layered structural framework.

Fig. 1 XRD patterns of the M–Cr materials before and after exposure to the aqueous solution of 10⁻¹ M of I⁻, IO₃⁻, or IO₄⁻. (A) Co–Cr, (B) Ni–Cr.

XPS measurements were conducted to identify the oxidation states associated with the key components of the prepared composites. Due to the high redox mobility of Co and Cr transition metals (at pH = 14, standard electrode potentials of the Co³⁺/Co²⁺ and CrO₄²⁻/Cr(OH)₃ couples are 0.17 V and −0.13 V, respectively³⁸), oxidation of the starting Co²⁺ and Cr³⁺ ions can be expected during the hydrothermal synthesis. The obtained XPS spectra for the M–Cr composites are shown in Fig. 2 (black traces), observed atomic compositions and the binding energy assignments are given in Tables 1 and S2,† respectively. For the Co–Cr composite, the Co 2p_3/2 spectrum shows a broad peak, which envelops the 780.9 and 779.8 eV regions corresponding to Co²⁺ (87%) and Co³⁺ (13%), respectively (Fig. 2 and Table 2).^39,40 Likewise, Cr 2p_1/2/2p_3/2 spectral region shows bands with the maxima at 586.8 and 577.2 eV suggesting presence of Cr³⁺ (77%)⁴¹ as well as shoulders at 588.7 and 579.2 eV suggesting presence of Cr⁶⁺ (23%)⁴² (Fig. 2 and Table 2). The O 1s spectrum (Fig. S1A,† black trace; Table S2†) shows a broad peak enveloping bands at 531.4 eV assigned to a mixed +2/+3 cobalt oxide⁴³ and 530.7/530.6 eV assigned to Cr₂O₃ (ref. 44) and CrO₃.⁴² The low intensity Cl 2p_3/2 spectrum at 198.7 eV (ref. 45) is assigned to the presence of chloride in the hydrotalcite phase. For the Ni–Cr material, the Ni 2p_3/2 spectrum is attributed to Ni(OH)₂ (Fig. 2 and Table 2).²⁸ The Cr spectral region is nearly identical to that of observed for the Co–Cr material and indicates that Cr³⁺ and Cr⁶⁺ contribute respective 78 and 22% to the overall intensity (Fig. 2 and Table 2).^41,42 The O 1s spectrum (Fig. S1B,† black trace; Table S2†) shows a broad peak as well, centered around 530.7/530.6 eV assigned to Cr₂O₃ (ref. 44) and CrO₃ (ref. 42) with a shoulder at 532.0 eV corresponding to Ni(OH)₂.⁴⁶

Fig. 2 XPS spectral overlays for the M–Cr materials before (black traces) and after 24 hour exposure to 10⁻¹ M I⁻ (violet traces), IO₃⁻ (red traces) or IO₄⁻ (orange traces). (A) Co 2p_3/2 region for Co–Cr, (B) Cr 2p_1/2 and 2p_3/2 regions for Co–Cr, (C) I 3d_5/2 region for Co–Cr, (D) Co 2p_3/2 region for Ni–Cr, (E) Cr 2p_1/2 and 2p_3/2 regions for Ni–Cr, (F) I 3d_5/2 region for Ni–Cr. The respective composition percentages obtained from deconvoluting the spectra are presented in Table 2.

Table 1 Obtained by XPS atomic composition of the M–Cr samples before and after 24 hour exposure to the aqueous solution of 10⁻¹ M of I⁻, IO₃⁻, or IO₄⁻

Composite	C 1s	O 1s	Cl 2p	Co 2p	Cr 2p	I 3d
Co–Cr	24.3	50.7	3.2	17.2	4.5	NA
Co–Cr/I^−	23.4	52.4	1.4	17.4	4.8	0.6
Co–Cr/IO3^−	27.0	50.6	1.5	15.5	4.2	1.3
Co–Cr/IO4^−	26.0	51.4	1.7	14.8	3.1	3.1

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Composite	C 1s	O 1s	Ni 2p	Cr 2p	I 3d
Ni–Cr	27.1	52.8	16.6	3.4	NA
Ni–Cr/I^−	21.7	55.6	18.9	3.2	0.0
Ni–Cr/IO3^−	23.0	54.3	17.0	3.9	1.7
Ni–Cr/IO4^−	23.5	53.9	16.8	3.6	2.3

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Table 2 Atomic percentages obtained from XPS (standard division = 5%) of oxidation states of Co, Ni, Cr and I in the M–Cr samples before and after 24 hour exposure to the aqueous solution of 10⁻¹ M of I⁻, IO₃⁻, or IO₄⁻

Sample	Co%		Cr%		I%
	Co^2+	Co^3+	Cr^3+	Cr^6+	IO4^−	IO3^−	I2	I^−
Co–Cr	87	13	77	23	NA
Co–Cr–I^−	91	9	76	24	—	—	—	100
Co–Cr–IO3^−	78	22	75	25	—	46	26	28
Co–Cr–IO4^−	87	13	80	20	100	—	—	—

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Sample	Ni%	Cr%		I%
	Ni^2+	Cr^3+	Cr^6+	IO4^−	IO3^−	I2	I^−
Ni–Cr	100	78	22	NA
Ni–Cr–I^−	100	70	30	—	—	—	—
Ni–Cr–IO3^−	100	69	31	—	44	24	32
Ni–Cr–IO4^−	100	60	40	48	27	13	12

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SEM images of the Co–Cr composite (Fig. 3A) shows powdery particles that aggregate together to form a cake-like heterogeneous matrix. The powdery morphology of the composite indicates a large surface area facilitating the uptake of anions. Upon further magnification, microcrystalline granules embedded onto the heterogeneous matrix are observed. EDS measurements indicate that the composition of the hydrotalcite-like heterogeneous matrix is distinctly different from the granules largely comprised of the Co₃O₄ spinels (Table 3). The granules are enriched in Co so that Co : Cr ratio is about 6.4 : 1 compared to the heterogeneous phase with Co : Cr ratio of about 2.4 : 1. The hydrotalcite phase showed the presence of Cl, suggesting intercalated chloride anion (Table S3†). These results are further supported by the TEM structural data (Fig. 4), which reveal a two-phase system comprising of particles with cuboidal microcrystalline structure embedded in the heterogeneous matrix. The heterogeneous matrix with significant presence of both Cr and Co, shows a highly diffused diffraction (Fig. 4F) further supporting the polycrystalline structure, characterized by small (>5 nm), highly distorted polycrystalline sheets observed by HAADF imaging (Fig. 4E). The HAADF and diffraction patterns of the Co-enriched cuboids confirm a highly ordered Co₃O₄ spinel structure with Co²⁺ and Co³⁺ being located at the tetrahedral and octahedral sites respectively (Fig. 4C and D).⁴⁷ Small amounts of Cr and Cl associated with the Co₃O₄ spinel phase is presumably due to their bleeding from the adjacent heterogeneous matrix. The cuboidal particles are observed to agglomerate and form porous regions (Fig. 4B).

Fig. 3 Representative SEM images of the M–Cr composites: untreated Co–Cr (A) and Ni–Cr (E) or exposed for 24 hour to the aqueous solutions of 10⁻¹ M IO₄⁻ (Co–Cr: B and Ni–Cr: F); 10⁻¹ M IO₃⁻ (Co–Cr: C and Ni–Cr: G); and 10⁻¹ M I⁻ (Co–Cr: D and Ni–Cr: H).

Table 3 Summary of EDS and elemental mapping of the M–Cr samples before and after exposure to the aqueous solution of 10⁻¹ M of I⁻, IO₃⁻, or IO₄⁻

M–Cr	Analyte	Dominant phase	M : Cr : I ratio	Details of elemental mapping
a The region measured was cracked on the surface.
Co–Cr	—	Hydrotalcite	2.4 : 1 : NA	Regions 1 and 2 in Fig. S2, Table S3
		Spinel	6.4 : 1 : NA	Regions 3, 4 and 5 in Fig. S2, Table S3
Ni–Cr	—	Hydrotalcite	2.6 : 1 : NA	Regions 6 and 7 in Fig. S2, Table S3
		Theophrasite	6.3 : 1 : NA	Region 8 in Fig. S2, Table S3
Co–Cr	IO4^− (24 hours)	Hydrotalcite	2.5 : 1 : 0	Region 1 in Fig. S4, Table S4
		Spinel	6.1 : 1 : 0	Region 2 in Fig. S4, Table S4
		Not identified^a	1.9 : 1 : 0.4	Region 3 in Fig. S4, Table S4
	IO4^− (7 days)	Hydrotalcite	1.7 : 1 : 0.4	Regions 5 and 6 in Fig. S4, Table S4
		Spinel	6.9 : 1 : 0.4	Region 4 in Fig. S4, Table S4
	IO3^− (24 hours)	Hydrotalcite	2.7 : 1 : 0.5	Regions 1 and 2 in Fig. S6, Table S5
		Spinel	6.6 : 1 : 0.1	Region 3 in Fig. S6, Table S5
	IO3^− (7 days)	Hydrotalcite	1.7 : 1 : 0.4	Regions 5 and 6 in Fig. S6, Table S5
		Spinel	6.7 : 1 : 0.3	Region 4 in Fig. S6, Table S5
	I^− (24 hours)	Hydrotalcite	2.6 : 1 : 0.1	Regions 1 and 2 in Fig. S8, Table S6
		Spinel	6.4 : 1 : 0.1	Regions 3 and 4 in Fig. S8, Table S6
	I^− (7 days)	Hydrotalcite	2.5 : 1 : 0.6	Regions 5 and 6 in Fig. S8, Table S6
Ni–Cr	IO4^− (24 hours)	Hydrotalcite	2.6 : 1 : 0.2	Regions 1 and 3 in Fig. S10, Table S7
		Theophrasite	6.7 : 1 : 0.5	Region 2 in Fig. S10, Table S7
	IO4^− (7 days)	Hydrotalcite	5.5 : 1 : 2.3	Regions 4 and 5 in Fig. S10, Table S7
	IO3^− (24 hours)	Hydrotalcite	2.6 : 1 : 0.4	Regions 1 and 2 in Fig. S11, Table S8
		Theophrasite	7.0 : 1 : 0.1	Region 3 in Fig. S11, Table S8
	IO3^− (7 days)	Hydrotalcite	5.1 : 1 : 1.8	Regions 4–6 in Fig. S11, Table S8
	I^− (24 hours)	Hydrotalcite	2.8 : 1 : 0	Region 2 in Fig. S12, Table S9
		Theophrasite	6.4 : 1 : 0	Regions 1 and 3 in Fig. S12, Table S9
	I^− (7 days)	Hydrotalcite	4.4 : 1 : 0	Regions 4 and 5 in Fig. S12, Table S9

---

Fig. 4 (A) Representative TEM image of the Co–Cr composite showing cuboidal microcrystalline structures embedded in a heterogeneous matrix. (B) Higher magnification TEM image showing cuboidal particles to agglomerate and form porous regions. (C) HAADF of cuboids consistent with spinel Co₃O₄ structure. (D) Diffraction map of cuboids consistent with spinel Co₃O₄ structure. (E) HAADF image of heterogeneous embedding matrix showing small (<5 nm), highly distorted and poorly crystalline sheets. (F) Diffraction map of heterogeneous matrix showing highly diffused rings.

SEM image of the Ni–Cr composite (Fig. 3E) shows a cake-like morphology similar to the Co–Cr composite, displaying microcrystalline granular particles being embedded onto a hydrotalcite like heterogeneous matrix. The Ni : Cr ratio in the heterogeneous matrix and in the granules is about 2.6 : 1 and 6.3 : 1, respectively (Table 3). The microscopy results are consistent with the X-ray diffractograms for both M–Cr composites and demonstrate the presence of two distinct phases in each material, the highly crystalline phase being consistent with a cubic spinel Co₃O₄ for Co–Cr material or with Ni(OH)₂ theophrastite phase for Ni–Cr, while the less crystalline heterogeneous phase is consistent with a hydrotalcite like structure.

Overall, based on the elemental analysis XPS, and microscopy results, as well as structural similarity of the heterogeneous M²⁺–Cr³⁺ phases of the materials obtained in this work with the corresponding literature hydrotalcite carbonates,³⁰ it is concluded that hydrotalcite phases mainly have general composition of [M_1−xCr_x(OH)₂]A_x/2·nH₂O, where A is intercalated anion; predominantly carbonate for Ni²⁺–Cr³⁺ and carbonate/chloride mixture for Co²⁺–Cr³⁺. These polycrystalline phases incorporate nearly entire stoichiometric amount of Cr used in the synthesis of M–Cr materials, with a negligible amounts getting incorporated in the second crystalline phases (as evidenced by the XPS and TEM results above).

Kinetics of iodine uptake

The affinity of the synthesized materials toward iodine and kinetics of its uptake was evaluated using IO₄⁻ and IO₃⁻ anions. In these experiments, uptake of IO₄⁻ and IO₃⁻ anions from the DI water or Hanford groundwater was monitored by Raman spectroscopy. In these experiments, about 100 mg of the composite was contacted with 1 mL of 25 mM solution of IO₄⁻ or IO₃⁻. The anion uptake with time was quantified by the reduction of the intensity of the corresponding Raman band at 793 (ref. 26) or 803 cm⁻¹ (ref. 25) in the aqueous contact phase. Both materials exhibited highly efficient and fast uptake of IO₄⁻ and IO₃⁻ anions so that about 90% of iodine was removed during the first 3 hours of contact and sorption was complete in or before 24 hours. Nearly identical sorption results were obtained for both anions regardless of the solution matrix (DI water or Hanford groundwater). Fig. 5 shows representative plots for IO₃⁻ uptake from the Hanford groundwater by Co–Cr and Ni–Cr materials. The insets of Fig. 5 show that concentration of IO₃⁻ remaining in solution can be fit to a logarithmic behaviour with respect to time, suggesting a first order of IO₃⁻ removal kinetics. The first order kinetic rate constants, obtained from the respective slopes, are −7.02 × 10⁻⁴ s⁻¹ for the Co–Cr composite and −8.06 × 10⁻⁴ s⁻¹ for the Ni–Cr composite respectively.

Fig. 5 Kinetics of IO₃⁻ removal by 100 mg composite from 1 mL of 0.025 M IO₃⁻ solution in Hanford ground water monitored by the progressive reduction of the Raman intensity of 803 cm⁻¹ IO₃⁻ band with time; the insets show a plot of logarithm of IO₃⁻ molar concentration in the contact solutions vs. time. (A) Co–Cr, regression analysis for the inset plot: ln[IO₃⁻] = −7.02 × 10⁻⁴t − 3.73, R² = 0.99 (B) Ni–Cr, regression analysis for the inset plot: ln[IO₃⁻] = −8.06 × 10⁻⁴t − 3.70, R² = 0.99.

Characterization of iodine-loaded composites

X-ray diffraction. The XRD patterns of the Co–Cr composite exposed for 24 hours to the 0.1 M aqueous solutions of individual iodine anions are nearly identical to the diffraction pattern of unexposed Co–Cr (Fig. 1A), with the only noticeable change being reduced crystallinity of the hydrotalcite phase in periodate – exposed material. Prolonged exposure to the iodine anions (7 days) resulted in significant changes in the both spinel and hydrotalcite phases of the Co–Cr material. The Co₃O₄ phase almost completely disappeared upon I⁻ exposure or became less crystalline upon IO₃⁻ and IO₄⁻ exposure as evident from Fig. 1A. On the other hand, the diffraction pattern of the hydrotalcite-like phase showed little change for the Co–Cr composite exposed to I⁻ but lost both intensity and crystallinity after prolonged exposure of the composite to IO₃⁻ and IO₄⁻ anions. Likewise, significant changes were observed in the mean crystallite size of the hydrotalcite phase for the materials exposed to either IO₄⁻ or IO₃⁻ anion. However, not much change was observed in the crystallite sizes upon prolonged I⁻ exposure (Table 4). These results imply distinct sorption mechanisms for each iodine anion by the Co–Cr composite. As observed by other techniques, both IO₄⁻ and IO₃⁻ anions are preferentially incorporated by the hydrotalcite phase. Incorporation of the periodate anion reduces the planar distance within the layers of the hydrotalcite phase indicating the strong interaction between this anion and Co–Cr cationic layers. On the other hand, uptake of the iodate anion results in nearly complete loss of the hydrotalcite structure possibly suggesting modification of the cationic layers due the redox reaction with IO₃⁻. Iodide reacts with both hydrotalcite and Co₃O₄ phases. Taken together, these results imply that the Co–Cr polycrystalline phase is involved in the uptake of all the three iodine anions in general, while the Co₃O₄ spinel is important for the sorption I⁻.

Table 4 Effect of 7 day exposure to the aqueous solution of 10⁻¹ M of I⁻, IO₃⁻, or IO₄⁻ on the crystallite sizes of the M²⁺–Cr³⁺ hydrotalcite phases

Composite	Crystallite sizes, nm. Standard deviation is given in the parenthesis
	Untreated	IO4^− treated	IO3^− treated	I^− treated
Co–Cr	3.6(1)	1.3(2)	6(1)	3.4(2)
Ni–Cr	2.9(1)	1.7(2)	2.8(1)	2.6(1)

---

Exposure of Ni–Cr composite to IO₄⁻, IO₃⁻ and I⁻ for 24 hours results in lowering of intensity of the XRD pattern attributed to Ni(OH)₂, along with broadening of the hydrotalcite phase pattern (Fig. 1B) indicating decrease in crystallinity of the entire composite material upon exposure to these anions. Upon prolonged exposure to the iodine anions for 7 days, the hydrotalcite phase is observed to further reduce crystallinity, while the Ni(OH)₂ phase is completely absent suggesting that both phases are involved in the iodine uptake. As in the case of Co–Cr, periodate incorporation reduces the planar distance within the layers of the Ni–Cr hydrotalcite phase indicating the strong interaction between this anion and Ni–Cr cationic layers (Table 4). On the other hand, no significant changes are observed in the crystallite sizes of this phase upon IO₃⁻ or I⁻ exposure.

X-ray photoelectron spectroscopy. X-ray photoelectron spectroscopy measurements were conducted using the M–Cr composites exposed to 10⁻¹ M IO₄⁻, IO₃⁻ or I⁻ for 24 hours to monitor redox changes in the composites upon the uptake of these anions and to get a qualitative measure of the uptake. The elemental composition of the iodine-loaded Co–Cr composite shows a significant fraction of iodine, indicating that the Co–Cr composite sorbs all three anions, and the sorption efficiency decreases in the order IO₄⁻ > IO₃⁻ > I⁻ (Table 1). Uptake of IO₄⁻ or I⁻ anion was not accompanied by a redox process as evidenced from the I 3d_5/2 spectrum showing a single IO₄⁻ line at 624.2 eV (ref. 48) or a single I⁻ line at 618.8 eV;⁴⁹ this result is consistent with the absence of the shifts in the binding energies of Co 2p_3/2 and Cr 2p_1/2 and 2p_3/2 bands (Fig. 2A–C and Table 2). On the other hand, uptake of IO₃⁻ was accompanied by the redox process. Iodine 3d_5/2 spectrum shows two new peaks in addition to IO₃⁻ assigned to I₂ (620.5 eV)⁵⁰ and I⁻ (618.8 eV),⁴⁹ respectively, accounting for the respective 26 and 28% of total iodine sorbed by the Co–Cr composite (Fig. 2C and Table 2). The concomitant broadening of the Co³⁺ 2p_3/2 line due to ingrowth of the Co³⁺ 779.8 eV band^37,38 was observed suggesting oxidation of the Co²⁺. The Co³⁺ content in the Co–Cr composite exposed to IO₃⁻ increased by about 9% in comparison with the original material (Table 2). Interestingly, no changes in the Cr 2p_1/2 and 2p_3/2 spectral profile were found, suggesting that the Cr³⁺ does not undergo oxidation upon IO₃⁻ uptake. On the other hand, the uptake of all these anions were accompanied by the lowering of the Cl 2p_3/2 band at 198.7 eV,⁴⁵ suggesting the replacement of chloride by the anions of iodine.

The Ni–Cr composite exhibited high affinity for IO₄⁻ and IO₃⁻ anions but uptake of I⁻ was negligible (Fig. 2D–F and Table 1). Sorption of IO₄⁻ and IO₃⁻ anions were accompanied by their partial reduction to IO₃⁻/I₂/I⁻ and I₂/I⁻, respectively, and oxidation of Cr³⁺ to Cr⁶⁺, whose content increased by about 10–20% (Table 2). As expected, there was no change in the oxidation state of Ni²⁺ upon iodine exposure.

Microscopy/EDS. Short summary of the SEM/EDS measurements of the iodine-exposed M–Cr materials is given in Table 3, which references the experimental results provided in the ESI.† The elemental mapping of the exposed Co–Cr materials reveals a general reduction in the amount of Cl in the EDS, showing the replacement of Cl with the anions of iodine.

SEM analysis on Co–Cr demonstrates that exposure of the composite to IO₄⁻, IO₃⁻ or I⁻ for 24 hours results in noticeable changes in the morphology of the heterogeneous phase (Fig. 3B–D). Exposure to IO₄⁻ results in pancake like morphology where the individual powdery units are stuck together. Exposure to IO₃⁻ makes the matrix more compact while giving it a coarser landscape. Exposure to I⁻ generates a crust-like morphology. Such significant morphological changes support chemisorption of iodine by the heterogeneous matrix.

Elemental mapping of 24 hour IO₄⁻ exposed Co–Cr composites shows no presence of iodine in the bulk material (Table 3). However, elemental analysis of the cracks on the surface shows a significant concentration of iodine. The TEM analysis (Fig. S5†) is consistent with the SEM results and demonstrates that neither Co₃O₄ nor Co–Cr hydrotalcite phases incorporate IO₄⁻ during first 24 hours of exposure. This, in unison with the fact that XPS of IO₄⁻-exposed Co–Cr composite exhibits a significant fraction of iodine suggests that initially sorption of IO₄⁻ is surface-limited.

Elemental analysis of 24 hour IO₃⁻-exposed Co–Cr composite demonstrates a high concentration of iodine in the regions dominated by the hydrotalcite matrix, while only small iodine amounts are observed to be associated with spinels (Table 3). This is further illustrated by elemental mapping, suggesting that IO₃⁻ has high affinity to the polycrystalline matrix, with the Co₃O₄ cuboids showing insignificant, presumably surface, uptake. The TEM analysis of IO₃⁻-exposed Co–Cr composite (Fig. S7†) confirms the SEM/EDS results (Fig. S6†) and reveals that the elemental composition and morphology of Co₃O₄ spinels remain largely intact upon IO₃⁻ exposure.

Elemental analysis of 24 hour I⁻-exposed Co–Cr composites reveals that both the polycrystalline hydrotalcite phase and the spinels demonstrate affinity to iodide, with the affinity of the hydrotalcite phase being slightly higher (Table 3). TEM analysis of the polycrystalline layer (Fig. S9†) shows small (∼1 nm) particles dispersed within the matrix. High diffraction contrast suggests that these are amorphous iodide-containing particles. Elemental mapping also indicates a significant absorption of iodine around the spinel nano-particles (Table 3). As discussed above, this high affinity of the Co₃O₄ phase towards I⁻ is also reflected in its complete disappearance from the X-ray diffractogram of Co–Cr composite after 7 day exposure to I⁻ due to its presumable reaction with I⁻ (Fig. 1A).

The microscopy of the Co–Cr composite after exposure to IO₄⁻ or IO₃⁻ for 7 days displays very little visible change either in heterogeneous matrix or the cuboidal granules (Fig. S3†). Elemental mapping results reveal that the composition of the cuboidal spinel phase remains almost invariant upon prolonged exposure to these anions, with slight incorporation of iodine presumably due to surface sorption. On the other hand, significant changes are observed in the elemental composition of the hydrotalcite phase (Table 3). This result is consistent with X-ray diffractograms (Fig. 1A) demonstrating that Co₃O₄ phase is not affected by prolonged exposure to these anions, while crystallinity of the heterogeneous hydrotalcite phase is greatly reduced. It is noteworthy that while there was no incorporation of IO₄⁻ in the Co–Cr hydrotalcite phase during first 24 hour of exposure, significant IO₄⁻ incorporation occurs in 7 days. This is indicative of a diffusion-limited ion exchange mechanism. In case of exposure to IO₃⁻, fast incorporation of this anion into the hydrotalcite phase occurs so that significant iodine loading is achieved within first 24 hours and only small increase of iodine loading is measured during the following days. These points out a different sorption mechanism involving chemical or redox interaction between IO₃⁻ and hydrotalcite phase. This redox mechanism is supported by the XPS results. The elemental mapping of the Co–Cr composite exposed to I⁻ for 7 days shows the disappearance of Co₃O₄ phase. For example, region 5 in the microscopic profile that resemble granular particles (Fig. S8†), shows near identical elemental composition as that of the hydrotalcite matrix (region 6 in Fig. S8† and Table 3), thereby supporting the chemical reaction of Co₃O₄ phase with I⁻. This, in unison with the XRD results on the Co–Cr composite after 7 day I⁻ exposure suggests that I⁻ uptake by the composite is associated with a chemical reaction of I⁻ with both the hydrotalcite as well as the Co₃O₄ phases of the composite.

SEM analysis performed on the Ni–Cr composite exposed to 10⁻¹ M IO₄⁻, IO₃⁻ or I⁻ for 24 hours shows very little visible changes in the morphologies either in the polycrystalline matrix or on the embedded granules (Fig. 3F–H). However, the elemental mapping clearly indicates prominent changes in the elemental composition upon exposure to IO₄⁻ and IO₃⁻ resulting in iodine incorporation in the composite.

Elemental mapping of the 24 hour IO₄⁻-exposed Ni–Cr composite shows almost equal distribution of iodine between the bulk hydrotalcite phase and the granules enriched with theophrasite phase (Table 3). On the other hand, elemental mapping of the 24 hour IO₃⁻ exposed composite shows a greater tendency of the absorbed iodine to be distributed in the bulk hydrotalcite matrix as compared to the granules matrix. Elemental analysis of 24 hour I⁻-exposed Ni–Cr show no incorporation of iodine either on the granules or within the heterogeneous matrix (Table 3) confirming weak affinity of this composite toward uptake of the I⁻ ions observed by XPS.

The elemental mapping of the Ni–Cr composite after exposure to IO₄⁻ or IO₃⁻ for 7 days shows complete disappearance of the Ni(OH)₂ rich granular phases upon prolonged iodine exposure. While the microscopic profiles shown in Fig. S10† for IO₄⁻ and Fig. S11† for IO₃⁻ depict profiles similar to the starting materials with granular particles embedded in the hydrotalcite, elemental mapping demonstrates a nearly identical elemental composition of the granules and the hydrotalcite phases (Table S8† for IO₄⁻ and Fig. S11† for IO₃⁻), indicating that the granules are merely morphological variations of the matrix, and not a different phase or a different chemical form. Prolonged exposure of the Ni–Cr composite to I⁻ solution did not facilitate sorption of this anion (Fig. S12 and Table S9†).

Batch sorption of IO₃⁻ and I⁻ from Hanford ground water (299-W19-36 well)

In the batch contact samples, the same weight of the composite (70 ± 5 mg) and the groundwater volume (15 mL) were used so that solution-to-solid ratio was 214 mL g⁻¹ in all samples. The concentration of IO₃⁻ or I⁻ in the Hanford groundwater collected from well 299-W19-36 was adjusted in the 15 to 130 000 μg L⁻¹ range (Table 5). Periodate anion is not expected to be present in the contaminated Hanford groundwater⁴ and thereby was not included in this testing. In addition, non-modified groundwater was subjected to batch contact testing; the total iodine concentration in groundwater was determined to be 8.6 ± 0.9 μg L⁻¹ by ICP-MS analysis. The 24 hour contact solutions were separated from the composite material and subjected to the ICP-MS analysis to determine iodine concentration. The IO₃⁻ and I⁻ batch sorption results are summarized in Table 5.

Table 5 Iodine uptake from the Hanford groundwater (well 299-W19-36) by the M–Cr composite material. The groundwater was tested as received without any pre-processing. Alternatively, the concentration of iodate or iodide in groundwater was adjusted by either KIO₃ or NaI. Each batch contact sample contained 70 ± 5 mg of sorbent suspended in 15 mL of groundwater

Iodine concentration in groundwater, μg L^−1			Kd(Iodine) mL g^−1		Percent iodine sorbed, %
Initial	Post contact		Co–Cr	Ni–Cr	Co–Cr	Ni–Cr
	Co–Cr	Ni–Cr
Iodate (IO3^−)
130 000	48 100	37 500	363	528	63	71
96 500	32 100	24 800	424	615	67	74
62 000	18 300	13 900	507	740	70	78
32 600	7340	6090	740	933	77	81
12 500	2420	1600	885	1450	81	86
10 100	1970	1130	888	1680	80	87
6400	1190	676	937	1810	81	89
2330	375	314	1120	1890	83	90
1240	212	133	1020	1770	86	89
1010	137	88	1350	2240	84	91
613	99	58	1090	2050	84	91
130	18.2	13	1320	1950	86	90
74	8.9	7.4	1540	1940	88	90
Iodide (I^−)
120 000	83 200	104 000	94	33	31	13
79 400	58 400	78 800	77	2	26	1
54 900	28 600	46 800	199	37	48	15
27 800	2780	24 700	1900	27	90	11
13 000	17.6	Not measured	156 000	—	∼100	—
10 800	5.7	9890	401 000	20	∼100	8
8710	12.5	7550	147 000	33	∼100	13
5480	5.5	5130	212 000	15	∼100	6
2910	5.4	2390	113 000	46	∼100	18
1050	4.3	948	51 000	23	∼100	10
837	2.3	743	79 000	27	∼100	11
584	2.1	Not measured	57 000	—	∼100	—
292	2.2	240	28 000	46	99	18
116	1.8	98	13 400	41	98	16
59	1.6	51	7800	32	97	13
Total iodine in unmodified groundwater
8.6 ± 0.9	1.5	3.1	1010	390	82	64

---

Both M–Cr composite materials exhibited efficient sorption of the IO₃⁻ anion. The Ni–Cr composite was found to be more effective sorbent as evident from the K_d(IO₃⁻) values across entire IO₃⁻ concentration range. The maximum K_d(IO₃⁻) values observed for the Co–Cr and Ni–Cr composites are about 1500 and 2200 mL g⁻¹, respectively. As the IO₃⁻ concentration in the groundwater decreases, K_d(IO₃⁻) values first increase due to the reduced loading of the sorbents and then gradually plateau. Overall it is observed that Co–Cr and Ni–Cr composites remove respective 80–88 and 85–90% IO₃⁻ from the groundwater initially containing 74–12 500 μg L⁻¹ IO₃⁻. This confirms high selectivity of the M–Cr materials for IO₃⁻ anion as the groundwater contains large access of other anions including nitrate (317 000 μg L⁻¹), chloride (181 000 μg L⁻¹), carbonate (116 000 μg L⁻¹), and sulfate (50 000 μg L⁻¹) (Table S1†).

The Co–Cr composite demonstrated very high affinity and selectivity for I⁻, so that its nearly quantitative removal from the groundwater was observed for the concentration below 13 000 mL g⁻¹ (Table 5) The Ni–Cr exhibited poor sorption of iodide as reflected by low K_d(I⁻) values. This result is consistent with the XPS and SEM/EDS measurements demonstrating absence of iodine in the I⁻ exposed Ni–Cr.

The performance of the inorganic composite materials tested for the uptake of iodine from the unmodified Hanford groundwater is of particular interest and can help to predict the removal of I-129 isotope. The K_d(iodine) values for the total iodine obtained using the unmodified groundwater generally follow the combination of the IO₃⁻ and I⁻ sorption profiles. To this end, the Co–Cr composite exhibited a K_d(Iodine) value of ∼1000 mL g⁻¹ and removed 82% of total iodine. Ni–Cr composite exhibited modest performance with K_d(Iodine) value of ∼400 mL g⁻¹ and removal of 64% of total iodine. These results can be attributed to the iodine speciation in the Hanford groundwater. Previous literature indicates that the Hanford groundwater contains predominantly IO₃⁻, accounting for up to 70% of total iodine.⁴ The minor iodine species include iodide and organoiodine. It is presumed that Co–Cr composite, which has high affinity for both IO₃⁻ and I⁻ anions, effectively uptakes them from groundwater accounting for 82% removal of total iodine. The iodine not removed from groundwater most likely consists of organoiodine species. This is consistent with the Ni–Cr performance, which exhibits high affinity for IO₃⁻ but not I⁻ anion, and removes only 64% of total iodine from the groundwater.

These results also demonstrate high selectivity of the M–Cr composites towards the anions of iodine (total iodine concentration in groundwater is 8.6 μg L⁻¹, Table S1†) even in presence of the bulk concentrations of coexisting anions including Cl⁻ (181 000 μg L⁻¹), NO₃⁻ (317 000 μg L⁻¹), CO₃²⁻ (116 000 μg L⁻¹) and SO₄²⁻ (50 000 μg L⁻¹) that are present in Hanford ground water as shown in Table S1.† This demonstrates the high selectivity of the composites and emphasizes the importance of the redox mechanism in the uptake of IO₃.

Summary

The M–Cr composites obtained by the hydrothermal synthesis consist of structurally well-defined phases as evidenced from the higher resolution and intensity of the XRD patterns. Co–Cr and Ni–Cr materials contain structurally very similar polycrystalline hydrotalcite phase with the general composition of [M_1−xCr_x(OH)₂A_x/2·nH₂O incorporating nearly entire amount of Cr present in the composites. Both Co–Cr and Ni–Cr composites contain a second highly crystalline phase with the respective structures corresponding to the Co₃O₄ spinel and Ni(OH)₂ theophrastite.

Overall sorption efficiency of IO₄⁻, IO₃⁻ and I⁻ anions by Co–Cr and Ni–Cr composites is summarized in Table 6. Structural characterization study of the composite materials before and after loading with iodine revealed distinct uptake mechanism for each anion/composite system as evident from the mutually consistent XRD, XPS, and microscopy analyses.

Table 6 Qualitative summary of IO₄⁻, IO₃⁻ and I⁻ uptake and resulting iodine speciation in the M–Cr composites and the observed changes in the oxidation states of the transition metals comprising the composites

Composite	Anion	Sorption	Variation of Cr oxidation state	Variation of Co or Ni oxidation state	Iodine speciation in the composites
Co–Cr	IO4^−	Moderate	No change	No change	IO4^−
	IO3^−	High	No change	Decrease in Co^2+ increase in Co^3+	IO3^−, I2, I^−
	I^−	High	No change	No change	I^−
Ni–Cr	IO4^−	High	Decrease in Cr^3+ increase in Cr^6+	No change	IO4^−, IO3^−, I2, I^−
	IO3^−	High	Decrease in Cr^3+ increase in Cr^6+	No change	IO3^−, I2, I^−
	I^−	None	No change	No change	—

---

The Co–Cr and Ni–Cr hydrotalcite phases showed effective uptake of the IO₃⁻ anion, the most abundant chemical species of iodine in the contaminated Hanford groundwater. It is significant to note that IO₃⁻ uptake by both Co–Cr and Ni–Cr composites was accompanied by its partial reduction to elemental iodine and iodide. In the Co–Cr hydrotalcite, Co²⁺ served as a reductant oxidizing to Co³⁺. Based on the relative reduction potentials for Co(OH)₃/Co(OH)₂ (0.42 V vs. NHE) and IO₃⁻/I₂ (1.2 V vs. NHE) couples, this process is not entirely surprising. However, it is surprising that there is no accompanying oxidation of Cr³⁺ in this process in spite of IO₃⁻ being a strong enough oxidant to oxidize Cr³⁺ to Cr⁶⁺ [E⁰′(CrO₄²⁻/Cr³⁺) = 0.11 V vs. NHE].³⁸ In fact, when an aqueous solution of Cr(NO₃)₃ was allowed to react with an aqueous stock solution of IO₃⁻ as an experimental control, rapid oxidation of Cr³⁺ to Cr⁶⁺ was observed (data not shown). The fact that a similar oxidation process is not observed in the Co–Cr composite implies that Cr³⁺ incorporation in composite raises its oxidation potential by such a significant amount that IO₃⁻ is no longer able to oxidize it. The uptake of the IO₃⁻ anion generates large distortion of the hydrotalcite phase so that its crystallite size is significantly increased (Table 4) presumably due to oxidation of Co²⁺. The IO₃⁻ uptake mechanism by the Ni–Cr hydrotalcite is different, as for this phase the reduction of IO₃⁻ was accompanied by oxidation of Cr³⁺ to Cr⁶⁺. Interestingly, the hydrotalcite crystallite size is preserved. Overall these results suggest that chemical reduction is an important driving force in the uptake of IO₃⁻ by the redox mobile hydrotalcite phases resulting in highly efficient and selective removal of this anion from such a complex solution matrix as contaminated groundwater. Under similar conditions, redox inactive Ni²⁺–Al³⁺ and Mg²⁺–Al³⁺ composites demonstrates low affinity for IO₃⁻.⁵¹ The specificity of the IO₃⁻ uptake by the composites in presence of bulk concentrations of coexisting redox inactive anions NO₃⁻, CO₃²⁻ and SO₄²⁻ is illustrated from the uptake data from Hanford groundwater, and further emphasizes the importance of redox affinity for this uptake. The mechanism of the iodate uptake can be characterized as redox-controlled ion exchange.

The M–Cr composites exhibit a different uptake mechanism for IO₄⁻ anion. For the Co–Cr composite, SEM/EDS analysis suggests that the IO₄⁻ is initially adsorbed on the surface only and slowly penetrates into the core of the material with time. This process is not associated with reduction of IO₄⁻ or oxidation of Co²⁺/Cr³⁺ indicating a diffusion-limited ion exchange mechanism. On the other hand, uptake of IO₄⁻ by the Ni–Cr composite is accompanied by the reduction of IO₄⁻ to IO₃⁻, I₂ and I⁻ so that the significant fraction of Cr³⁺ is oxidized to Cr⁶⁺.

Iodide exhibited affinity to both hydrotalcite and spinel phases of the Co–Cr composite. On the other hand, Ni–Cr composite failed to incorporate I⁻. It is an interesting observation that while the hydrotalcite phase of Co–Cr can uptake IO₄⁻, IO₃⁻ and I⁻, the Ni–Cr hydrotalcite phase can uptake only IO₄⁻ and IO₃⁻. It is likely that the geometries of IO₄⁻ (distorted tetrahedral), IO₃⁻ (trigonal pyramidal) and I⁻ (spherical) influence the complimentarily of the uptake.

Both Co–Cr and Ni–Cr composites performed remarkably well for the removal of IO₃⁻ from the Hanford groundwater demonstrating exceptional selectivity over other anions common in the subsurface. Only Co–Cr demonstrated effective uptake of I⁻. Overall the Co–Cr composite removed major fraction of the total iodine from the unmodified Hanford groundwater performing significantly better than previously reported LDH materials in the complex solutions.^8–10

This work presents a new approach of multi-anion removal where the sorbent can remove multiple anions of interest through different mechanisms; where some can be removed through a simple ion exchange mechanism while others can be removed through a redox based approach. This work also concludes that Co-based hydrotalcite/spinel composites are highly promising candidates for the selective removal of the dominant inorganic forms of iodine from groundwater, as borderline-soft Co ion present in the crystalline different phases is responsible for the uptake of both IO₃⁻ and I⁻ species. This work also emphasizes the importance of designing redox active composite materials preferentially containing dissimilar structures to ensure uptake of various iodine species and maximize removal of total iodine from the groundwater. Layered hydrotalcites containing redox active components activate two uptake mechanisms driven by ion exchange and the redox reactions, greatly enhancing affinity and selectivity toward redox mobile anions. It appeared that for Co–Cr material, Cr does not interact with incorporated iodine maintaining a +3 oxidation state and most likely plays only indirect role in iodine uptake serving as a building block of hydrotalcite phase. Future investigation of Co-based composites containing environmentally friendly trivalent transition metal ions, e.g. Fe³⁺ and Al³⁺ is warranted.

Acknowledgements

This research was supported by (1) the Laboratory Directed Research and Development Program at the Pacific Northwest National Laboratory operated by Battelle for the U.S. Department of Energy under Contract DE-AC05-76RL01830 and (2) the U.S. Department of Energy Richland Operations Office. This manuscript was prepared by the Deep Vadose Zone – Applied Field Research Initiative at Pacific Northwest National Laboratory. Part of this research was performed at EMSL, a national scientific user facility at PNNL managed by the Department of Energy's Office of Biological and Environmental Research.

Notes and references

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Footnote

† Electronic supplementary information (ESI) available: Composition of 299-W19-36 Well of Hanford, WA; assignment of XPS data as well as additional XPS spectra; microscopy and elemental mapping of the M–Cr composites after exposure to anions of iodine. See DOI: 10.1039/c6ra13092e

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