Selective capture of radionuclides (U, Pu, Th, Am and Co) using functional nanoporous sorbents.

Abstract

This work evaluated sorbent materials created from nanoporous silica self-assembled with monolayer (SAMMS) of hydroxypyridinone derivatives (1,2-HOPO, 3,2-HOPO, 3,4-HOPO), acetamide phosphonate (Ac-Phos), glycine derivatives (IDAA, DE4A, ED3A), and thiol (SH) for capturing of actinides and transition metal cobalt. In filtered seawater doped with competing metals (Cr, Mn, Fe, Co, Cu, Zn, Se, Mo) at levels encountered in environmental or physiological samples, 3,4-HOPO-SAMMS was best at capturing uranium (U(VI)) from pH 2–8, Ac-Phos and 1,2-HOPO-SAMMS sorbents were best at pH <2. 3,4-HOPO-SAMMS effectively captured thorium (Th(IV)) and plutonium (²³⁹Pu(IV)) from pH 2–8, and americium (²⁴¹Am(III)) from pH 5–8. Capturing cobalt (Co(II)) from filtered river water doped with competing metals (Cu, As, Ag, Cd, Hg, Tl, and Pb) was most effective from pH 5–8 with binding affinity ranged from IDAA > DE4A > ED3A > Ac-Phos > SH on SAMMS. Iminodiacetic acid (IDAA)-SAMMS was also outstanding at capturing Co(II) in ground and seawater. Within 5 minutes, over 99% of U(VI) and Co(II) in seawater was captured by 3,4-HOPO-SAMMS and IDAA-SAMMS, respectively. These nanoporous materials outperformed the commercially available cation sorbents in binding affinity and adsorption rate. They have great potential for water treatment and recovery of actinides and cobalt from complex matrices.

1. Introduction

The ability to selectively capture toxic heavy metals from complex aqueous mixtures is a topic of significant interest, and one that has application in multiple areas such as environmental remediation, pollution prevention, sensing and detection, and medical diagnostics and treatment. For radioactive isotopes, the concern is not limited to their chemical toxicity, but also their radiological hazards [1]. These radionuclides can be released into the environment through a variety of vectors, such as mining activities, leaks or spills of nuclear waste, explosion of a nuclear weapon, or terrorist activities (e.g., “dirty bombs”) [2]. The ability to efficiently and selectively capture radionuclides offers clear benefits to water treatment/purification, industrial hygiene for ore processing and extraction, and perhaps even new therapeutics for patients suffering from radiological contamination. Nuclear power remains important in our energy landscape, accounting for 20% of the US’s electricity each year since 1990 and 56% of the US’s carbon-free electricity in 2017 [3]. This places emphasis on radiological stewardship, and separation capability is a key component of that.

We have had long-standing interest in the development of functional nanoporous sorbent materials for selective chemical separations. As a part of these efforts, we have developed several classes of self-assembled monolayers on mesoporous supports (SAMMS) specifically for the capture of various metal ions. SAMMS materials are made by constructing a dense monolayer of covalently anchored organosilanes on the surface of nanoporous silica supports. The nanoporous silica provides a rigid, well-ordered pore structure that allows rapid metal ion transport and binding kinetics and a large surface area (500 to over 1000 m²/g) that provides high sorption capacity. This synthesis can be carried out in routine condensed phase solvents, or supercritical fluid media [4, 5]. The posture of the functional organosilane leaves the active ligand at the monolayer interface, where it is able to chelate the metal ion of interest [6]. If these organosilanes contain suitable ligands at their terminus, then they can make excellent sorbent materials for heavy metals [7–9], transition metals [10, 11], lanthanides and actinides [12–17], oxometallic anions [18, 19], and cesium and thallium [20, 21]. Multiple metal-ligand interactions are possible due to the close proximity of these ligands on the mesoporous silica, creating attractive strategy for capturing the actinide cations, with their high coordination numbers. Previous work has shown that chelation of 4 ligands per a rare earth cation is possible in systems with dense monolayer coverage [15, 22]. This “macromolecular chelation” contributes to the high binding affinity of these sorbent materials. The materials are not only effective for water treatment, but they are also effective at capturing MeHg/Cd/Pb [8] and Cs [21] in rodent animals.

Our previous work with the actinides was typically carried out in laboratory buffer solutions [16]. While this is a very useful exercise for mapping out the fundamental utility of a ligand for a specific analyte as a function of pH, it is not necessarily representative of a sorbent’s actual performance in the field due to the matrix effects arising from variable ionic strength, competing ions, complexants and dissolved organic matter, among others. It is critically important to evaluate sorbent performance in a matrix that is similar to what will be encountered in the field in order to know how to best deploy the technology. In this work, we have chosen to use pH adjusted seawater as a test-matrix due to its high ionic strength, the presence of an assortment of competing ions, and dissolved organic matter. The high ionic strength of seawater also enables it to be used as a simulant for some biological fluids, such as extracellular fluid and blood plasma. This manuscript summarizes our efforts to capture important actinides (U, Th, Pu, and Am) from seawater using a suite of functional nanoporous sorbent materials.

The actinides are not the only radionuclides of concern to human health and the environment. Other radionuclides, due to their commercial significance, are widely distributed and potentially easier to gain access to. For example, ⁶⁰Cobalt (⁶⁰Co), a beta and gamma emitter with a 5.27 year half-life [23], is widely used in radiotherapy cancer treatment [24], food irradiation [25], and commercial sterilization devices [26]. Radioactive materials can infiltrate the environment and come in contact with humans via nuclear fallout, malevolent activities such as an improvised radiological dispersal device or accidental processes for peaceful commercial uses of the isotopes. Therefore, it is important that we develop not only the ability to selectively capture the actinides, but also ⁶⁰Co that could have an impact on human health and quality of life.

The chemistry of Co(II) is significantly different from that of the actinides. In contrast to the high coordination numbers and variable coordination geometries of the actinides, Co(II) tends to form hexacoordinate, octahedral complexes. In water, Co(II) forms stable complexes with amines (e.g. NH₃, ethylenediamine, etc.) and chelating carboxylates (e.g. oxalate, glycine, etc.), and it forms strong complexes with EDTA (log K = 16.3 [27]). These observations led us to look towards various glycine derivatives, similar to EDTA, as ligands for the capture of Co(II) from natural waters. We evaluated three nanoporous sorbents that integrated ligands derived from glycine – iminodiacetic acid (IDAA, a tricoordinate ligand), ethylenediamine triacetic acid (ED3A, a pentacoordinate ligand), and diethylenetriamine tetracetic acid (DE4A, a heptacoordinate ligand). Although we have previously reported these sorbents [7, 11], this is the first time they are evaluated for capturing of Co(II) in highly complex matrices with large pH range from pH 0–8, and compared to identify the best material under the given conditions. This is also the first time we report the stability of these materials as a function of solution pHs.

2. Experimental

2.1. Reagents and test matrices

Chromium (Cr³⁺), manganese (Mn²⁺), iron (Fe³⁺), cobalt (Co²⁺), copper (Cu²⁺), selenium (Se⁴⁺), molybdenum (Mo⁶⁺), uranium (U(VI)), thorium (Th(IV)), lead (Pb²⁺), mercury (Hg²⁺), silver (Ag⁺), arsenic (As³⁺), thallium (Tl⁺), and zinc (Zn²⁺) were purchased as standard solutions containing 1,000 or 10,000 mg/L of appropriate element in ~ 2–5% HNO₃ or HCl. ²³⁹Pu(IV) stock was kept in 2.0 M HNO₃ and ²⁴¹Am(III) stock was kept in 0.5 M HCl. Seawater was obtained from the Sequim Bay, Washington, river water was collected from the Columbia River, near Richland, Washington, and ground water was collected from the Hanford site in Richland, Washington. Seawater had a pH of 7.9, salinity of 23%, total organic carbon (TOC) of 1.4 mg/L, and contained Na⁺ and Cl⁻ at about 500 mM, Mg²⁺ at 50 mM, Ca²⁺/K⁺ at 10 mM, SO₄²⁻ at 28 mM and HCO₃⁻ at 2.4 mM. River water had a pH of 7.8, TOC of 5.4 ± 0.2 mg/L and contained 0.9 mM HCO₃⁻ and low concentrations (<0.3 mM) of Na⁺, Cl⁻, Mg²⁺, Ca²⁺, and SO₄²⁻ [28]. Ground water had a pH of 8.1, TOC of 0.5± 0.3 mg/L, and contained low concentrations (<0.3 mM) of Cl⁻, Mg²⁺, K⁺, and SO₄²⁻, 0.4 mM of Ca²⁺, and 2 mM of Na⁺ and HCO₃⁻ [29]. The waters were filtered with 0.45 μm cellulose acetate membranes prior to use. Commercially available reagents of highest purity grade (Aldrich Co) were used throughout this study.

2.2. Sorbents

Synthesis and characterization of SAMMS materials have been described elsewhere, including SAMMS functionalized with three analogs of hydroxypyridinones (1,2-HOPO, 3,2-HOPO, or 3,4-HOPO) [16], acetamide phosphonic acid (Ac-Phos) [30], thiol (SH) [31], and EDTA-analogs (IDAA, DE4A, and ED3A) [11]. The commercial iminodiacetate sorbent, Chelex®−100 Resin, was obtained from Bio-Rad (Hercules, California) and activated carbon (Darco KB-B) was obtained from Sigma Aldrich.

2.3. K_d measurements

The metal sorption performance was evaluated in terms of the distribution coefficient (K_d, mL/g), which is a mass-weighted partition coefficient of the metal ion between solid phase and liquid supernatant phase. We first screened the materials for best capture of actinides from filtered seawater using U(VI) as a model actinide. The pH of the seawater was adjusted to desired pH values (pH 0–8) with HNO₃ and/or NaOH solutions. The seawater was spiked with U(VI) and competing cations (Mn, Fe, Co, Cu, Zn, Se and Mo) to achieve 0.5 mg/L concentration (each). The K_d values were measured in batch experiments with 0.01 gram of sorbent and 10 mL of solution (liquid per solid ratio (L/S) of 1000 mL/g). The suspension was shaken in a 20-mL polypropylene bottle at a speed of 200 rpm for 2 hrs at room temperature. After the batch contacts, the metal-laden sorbents were filtered through 0.22 μm Nylon filter in a polypropylene housing. Both initial (also filtered to eliminate the potential precipitates) and final solutions were analyzed by an inductively coupled plasma-mass spectrometer (ICP-MS, Agilent 7500ce, Agilent Technologies, CA). The detection limits (DL) were 0.09 part per billion (ppb) for Th, 0.02 ppb for U, 0.05 ppb for Co(II), and 3.0 ppb for Si. The measurements were carried out in duplicates or triplicates and the average values were reported. Similar experiments were conducted for Th(IV) under the same conditions.

²³⁹Pu and ²⁴¹Am stocks were diluted in 0.1 M citrate buffer to increase the pH to 2 before spiking to the seawater. Experiments were carried out in the same manner with U(VI) and Th(IV) except that the initial actinide concentration was 500 dpm/mL and the L/S ratio of 5000 mL/g. ²³⁹Pu was analyzed by liquid scintillation counter (Wallac 1414 Liquid Scintillation Counter) and ²⁴¹Am was analyzed using gamma counter (Wallac Wizard 1480 Gamma counter, Perkin-Elmer, Waltham, MA, USA).

Similar K_d measurements were carried out with Co(II) in filtered river water that was adjusted to pH 0–8 and spiked with competing cation metals (Cu, As, Ag, Cd, Hg, Tl, and Pb) to achieve 0.1 mg/L concentration (each) and the L/S ratio was 5000 mg/L. Co(II) concentrations were analyzed by ICP-MS.

2.4. Sorption kinetics

Kinetics experiments were carried out similarly to batch experiments. For U(VI) binding kinetics, a 0.01 g quantity of 3,4-HOPO-SAMMS was dispersed in a 50 mL of filtered seawater containing 0.1 mg/L of U(VI) (L/S of 5000 mL/g). For Co(II) binding kinetics, a 0.05 g quantity of IDAA-SAMMS or Chelex-100 was dispersed in 50 mL of filtered seawater containing 0.1 mg/L of Co(II) (L/S of 1000 mL/g). The two L/S ratios were selected based on the K_d values of the materials for the selected metal ions (higher the K_d values, higher the L/S is). The aliquots were removed at 1, 2, 3, 5, 10, 30, 60 min, 2, 8, and 24 hr, filtered through a 0.22 μm Nylon filters and subjected to ICP-MS analysis along with the initial solution (zero min point).

2.5. Material stability

Along with the K_d measurements, Si (leached out from the sorbent) was measured in the solutions before and after batch contacting with SAMMS. The mg of Si dissolved per gram of material was reported at various solution pHs as the average value of three triplicates.

3. Results

3.1. Screening for most effective SAMMS materials for U(VI) capture.

We screened for the best materials for actinide capture in filtered seawater (pH 0–8) using U(VI) as a representative actinide. It is highly desirable to capture actinides such as uranium from seawater since oceans contain billions of tons of uranium. Understanding how sorbent materials perform in this large pH window (pH 0–8) will have many applications. Actinides normally form insoluble hydroxides and polymeric oxide species in alkaline pHs. Actinides normally form insoluble hydroxides and polymeric oxide species in alkaline pHs and separation of actinides from glass leachates and nuclear waste sludge will require acidification to pH lower than pH 4[16]. To be useful, the sorbent materials must be effective in this low pH range. High ionic strength seawater in the pH range of 0–8 is also a good model for GI tract fluids in our effort to develop safe sorbents for capturing of radionuclides in the GI tract. The dopants (Mn, Fe, Co, Cu, Zn, Se and Mo) were chosen because they are common minerals found in the body and were used at the same trace concentration with that of U(VI). They are also common trace metals in natural waters and wastewaters.

The initial series of experiments were carried out with a suite of SAMMS sorbent materials that have shown promise for actinide binding. The structures of the ligands used to build these sorbent materials are summarized in Figure 1. The seawater was also “doped” with competing cation metals at levels representative of what might be encountered in environmental or physiological samples. Each of the sorbent materials was evaluated by measuring its binding affinity at a series of different pHs. The distribution coefficient (K_d) is a useful measure of a sorbent’s binding affinity for a specific analyte, typically in a diluted condition. K_d is a mass-weighted partition coefficient, as defined in Equation 1 (with units of mL/g):

$${\text{K}}_{\text{d}}=\left[\left({\text{C}}_{\text{i}}-{\text{C}}_{\text{f}}\right)/{\text{C}}_{\text{f}}\right]\times \left(\text{V}/\text{M}\right)$$ Eq. 1

where C_i and C_f are the initial and final concentrations of the analyte respectively (at equilibrium), V is the volume of solution in the experiment and M is the mass of sorbent used. The higher the K_d value, the more effective the sorbent is for capturing that metal ion. Generally speaking, a K_d value of 5,000 and above is considered very good, and a K_d value greater than 50,000 is considered outstanding [32]. The K_d values for U(VI) from pH 0–8 are shown in Figure 2 for 6 functionalized nanoporous sobents and Chelex 100 Resin.

Figure 1.

SAMMS materials having various organic moieties attached on well-ordered mesoporous silica.

Figure 2.

K_d of uranium (U(VI)) on various SAMMS materials and Chelex-100, measured in pH-adjusted seawater. Initial U(VI) of 0.5 mg/L in the presence of competing cations (Cr, Mn, Fe, Co, Cu, Zn, Se and Mo) at 0.5 mg/L (each). L/S of 1000 mL/g.

U(VI) exists mainly as UO₂²⁺ at a highly acidic pH (pH < 3), and mainly as (UO₂)₂CO₃(OH)₃⁻ at pH 3–6 [33]. In seawater (pH 7 and above), UO₂²⁺ forms complexes with calcium and carbonates mainly as Ca₂UO₂(CO₃)₃ and CaUO₂(CO₃)₃² [34]. Ac-Phos-HOPO-SAMMS were shown to have K_d values from 10⁵ – 10⁷ in the near neutral pH range found in most natural waters indicating that these ligands on SAMMS can displace the carbonate ligands from the uranium center. This level of performance, in high ionic strength of seawater and a large excess of competing species, clearly indicates the high degree of selectivity that these SAMMS sorbents have for the actinides over the more ubiquitous metal cations and those doped into the solutions. While the HOPO ligands were found to be generally superior to other ligands over a pH range of 3–8, their binding affinity dropped off markedly at below pH 2 due to proton competition and pH above 8 due to some insolubility of U(VI). Note that the K_d values of 10⁵–10⁷ mean that 99.5–100% of U(VI) was removed from the solutions and the drop in K_d values at pH above 8 was due to lower concentration of U(VI) in the initial solution after the U(VI) precipitates was filtered out by the 0.22 μM filters. The acidity of HOPO ligands ranges from 1,2-HOPO > 3,2-HOPO > 3,4-HOPO (pKa of 5.78, 8.66, and 9.01, respectively [35]), which corresponds to the affinity for U(VI) that ranged from 1,2-HOPO > 3,2-HOPO > 3,4-HOPO at pH <1. However, at pH < 1, Ac-Phos-SAMMS that contains a carbamoylmethylphosphine oxide (CMPO) based ligand which effectively coordinate actinides with carbonyl and phosphoryl oxygen atoms even in nitric solution [36], had the highest affinity among all sorbents tested (K_d of 5×10⁴).

SAMMS with glycine derived ligands performed moderately (K_d of 10⁴–10⁵) at pH near neutral range because they are weaker than Ac-Phos and HOPO at competing with carbonate species for uranium. The DE4A is an analog of diethylenetriamine pentaacetate (DTPA), suggesting that the DTPA chelating agent used in clinics is still suboptimal for actinide decorporation. Included in this plot are the K_d values obtained using Chelex-100 (a cross-linked polystyrene resin with pendant iminodiacetic acid groups), which was less effective than Ac-Phos and HOPO-SAMMS at capturing U(VI).

3.2. Effectiveness of 3,4-HOPO-SAMMS at capturing other actinides

As can be observed in Figure 2, 3,4-HOPO-SAMMS had the highest K_d in neutral pH range and consequently was selected for further study in filtered seawater with other actinide cations, specifically Th(IV), ²⁴¹Am(III) and ²³⁹Pu(IV). Th(IV) can form weak complexes with Cl⁻ and NO₃²⁻ at low pH [37]. In seawater with high salinity and low contents of organic species, hydrolysis of Th(IV) forms Th(OH)₄⁰ as the most prevalent species [38]. At high pH and in the presence of carbonate, Th(IV) may also exist as Th(CO₃)₅⁶⁻ [39]. Th(IV) is also known to form strong complexes with organic species in seawater [40]. In seawater, Am(III) exists as Am(III) as a predominant species at acidic pH, AmSO₄⁺ at pH 7.0, Am(CO₃)⁺ at pH 7.8, and Am(CO₃)₂⁻ and Am(OH)₂⁺ at pH 9.0 [41]. As reviewed by Choppin and Morgenstern [42, 43], in natural water, Pu can exhibits four valences (III, IV, V, and VI); low valence stage (Pu(III) and Pu(IV)) at low pH and higher valence stages (e.g., Pu(IV) and Pu(V)) at neutral and higher pH. They suggest that Pu(III) is readily oxidized to Pu(IV) in oxic water, and hydrolyzed Pu(IV) can bind to colloidal and suspended materials in natural water and form less soluble Pu(IV)(OH)₄. In addition, Pu(V) as PuO₂⁺ is the most stable oxidation state in sea water and oxic water, while Pu(VI) as PuO₂²+ is easily reduced to PuO₂⁺.

Figure 3 shows the K_d values of 3,4-HOPO-SAMMS for the four actinides from pH 0 to 8. Figure 3 shows that K_d values for Pu were generally in between 10⁴–10⁵ in the pH range of 1.9–7. The 3,4-HOPO-SAMMS bound Th(IV) across a wide range of pH values with K_d of 3×10⁶, but was ineffective under highly acidic conditions (pH ~ 0). ²⁴¹Am(III) was not captured as effectively as the other three actinides, especially at pHs below 4. Am(III) is the “softest” of the typical actinide cations. The HOPO ligands are fairly “hard” ligands for binding hard Pu(IV) and U(VI). The electronic mismatch between Am(III) and 3,4-HOPO ligand led to lower binding affinity and inability to compete with ligand protonation under the more acidic conditions (H⁺ is a very “hard” acid and hence out competes Am(III) for the 3,4-HOPO ligand).

Figure 3.

K_d of thorium (Th(IV)) uranium (U(VI)), americium (²⁴¹Am(III)), and plutonium (²³⁹Pu(IV)) on 3,4-HOPO-SAMMS measured in pH-adjusted seawater. Initial ²⁴¹Am or ²³⁹Pu of 500 dpm/mL (measured separately). Initial U and Th of 0.5 mg/L (measured together). L/S of 5000 mL/g (for Am and Pu) or 1000 mL/g (for U and Th).

The K_d of 3,4-HOPO-SAMMS for all actinides dropped sharply at pH 0 and at pH > 7 due to protonation of the 3,4-HOPO ligand-complex at very low pH and the poor solubility of actinides at very high pH, respectively. Actinides tend to form strong complexes with carbonate and hydroxide species under basic conditions [44] reducing the effectiveness of all solid phase sorbent materials. In any event, the 3,4-HOPO-SAMMS appears to be exceptionally good for capturing U(VI) and Th(IV) from seawater over a broad range of pH from 2 to 7, and very good for Pu over a similar range of pH values. Under near-neutral conditions, 3,4-HOPO SAMMS produced K_d values of ~10⁷ for U(VI), ~10⁶ for Th(IV), ~10⁵ for Am(III) and ~10⁵ for Pu(IV).

3.3. Rapid capture of U(VI) by 3,4-HOPO-SAMMS

The 3,4-HOPO SAMMS captured U(VI) from seawater rapidly as shown in Figure 4. Equilibrium was achieved very quickly, with >99% of the U(VI) in solution being captured in less than 5 minutes at a L/S ratio of 5000. This is attributed to the rigid and well-ordered porous structure of the MCM-41 (Mobil Composition of Matter No. 41) substrate, allowing easy transport of U(VI) to the HOPO binding sites and the high affinity of the HOPO to U(VI).

Figure 4.

Sorption kinetics of U(VI) on 3,4-HOPO-SAMMS measured in seawater, pH 7.7, at L/S of 5000 mL/g.

3.4. Th(IV) capture by 1,2-HOPO-SAMMS versus 3,4-HOPO-SAMMS

In addition to 3,4-HOPO SAMMS, 1,2-HOPO-SAMMS bound U(VI) quite well, across the entire range of pH values studied (see Figure 2). Being a more acidic ligand, 1,2-HOPO-SAMMS was better than 3,2-HOPO-SAMMS and 3,4-HOPO-SAMMS at capturing U(VI) at pH around 1.0 and below. In order to see if this held true for Th(IV), we performed a series of experiments using 1,2-HOPO-SAMMS to bind Th(IV) across the pH ranging from 0 to 7 as shown in Figure 5. When compared with 3,4-HOPO-SAMMS which had no Th(IV) affinity at pH ~ 0, the more acidic 1,2-HOPO-SAMMS retained high binding affinity across the entire pH range (K_d of 3×10⁶ from pH 0 to 6.7). At above pH 7, Th(IV) has poorer solubility resulting in lower K_d values.

Figure 5.

K_d of Th(IV) on 1,2-HOPO-SAMMS and 3,4-HOPO-SAMMS measured in pH-adjusted seawater. Initial Th(IV) of 0.5 mg/L and L/S of 1000 mL/g.

3.5. Co(II) capture with SAMMS functionalized with glycine derivatives

Figure 6 shows the K_d of Co(II) on SAMMS sorbents in filtered river water that had been doped with 0.1 mg/L of Co(II) and a variety of competing ions (e.g. Cu, As, Ag, Cd, Hg, Tl, and Pb of the same concentration). These competing metals at the same trace concentration with that of Co(II) were selected because they are considered important contaminants in natural waters and may have affinity for the glycine derivatives. Cobalt exists at Co(II) in natural waters as free species and can form a complex with Cl⁻ in seawater or with carbonate in river water at high pH [45, 46]. As can be seen from the data in Figure 6, between a pH of about 5 and 8.5, the three glycine derivatives (IDAA, DE4A, ED3A) all displayed similar behavior, with a high binding affinity for Co(II) (K_d of 10⁵ – 10⁶). Capture of Co(II) is by amino groups (pKa of amino groups ~ 9.6 [47] and the pKa of IDAA was determined to be 5.66 [48], which should be similar to that of DE4A and ED3A). Not surprisingly, below a pH of ~4 the binding affinity for Co(II) dropped off markedly due to protonation of the ligands (typical for this class of ligand [11]). Thiols and Ac-Phos, which are well-suited for soft heavy metals and rare earth cations, respectively, were ineffective for capturing Co(II) from the river water (K_d values from 10²–10³). Table 1 summarizes the K_d values of SAMMS materials versus commercially available sorbents for metal cations such as Chelex-100 and Darco activated carbon in river, ground and sea water. The three glycine based SAMMS (IDAA, DE4A, ED3A) outcompeted the two commercial sorbents, especially in challenging matrices like ground water with high bicarbonate and sea water with high chloride content. Among the three SAMMS, IDAA-SAMMS performed as well as DE4A-SAMMS, and both are better than ED3A-SAMMS, which agrees with Figure 6 (measured in river water).

Figure 6.

K_d of Co(II) as a function of pH (measured in pH adjusted river water) on various sorbent materials. Initial Co(II) concentration of 0.1 mg/L in the presence of competing cations including Cu, As, Ag, Cd, Hg, Tl, and Pb at 0.1 mg/L (each). L/S of 5000 mL/g.

Table 1. K_d of various sorbents for Co(II) measured in river water, ground water, and sea water.

Initial Co(II) of 0.1 mg/L in the presence of competing cations including Cu, As, Ag, Cd, Hg, Tl, and Pb at 0.1 mg/L (each). L/S of 5000 mL/g.

Sorbent	River water (pH 7.75)	Ground water (pH 8.06)	Sea water (pH 7.82)
IDAA-SAMMS	730000	560000	800000
DE4A-SAMMS	620000	580000	510000
ED3A-SAMMS	250000	210000	230000
Chelex-100	360000	92000	28000
Darco AC	1400	1500	100

3.6. Rapid capture of Co(II) by IDAA-SAMMS

Due to the similarity of IDAA and DE4A-SAMMS sorbents in the near neutral pH region, we chose IDAA-SAMMS to represent this group of sorbent materials for a comparison of the sorption kinetics to those of Chelex-100. The data are shown in Figure 7. Both sorbents were capable of capturing >99% of the Co(II) in filtered seawater (i.e., in 5 minutes by IDAA-SAMMS and 10 minutes by Chelex-100). However, the initial Co(II) sorption with IDAA-SAMMS was notably faster than with the Chelex-100 resins containing IDAA ligands (i.e., 95% of Co(II) was captured by IDAA-SAMMS vs. 62% by Chelex-100 in 1 minute). The rigid structure of SAMMS supports faster metal transport through the pores than polymeric Chelex-100, which may suffer from solvent swelling and shrinking at high ionic strength conditions. Similar rapid kinetics has been found with other SAMMS materials such as thiol-SAMMS for capturing lead [9], copper ferrocyanide-SAMMS for capturing cesium [20], iron-ethylenediamine-SAMMS for capturing phosphate [18], ethylenediamine-SAMMS for capturing copper [10].

Figure 7.

Sorption kinetics of Co(II) on IDAA-SAMMS and Chelex-100 measured in seawater, pH 7.8, at L/S of 1000 mL/g.

3.7. Stability of SAMMS materials vs. solution pH

Sorbent stability is an important issue for chemical separations. A series of experiments were carried out to evaluate the chemical stability of these SAMMS sorbents in pH adjusted seawater. Figure 8 shows the stability (as measured in terms of dissolved Si in the supernatant per mass of sorbent) of five SAMMS materials after being vigorously stirred in pH-adjusted seawater (pH 0–8.1) for 2 hours. The Si may come from both organosilane layer and very fine silica particulates that passed through the 0.22 micron filters. Figure 8 shows that most of the sorbents studied demonstrated good stability across pH 2–7.5, typically experiencing less than 2% Si leaching from the sorbents. Most notable was the 3,4-HOPO-SAMMS, which displayed well below 0.4% Si leaching across the entire pH range [13]. Although Si leaching of IDAA-SAMMS was <1% at pH 3.9 and above, the leaching was increased to 2.6% at pH 2.2 and 16% at pH 0.11. It is worth noting that the IDAA headgroup contains a basic N atom, and such functionality is known to interact strongly with the surface silanols (on the silica support structure) and can introduce defects into the monolayer structure [49] by loosely bonding to the surface in and inverted configuration and not a strong silane grafting (with the IDAA moiety away from the surface). Under acidic conditions, this amine ion pair will be protonated, thereby releasing the ligand and exposing these defect sites, rendering the adjacent silanes more vulnerable to hydrolysis, and ultimately silane loss. Similar behavior was found with DE4A-SAMMS but to a lesser degree (8% leachate at pH 0.0). Stability for this class of materials at low pH could be improved with synthetic processing conditions that release the amines bound to the silica surface (with a protonating acid wash) and filled the holes in the monolayer with organosilane (with or without and active ligand headgroup). Clearly, the stability of SAMMS sorbents is related to the chemistry of the head groups and processing conditions. Nevertheless, the majority of these sorbents display very good stability in seawater and across a wide range of pH values from 2 to 7.5.

Figure 8.

Leachate of various SAMMS materials measured in pH-adjusted seawater.

4. Conclusions

This work has demonstrated that chemically modified nanoporous sorbent materials (SAMMS) are effective at capture of selected actinides and cobalt from natural waters. Most of the work was done in seawater, which is challenging due to its high ionic strength. While performance is strongly dependent upon surface chemistry, the SAMMS sorbents were shown to be effective across a wide range of pH. The high ionic strength, and presence of competing ions, did not prevent effective radionuclide capture. In the case of U, a number of ligands were found to be effective, with many K_d values in excess of 10⁶. Generally speaking the best ligand was the 3,4-HOPO ligand over a pH range from 7.5 down to 2.1. Under highly acidic conditions (pH < 1), Ac-Phos- and 1,2-HOPO-SAMMS were found to be more effective for U capture. 3,4-HOPO SAMMS was also found to have good affinity for Th(IV) and Pu(IV) across a wide range of pH (2–8), as well as being effective for binding Am(III) between a narrower pH window from about 5 and 8. At pH < 1, only 1,2-HOPO-SAMMS and not 3,4-HOPO-SAMMS retained high binding affinity for Th(IV). Sorption kinetics for U(VI) in filtered seawater were found to be rapid, with >99% of the U being captured in less than 5 minutes.

All three of the chelating glycine derivatives tested (e.g. IDAA, ED3A, and DE4A) were found to be effective for capturing Co(II) from filtered seawater between a pH of ~5 and ~8. Kinetics were found to be faster than the Chelex-100 resin, with equilibrium being achieved in a minute. Not surprisingly, below a pH of ~4 these carboxylic acid ligands were found to be ineffective due to ligand protonation from solution. Ligands tailored for other classes of metal ions (e.g. thiols, Ac-Phos) were also found to be ineffective for binding Co(II) from filtered seawater.

The physical stability of these sorbents was generally found to be good to excellent across a wide pH range (0–8), with minimal Si leaching (typically < 2 wt.%) taking place during the batch contact experiments. The one exception to this observation was the IDAA-SAMMS at highly acidic conditions because it has a basic N atom that is known to interact with the silica surface. SAMMS sorbents are highly scalable and our findings demonstrate that they have great potential for capturing actinides and cobalt from complex real-world matrices such as natural waters (sea, river, ground waters) and nuclear waste water and may be used as radionuclide decorporation in the gastrointestinal tract.

Highlights:

• 3,4-HOPO-SAMMS was best at capturing uranium from seawater from pH 2–8

• Ac-Phos-SAMMS was best at capturing uranium at below pH 2

• 3,4-HOPO-SAMMS was effective at capturing Th and Pu from pH 2–8, and americium from pH 5–8

• IDAA-SAMMS was most effective at capturing Co from river water from pH 5–8

• Over 99% of U and Co was captured by 3,4-HOPO- and IDAA-SAMMS from seawater within 5 minutes