Iron oxide waste form for stabilizing 99Tc

Journal of Nuclear Materials 429 (2012) 201–209

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Iron oxide waste form for stabilizing

99

Tc

Wooyong Um a,⇑, Hyunshik Chang a,1, Jonathan P. Icenhower b, Wayne W. Lukens b, R. Jeffrey Serne a, Nik Qafoku a, Ravi K. Kukkadapu a, Joseph H. Westsik Jr. a a b

Pacific Northwest National Laboratory, USA Lawrence Berkeley National Laboratory, USA

a r t i c l e

i n f o

Article history: Received 23 February 2012 Accepted 1 June 2012 Available online 3 May 2012

a b s t r a c t Crystals of goethite were synthesized with reduced technetium [99Tc(IV)] incorporated within the solid lattice. The presence of 99Tc(IV) as a substituting cation in the matrix and ‘‘armoring’’ by an additional layer of precipitated goethite isolated the reduced 99Tc(IV) from oxidizing agents. These products were used to make monolithic pellets to quantify an effective diffusion coefficient for 99Tc from goethite waste form contacted with a synthetic Hanford IDF (Integrated Disposal Facility) pore water solution (pH = 7.2 and I = 0.05 M) at room temperature for up to 120 days in static reactors. XANES analysis of the goethite solids recovered post-run demonstrated that the 99Tc in the goethite crystals remains in the reduced 99 Tc(IV) state. The slow release of pertechnetate concentration with time in the static experiments with the monolith followed a square root of time dependence, consistent with diffusion control for 99Tc release. An apparent diffusion coefficient of 6.15  1011 cm2/s was calculated for the 99Tc–goethite pellet sample and the corresponding leaching index (LI) was 10.2. The results of this study indicate that technetium can be immobilized in a stable, low-cost Fe oxide matrix that is easy to fabricate and these findings can be useful in designing long-term solutions for nuclear waste disposal. Published by Elsevier B.V.

1. Introduction Neutron-induced fission of 235U-enriched nuclear fuel yields technetium isotopes in relatively large amounts, the most important of which is technetium-99 (99Tc) [1–3]. Approximately 1 kg of 99Tc with a long half-life (2.13  105 years) is produced for every ton of nuclear fuel ‘‘burned’’ in a typical reactor [4]. The radiological and chemical properties of 99Tc present some unique problems for nuclear waste disposal. In virtually all near surface conditions, 99 Tc exists in the highly soluble and mobile pertechnetate form ½TcðVIIÞO 4  [5,6]. Environmental concerns have been raised because of the long half-life and high mobility of 99Tc in oxidizing subsurface environments [7]. The highly soluble pertechnetate oxyanion, 99 TcðVIIÞO 4 , does not sorb onto most terrestrial sediments [8], so 99 TcO 4 migrates at nearly the same velocity as groundwater [9] under common subsurface conditions (i.e., pH close to neutral or slightly alkaline and suboxic conditions). In the natural environment, there may be cases in which pertechnetate may react in locally reducing conditions, caused by microbial activity or with reduced inorganic metal phases, primarily magnetite and sulfides,

⇑ Corresponding author. Address: Pacific Northwest National Laboratory, PO Box 999, P7-54, 902, Battelle Boulevard, Richland, WA 99354, USA. Tel.: +1 509 372 6227; fax: +1 509 371 6919. E-mail address: [email protected] (W. Um). 1 Present address: Savannah River Ecology Laboratory, P.O. Drawer E, Aiken, SC 29802, USA. 0022-3115/$ - see front matter Published by Elsevier B.V. http://dx.doi.org/10.1016/j.jnucmat.2012.06.004

that will induce reduction and retard the mobility of 99Tc. Under reducing conditions, pertechnetate can precipitate as 99 Tc(IV)O22H2O [10,11], sorb to mineral phases [7], and be retained in different natural environments [12–14]. However, even when reduced to more insoluble forms, the reoxidation of 99Tc(IV) by changing redox conditions, such as contact with oxygen, can result in release of pertechnetate back into the environment [14], leading to the prediction of high 99Tc release rates in many performance assessments [15–17]. For example, Lee and Bondietti [10] found that oxidation of an FeS/Tc(IV) precipitate allowed 70% of the 99Tc to return to solution over a period of 8 months, which could be 99 due to reoxidation back to 99 TcO Tc(IV) 4 , or to solubilization of species at the low pH associated with oxidation of the sulfide [18]. Further, the long half-life of 99Tc provides ample opportunity for 99 Tc(IV) to reoxidize and frustrates efforts to fashion a long-term stable immobilization matrix. To prevent 99Tc release into the environment, plans are being drawn to sequester technetium into a stable waste form that will be inert to dissolution and oxidation reactions. Numerous waste forms have been proposed, including cement/grout [16,19], glass [20], hydroceramics [21,22], phosphate-bonded ceramics [21,23], and metal alloys [24]. Evaluating the efficacy of these proposed waste forms is difficult, because data that bear on the long-term immobilization of 99Tc are sparse. Short-term laboratory tests have, however, underscored a potential pitfall for many of these materials: 99Tc reduced to the more immobile 99Tc(IV) form is vulnerable to the infiltration of oxidizing agents, especially oxygen.

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Upon contact with oxygen, reduced 99Tc(IV) has been found to reoxidize rapidly to pertechnetate [16,18]. Only the Savannah River Site ‘‘saltstone’’ [17] has been shown to hinder reoxidation of 99 Tc(IV), but mainly because of slow diffusion of oxygen into the immense size of waste form slab, and not because of any inherent reoxidation resistant qualities of the material. Thus, formulating a waste form that isolates 99Tc(IV) from contact with O2 is key to the success of any proposed immobilization strategy. Several previous studies have shown that when 99Tc(IV) is substituted into a mineral lattice, technetium is effective sealed from oxygen. For example, investigations into the interaction of pertechnetate with Fe (oxy) hydroxide and sulfide minerals have shown that once 99Tc(VII) is reduced, the association of 99Tc(IV) with iron is strong and limited reoxidation of 99Tc(IV) is found because direct substitution of 99Tc(IV) for Fe in the Fe oxide structure is possible during precipitation and crystal growth reactions [25– 28]. With these studies as a guide, we recently reported successful experiments in which pertechnetate was reduced through interaction of Fe2+ with the oxy-hydroxide phase, goethite [a-FeO(III)OH]. We demonstrated that 99Tc(IV) was incorporated within the goethite mineral lattice during mineral growth and following additional goethite overgrowth or ‘‘armoring’’ limited reoxidation of 99 Tc(IV), despite exposure of 99Tc(IV)-bearing goethite was to solutions in ambient oxidizing conditions [28]. There are additional reasons why goethite-based waste forms should be considered for immobilizing technetium. As an endstage weathering product in temperate environments, goethite is an extremely stable mineral phase at or near the Earth’s surface [29]. Previous studies have shown that hydrogen and oxygen atoms in the goethite structure are resistant to exchange with the environment for up to 50 Myr [30], attesting to the thermodynamic and kinetic stability of goethite. In addition, goethite can be readily synthesized in the laboratory with particle sizes and surface areas that can be manipulated readily [31], processes that can be easily and cheaply up-scaled to industrial levels. Furthermore, iron in goethite is in the stable trivalent (III) state so that in the oxidizing environments that will likely prevail in the disposal repository setting, there will be minimal chemical potential for redox reactions that may affect greater mobility of 99Tc(IV). Therefore, a waste form based on goethite appears to have the potential to be a simple, durable, cost-efficient, and effective candidate for disposing of 99Tc from radioactive wastes. Despite the arguments marshaled above and from our previous study on a goethite-based waste form, we still lack quantitative data on the release of technetium to the environment. Our previous work demonstrated that the release of 99Tc to the environment due to dissolution of goethite is very slow under a range of geochemical conditions, and we postulated that the mechanism of release is transport limited in circum-neutral solutions. Diffusion of 99Tc(VII) through the goethite lattice must be preceded by the incursion of oxygen. Isolation of 99Tc(IV) from oxidizing agents appears to be key to its immobilization, as has been argued by previous investigators [32]. Accordingly, the experiments described herein investigated 99Tc diffusion coefficient from the 99Tc–goethite waste form in batch experiments, and evaluated mineralogical changes and reoxidation of 99Tc(IV) using macroscopic and spectroscopic analyses.

2. Material and methods 2.1. Synthesis of iron oxides Goethite was synthesized based on a scaled-down procedure of Schwertmann and Cornell [31]. To summarize, ferric nitrate [11.4 g of Fe(NO3)39H2O] was dissolved into NANOpureÒ water (100 mL)

and reacted with 2-M sodium hydroxide (NaOH) (150 mL). The slurry was heated in an oven (80 °C) for 7 days. The solid product was filtered from solution by vacuum filtration, washed with fresh NANOpureÒ water two times, and air dried overnight. The air-dried solid was then gently crushed to a powder form. In addition to goethite, two-line ferrihydrite (Fe5HO84H2O) was also synthesized using a modified method of Schwertmann and Cornell [31]. The synthesis method for ferrihydrite was similar to that of goethite without the heating step. Briefly, ferric nitrate [8.0 g of Fe(NO3)39H2O] was dissolved into NANOpureÒ water (100 mL) in a polyethylene bottle (250 mL); 1 M NaOH was added dropwise while stirring the slurry, until a pH of 6 was obtained. To obtain a pH of 7–8, additional low-concentration NaOH (i.e., 0.01 M) was added while continuously stirring. The final solid product was filtered from solution by vacuum filtration, washed with fresh NANOpureÒ water four times, and air dried overnight before use.

2.2. Technetium removal by Fe(II)-iron oxides Batch experiments to determine the extent of Fe(II) adsorption onto either goethite or ferrihydrite and the subsequent ability of the iron oxides to sequester 99Tc were conducted at different pH levels. These experiments were conducted to study the effect of pH on dissolved Fe(II) concentration and subsequent 99Tc(VII) removal by the Fe(II)–goethite or –ferrihydrite solids. Some of the pH variation tests were performed with only Fe(II) in solution (i.e., no goethite or ferrihydrite solid present) to determine empirically the solubility of Fe(II/III) oxyhydroxides. Solubility diagrams for Fe(OH)3(s) and Fe(OH)2(s) were also constructed based on the stability constants of Fe(II)/Fe(III) hydrolysis species [33] and compared with measured Fe(II) concentrations in the effluent solutions from the various batch tests. At each sub-sampling step, a small aliquot of filtered solution was used to measure the pH and the concentrations of 99Tc(VII), Fe(II) and total Fe. Concentrations of 99 Tc(VII) and total Fe in the supernatants were determined using ICP-MS and ICP-OES, respectively. The dissolved ferrous Fe(II) concentration was determined using the ferrozine colorimetric method [34] with a HACH DR/890 colorimeter. Details of the synthesis procedures and the strategy for incorporating reduced 99Tc(IV) into goethite can be found in the previous paper (similar to Sample 2–5 preparation) [28]. Even though our previous study worked for waste form development using both off-gas secondary waste stream [28] and simple de-ionized (DI) water conditions, this study focused more for 99Tc diffusive leachability from Tc–goethite waste form prepared in simple DI solution. Succinctly, synthesized goethite powder (2.75 g) was re-suspended in de-aerated and DI water (250 mL) and the pH was lowered to 62.0 by adding 2 M nitric acid. Powdered FeCl24H2O (3.48 g) was directly added to the goethite slurry as the Fe(II) source and reacted for 1 day in an anoxic chamber (Coy Laboratory Products) equipped with a H2/O2 gas analyzer and palladium-coated alumina catalyst. A mixture of N2 (97%) and H2 (3%) was used as the anaerobic gas in the chamber. Following this, 0.25 mL of 99Tc(VII) from a NaTcO4 standard solution (2.2  102 M) was prepared in Pacific Northwest National Laboratory (PNNL) using NaTcO4 salt and added to make a total 2.2  105 M of 99Tc in the Fe(II)–goethite slurry (250 mL) and homogenized for 1–2 days on a platform shaker. After mixing, sodium hydroxide (150 mL of 2 M NaOH) was added to facilitate 99 Tc–goethite precipitates. The 99Tc–goethite solids were further modified to armor the 99Tc–goethite solids with additional goethite precipitates using separately prepared Fe(NO3)39H2O (11.4 g/ 100 mL) solution. After 1–2 days of reaction with added ferric nitrate and sodium hydroxide solutions, the bottle containing the final slurry was removed from the anaerobic chamber and placed inside an oven at 80 °C for 7 days. The final 99Tc–goethite solid products

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were separated by filtration, dried in air, and used for additional analysis. 2.3. Solid phase characterization The initial goethite and ferrihydrite substrates and the final solid product before and after diffusion test were first characterized using a Scintag XRD unit equipped with a Cu K-alpha radiation (40 kV, 35 mA) source. The bulk 99Tc–goethite samples were homogenized by grinding in an agate mortar and pestle, and mounted into a small circular sample holder before scanning from 2 to 75 degrees 2-theta. Data reduction and phase identification were done by JADE software with PDF X-ray diffraction (XRD) database. Acid extraction, using 8 M HNO3 with heat at 90 °C was used to determine the total 99Tc in the final 99Tc–goethite solid. Transmission electron microscopy (TEM) samples were prepared by dispersing a small amount of the 99Tc-laden Fe(II)–goethite slurry in methanol and depositing this onto a lacey carbon TEM copper (Cu) grid. TEM characterization was performed using a FEI Tecnai T30 operated at 300 keV equipped with a Gatan ORIUS digital camera. Analysis was also performed by identifying the mineralogy with selected area electron diffraction (SAED). The Mössbauer sample was prepared by mixing a dried 99Tc– goethite sample with Vaseline in a Cu holder sealed at one end with clear tape. After mixing the sample in the holder, the opened end of the holder was sealed with clear tape. The Mössbauer spectra were collected at room temperature (RT; data not shown) and liquid nitrogen (77 K) using a 50-mCi (initial strength) 57Co/Rh single-line thin source. The velocity transducer, MVT-1000 (WissEL) was operated in a constant acceleration mode (23 Hz, 5± or ±12 mm/s). An Ar–Kr proportional counter was used to detect the radiation transmitted through the holder, and the counts were stored in a multichannel scalar as a function of energy (transducer velocity) using a 1024-channel analyzer. Data were folded to 512 channels to give a flat background and a zero-velocity position corresponding to the center shift (CSd) of a metal Fe foil at room temperature (RT). Calibration spectra were obtained with a 25-lm-thick metal Fe foil (Amersham, England) placed in the same position as the samples to minimize any errors due to changes in geometry. A closed-cycle cryostat (ARS, Allentown, Pennsylvania) was used for 77 K measurement. The Mössbauer data were modeled with the Recoil software (University of Ottawa, Canada) using a Voigt based structural fitting routine. The coefficient of variation of the spectral areas of the individual sites generally ranged between 1% and 2% of the fitted values. 2.4. X-ray Absorption Near Edge Structure (XANES) spectroscopy Solid standards of KTcO4, NaTcO4, TcO 4 adsorbed on ReillexHPQ resin and TcO22H2O, and final 99Tc–goethite sample before and after diffusion tests were analyzed to determine the 99Tc oxidation state change. The XANES spectra were collected on beamline 4-1 at the Stanford Synchrotron Radiation Laboratory (SSRL). The solid samples were mounted on Teflon sample holders and sealed with Kapton tape. A Si(2 2 0) double-flat crystal monochromator was used and the energy was calibrated by using the first inflection point of the 99Tc K edge spectrum of the 99Tc(VII) standard, defined as 21.044 keV. The 99Tc-standards and 99Tc–goethite spectra were collected in transmission and fluorescence mode, respectively, at room temperature using a 13-element germanium detector. Data reduction and analysis were performed using the software IFEFFIT [35] and ATHENA/ARTEMIS [36] after correction for detector dead-time. The XANES spectra for the 99Tc–goethite samples were fit using a linear combination of the XANES spectra of 99Tc(IV) and 99Tc(VII) standards collected by Lukens et al. [11].

2.5. Effective diffusion coefficient measurement for

99

Tc

The ground-powder 99Tc–goethite sample was compacted into cylindrical monolithic pellets by mixing with dissolved polyethylene glycol organic binder at about 3 wt%, and was prepared for diffusion leach testing. The 99Tc–goethite–organic binder slurry was dried overnight in a heated bath at 80 °C and then pressed with up to 2500 kg of load force in a 1.3-cm-diameter die with a Carver Press, giving a resultant pressure of 150 MPa. A minor amount of oleic acid was used on the final monolithic pellet surfaces to aid in preserving the pellet integrity. The resultant monolithic pellets containing the 99Tc goethite solids were used to determine the effective diffusion coefficient for 99Tc leached from the monolith immersed in Integrated Disposal Facility (IDF) pore water solution (500 mL) in the Hanford Site 200 East Area at room temperature [37]. The IDF pore water solution is germane to the expected aqueous solution compositions coursing through the solidified radioactive wastes disposal environment. The chemicals used to prepare the IDF pore water leaching solution are listed in Table 1. Photographs of the monolithic pellets with 99Tc–goethite with a diameter of 1.3 cm and a height of 0.32 cm before and after the diffusion leach testing are shown in Fig. 1. For the monolith diffusion tests, a subsample (1 mL) of leachate was periodically collected from 30 min after the test commenced to 120 days through the leaching period using a 0.45-lm Nalgene syringe filter and submitted for analyses of dissolved Fe(tot) and 99 Tc. The pH was directly measured in the slurry solution when each of the leachate samples was collected. Analysis methods for Fe(tot) and 99Tc were the same as described above. After the diffusion tests were completed, the monolithic 99Tc–goethite sample was collected and prepared for characterization of the solids. Because goethite has a low solubility at neutral pH values similar to that of the IDF pore water (see the leaching results below), the contaminant-specific, effective diffusion coefficient is commonly used to quantify the migration rate of 99Tc out of the monolithic waste form. The observed effective diffusivity for 99Tc out of the monolithic waste form was calculated using the analytical solution, Eq. (1), for simple radial diffusion from a cylindrical monolithic 99Tc–goethite pellet as presented by Crank [38]:

Di ¼ p



2 Mt pffiffiffiffi i pffiffiffiffiffiffiffiffi 2qC o ð t i  t i1 Þ

ð1Þ

where Di is the observed diffusivity of 99Tc for the leaching interval, i [m2/s]; Mti the Tc mass released during the leaching interval i [mg/ m2]; ti the cumulative contact time after the leaching interval, i [s]; ti1 the cumulative contact time after the leaching interval, i  1 [s]; Co the initial leachable 99Tc content [mg/Kg] in a monolithic pellet; and q is the pellet density [Kg-dry/m3]. The mean observed 99Tc diffusivity was determined by taking the average of each of the interval-observed diffusivities.

Table 1 Chemical composition of the solution and the list of ingredients used to prepare the simulated IDF pore water from the Hanford 200-East Area. Chemicals

Concentration (M)

CaSO4 NaNO3 NaHCO3 NaCl MgSO4 MgCl2 KCl Ionic strength pH Alkalinity

1.2  102 3.4  103 3.0  104 2.1  103 2.6  103 2.4  103 7.0  104 0.05 M 7.2 29 mg/L

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Fig. 1. Cylindrical monolithic pellets of 99Tc–goethite samples. (Left) pellet sample before diffusion leach testing by IDF pore water; (right) pellet sample after diffusion leach testing for 120 days by IDF pore water. The difference in color is from the wetted pellet after leaching. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

The leachability index (LI), which is a parameter derived directly from diffusion test results, is also used to evaluate diffusion-controlled 99Tc release with respect to time. The LI of 99Tc is calculated with:

  Dn LIn ¼ log cm2 s

ð2Þ

where LI is the leachability index and Dn is the effective diffusivity for 99Tc (cm2/s) during the leach interval n. Because the LI is the negative logarithm of the measured effective diffusivity of 99Tc, a larger LI value indicates that less 99Tc has leached during the diffusion test. 3. Results and discussion 3.1. 99Tc(VII) and Fe(II) removal in solutions containing no Fe oxide solid Changes in dissolved Fe(II) and 99Tc concentrations in the final supernatant solution were investigated as a function of pH without the presence of Fe-oxide minerals. Dissolved Fe(II) and Fe(tot) concentrations in solution without 99Tc(VII) addition were similar to each other and stable (remained at the initial concentrations) when pH was below 4.5, thereby indicating that most of the Fe in solution existed as ferrous iron, Fe(II) (Fig. 2, left). However, the dissolved Fe(II) concentration decreased with increasing pH (>5.0) and approached zero in solutions at pH higher than 8.0. Total

dissolved Fe concentrations also showed the same decreasing trend with increasing pH and no detectable Fe(II) and Fe(tot) concentrations were found in solution containing no 99Tc(VII) at pH higher than 8.0. There were minor differences between total Fe and Fe(II) concentrations in solutions containing no 99Tc(VII) when the pH was below 4.5, suggesting that minor amounts of oxidized Fe(III) could be present even in these more acidic solutions. All the experiments were conducted inside an anaerobic chamber that was monitored for oxygen concentrations. Despite these precautions, small amounts of oxygen could be present in the chamber during the experiments; thus, one cannot eliminate the possibility that small amounts of Fe(II) were oxidized to Fe(III). Nearly complete removal (100%) of dissolved Fe(II) at pH values higher than 7.0 was attributed to precipitation of Fe(II) hydroxide as Fe(OH)2(s) (white rust) and magnetite. Independently prepared Fe(II) precipitate at pH = 7.5 from a similar solution composition showed an identical XRD pattern to crystalline Fe(OH)2(s) [28]. Dissolved Fe(II) concentrations were slightly different in the absence or presence of 99Tc(VII) in solution. Lower concentrations of dissolved Fe(II) than Fe(tot) were found even at low pH values (3.2 and 4.1) when 99Tc(VII) was present (Fig. 2, left), in comparison to the same acidic solutions containing Fe(II) only below pH 5.0. When 99Tc(VII) was not present in solution, dissolved Fe(II) concentrations at these low pH levels were stable and similar to both the initial Fe(II) and total Fe concentrations. Because no 99Tc(VII) had been added to the Fe(II)-only solutions, no significant redox reactions were expected to occur in which Fe(II) would be oxidized to Fe(III) (Fig. 2, left). As mentioned, there could have been some

Fig. 2. Changes in Fe or 99Tc concentrations in experiments over pH. (Left) Fe(II) and Fe(total) for experiments without Fe(III) oxide solids; (right) solution without the presence of Fe(III) oxide solids.

99

Tc in Fe(II)-bearing

205

minor oxidation of Fe(II) from oxygen absorption in the solution, even if no 99Tc(VII) was added. However, when 99Tc(VII) and oxygen contamination were present together in solution, there was a strong driver for oxidation of Fe(II) as shown by the measurable difference between measured Fe(II) and Fe(tot) concentrations even at low pH (<5.0). Because the Fe(tot) concentration in Fe(II) + Tc(VII) test solution at low pH values (<5.0) showed the same concentration as the initial Fe(II) concentration, the concentrations of dissolved ferric iron, Fe(III), could be determined by the difference between measured Fe(II) and Fe(tot) concentrations in Fe(II) + Tc(VII) test solutions. The calculated Fe(III) concentration in the Fe(II) + Tc(VII) test solutions was about 50% and 40% of initial Fe(II) concentration added to the test tubes at pH = 3.2 and 4.1, respectively. The mass of 99Tc(VII) added to the solutions was less than 0.1% of the mass of starting Fe(II) so that even if all the 99 Tc(VII) was reduced to 99Tc(IV), the electrons that each mole of 99 Tc(VII) requires to be reduced would be supplied easily by the oxidation of minor amounts of Fe(II). Complete reduction of all the 99Tc(VII) present in the solution would not require nearly as much Fe(II) to be oxidized to Fe(III) as suggested by the apparent ingrowth of Fe(III) calculated from the difference in the measured total Fe and measured Fe(II) remaining in these solutions. Therefore, transformation of the Fe(II) to Fe(III) could also result from redox reactions between Fe(II) and the oxygen contamination within the 99Tc(VII)-spiked solution. Negligible 99Tc removal from solution at these low pH values (Fig. 2, right) also suggests that most of the formation of Fe(III) was caused by oxygen contamination introduced with 99Tc addition. Independent saturation index (SI) calculations indicate that at the concentrations of 99Tc in these tests, precipitation of TcO2xH2O would not occur until pH values exceeded 11. In addition, the slight decrease in Fe(III) concentration at pH (4.1) compared to solution at pH 3.2 (Fig. 2, left) indicates that minor amounts of Fe(OH)3(s) or magnetite containing both Fe(II) and Fe(III) may have formed, even though no visible precipitate was observed in these experiments. In contrast, dark green magnetic precipitates were observed in solutions above pH 7 containing both Fe(II) and 99 Tc(VII). These neophases exhibited magnetic properties and were attracted to the magnetic stirring bar used to agitate the slurry [28]. The dark green precipitate was likely a mixture of green rust (layered hydrous oxides containing Fe(II)/Fe(III) with interlayer anions) and magnetite that resulted from mineral transformation of initial precursor, white rust, Fe(OH)2(s), after reacting with dissolved Fe(III) at higher pH values. The green precipitate (green rust) also can be formed from solution containing both dissolved Fe(II) and Fe(III), which transform quickly to more crystalline ferric oxides dominated by magnetite with minor contents of goethite, as were found in the SEM/TEM photomicrographs for similar sample condition [28]. The solubility curves for Fe(OH)3(s) and Fe(OH)2(s), as well as the aqueous Fe(II) concentrations measured in solution, are shown in Fig. 3. The saturation index (SI) for the solutions was calculated based on the solubility product (Ksp) for Fe(OH)2(s) of 1015. The calculated SI was supersaturated with respect to ferrous hydroxide [Fe(OH)2(s)] at pH higher than 7.0, suggesting that it precipitates at these conditions. These results are in accord with the XRD patterns of centrifuged solids in tests prepared separately without goethite being present [28], and indicates the likelihood of white rust, Fe(OH)2(s) at pH = 7.5. Rapid precipitation of Fe(OH)2(s) from the 0.07 M Fe(II) solution can occur when the solution pH is raised to higher than 7.0, even without goethite being present. Because white rust, Fe(OH)2(s), is very unstable towards oxidation, the precipitate rapidly transforms into magnetite and/or goethite with increasing pH. However, this mineral transformation sequence among various Fe oxides occurs so fast that it cannot be distinguished in this batch test. For high pH solutions (i.e.,

Log Fe (total) Concentration (M)

W. Um et al. / Journal of Nuclear Materials 429 (2012) 201–209

0 -1

Fe(II)-sol only

-2 -3

Fe(II)-sol+Tc

-4

FeOH2(s)

-5

Fe(II)-sol+G

FeOH3(s)

-6 -7

Fe(II)-sol+F

-8 -9 -10

Initial Fe(II)

1

2

3

4

5

6

7

8

9 10 11 12

pH Fig. 3. Solubility curves for Fe(OH)3(s) and Fe(OH)2(s) along with measured Fe(II) solution concentrations at different pHs and with different Fe oxides such as goethite (G) and ferrihydrite (F). The brown line is the calculated solution concentration of Fe(total) controlled by the solubility of Fe(OH)3(s) and the blue line is the Fe(total) controlled by Fe(OH)2(s).

pH > 7.5), dissolved Fe(II) and Fe(tot) concentrations in Fe(II) + 99Tc(VII) solutions were identical and very low, suggesting that essentially all the Fe was precipitated and removed from solution as solid phases. The concentration of 99Tc in solution followed a similar trend to that of dissolved Fe(tot) concentrations with varying pH levels up to 9.0 (Fig. 2, right). There was no 99Tc removal in test solutions at low pH values (3.2 and 4.1). However, at higher pH (7.5–9.5), most of the 99Tc(VII) in solution was removed by the newly precipitated Fe(II/III) oxides. An exception to these trends occurred at pH 11.2. The lower percentage of 99Tc removed from solution after 1 day of reaction (only 69% of 99Tc removal from solution; see Fig. 2, right) was most likely caused by 99Tc(IV) reoxidation from minor oxygen contamination entering the test tube. Because we found no detectable aqueous iron [Fe(II) or Fe(tot)] concentrations, dissolution of newly precipitated 99Tc(IV)-bearing Fe(III) oxides did not contribute to these higher 99Tc(VII) concentrations. We speculate that 99Tc(IV) was present as TcO2xH2O(s) that became vulnerable to oxygen diffusion into the tubes at longer contact time.

3.2. 99Tc(VII) and Fe(II) removal from solution containing goethite or ferrihydrite Removal of Fe(II) and 99Tc(VII) from solution in the presence of goethite or ferrihydrite was investigated over a range of pH values. At low pH (2.0), after 1 day of contact, Fe(II) concentrations remained unchanged, suggesting that Fe(II) adsorption onto either goethite or ferrihydrite was negligible (Fig. 4, left). Adsorption of Fe(II) onto goethite was insignificant up to pH = 6.5 and equivalent to results from ‘‘blank’’ tests lacking goethite. On the other hand, a decrease of Fe was observed at low pH when Fe(OH)2(s) precipitation occurred. Even above pH 6.5, where goethite becomes stable, concentrations of Fe [both as Fe(II) and Fe(tot)] decrease rapidly in some experiments, whether or not goethite is present (Figs. 2 and 4), consistent with rapid precipitation of Fe(OH)2(s) (Fig. 3). For cases in which ferrihydrite was present, a steeper decrease in Fe concentrations with time occurred as the pH values approached 5.7. In the presence of ferrihydrite, these tests showed about 90% Fe(II) removal at pH = 5.7 by adsorption processes (Fig. 4, left). Removal of Fe(II) in solution by adsorption on ferrihydrite was also much higher at a pH of 6.5 than in the systems without Fe oxide (Fig. 2, left) and with goethite present in solution (Fig. 4, left). The observed enhanced drop in both Fe(tot) and Fe(II) concentrations

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Fig. 4. Dissolved Fe(II) and Fe(tot) concentrations in experiments conducted at different pHs with Fe oxides present (G = goethite; F = ferrihydrite) (left) and 99Tc removal in Fe(II) solutions with Fe oxides present at different pHs (right). The solid concentration in these batch experiments was 0.3 g/50 mL.

in the ferrihydrite slurries at pH values between pH 5.5 to 6.5, conditions in which precipitation of Fe(OH)2(s) also occurs, confirms that the dominant mechanism of Fe(II) initial removal in the ferrihydrite slurries was Fe(II) adsorption on ferrihydrite, not precipitation of Fe(OH)2(s). In general, goethite precipitation occurred only above pH 6.5 and was a significant cause of Fe loss from solution under these conditions. Thus, Fe(OH)2(s) and ferrihydrite are the main arbiters of Fe removal from solution at low pH, where Fe(II) is stable, and these solid phases are precursors to goethite formation as the pH of the system was increased in the experiments. The calculated equilibrium solubility curve for Fe(OH)2(s) plotted in Fig. 3 matched very well with the measured Fe(II) concentrations in the sample with goethite present at pHs higher than 7.5 (Fig. 3). However, the measured Fe(II) solution concentrations in the presence of ferrihydrite fell to the left of the calculated equilibrium solubility curve line in the region of undersaturation, supporting our hypothesis that Fe(II) removal in these solutions resulted from adsorption of Fe(II) onto ferrihydrite and not precipitation of Fe(OH)2(s). Removal of 99Tc from solution was also measured in the batch tests at different pH values containing either goethite or ferrihydrite. When 99Tc(VII) was spiked in Fe(II) solutions containing goethite or ferrihydrite solids at low pH (<3.0), the concentration of 99 Tc was constant and similar to the initial 99Tc concentration

Fig. 5. (A) TEM image and electron diffraction pattern (B) of goethite crystals from 99 Tc–goethite powder sample. The specific crystallographic planes on the goethite crystallite have been identified on the figure (A). The electron diffraction pattern shows the presence of some overlapping crystals (B).

(Fig. 4, right). Because Fe(II) adsorption onto goethite or ferrihydrite was insignificant and the solution was undersaturated with respect to Fe(OH)2(s) at these low pH values (<3.0), we infer that the key to 99Tc(VII) reduction is sorption of Fe(II) onto solid Fe (oxy)hydroxide phases, where the heterogeneous catalysis of technetium reduction occurs. The constant concentrations of 99Tc(VII) near those at the beginning of the experiment also indicates that TcO2xH2O(s) is not precipitating at this pH. However, 99Tc(VII) removal was fast and complete even after 1 day of reaction in solution with ferrihydrite or goethite present at higher pH values (>5.0), most likely resulting from 99Tc-reduction during reaction with either adsorbed Fe(II) on the ferrihydrite or the Fe(OH)2 precipitates that also formed in both systems. 3.3. Solid phase characterization of

99

Tc–goethite

The mineralogy of the initial goethite and ferrihydrite, 99Tc–goethite powder sample, and the final 99Tc–goethite pellets after diffusion leach test was determined and compared using XRD analysis. Initially synthesized solid substrate was identified as goethite or 2-line ferrihydrite [39]. The measured XRD pattern of the 99Tc–goethite powder was also found to be solely goethite, which is the same as Sample 2–5 in Um et al. [28] based on the exact match to the XRDJADE program. Acicular goethite grains were observed by TEM in the technetium-bearing sample (Fig. 5). The SAED data for this sample also clearly indicated that goethite was well crystallized (Fig. 5b).

Fig. 6. 57Fe transmission Mössbauer spectra of temperature.

99

Tc–goethite sample at 77 K

W. Um et al. / Journal of Nuclear Materials 429 (2012) 201–209

207

Fig. 7. Calculated 99Tc effective diffusivities for 99Tc–goethite pellet as a function of cumulative leaching time (a) and the linear dependence between cumulative 99Tc release and cumulative leaching time, t½.

The Mössbauer spectral features of room temperature (data not shown) and 77 K (Fig. 6) are similar to goethite [40]. More importantly, Mössbauer spectral analysis showed the sample to be free of adsorbed Fe(II) species that exhibits distinct peaks around 0 and 2.5 mm/sec respectively [41], confirming that all the Fe present in the final 99Tc–goethite product was Fe(III) in 99Tc–goethite sample. More solid characterization results for the 99Tc–goethite using SEM and TEM can be found in our previous paper [28]. The final 99Tc–goethite pellet sample after 120 day reaction with the simulated IDF pore water solution also showed identical XRD patterns to the initial 99 Tc–goethite sample [42]. The lack of mineralogical change even after 120 days contact with oxygenated aqueous solution indicates that goethite is very stable at circumneutral pH conditions, which is germane to waste form storage at the Hanford Site. 3.4. Diffusion coefficient and leachability index for

99

Tc

Monolithic and cylindrical pellets of the 99Tc goethite waste form were used to determine an effective diffusion coefficient (or leachability index) for 99Tc immersed in IDF pore water. A small aliquot of leachate was removed from the leaching container at time intervals from 30 min to 120 days throughout the leaching period. The small aliquot was analyzed for 99Tc concentration and pH. The pH remained constant at a value close to the initial pH (7.2) of the IDF pore water. The diffusivity of 99Tc from 99 Tc–goethite pellet sample was calculated with Eq. (1), and the results are displayed in Fig. 7. The first value of 99Tc diffusivity after 30 min of contact was low, likely because of effects imposed by the oleic acid used on the final monolithic pellet surfaces to keep the pellets intact. Thereafter, the data showed a low effective diffusion constant for 99Tc over the course of the experiment (Fig. 7). The calculated individual interval diffusivity values for the 99 Tc–goethite pellet showed a gradually decreasing trend as the contact time increased, except one value that was determined at early 30 min of contact (Fig. 7a). Based on the assumptions and boundary conditions for simple radial diffusion from a cylinder into an infinite bath presented by Crank [38], the mass release of 99 Tc should follow a linear dependence on the square root of time, t½. Excluding the first data point collected at 30 min, the 99Tc–goethite pellet sample showed a mass release proportional to t½ with a linear correlation coefficient (R2 = 0.90) in Fig. 7b. The calculated average 99Tc diffusivity for 99Tc–goethite pellet sample is 6.15  1011 cm2/s, and the leachability index (LI) is 10.2 accordingly. The low value for 99Tc diffusivity and the high LI value are comparable to those found in several different Hanford grouts that ranged from 1012 to 109 cm2/s with LIs from 9 to 12 [43]. The LI is used as a performance criterion for decision of use and disposal of treated waste, and in most cases, treatment is considered effective

for solid waste forms when the LI value is equal to or greater than 9.0 [44]. The value of LI for the 99Tc–goethite pellet is also similar to or slightly higher (better performance) than those of Cast Stone (LI = 9.4–10.3 for 99Tc) prepared with a low-activity waste simulant [19]. In addition, if 99Tc sequestered-goethite powders were used as a part of ingredient material to produce the solidified waste forms, the values of diffusivity and LI for 99Tc would be even lower and higher, respectively than those values determined using 99Tc– goethite pellet here. 3.5. XANES analysis for 99Tc oxidation state on the 99Tc–goethite pellet Because the XANES spectra for the 99Tc(VII) standards (NaTcO4, KTcO4, and TcO 4 adsorbed on ion exchange resin) are very similar and are characterized by a strong pre-edge feature due to the 1s to 4d transition for the tetrahedral 99 TcO 4 anion, only one XANES 99 spectrum for 99 TcO collected from Tc-adsorbed resin is provided 4

Fig. 8. Normalized XANES spectra for 99Tc(VII) and 99Tc(IV) standards as well as 99 Tc goethite samples before and after diffusion leach test. The rectangular black symbols and red lines indicate the measured data and a linear combination fit, respectively. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

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in Fig. 8. The XANES spectrum of TcO22H2O standard for 99Tc(IV) is very distinctive and characteristic for 99Tc(IV) coordinated by six oxygen atoms in an octahedral geometry. The oxidation states of 99Tc in the 99Tc–goethite powder sample both before and after contact with oxygenated aqueous solution for 120 days reaction were determined by fitting their collected XANES spectra by a linear combination fit using the spectra for the 99Tc(VII) and 99Tc(IV) standards. In all cases, the fitting results indicated that only 99Tc(IV) was present in both unleached and leached 99Tc–goethite samples. The best fit results in Fig. 8 showed 99 that the fraction of 99Tc present as 99 TcO Tc–goethite pel4 in the let sample even after 120 days diffusion leach test was less than 1%, demonstrating that the reoxidation of the 99Tc(IV) initially incorporated within the Fe(II)-treated goethite mineral lattice was minimal and consequently diffusive transport of 99Tc(VII) was limited in the goethite waste form. 4. Conclusions Based on the results of this study, the following conclusions were drawn: (1) The observed high-percentage 99Tc incorporation within the Fe(II)-treated Fe-oxide mineral (mainly goethite) structure provides a viable option for treating waste streams containing 99Tc(VII) and forming stable 99Tc-bearing Fe oxide solid waste forms. (2) Reduced and incorporated 99Tc within the goethite is unlikely to be released, even when the final 99Tc–goethite product is exposed to oxidizing conditions. In the circumneutral Hanford pore water solution, the concentration and an effective diffusion coefficient (6.15  1011 cm2/s) of leached 99Tc from the goethite pellet waste form were very low, suggesting potential use of goethite as a waste form. (3) A previous study of the evolution of 99Tc oxidation states in cement waste forms showed that the fraction of 99Tc(VII) rapidly increased from about 10% to 40–50% after exposure to atmospheric oxygen within 4 months [16]. Even though 99 Tc(VII) can be initially reduced to form TcO22H2O(s) in the solidified waste forms, reoxidation to 99Tc(IV) occurs rapidly when it reacts with oxygen because 99Tc(IV) as TcO22H2O(s) is not chemically and structurally stable when contacted with oxygen. However, reduced 99Tc(IV) that is coprecipitated within goethite lattices and subsequently armored with additional goethite layers is capable of protecting 99Tc(IV) from future oxygen contact and reduces the rate of 99Tc reoxidation from 99Tc(IV) to 99Tc(VII) due to the chemical and structural stability of goethite.

Acknowledgments Funding was provided by the DOE Environmental Management (EM-31) Program. PNNL is operated for the DOE by Battelle Memorial Institute under Contract DE-AC05-76RL0 1830. Part of this research was performed at Lawrence Berkeley National Laboratory and was supported by the Director, Office of Science, Office of Basic Energy Sciences, Chemical Sciences, Geosciences, and Biosciences Division, of the U.S. Department of Energy and by the Director, Office of Science, of the U.S. Department of Energy under Contract No. DE-AC02-05CH11231. The XANES data collection was carried out at the SSRL, a national user facility operated by Standard University on behalf of the US DOE. A portion of funding was provided for this research by WCU (World Class University) program at the Division of Advanced Nuclear Engineering (DANE) in POSTECH through the

National Research Foundation of Korea funded by the Ministry of Education, Science and Technology (R31-30005). A portion of solid characterization analyses was also performed using EMSL, a national scientific user facility sponsored by the Department of Energy’s Office of Biological and Environmental Research and located at Pacific Northwest National Laboratory.

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