Solution Speciation of Plutonium and Americium at an Australian

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Solution Speciation of Plutonium and Americium at an Australian Legacy Radioactive Waste Disposal Site Atsushi Ikeda-Ohno,*,†,‡,§ Jennifer J. Harrison,‡ Sangeeth Thiruvoth,‡ Kerry Wilsher,‡ Henri K. Y. Wong,‡ Mathew P. Johansen,‡ T. David Waite,† and Timothy E. Payne†,‡ †

School of Civil and Environmental Engineering, The University of New South Wales, Sydney, New South Wales 2052, Australia Institute for Environmental Research, Australian Nuclear Science and Technology Organisation, Locked Bag 2001, Kirrawee DC, New South Wales 2232, Australia

‡

S Supporting Information *

ABSTRACT: During the 1960s, radioactive waste containing small amounts of plutonium (Pu) and americium (Am) was disposed in shallow trenches at the Little Forest Burial Ground (LFBG), located near the southern suburbs of Sydney, Australia. Because of periodic saturation and overflowing of the former disposal trenches, Pu and Am have been transferred from the buried wastes into the surrounding surface soils. The presence of readily detected amounts of Pu and Am in the trench waters provides a unique opportunity to study their aqueous speciation under environmentally relevant conditions. This study aims to comprehensively investigate the chemical speciation of Pu and Am in the trench water by combining fluoride coprecipitation, solvent extraction, particle size fractionation, and thermochemical modeling. The predominant oxidation states of dissolved Pu and Am species were found to be Pu(IV) and Am(III), and large proportions of both actinides (Pu, 97.7%; Am, 86.8%) were associated with mobile colloids in the submicron size range. On the basis of this information, possible management options are assessed.

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INTRODUCTION Concerns about radioactive contamination originating from nuclear waste disposal have led to an extensive range of research investigations into the confinement of radioactive waste,1−3 the migration of radionuclides in various natural environments,4−9 the influence of microbial effects10−12 and ultimately radionuclide uptake in exposed human populations.13,14 The mobility and fate of radioactive contaminants in engineered and natural environments, as well as in living organisms, is influenced by their chemical speciation in relevant systems, such as chemical forms, oxidation states, and colloid formation. For this reason, the chemical speciation of radioactive contaminants has been extensively investigated.15−21 The Little Forest Burial Ground (LFBG) is a near-surface radioactive waste disposal site located near Lucas Heights in Australia, approximately 30 km from the center of Sydney (Figure 1a). From 1960 to 1968, the site was used for the disposal of low-level radioactive solid and liquid wastes generated at the Lucas Heights research facility during activities of the Australian Atomic Energy Commission (AAEC).22,23 Since the cessation of disposal operations in 1968, the site has been maintained and monitored by the AAEC and its successor, the Australian Nuclear Science and Technology Organisation (ANSTO). Because of the range of research activities undertaken at the time, the buried wastes contained a variety of radionuclides including 3H, 60Co, 90Sr, 137Cs, 232Th, and 235,238 U, as well as over a tonne of beryllium.22,23 To assess the environmental impact and the selection of possible manage© XXXX American Chemical Society

Figure 1. Geographical location of the Little Forest Burial Ground on an Australian/Sydney map (a), on a local map of Lucas Heights (b) and an aerial view of LFBG (c).

ment strategies for the LFBG, the distribution and behavior of radioactive contamination at the site is being investigated.22−25 Received: January 31, 2014 Revised: July 8, 2014 Accepted: August 5, 2014

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were monitored in situ using an in-line flow cell equipped with a multiprobe system (YSI 556 MPS). The water samples were collected both with and without in-line membrane filtration (0.45 μm, Waterra FHT-Groundwater) and stored in an HDPE container prewashed with the trench water. The in-line membrane filter was made of polyethersulphone and presaturated with the trench water before sample collection. The samples were transported from LFBG directly to a nearby ANSTO laboratory for further chemical analysis immediately after the collection. Because LFBG is situated at a short distance from the ANSTO site (Figure 1b), the time lag between the start of sample collection and the beginning of chemical analysis in the laboratory was approximately 1 h and 30 min (i.e., 1 h of sample collection and 30 min of sample transport from LFBG to the ANSTO laboratory), which minimized possible deterioration and transformation of the collected water samples during transportation, in particular, the possible oxidation of reduced actinide species (section 2.4 in SI) and the generation of subsequent particles.26,27 The soil chemistry at the site was summarized in a previous study.28 Chemical Analysis of Actinides in LFBG Trench WaterOverview. The following information on Pu and Am in the LFBG trench water was sought through this study: activity/concentration, oxidation states, formation of colloidal species, and particle size distribution. Following the separation and purification processes using Eichrom TEVA and TRU resins,29 the activity of Pu and Am nuclides in the samples was determined by alpha spectroscopy using a Canberra Alpha Analyst system equipped with Passive Implanted Planar Silicon (PIPS) detectors (Canberra). Instrument setup, detector calibration and measurement parameters for the alpha spectroscopy have been reported previously.29 The determination of the concentrations of inorganic cations was carried out by ICP-AES (Varian VISTA AX CCD Simultaneous ICP-AES) and ICP-MS (Varian 820-MS Quadrupole ICP-MS) while the concentrations of inorganic anions were determined by ion chromatography (Dionex DX-600 IC System). All the samples were triplicated to minimize experimental errors. Chemical Analysis of Actinides in LFBG Trench Water. 1. Oxidation State Analysis. The oxidation states of actinide (An) species in the collected trench water (filtered on-site with a 0.45 μm in-line membrane filter) were determined by two different coprecipitation and solvent extraction techniques; first, coprecipitation with NdF3 and, second, extraction using thenoyltrifluoroacetone (TTA) and di(2-ethylhexyl) phosphoric acid (HDEHP). These two methods have been widely applied to the oxidation state analysis of actinides in seawater30−32 and groundwater.19,31,33−35 The combination of these two different methods provides more reliable results than a single method. Full experimental details of these oxidation state analyses are described in sections 2.1 and 2.3 in SI. The NdF3 coprecipitation method36 can differentiate only between reduced actinide species (i.e., An(III)/(IV)) and oxidized species (i.e., An(V)/(VI)), while the three-step solvent extraction using TTA and HDEHP can separate the four different oxidation states individually.37 The initial step of the TTA/HDEHP solvent extraction is the separation of An(V) and colloidal species from An(III/IV/VI) by TTA (Figure S2 in SI), the separation efficiency of which is maximized at weakly acidic conditions (pH = 5−6).38,39 As the trench water from the LFBG was already weakly acidic (pH = 5.1), the TTA/ HDEHP solvent extraction was applied directly to the collected water samples without pH adjustment. The validity of the

A specific concern at the LFBG is the presence of transuranic nuclides including plutonium (239,240Pu) and americium (241Am).22,23 Although the amount of Pu and Am is reported to be small (several grams of Pu in total and a similar activity of Am),22 the presence of these long-lived alpha-emitters in the buried wastes motivates investigation of their potential impacts on natural and human environments, including animals and plants that utilize the site,24 local rivers and creeks25 (Figure 1b), and developing residential areas within a few kilometers of the LFBG.25 Therefore, there is a need to understand the status of radioactive contamination at the site. A recent paper has described the phenomenon of trench “bathtubbing” which has caused dispersion of Pu into surrounding soils.25 Although the trenches are perched within clay-rich soils that are typically unsaturated, periodical infiltration into the trenches has led to the dispersion of Pu contamination. This previous study was undertaken using a sampling point which had been installed in a cavity which developed after subsidence of a trench due to intense rainfall in 2011. This sampling point permits the sampling of waters within a former trench, which have unusually high levels of Pu and Am due to the proximity of the buried wastes.25 The present study builds on the previous study to elucidate the chemical forms of Pu and Am in the LFBG trench water, to better understand the possible migration of these radioactive contaminants and to propose suitable management strategies to limit actinide contamination at the site. The primary aim of the present study was to determine the chemical speciation of Pu and Am (i.e., chemical forms and oxidation states) and size distribution of Pu/Am-containing colloids (if present) in the water collected from the sampling point in the former waste trench (Figure 1c). The results are compared with thermochemical predictions based on the observed water chemistry at LFBG, as well as with results from other waste disposal sites contaminated with these actinides.

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EXPERIMENTAL SECTION Sample Collection. Trench water samples were collected from the trench-sampler within the waste trench area at LFBG, which was installed for trench water collection on 1 August 2011.25 The pipe penetrates 1.55 m below the present ground surface, with the water chemistry likely to be strongly influenced by the radioactive wastes buried in the trenches, which are reportedly covered by approximately 1 m of the local clay soil.22 Significant activity concentrations of Pu and Am (18.5−31.0 Bq/L for 239,240Pu and 10.1−19.9 Bq/L for 241Am in unfiltered trench water) have been observed in the trench water.25 Since the water level in the trench is often significantly below the ground surface, a considerable amount of rainfall is required prior to sampling, in order to reach a water level sufficient for sample collection.25 Therefore, trench water was obtained on 31 January 2013, following a period of frequent rainfall (Figure S1 in Supporting Information (SI)). The initial water level before the sample collection was 82 cm below the ground surface. Before collecting trench water samples, the sampling bore was purged using a peristaltic pump until the water parameters stabilized and the subsequent four liters of the accumulated water was discarded. After this purging process, the sample collection was performed at a steady water level of 1.1 m below the ground surface. During the sample collection, chemical parameters of the collected water samples (temperature, pH, oxidation−reduction potential (Eh), specific conductance, and dissolved oxygen (DO) concentration) B

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isotopes detected were the same as those reported for samples taken in late 201125 although the concentrations were slightly lower. The wastes buried at LFBG are known to have contained radioactive materials associated with research on nuclear power reactor design22 which involved 239,240Pu. Although the source of Am is not known definitively, 241Am is typically generated either by neutron activation of 239U/239Pu in nuclear reactors or by the alpha bombardment of 238U.44 241Am has commonly been detected in association with 239,240 Pu in LFBG samples.22,23,25 Particular attention is focused on the 239,240Pu and 241Am speciation in LFBG trench water in the following sections. Oxidation States of Actinides in LFBG Trench Water. Accurate information about the oxidation states of actinides in the trench water is essential for the reliable assessment of their environmental fate and possible management strategies as their migration behavior is greatly influenced by their oxidation states.17,18,21,45 Both Pu and Am can possess four oxidation states (III, IV, V, and VI) in aqueous solutions.46 Pu can be further oxidized to a heptavalent state (Pu(VII)) under highly alkaline conditions46 which are not relevant to natural aquatic environments. While the redox behavior of Am is typically simpler than Pu, being generally restricted to Am(III), it was decided that both Pu and Am would be the subject of detailed oxidation state analysis in this study. Figure 2a shows the distribution of 239,240Pu in the trench water (0.45 μm filtered) using the NdF3 coprecipitation method, indicating that almost all the Pu species (97.2%) are in the reduced states (i.e., Pu(III) and/or (IV)). The NdF3 coprecipitation for 241Am suffered from a low recovery (3.4%) arising from the chemical similarity between Am and coexisting Nd (see section 2.2 in SI for more detail). Therefore, the 241Am distribution data by the NdF3 coprecipitation are not taken into consideration in this study. The separation of individual oxidation states of Pu in the trench water (0.45 μm filtered) by the TTA/HDEHP solvent extraction further indicated that 239,240Pu is distributed primarily in the TTA-Aqueous (Pu(V) and/or colloidal, 35.5%) and in the TTA-Organic (Pu(IV), 64.5%) fractions (Figure 2b). The amounts of Pu in the HDEHP-Organic (Pu(VI)) and HDEHPAqueous fractions (Pu(III)) were insignificant. On the other hand, 241Am was partitioned mainly in the TTA-Aqueous (Am(V) and/or colloidal, 32.3%) and in the HDEHP-Aqueous (Am(III), 67.7%) fractions. In the case of Am, there was no significant amount detected in either TTA-Organic (Am(IV)) or HDEHP-Organic (Am(VI)) fractions. The thermochemical simulation for Pu based on the chemical conditions of the LFBG trench water indicates that the dominant form of Pu species under the relevant conditions is an aquo species of Pu(III) (indicated by a red circle in Figure 3a). However, the pH−Eh condition of the LFBG trench water (pH = 5.1 and Eh = 0.14 V) is close to the boundary between Pu(III) (purple area in Figure 3a) and Pu(IV) (orange area in Figures 3a), suggesting that even a slight change of chemical conditions may alter its oxidation state. It is likely that such changes could occur during the cycling of water level in the trench, including the slow decline of water level in dry periods and the influx of water during rainfall events. Considering that no detectable amount of Pu(III) species was separated by the TTA/HDEHP solvent extraction, the oxidation state of Pu in the reduced fraction from the NdF3 coprecipitation must be tetravalent (i.e., Pu(IV)). In the TTA/ HDEHP solvent extraction, a considerable amount of Pu was

TTA/HDEHP solvent extraction for An oxidation state separation was verified using 232U(VI), 242Pu(IV), and 243Am(III) tracers dissolved in the LFBG groundwater taken from another sampling point (section 2.4 in SI), which has similar chemical composition without An contamination. After the treatment and purification processes,29 the activity of Pu and Am in the separated fractions was determined by alpha spectroscopy. Chemical Analysis of Actinides in LFBG Trench Water. 2. Particle Size Fractionation. The size distributions of colloidal Pu and Am species in the collected LFBG trench water were determined by particle-size fractionation using centrifugal filters (Sartorius Vivaspin 20, pore size; 0.20 μm, 1000 kDa, 100 kDa, and 10 kDa-MWCO). The trench-water, filtered in-line with a 0.45 μm membrane filter, was transferred into a centrifugal filter and centrifuged at 4000 rpm (3350 RCF) for 10 min at 293 K. The filtrates were sequentially transferred into smaller filters and centrifuged under the same conditions. The filters holding retentates were gently heated on a hot plate with aqua regia to dissolve the retentates as well as to digest the filter materials. The obtained samples were further treated according to the previously mentioned treatment and purification processes. The activity of Am and Pu in the obtained samples was determined by alpha spectroscopy. Thermochemical Modeling. The chemical speciation (Eh−pH diagrams (Pourbaix diagrams) and speciation distribution) of Pu and Am under water conditions relevant to LFBG shallow groundwaters was simulated using the speciation modeling software JCHESS (Version 2.0)40 based on the obtained activity data for these nuclides and concentration data for inorganic ions. The OECD-NEA Chemical Thermodynamic database41 was used as the source of thermochemical parameters for Pu and Am nuclides while the EQ3/6 database42 was employed for simulating Fe and Mn speciation. Redox reactions were taken into account in the speciation calculations, while sorption was not considered. Activity corrections were performed using the Davies equation.43

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RESULTS Activity Concentration of Pu and Am in LFBG Trench Water. The activities of Pu and Am in the unfiltered trench water collected at the LFBG are summarized in Table 1. Most of the detected alpha activity is due to 239,240Pu and 241Am (9.88 and 9.20 Bq/L, respectively) with a small contribution from 238 Pu (0.04 Bq/L). No detectable amounts of Th isotopes, U isotopes, 231Pa, 237Np, 242Pu and 243Am were measured. The Table 1. Activity Concentrations (Bq/L) of Actinides in Unfiltered Trench Water Collected from LFBG in January 2013a element 90Th 92U 94Pu

95Am

a b

isotopes measured 228

Th, 229Th, 230Th, and 234 U, 235U, and 238U 238 Pu 239,240 Pu 242 Pu 241 Am 243 Am

activity (Bq/L) 232

Th

BDLb BDLb 0.04 ± 0.02 9.88 ± 0.27 BDLb 9.20 ± 0.17 BDLb

Errors are given as a 95% confidence interval for triplicate samples. BDL: below detection limit of 0.02 Bq/L. C

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method, which does not differentiate colloidal and noncolloidal (i.e., dissolved) species,36 suggests that the Pu species in the LFBG trench water are dominated by the reduced forms. Therefore, it is reasonable to conclude that the Pu species retained in the TTA-Aqueous fraction are colloidal forms of reduced Pu (i.e., Pu(III) or -(IV)) rather than the pentavalent state. Compared with the results for 239,240Pu, the distribution of 241 Am is relatively straightforward. As illustrated in Figure 3b, the trivalent state (Am(III)) is the only oxidation state which is feasible under the LFBG trench water conditions. This is consistent with the fact that the majority of Am species in the LFBG trench water were found in the HDEHP-Aqueous fraction (Am(III) fraction). Since the presence of Am(V) species is highly unlikely, the Am species retained in the TTAAqueous fraction must be a colloidal form of Am(III). The oxidation states of actinides in natural aquatic conditions have been previously investigated in only a limited number of studies. The Pu in trench leachates from the Maxey Flats radioactive waste disposal site, Kentucky, USA, was found to be mostly dissolved Pu(IV) species over a wide range of pH from 1.9 to 12.2,19 with codisposed organic contaminants implicated in Pu solubilization. The Pu nuclides released into the groundwater (pH = 6.4−10.1) at Mortandad Canyon, New Mexico, USA, were also dominated by the reduced states.33 On the other hand, a high abundance of oxidized Pu species (Pu(V/VI)) was observed in groundwater (pH = 3.9−6.8) at the Savannah River Site, South Carolina, USA,34 and in groundwater (pH = 7.0) collected at Yucca Mountain, Nevada, USA.37 These studies indicate that, as well as pH, other parameters, particularly Eh, are important factors to consider when investigating Pu oxidation states. The Eh value of groundwater provides an indication of the potential of dominant redox-active species existing in the system, although only a few elements (e.g., C, N, O, S, Fe, and Mn) are generally considered to be predominant participants in the redox processes of natural aquatic systems.47 The trench water collected at LFBG contained high concentrations of Fe and Mn (8.91 and 0.378 mg/L, respectively), possibly because of the metallic items disposed and because the wastes were disposed of above a weathered

Figure 2. Distribution of 239,240Pu in NdF3 coprecipitation (a), and distribution of 239,240Pu (b) and 241Am (c) in the TTA/HDEHP solvent extraction. The activity of 239,240Pu in HDEHP-Organic and HDEHP-Aqueous fractions and 241Am in TTA-Organic and HDEHPOrganic fractions was below the detection limit. The relevant radioactivity data are provided in Table S1 in SI. The chart colors in panels b and c are as shown in the separation scheme in Figure S2 in SI (i.e., green (An(III)), yellow (An(IV)), blue (An(V) and colloidal), and red (An(VI))).

found in the TTA-Aqueous fraction which retains Pu(V) and/ or colloidal species. As illustrated in Figure 3a, the pH−Eh condition of the LFBG trench water is too anoxic to form Pu(V) (as PuO2+) species. Besides, the NdF3 coprecipitation

Figure 3. Pourbaix (Eh−pH) diagrams of Pu (a) and Am (b) for trench water at the LFBG. The red circles indicate the actual pH−Eh condition of trench water collected in January 2013. [Pu]total = 1.80 × 10−11 mol/L and [Am]total = 2.99 × 10−13 mol/L based on the radioactivity data in Table 1. Details of the trench water chemistry and calculation conditions are described in section 3 in SI. D

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mineral-rich siltstone layer.25 The Pourbaix diagrams of Fe and Mn species under the LFBG trench water conditions are given in Figures S7 and S8 in SI. The results indicate that Fe could exist as a mixture of Fe(II) and -(III) phases, the latter of which potentially oxidizes Pu(III) to Pu(IV) (but not to Pu(V) or higher) with the formation of other Fe(II) minerals (Figure S7 in SI). In contrast, Mn forms nonoxidative aquo Mn2+ species over a wide range of pH/Eh which are unlikely to be involved in the redox processes of Pu(III/IV) (Figure S8 in SI). The groundwaters at both the Maxey Flats disposal site and at Mortandad Canyon, in which Pu species were found in the reduced states, contained high concentrations of Fe (up to several mg/L),19,33 while the groundwater at the Savannah River Site and at Yucca Mountain, in which Pu species were mainly in the oxidized form, was found to be iron-poor (in the μg/L range).34,48 Although these facts do not directly explain why Pu tends to form the oxidized species in iron-poor conditions, the presence of Fe in high concentration, which is one of the major determinants of the Eh of groundwater, is probably a significant parameter affecting the redox behavior of Pu in groundwater. Colloidal Species of Actinides in LFBG Trench Water. The presence of submicrometer-sized colloids potentially enhances the transport of insoluble or less-mobile chemical contaminants,49,50 including actinides,20,51 in natural aquatic environments. Therefore, information about colloidal species is indispensable for predicting the mobility of Pu and Am at LFBG. As discussed above, the LFBG trench water treated with a 0.45 μm filter on-site was subsequently passed through membrane filters of different pore sizes and the activity of 239,240 Pu and 241Am in the retentates was measured by alpha spectroscopy. Figure 4 shows the resulting size distribution of 239,240 Pu and 241Am, indicating that most of the Pu (97.7%) and Am (86.8%) in the LFBG trench water was found in colloidal particulates larger than 0.20 μm in diameter. For the remainder, Am was distributed uniformly in the smaller size fractions, while most of the Pu residue was found in the filtrate of 10 kDa filtration (“0.20 μm). Given the fact that eigencolloids (i.e., colloids primarily consisting of actinides) are usually in the nanometer size-ranges,53 the colloidal Pu and Am particles formed in the LFBG trench water are most likely to be pseudocolloids consisting mainly of other minerals and/or organic substances. The groundwater at LFBG is iron-rich across the site (up to 48 mg/L, Table S4 in SI), resulting in the potential formation of Fe(III) oxides such as ferrihydrite (amorphous-FeOOH) and lepidocrocite (γ-FeOOH). These minerals are eventually transformed to more crystalline forms such as goethite (α-FeOOH) (section 3.3 in SI) with these minerals potentially forming particles in the colloidal size range.47 Natural organic substances, such as humic acid, could also generate colloidal particles in the natural aquatic environment.47,54 Both Fe- and organic-based colloidal particles can adsorb actinide species, forming pseudocolloids.55−57 Besides, organic substances can enhance the colloidal stability of Fe-based particles in aqueous solutions58 and potentially affect the adsorption behavior of Pu/Am onto the Fe-based colloids.59 The particle size of Fe- and organic-based colloids typically ranges from tens-of-nanometers to submicron order in natural aquatic systems,60,61 which is in line with the observed size distribution of Pu and Am in Figure 4. Given the fact that no suspension was observed in the collected trench water, the upper size limit of the actinide-retaining particles would be less than tens of microns.47 Although further study is required to fully characterize the observed colloids, the present results indicate that Pu and Am are primarily associated with submicron−micron scale pseudocolloids consisting of mineralor organic particles in the trench waters at LFBG, possibly Fedominated particles. There was a 1.5-h interval between the sampling and the lab analysis, which could potentially produce secondary colloids by air intrusion into the water samples. Assuming that the observed colloids are primarily composed of Fe minerals, the most probable pathway to generate secondary colloids is the air oxygenation of Fe(II).62 However, even assuming a significant amount of Fe(II) in the trench water, its oxygenation reaction should be slow enough to maintain iron in the Fe(II) form during the 1.5-h period under the weakly acidic condition of the trench water (pH = 5.1).63 Therefore, the scenario that secondary colloids formed during the 1.5-h sample transport is unlikely. As previously mentioned, the

Figure 4. Size distribution of colloids attaching 239,240Pu and 241Am in LBFG trench water. The concentration (y-axis) is relative to the total concentration in the unfiltered trench water. The error bars indicate 95% confidence interval of triplicate samples. The relevant radioactivity data are provided in Table S3 in SI. E

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Figure 5. Speciation distribution of Pu as functions of pH (a) and total Fe(III) concentration (b) under shallow groundwater conditions relevant to LFBG. The orange-colored area in (a) indicates a potential pH variation at LFBG, while that in (b) highlights the critical region where Fe(III) could oxidize Pu(III) to Pu(IV). Details of the chemistry of shallow groundwater at LFBG and calculation conditions are described in section 3 in SI.

Pu species remaining in the TTA-Aqueous fraction (35.5%, blue data in Figure 2) are the solid or colloidal form of PuO2(c), dissolved species are still expected to dominate the Pu speciation. Therefore, the production of minerals and precipitates is not considered in the further calculations of Pu speciation. Given the absence of PuO2(c) in the system, cationic Pu(III) species (Pu3+, [PuSO4]+, and [PuOH]2+) constitute the major Pu species below pH = 6, while hydroxide species of Pu(IV) ([Pu(OH)3]+ and Pu(OH)4(aq)) become dominant above pH = 6 (Figure 5a). As has been discussed above, the results of oxidation state analysis indicate that the dissolved species of Pu in the collected trench water are mainly in the tetravalent form (Pu(IV)) which is possibly generated by the oxidation of Pu(III) via interaction with Fe(III) minerals. The Fe concentrations measured across the site are well above the solubility limit of iron minerals, resulting in the dominant formation of an Fe(III) oxide (goethite) (section 3.3 in SI). The concentration of Fe(III) in the shallow groundwater significantly affects the oxidation states of Pu species. That is, an increase in total Fe(III) concentration induces the oxidation of Pu(III) to Pu(IV) and -(V) through the heterogeneous redox process with Fe(III) minerals (Figures 5b and S11 in SI). It should be also noted that the Fe concentration in the vicinity of the trench area (indicated with a black-dash line in Figure 5b) is close to the critical point where Fe(III) vigorously oxidizes Pu(III) species to form hydroxide Pu(IV) species. Therefore, the oxidation of Pu(III) to Pu(IV) by Fe(III) is a possible scenario in the vicinity of the trench area. Figure 5b also suggests that a considerable amount of Pu(V) species (PuO2+) is formed in the higher Fe(III) concentration range. As PuO2+ is considered to be the most mobile form of Pu in aquatic systems,21 the formation of PuO2+ in the LFBG shallow groundwater potentially facilitates the spreading of Pu contaminants from the trenches at the site, although the subsurface transport of PuO2+ is still expected to be inhibited by the clay-rich soils surrounding the trenches which have relatively high cation adsorption capacity. In comparison with Fe(III), Fe(II) is not expected to play a significant role in the redox behavior of Pu (Figure S9e in SI). In addition to pH, the interaction with coexisting inorganic ligands could also influence the solution speciation of An. As summarized in Table S2 in SI, the LFBG trench water is fairly

LFBG trench water undergoes a cycling in water level, including a slow decline of water level in dry periods and the influx of water during rainfall events.25 This probably affects the chemical characteristics of the trench water, such as pH and dissolved oxygen concentration, with resultant changes in Fe speciation including the formation (and possibly dissolution) of Fe-based colloids.

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DISCUSSION Potential Change of Pu and Am Speciation upon Local Migration at LFBG. It has been shown in the previous sections that the majority of Pu and Am species present in the trench water at LFBG are associated with colloidal particles. However, a substantial amount of the Pu and Am species attached to colloids can be detached from the colloid matrix relatively easily by, for example, interacting with inorganic/ organic ligands, finally forming dissolved species. This is substantiated by the fact that more than 60% of Pu and Am in the LFBG trench water filtered with 0.45 μm are extracted by TTA and HDEHP, both of which are strong organic ligands (Figure 2). Given that the trench water at LFBG contains significant concentrations of strong inorganic ligands (e.g., carbonate or sulfate, Tables S2 and S4 in SI) and the potential presence of organic ligands (e.g., EDTA), the speciation of dissolved actinide contaminants is one of the key factors to be considered for assessing the potential migration behavior of radionuclides at LFBG. Such local migration could potentially occur by the long-term iteration of “bathtubbing” events.25 To predict the potential change of the speciation of dissolved Pu and Am in the event of local subsurface migration at LFBG via shallow groundwater, thermochemical calculations were performed under various conditions relevant to the shallow groundwaters at LFBG (Tables S2 and S4 in SI). The pH value is one of the key chemical parameters which affect the speciation. The pH of shallow groundwaters at LFBG varies around the weakly acidic-circumneutral range. When the production of minerals is taken into account in the calculation of Pu speciation, a solid phase of PuO2(c) is expected to precipitate above pH = 4.5 (Figure S9a in SI). However, as shown in Figure 2, a substantial amount of Pu was extracted from the trench water by TTA, which is incapable of extracting solid/precipitate and colloidal species. Even assuming that the F

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abundant in bicarbonate (HCO3−), sulfate (SO42−), and chloride (Cl−) species, all of which can potentially form aqueous complexes with An. Assuming the dominant presence of Pu(IV), the solution speciation of Pu in the trench area is mainly characterized by its hydroxide species (Figure 5). When the bicarbonate concentration varies through geochemical interactions within the local surroundings, such as interactions with atmospheric or biologically derived CO2, the Pu speciation is also expected to vary. That is, a neutral aquo species of Pu(OH)4(aq) becomes predominant with increase in the bicarbonate concentration (Figure S9b in SI). In the basic pH range, bicarbonate is transformed into carbonate (CO32−) which shows a strong coordination to An.15 However, carbonate concentration will be very low at the measured groundwater pH of 5.1, with the result that the Pu speciation is dominated by the hydroxide species regardless of the bicarbonate concentration. The variation in bicarbonate concentration will primarily affect the pH of the shallow groundwater, which indirectly induces a change in Pu speciation. In comparison with the bicarbonate case, variation in sulfate and chloride concentrations does not impact Pu speciation significantly (Figure S9c and S9d in SI). These results indicate that the solution speciation of Pu at LFBG is governed exclusively by the strong hydrolysis of Pu(IV),46 resulting in the formation of cationic and neutral Pu(IV) species. In comparison with the Pu speciation which is dominated by hydroxide species, the solution speciation of Am is more complicated, as the hydrolysis of Am(III) is much weaker than that of Pu(IV) and, accordingly, other coexisting ligands can potentially interact with Am(III). In the trench water, an aquo species of Am(III) (Am3+) and a monosulfate species ([AmSO4]+) are expected to dominate Am speciation (Figure S10a in SI). With an increase in pH from circumneutral to basic, the Am speciation tends toward formation of carbonate and hydroxide species (Figure S10a in SI). The formation of sulfate and carbonate species is promoted by the potential increase in sulfate and bicarbonate concentrations, respectively, while the variation in chloride concentration leads to no significant change in the Am speciation (Figure S10b−S10d in SI). On the basis of these results, the solution speciation of Am at LFBG is expected to be dominated primarily by cationic species of Am3+, [AmSO4]+, and [AmCO3]+. Implications for Possible Management Options for Pu/Am Contamination at LFBG. Management strategies for contaminated waste sites typically aim to prevent or at least minimize releases (in cases of shallow ground disposal), and to reduce the spread of any contamination after release events that may occur. Given the fact that the trench area is enclosed with a clay-rich layer which is expected to impede the lateral flow of the trench water,25 the simplest way to control the release of contaminants at LFBG appears to be to reduce infiltration as much as practicable, for example by constructing an effective engineered cover. Additionally, given that the chemical form of Pu and Am contaminants in the trench water at LFBG is primarily colloidal, it may be possible for Pu and Am to be effectively retained by porous media if an interception barrier is constructed or a groundwater extraction/treatment option is implemented. For simple decontamination of the LFBG trench water, filtration with a 0.2 μm pore size filter would remove more than 85% of actinide activity (97.7% for Pu and 86.8% for Am). The potential for the Pu/Am colloids to be retained on porous media could also contribute to the use of artificial

barriers as a possible measure to intercept the migration of Pu/ Am contaminants in shallow groundwater-runoff, or in water that moves in the shallow subsurface.25 Our findings also indicate that Pu and Am could be potentially desorbable (for example, by the reagents used in the chemical separation scheme) which implies that they may become detached from the colloids with resultant formation of dissolved species (primarily in the forms [Pu(OH)3]+, Am3+, [AmSO4]+, and [AmCO3]+). The potential for detachment of Pu and Am from the colloids would need to be considered in cases where reactive barriers are evaluated for use at the site. Although porous barriers may capture colloids, an additional reactive media would be needed to collect and retain the dissolved species. Retention capacity that is both strong, and long-lasting, is desirable as it is typically impractical to frequently replace the reactive media on a routine basis. For conditions at the LFBG, an effective reactive media might contain iron minerals to sorb the dissolved Pu and Am species and potentially transform them into the less-mobile solid oxide phase.64,65 Our findings to date provide a specific explanation as to the limited groundwater transport of actinide contaminants from the LFBG trenches compared to the surface pathway via the “bathtubbing” mechanism.25 That is, the “bathtubbing” events are the major process spreading the actinide contamination around the trench area, while the underground lateral flow is restrained because the clay-rich layers surrounding the trench area retain the actinide contaminants. These data can directly improve the technical- and cost-effectiveness of any future management or remediation strategies. However, the precise nature of the colloids attaching Pu/Am has yet to be fully defined with respect to their potential behavior in treatment/ remediation scenarios (e.g., reactive barriers), as well as their long-term fate in ambient environmental conditions. Detailed studies on the physicochemical forms of the colloid-associated actinides are currently underway.

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ASSOCIATED CONTENT

S Supporting Information *

Rainfall data at LFBG, experimental details of oxidation state analysis, particle size fractionation, chemical analysis of sampled water, detailed descriptions of thermochemical modeling, and additional radioactivity data. This material is available free of charge via the Internet at http://pubs.acs.org.

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AUTHOR INFORMATION

Corresponding Author

*Phone: +49-351-260-3156. E-mail: [email protected] or [email protected]. Present Address

§ A.I.-O.: Institute of Resource Ecology, Helmholtz-Zentrum Dresden-Rossendorf (HZDR), Bautzner Landstrasse 400, 01328 Dresden, Germany

Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS We acknowledge the following ANSTO staff: B. Rowling for support with field sampling, S. Hankin for geographical data, D. I. Cendón and C. E. Hughes for discussion on the hydrogeology at LFBG, and L. Dyer and J. Crawford for G

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