Radionuclide behaviour in hyperalkaline systems relevant to geological disposal of. radioactive waste

Transcription

1 Radionuclide behaviour in hyperalkaline systems relevant to geological disposal of radioactive waste A thesis submitted to the University of Manchester for the degree of Doctor of Philosophy in the Faculty of Engineering and Physical Sciences 2014 Kurt Smith School of Earth, Atmospheric, and Environmental Science

2 Abstract In many countries the current plan for the management of intermediate and high level radioactive wastes is to dispose of the radioactive materials underground in a Geological Disposal Facility (GDF) to prevent release of radioactivity to the environment. In the UK, the repository for intermediate level waste may be backfilled with cementitious material and it is clear that grout and cement will be used during many disposal concepts. Upon saturation, the cement will react creating a region of hyperalkaline geochemical conditions extending away from the GDF, within which, significant changes in radionuclide behaviour are expected. Therefore, this thesis utilises a range of experimental and analytical techniques to try to gain a mechanistic understanding of the behaviour of some key radionuclides (U(VI), Np(V) and Eu(III) as an analogue for Cm(III)/Am(III)) in a range of high ph systems of direct relevance to any cementitious GDF. U(VI) interaction with calcite (calcium carbonate, a common component in high ph cements and the natural environment) surfaces was studied in the 'old' (Ca(OH) 2 solution; ph 10.5) and 'young' (Na +, K +, Ca 2+ ; ph 13.3) leachates. In the 'old' leachate, luminescence spectroscopy, batch experiments and kinetic modelling suggested that at low concentrations ( 0.42 µm) a Ca 2 UO 2 (CO 3 ) 3 -like surface complex formed. At higher concentrations, batch experiments, extended X-ray absorption fine structure spectroscopy and luminescence suggested that a surface mediated precipitation mechanism was controlling U(VI) concentrations. Further TEM analysis confirmed that a calcium uranate (CaUO 4 ) solid phase was forming on the calcite surfaces. In the 'young' leachate, batch experiments showed that U(VI) had little affinity for the calcite surface, with no statistically relevant removal from solution observed over a 18 month period. Small angle X-ray diffraction data demonstrated that the U(VI) was probably present in the form of U(VI) intrinsic colloids. Np(V) solubility and sorption to calcite under hyperalkaline conditions were studied using batch, X-ray absorption spectroscopy, and geochemical modelling techniques. It was determined that Np(V) solubility in 'old' cement leachates was consistent with the literature. However, in 'young' cement leachates, an unidentified calcium containing phase was controlling solubility. It was demonstrated that sorption to calcite in 'old' leachates was controlled by the formation of a >CO 3 NpO 2 surface complex, whereas, in the 'young' leachates interaction with the calcite surface was controlled by a precipitation mechanism. Eu(III) sorption to a potential GDF backfill material, Nirex Reference Vault Backfill (NRVB) cement, was studied. The kinetics of removal were rapid with 98.5% Eu(III) removal within 24 hours. Ultrafiltration experiments indicated that all Eu(III) remaining in solution was associated with NRVB derived colloids. Additional experiments using ethylenediaminetetraacetic acid (EDTA) as a competing ligand show that removal from solution was significantly reduced at high concentrations (>0.01 M). These EDTA experiments also indicated some irreversibility in the systems, possibly caused by incorporation into the C-S-H or calcite structures. i

3 Contents Declaration Copyright Statement Acknowledgement List of Figures List of Tables Structure of Thesis 1. Introduction 1.1. The Nuclear Fuel Cycle 1.2. The Geological Disposal Facility 1.3. Backfill Material 1.4. Host Geology 1.5. Cementitious Backfill and the Chemically Disturbed Zone 1.6. Calcite Relevance Surface Charge Carbonation 1.7. Uranium Relevance Speciation and Environmental Behaviour U(VI) Interaction with Calcite 1.8. Neptunium Relevance Speciation and Environmental Behaviour Interaction with calcite ii

4 1.9. Europium Relevance Europium Speciation Interaction with Cement Aims and Objectives References 2. Instrumentation and Methodology 2.1. Experimental Approaches Batch Experiments Centrifugation Calculations Ultrafiltration Experiments Experimental Errors Synthetic Cement (BIGRAD) Leachates Carbonate Exclusion U Radiotracer Preparation 2.2. Geochemical Modelling 2.3. Elemental Analysis Inductively Coupled Plasma Mass Spectrometry 2.4. Radiometric Techniques Liquid Scintillation Counting Gamma Spectroscopy 2.5. Spectroscopy Ultraviolet - Visible Spectroscopy Luminescence Spectroscopy 2.6. Diffraction Techniques Bragg Diffraction Powder X-ray Diffraction 2.7. Microscopy Scanning Electron Microscopy Transmission Electron Microscopy iii

5 2.8. Synchrotron Techniques Synchrotron Light X-ray Absorption Spectroscopy Overview Experimental Analysis Small Angle X-ray Scattering 2.9. References 3. Accepted Research Paper - Geochimica et Cosmochimica Acta U(VI) behaviour in hyperalkaline calcite systems Supporting Information 4. Pre-submission Research Paper Np(V) sorption and solubility in high ph calcite containing systems Supporting Information 5. Published Research Paper - Mineralogical Magazine Europium interaction with a vault backfill at high ph 6. Conclusions and Future Work 6.1. References iv

6 Declaration No proportion of the work referred to in the thesis has been submitted in support of an application for another degree or qualification of this or any other university or institute of learning. 1

7 Copyright Statement 1) The author of this thesis (including any appendices and/or schedules to this thesis) owns certain copyright or related rights in it (the copyright ) and he has given The University of Manchester certain rights to use such copyright, including for administrative purposes. 2) Copies of the thesis, either in full or in extracts and whether in hard or electronic copy, may be made only in accordance with the Copyright, Designs and Patents Act 1988 (as amended) and regulations issued under it or, where appropriate, in accordance with licensing agreements which the University has from time to time. This page must form part of any such copies made. 3) The ownership of certain Copyright, patents, designs, trademarks and any or all other intellectual property (the Intellectual Property Rights ) and any reproductions of copyright works in the thesis, for example graphs and tables ( Reproductions ), which may be described in this thesis, may not be owned by the author and may be owned by third parties. Such Intellectual Property Rights and Reproductions cannot and must not be made available for use without the prior written permission of the owner(s) of the relevant Intellectual Property Rights and/or Reproductions. 4) Further information on the conditions under which disclosure, publication and commercialisation of this thesis, the Copyright and any Intellectual Property and/or Reproductions described in it may take place is available in the University IP Policy (see in any relevant Thesis restriction declarations deposited in the University Library s regulations (see and in The University s policy on presentation of Theses. 2

8 Acknowledgements After four years working on this thesis there are several people that have been essential and deserve thanks. Firstly, my supervisors, Nick Bryan and Kath Morris, have been very patient over the last four years correcting endless split infinitives, mixed up tenses, and general scientific misconception. Their guidance, advice and support have been fantastic and I could not have asked for more. Fay, my family, and friends (with a special mention to Jess) have provided endless emotional (and financial!) support over the last four years. I don't think I could have completed this PhD without them. I would like to thank all of radiochemistry, especially Dan, Jen, Kate, Ryan, and Tony. I would also like to thank the Nuclear First DTC (EP/G037140/1) and the BIGRAD consortium (NE/H007768/1) for funding and training opportunities. 3

9 List of Figures Figure Illustration of the multi-barrier disposal concept. 19 Figure Cutaway of a steel drum containing cementitiously immobilised solid waste. 20 Figure Change in ph of the aqueous solutions in presence of calcite and air. 25 Figure Modified electrical double layer model for the calcite surface. ζ = zeta potential; ψ = potential at the stern layer; and δ = represents the partial charge on the ions that were left behind at the bulk termination. At X = 0, a layer of H and OH are chemi-sorbed onto the bulks dangling bonds. At X < 0, potential is so high that attaching ions do not sorb, they precipitate. When X < 0 potential is so high they precipitate rather than sorb. 27 Figure The influence of ph on the zeta potential of calcite. 28 Figure Speciation of 1 ppm U(VI) and log total dissolved carbonate (broken line) as a function of ph. Log P CO2 = -3.5 in 0.1 M NaNO Figure Adsorption as a function of ph to subsurface sediments, where different symbols represent sediment collected from three different U.S. Department of Energy sites. Solid/solution ratio = 3.33 g/l in 0.1 M NaNO 3 with log P CO2 = Figure A) Local coordination of Ca 2+ in the calcite structure. B) The local structure of UO 2 2+ incorporated in calcite. 33 Figure (A) Ball and stick representation of positive vicinal region of calcite. (B) geometries of and + step regions along the <481> and <441> directions. (C) Ball and stick representation of negative vicinal region of calcite with uranyl complexed with the surface. 36 4

10 Figure Activity of radionuclides in HLW from reprocessing of Spent Nuclear Fuel (SNF) with time after processing. 37 Figure Np(V) speciation in solution (Np(V) = 1 x 10-5 M with I = 0.01 M) in a nitrogen atmosphere (top) and with log pco 2 = -3.5 log [CO 2 ] (bottom). 39 Figure Neptunium(V) absorption to montmorillonite as a function of ph, with and without atmospheric carbonate. 40 Figure Model of neptunyl ion incorporated into the calcite structure with neptunyl located on the Ca 2+ site with the two axial neptunyl oxygens substituting for two carbonate ions. 41 Figure Surface loading of 2 µm Np versus time with a solid to liquid ratio of 20 g L Figure Speciation of 1 x 10-5 M Eu 3+ in 0.5 M NaClO 4 as a function ph in carbonate free conditions. 45 Figure Speciation of Eu(III) as a function of ph in equilibrium with 0.03% CO 2(g) (thick lines) or 1% CO 2(g) (thin lines). 45 Figure Eu(III) sorption to calcite, in equilibrium with atmospheric CO 2. The hollow symbols represent sorption after 24 hours while the filled symbols represent after 30 days. The dashed lines represent the sorbed Eu associated with a particular surface complex whereas, the solid line represents total modelled sorption. 47 Figure Schematic of an ultrafiltration cell. 64 Figure The normal distribution. 65 Figure Schematic of CO 2 scrubbing system. 67 5

11 Figure LSC spectra of the five fractions (effluent) collected from the separation of 232 U and 228 Th recorded using a Quantulus ultra low background LSC using alpha beta discrimination to isolate the alpha activity. Fraction I, I and III contain 228 Th in concentrated HCl while and IV and V are 232 U containing fractions in 0.1 M HCl. 69 Figure Example PHREEQC input file to determine the speciation and saturation U(VI) (10 mg / kg water, ph 10.5, 25 C, in equilibrium with 10 moles of calcite solid). 69 Figure Schematic of an ICP-MS. 70 Figure Schematic of an ICP torch. 72 Figure Schematic of a LSC instrument. 74 Figure Example pulse height vs. pulse length for alpha and beta emitters. The PSA value defines the gradient that separates the events. 75 Figure Simplified Jablonski diagram illustrating the processes leading to florescence. 79 Figure Schematic of a typical fluorometer. 79 Figure Illustration of Bragg diffraction. 81 Figure Schematic of a powder XRD instrument. 82 Figure Schematic of a typical Synchrotron facility. 86 Figure Example U(VI) L III edge XAS spectrum with key regions of interest highlighted. 88 6

12 Figure Schematic of a typical beam line set up for the collection of XAS data in transmission and fluorescence modes. 89 Figure Example of scattering between an emitted photoelectron and scattering atoms. 90 Figure An example of Fe EXAFS data in K (bottom) and R (top) space. Blue lines represent experimental data whereas red lines represent model fit. The dashed lines represent the individual contribution to the spectrum from Fe-O and Fe-Fe scatterers. 91 Figure Illustration of some simple multiple scattering paths. 92 Figure Schematic of SAXS data collection set up. 93 Figure Idealised SAXS data for a monodisperse population of spheroid particles, showing key features of the pattern. Model data was generated using the Irena software macro for IgorPro Figure 3.1. Fraction of U(VI) (5.27 x 10-5 µm) remaining in solution versus contract time with calcite as a function of solid to solution ratio. Lines represent model predictions (calibrated with the 1 : 10 data set only). Error bars are 1σ of five replicates. 116 Figure 3.2. Fraction of U(VI) (5.27 x 10-5 µm) remaining in solution at equilibrium versus calcite concentration. Symbols represent experimental data, lines represent predictions calculated using the simple surface complexation model. Only the filled triangle and circle data points (0.1 g ml -1 calcite) were used in fitting; hollow points are for comparison to the model. Error bars are 1σ from five replicates

13 Figure 3.3. Fraction of U(VI) ( µm) remaining in OCL solution versus contract time with calcite as a function of solid to solution ratio (1:10 (A), 1:50 (B), 1:500 (C)). Lines represent predictions from the kinetic model for the 0.42 µm data calculated from the 1:10 data set with rate constants k 1 = 1.78 x 10-6 L mol -1 s -1 and k 2 = 1.23 x 10-6 s -1. Error bars are 1σ of three replicates. 120 Figure 3.4. A: An ungated luminescence spectrum collected from the OCL 0.21 µm U(VI)-calcite sample following 250 nm excitation. B: A luminescence spectrum of a U(VI)-calcite surface complex sample following 420 nm excitation. C: Luminescence spectrum (293 K) of a liebigite (Ca 2 UO 2 (CO 3 ) 3 ) standard following 420 nm excitation. 122 Figure 3.5. A: An ungated luminescence spectrum collected from the 0.42 µm U(VI) OCL calcite sample following 250 nm excitation. B: A 'short gate' ( ms) spectrum of 100 µm U(VI) sorbed onto calcite following 420 nm excitation from the literature. C: Ungated luminescence spectrum collected from U(VI) incorporated into calcite standard following 250 nm excitation. 123 Figure 3.6. A: An ungated luminescence spectrum from a 4.20 µm U(VI) OCL calcite sample following 250 nm excitation. B: Ungated luminescence spectrum from a 2.10 µm U(VI) OCL calcite sample following 250 nm excitation. C: Ungated luminescence spectrum from an (84.0 µm) OCL U(VI) precipitate following 250 nm excitation. D: Ungated luminescence spectrum a calcium uranate sample following 420 nm excitation from the literature. 126 Figure 3.7. A: Backscattering electron image of calcite crystal coated with an electron dense material, lighter areas indicate higher electron density. The white line indicates the area in which the TEM thin section was prepared. B: SEM image of the calcite thin section (approximately 10 µm wide and 100 nm at its thinnest). 128 Figure 3.8. Energy dispersive X-ray elemental map overlaid onto high angle annular dark-field image (z contrast) of the calcite surface showing a thin uranium rich (green) coating

14 Figure 3.9. High magnification image of a calcite surface coated with a uranium containing phase. The selected area electron diffraction image (Figure 11) was collected from site Figure A: High magnification TEM image of the uranium containing phase and a distorted calcite surface B: High magnification TEM image of a 'pristine' calcite surface. 131 Figure Selected area electron diffraction image collected from U(VI) precipitate (Figure 9, site 1) overlain with powder diffraction data calculated from crystallographic data for calcium uranate. 132 Figure k 3 -weighted χ functions with Fourier transforms for: (A) 0.10 g calcite reacted with 50 ml 21.0 µm U(VI) OCL solution; (B) 0.10 g calcite reacted with 50 ml 42.0 µm U(VI) OCL solution; (C) U(VI) coprecipitated with calcite standard. Solid lines represent experimental data while dashed lines are the fits using the parameters listed in Table Figure Normalised XANES spectra for the U L III edge of a 42.0 µm U(VI) reacted calcite sample (1 : 500) and Ca-uranate and schoepite standards. 135 Figure SAXS data collected from a 42.0 µm U(VI) calcite equilibrated YCL solution, aged for 18 months. Solid line represents the experimental data, the dashed line the modelled intensity, and the dotted lines indicate the relative contribution of the two particle populations. The gradient of the Porod region was calculated as Figure 4.1. Solubility of Np 2 O 5 (red) and, aged (blue) and fresh (green) Np(V)O 2 (OH) (am) solid phases as a function of ph in NaOH systems (PHREEQC predictions). The points at ph 10.5 and 13.3 represent observed Np(V) solubility in the calcite equilibrated YCL and OCL systems

15 Figure 4.2. Observed Np(V) solubility after 24 hours as a function of calcium concentration in ph 13.3 NaOH systems with an initial Np(V) concentrations of 4.20 µm (A) and 42.0 µm (B). Triangles/dotted lines represent Np(V) solubility against initial Ca concentration. Circles/solid lines represent Np(V) solubility, but rather then initial Ca concentrations show calculated [Ca 2+ ] after precipitation of the Np(V) phase. The horizontal line represents observed solubility in ph 13.3 NaOH solutions in the absence of added Ca Figure 4.3. XANES spectra recorded from (A): a Np(V) precipitate recovered from a 42.2 µm NpO + 2 calcite equilibrated YCL solubility experiment (red line) and reference spectrum of Np(V) in HNO 3 (blue line); (B) XANES reference spectra recorded from two UO 2+ 2 mineral phases, schoepite and calcium uranate (blue and red lines, respectively). The arrows indicate the O ax resonance feature in each spectrum. 175 Figure 4.4. Sorption isotherm from OCL and YCL batch sorption experiments. Lines represent linear fits to the data with gradients of 0.90 ± 0.12 and 1.69 ± 0.20 L Kg -1 for OCL and YCL systems, respectively. 179 Figure 4.5. Predicted Np(V) sorption at equilibrium against observed Np(V) removal from solution at apparent equilibrium in OCL systems. Circles show data that was used in calculation of model constants, triangles show points that were blind predictions of experimental data. 181 Figure µm Np(V) sorption to calcite solid phase in YCL solution with solid to solution ratios of 1 : 10 (circles), 1 : 50 (diamonds) and 1 : 500 (triangles) with an additional calcite free control (square). 182 Figure 7. Np(V) sorption to calcite in OCL with 1.62 x 10-3 (A), 1.62 x 10-2 (B), 0.16 (C) and 1.62 (D) µm Np(V) concentrations and solid to solution ratios of 1 : 500 (triangles), 1 : 50 (diamonds) and 1 : 10 (circles) with additional calcite free controls (squares). Lines are model calculations: A and B were used in the fittings while B and C were blind predictions

16 Figure 4.8. Schematic of processes used to describe the kinetics of Np(V) removal in OCL systems. 185 Figure 5.1. Plot of Eu 3+ remaining in solution relative to NRVB mass; equilibrated for 24 hrs; no ionic buffer; solution volume = 10 ml; [Eu T ] = 7.91 x M. 197 Figure 5.2. Pie Chart showing Eu size fractionation in NRVB water determined by ultrafiltration. 197 Figure 5.3. Plot of Eu 3+ sorption (plotted as a percentage of Eu remaining in solution) to NRVB versus equilibration time, as a function of NRVB mass (g); Solution volume = 10 ml ; I = 0.1 M; [Eu T ] = 7.91 x M. 197 Figure 5.4. Plot of Eu 3+ sorption as a function of EDTA concentration; solution volume = 10 ml; 0.3 g NRVB; [Eu T ] = 7.91 x M. 197 Figure 5.5. Eu 3+ desorption with the addition of EDTA (0.1 M) with varying Eu 3+ / NRVB (0.3 g) equilibration times; [Eu T ] = 7.91 x M; [NaClO 4 ] = 0.1 M. 197 Figure 5.6. Percentage of Eu remaining in solution: experimental data, (symbols) for systems with no added Na (squares) and with added [Na] = 0.1 M (diamonds); predictions (lines) of variation in Eu solution loading based on experimental values for the highest and lowest mass systems assuming a pure surface complexation mechanism (full line for systems with no added Na; dashed lines for systems with added [Na] = 0.1 M). 197 List of Tables Table UK radioactive waste and materials inventory. 17 Table 1.2. Activity of Cm and Am isotopes in UK HLW and ILW waste inventories,

17 Table 1.3. Ionic radii for six and eight coordinated trivalent Eu, Am and Cm. 44 Table 2.1. Synthetic cement leachate solution composition. 66 Table 3.1. Synthetic cement leachate solution composition. 108 Table 3.2. PHREEQC calculated saturation indices (SI*) for supersaturated U(VI) phases at 4.20 µm in YCL and OCL with and without calcite equilibration. 113 Table 3.3. PHREEQC calculated U(VI) speciation (4.20 µm) as fractional composition in YCL and OCL with and without calcite equilibration. 114 Table 3.4. EXAFS best fit parameters for U(VI)-calcite samples. 136 Table 3.5. Summary of fitted parameters in SAXS model. 138 Table 3.6. Summary of mechanisms of U(VI) uptake to calcite solids in calciteequilibrated OCL systems. 139 Table 4.1. Synthetic cement leachate solution composition. 167 Table 4.2. PHREEQC calculated saturation indices for supersaturated Np(V) phases at 42.0 µm in YCL and OCL with calcite equilibration. 170 Table 4.3. PHREEQC calculated Np(V) speciation (42.0 µm) as percentage composition in YCL and OCL with calcite equilibration. 176 Table 4.4. Summary of calcite-solution surface reactions and constants used in surface complexation model

18 Table 5.1. Europium speciation (expressed as percentages) predicted by PHREEQC for solutions in equilibrium with the phases in NRVB (See text for details of the calculations). Data given for systems with no added Na and for solutions with 0.1 M added NaClO4. The final column is for a calculation with final ph forced to Structure of thesis This thesis is being presented by papers, with a introduction (Chapter 1), a methodology/instrumentation chapter (Chapter 2), and three experimental chapters: Chapter 3 is a manuscript accepted for publication in Geochimica et Cosmochimica Acta (2014) that studies the interaction of uranyl(vi) with calcite solids in synthetic hyperalkaline cement leachates, designed to be representative of a young and an old cementitious GDF. This study involved a wide range of experimental methods, such as: batch experiments; XANES spectroscopy; EXAFS spectroscopy; luminescence spectroscopy; TEM; SAXS; and thermodynamic and kinetic modelling to gain a mechanistic understanding of U(VI) uptake to calcite. Kurt Smith was the primary author and researcher. Adam Swinburne and Pieter Bots assisted in luminescence and SAXS data acquisition, respectively. The remaining authors provided guidance and editorial support for the manuscript. Chapter 4 is a pre-submission manuscript, which investigates the solubility and sorption of the neptunyl(v) cation in contact with calcite solids in geochemical conditions, consistent with a old and a young leachate from a cementitious GDF. This study utilised a number of techniques including batch uptake, XAS and involved a significant thermodynamic/kinetic modelling component to gain an understanding of the mobility controlling mechanisms under high ph conditions. Kurt Smith was the primary author and researcher. The remaining authors provided guidance and editorial support for the manuscript. 13

19 Chapter 5 is a manuscript published in Mineralogical Magazine (2012) in collaboration with another PhD student, Ryan Telchadder. The paper investigates the uptake of Eu(III) to NRVB cement, one of the proposed backfill materials for a cementitious GDF. NRVB consists of approximately 20% calcite. This is done in contact with EDTA as a competing ligand in order to assess the strength of uptake and possible removal mechanisms. Kurt Smith co-wrote and edited the text, and contributed experimental data either alone (Figure 1, 5) or with Ryan Telchadder (Figure 2, 3) throughout the chapter and carried out all geochemical calculations (Table 1, Figure 6). The main conclusions of this thesis are then laid out in Ch. 6 (conclusions). 14

20 1. Introduction The UK has had a nuclear programme for many years, for weapons development through to current power generation activity. Furthermore, in very recent years, the Government has signalled the start of a new power reactor construction programme to maintain energy security and meet climate change obligations (Department for Environment Food and Rural Affairs, 2008). The Government has decided that continued surface storage of by-products from the nuclear fuel cycle cannot continue indefinitely, due to increased proliferation and terrorist risk. Therefore, it is UK Government policy to construct a deep subsurface geological disposal facility that can contain these wastes over geological time, until their activity has reduced to near background levels The Nuclear Fuel Cycle The term nuclear fuel cycle relates to the whole process of electricity generation via nuclear fission. The front-end of the nuclear fuel cycle consists of mining, which can be carried out by conventional or in situ leaching techniques, enrichment, fuel fabrication and power generation. Some older reactor designs can generate electricity from natural, ca. 0.7% 235 U uranium (specifically MAGNOX / CANDU reactors), however, this produces a low energy density and thus a large and relatively inefficient reactor core. In light of this, most modern reactors require uranium fuel to be enriched to up to 4 % in fissile 235 U, which allows a smaller reactor core to be engineered and constructed. Uranium can be enriched in one of two ways, by energy intensive gas diffusion or technically difficult centrifugation techniques (Wilson, 1996). The enriched uranium is then manufactured into a suitable fuel unit, which can be used in a reactor, typically ceramic UO 2 pellets. 15

21 The back end of the fuel cycle is concerned with treatment of the irradiated fuel, disposal of mining wastes, and decommissioning of obsolete nuclear facilities and the wastes associated with them. Following irradiation, the spent fuel is removed from the reactor and held for several years in a cooling pond, while the shortest lived isotopes decay. Depending on the country, the fuel cycle can then take two paths, an open or closed cycle. In countries that operate an open fuel cycle, the fuel will be isolated from the biosphere in geological storage directly as spent fuel and without reprocessing. In a closed fuel cycle (for example the UK runs a partially closed fuel cycle) the fuel is removed from the ponds, its cladding removed and the contents dissolved in nitric acid. It then undergoes the PUREX (Plutonium URanium EXtraction) process, where it is reprocessed to separate the uranium, plutonium, minor actinides and fission products (Wilson, 1999). In the UK, new build reactors have no commitment to reproccessing of spent fuel (DEFRA, 2008). The plutonium can then be recycled by the manufacture of mixed (uranium and plutonium) oxide fuel pellets although no power reactors in the UK are licensed to burn MOX. The separated fission products are contained in a High Activity Raffinate (HAR), which is generally vitrified into a durable glass wasteform. In the UK, Low Level Waste (LLW, < 4 GBq/Tonne ɑ and < 12 GBq/Tonne βγ) undergoes shallow burial at the Drigg LLW repository, Cumbria, UK (Fox et al., 2006), while Intermediate Level Waste (greater activity than LLW, with no heat generation) and High Level Wastes (greater activity than LLW with heat generation) are held in interim storage whilst a long term solution for these materials is found (DEFRA, 2008). The UK's inventory of nuclear materials is shown in Table 1.1. It is currently UK Government policy to construct an underground Geological Disposal Facility (GDF) which is intended to contain radioactive wastes from the nuclear cycle over geological time, until their activity has reduced to near background levels. 16

22 Table UK radioactive waste and materials inventory (NDA and DEFRA, 2008; DEFRA, 2008). Packaged Volume Radioactivity (April, 2040) Materials m 3 % TBq % HLW 1,2,3, x x ILW 1,2, x x LLW 1,2, x <1.00 x Spent Fuel 1,4, x x Pu 1,4, x x U 1,4, x x Total 4.77 x x Consistent with the 2007 UK Radioactive Waste Inventory (NDA and DEFRA, 2008). 2 Packaging assumptions for HLW, ILW and LLW not suitable for disposal at the existing national LLWR are taken from the 2007 UK Radioactive Waste Inventory (NDA and DEFRA, 2008). 3 The HLW packaged volume may increase when the facility for disposing the canisters, in which the vitrified HLW is currently stored, has been implemented. 4 Packaging assumptions for Pu and U and spent nuclear fuels are taken from the CoRWM Baseline Inventory (Committiee on Radioactive Waste Management, 2005) is the assumed start date for the geological disposal facility. The Government has started a public consultation process on the placement of a GDF and has asked for communities willing to host a repository to volunteer (Department for Environment Food and Rural Affairs, 2008). In January 2013, the west Cumbrian local authority withdrew from the GDF selection process. This highlighted the complex political, scientific and societal factors which make GDF implementation a unique challenge. Once a decision on GDF placement and design is made, implementation can begin. No new nuclear power generation will be allowed without a pathway for disposal. Any GDF is likely to contain large quantities of cement, because of the knowledge base that has been built around the long-term behaviour of cementitious waste forms (ILW, LLW) in a GDF and because of the intrinsic use of cements in construction projects. This knowledge and experience was developed as part of the 17

23 NIREX programme that ran from the 1970s to the 1990s and resulted in an unsuccessful planning application for underground rock laboratory in the 1990s. The NIREX disposal concept was based around a cementitious repository for ILW to be sited near Sellafield. Although this programme was cancelled, a significant amount of the UK ILW inventory has already been processed into cement wasteforms, and so it is highly likely that the eventual UK ILW repository will be cementitious to some extent. In the UK, the GDF will contain approximately 365,000 m 3 conditioned ILW waste (Table 1.1). Therefore, even if the backfill is not cement, there will be large quantities of grout and engineering cement within the facility and this is a driver for the research described in this thesis. In this generic ILW disposal model, upon waste emplacement and facility abandonment and resaturation, the cement inside the GDF will undergo hydrolysis and generate hyperalkaline (ph > 10) conditions in and around the repository, which will be likely to alter key radionuclide sorption and transport properties significantly (NDA, 2011) The Geological Disposal Facility It is likely that any UK GDF design will be based on the multi-barrier concept (Figure 1.01) and may be designed to contain higher activity wastes (both ILW and HLW) i.e. the codisposal concept (Chambers et al., 2003; Folger, 1996). The GDF will not contain LLW which currently undergoes shallow burial at the Drigg LLW repository. The core principle of the multi-barrier concept is to have a series of engineered barriers that complement the host geology to minimise radionuclide release to the environment over long timescales. As part of the GDF design, wastes will be incorporated / encapsulated in a suitable wasteform. In the case of high activity raffinate, vitrification will be undertaken, whereas with many other wastes a cementitious wasteform will be sufficient (Fox et al., 2006). The wasteforms will 18

24 then be placed inside a container, for example cementitious wasteforms will be placed into steel drums (Figure 1.02) in preparation for disposal. These wasteforms will then be disposed in a series of deep underground tunnels, shafts and drifts during waste emplacement which are ultimately abandoned under controlled conditions when waste emplacement ceases. In a UK context, the GDF will probably be built at a depth of 200 m to 1000 m (CoRWM, 2006; DEFRA, 2008). Figure Illustration of the multi-barrier disposal concept (NDA, 2010). 19

25 Figure Cutaway of a steel drum containing cementitiously immobilised solid waste (Chambers et al., 2003) Backfill Material As part of the multi-barrier concept, the GDF complex needs to be 'backfilled'. The backfill material is designed to retard radionuclide migration. Internationally, there are two categories of backfill materials; impermeable and alkali (Chambers et al., 2003; Folger, 1996). Both of these materials would retard radionuclide mobility, but through different mechanisms. An impermeable backfill requires a high cation exchange capacity and a low hydraulic conductivity (Madsen, 1998), such a backfill would be suitable for HLW (e.g. glass and spent fuel), and works by preventing the movement of water to and from the waste. An alkali backfill would instead promote hydrolysis and sorption by causing surfaces to become negatively charged. This increases the attraction between the surface and any cationic radionuclides (although it should be noted that decreasing attraction between the surface and negatively charged radionuclides will occur). The high ph will also promote hydrolysis and precipitation of metal ions. The typical impermeable backfills for HLW disposal are phyllosilicates, such as bentonite (Fernandez et al., 2006; Karnland et al., 2007; Madsen, 20

26 1998; Meleshyn et al., 2009; Pedersen, 1999; Perdrial et al., 2009; Smellie and Karlsson, 1999). In the UK, and as a result of the NIREX program, historically most research has focused on ILW GDF design and alkaline backfill material. Approximately 20% of the UK ILW stockpile has already been processed into cement wasteforms (NDA and DEFRA, 2008; Morris et al., 2011) and therefore it is clear that any ILW GDF will need to accommodate cement based materials (Braney et al., 1993; Chambers et al., 2003; Clark et al., 1994; Fernandez-Carrasco et al., 2000; McCarter et al., 2004; Pointeau et al., 2001; Swanton et al., 2009; Wieland and Spieler, 2001; Wieland et al., 2004) Host Geology The final component in the multi-barrier concept is the host geology, which is likely to have a large affect on the design of the GDF with hydrogeological, thermal, chemical and mechanical properties typically taken into account in any safety case (Baldwin et al., 2008). Many European communities are relatively advanced in their GDF programmes and are considering differing geologies for their host rock, including crystalline, sedimentary and evaporite deposits (Baldwin et al., 2008; Morris et al., 2010). However, the UK s plans for disposal are based on volunteerism (DEFRA, 2008). This process has not yet identified a suitable site, and so at the time of writing, the UK host rock type is not known. The engineered barriers are designed to contain the radionuclides for several thousands of years, but subsurface evolution will undoubtedly cause the barriers to decay over tens of thousands of years, and it is clear that a fraction of the remaining radioactivity is likely to migrate into the surrounding environment (Department for Environment Food and Rural Affairs, 2008). 21

27 1.5. Cementitious Backfill and the Chemically Disturbed Zone Cement is regularly discussed as a potential backfill material as it is a tried and tested engineering material (although, not over geological time), which is porous and permeable. This will allow relatively uniform geochemical conditions to form in the GDF, whilst allowing excess gasses formed by the radiolysis of water, microbial activity and chemical reactions in the waste to escape into the surrounding environment (preventing pressure increases; Chambers et al., 2003). In a cement based repository, the hyperalkaline conditions are the basis for containment. Therefore, it is important that the GDF has enough buffering capacity to maintain alkaline conditions for the lifetime of the repository (~ 1 x 10 6 a) (Baldwin et al., 2008). UK research to date has focused on Nirex Reference Vault Backfill (NRVB) cement which was the proposed backfill in the NIREX disposal programme (Chambers et al., 2003). NRVB consists of Ordinary Portland Cement, limestone flour and hydrated lime in a 256 : 291 : 100 ratio, mixed with water in a 1 : 8 ratio (McCarter et al., 2004). Upon hydration of cement with water, CSH (Calcium-Silica-Hydrate) gels form (xcao.sio 2.yH 2 O) with Ca/Si ratios between 0.7 and 1.7 (Pointeau et al., 2001). The Ca/Si ratio typically decreases with time hydrated (Pointeau et al., 2001). CSH phases typically have a high affinity for cations and anions (Pointeau et al., 2008; Pointeau et al., 2001; Tits et al., 2011; Viallis-Terrisse et al., 2002). This, coupled with high specific surface areas (200 m 2 g -1 ), means CSH gels can account for the majority of the reactivity of hydrated cements (Schlegel et al., 2004). Using models of dissolution in pure water and only considering Portlandite (Ca(OH 2 )) and CSH (Calcium-Silica-Hydrate) gel, Chambers et al. (2003) modelled the evolution of ph in a GDF. It is accepted that dissolution of KOH and NaOH will generate a ph of ~ 13 and then Ca(OH) 2 will generate an alkalinity of ph 12.5 before the incongruent dissolution of CSH gels decreases the ph continuously from ph 12.5 to

28 (NDA, 2010). Finally, the congruent dissolution of CSH gels generates a constant alkalinity of 10.5 after approximately 500,000 years (Chambers et al., 2003). There are many processes that can alter the effectiveness of the cement conditioning, including corrosion. Iron will be one of the most abundant elements in the repository due to its use in waste forms and engineering of the GDF. The likely presence of iron in the host rock and corrosion of iron metal will generate reducing conditions (Chambers et al., 2003). The formation of reducing conditions is hypothesised to significantly decrease the mobility of most redox sensitive radionuclides, such as the actinides (U, Np, Pu), which are more prone to hydrolysis and precipitation in their reduced oxidation states. However, coupling of ph and redox has been shown to increase the stability of higher oxidation states of some actinides in high ph environments (Gaona et al., 2012; Gaona et al., 2013) Calcite Relevance Calcite is the most common calcium carbonate mineral in the environment (in comparison to aragonite, vaterite and amorphous calcium carbonate). Calcite has a rhombohedral crystal structure with alternating planes of six-coordinated calcium cations and carbonate anions with unit cell parameters a and c of 4.99 and Å, respectively (Reeder, 1983). The properties of calcite give it strong sorption characteristics for various metal ions including: Co 2+ ; Cd 2+ ; Mn 2+ ; Zn 2+ ; Mg 2+ ; Ba 2+ ; Sr 2+ ; UO 2+ 2 ; Ni 2+ ; NpO + 2 ; La 3+ ; Eu 3+ ; Sm 3+ ; PuO + 2 ; and PuO 2+ 2 (Bailey et., 2005; Dong et al., 2005; Elzinga et al., 2004; Geipel et al., 1997; Hay et al., 2003; Heberling et al., 2008a; Parkman et al., 1998; Rihs et al., 2004; Sawada et al., 2005; 23

29 Zachara et al., 1991; Zavarin et al., 2005). Calcite is ubiquitous in the environment and will likely be present in the GDF host rock, regardless of site selection. It also makes up a significant quantity (29 % by mass) of the proposed NRVB backfill material (NDA, 2010). In addition, in an alkali based GDF the portlandite component of the cements used in construction and for the backfill will undergo carbonation to calcite (Dow and Glasser, 2003) via reaction with carbonic acid. Thus, the presence of calcite within a repository is likely to be ubiquitous and to increase with respect to time Surface Charge Surface charge affects the sorption of ions to mineral surfaces, and therefore an understanding of the surface potential of calcite is required. Thermodynamic considerations suggest that the main surface potential determining ions are Ca 2+, HCO - 3, CO 2-3, H + and OH - (Somasundaran and Agar, 1967). However, experimental results show that the major ions for determining surface potential are Ca 2+ and CO 2-3, rather than H + and OH -, over a ph range of 7-12 (Thompson and Pownall, 1989). The consumption of H + and OH - in a solution with calcite are balanced at approximately ph 8.2 by the hydrolysis of surface carbonate, calcium, and sorption at the mineral-water interface (Somasundaran and Agar, 1967), i.e. the point of zero charge (ph PZC ; Figure 1.03). 24

30 Figure Change in ph of the aqueous solutions in presence of calcite and air (Somasundaran and Agar, 1967). The ph PZC of calcite actually varies from sample to sample, depending on the mineral history and source. This is because carbonates experience surface charging differently from simple oxide minerals (Stipp, 1999). In oxide minerals, the PZC is controlled as a function of successive H + chemisorption onto surface O - functional groups to give OH and OH + 2. However, termination of bulk calcite (and carbonates in general) creates open bonds in surface Ca 2+ and O atoms, and thus an alternating cationic/anionic surface. Hydrolysis of a layer of OH - and H + is required to address the local charge imbalances (Stipp, 1999) with additional calcium and carbonate species sorbing in the Stern layer (Figure 1.04) creating the overall surface charge. Therefore, the point of zero charge for calcite differs from oxide minerals in that it is dictated by Ca 2+ and CO 2-3, rather than H + and OH - (Foxall et al., 1979; Stipp, 1999). Hence, in the context of carbonate phases, a ph PZC value is meaningless, and instead surface neutrality should be considered in terms of pca PZC/pCO 3 PZC, which is 25

31 indirectly controlled by solution ph. Stipp (1999) rearranged the Electrical Double Layer model (EDL) for the calcite surface, taking into account the differences in surface charging between oxides and carbonates (Figure 1.04). The zeta potential of the calcite surface is closely related to the charge that the ions will experience as a result of surface charging. Figure 1.05 shows how the zeta potential changes as a function of ph with different electrolyte solutions. 26

32 Figure Modified electrical double layer model for the calcite surface. ζ = zeta potential; ψ = potential at the stern layer; and δ = represents the partial charge on the ions that were left behind at the bulk termination. At X = 0, a layer of H and OH are chemi-sorbed onto the bulks dangling bonds. At X < 0, potential is so high that attaching ions do not sorb, they precipitate. Taken from Stipp (1999). 27

33 Figure The influence of ph on the zeta potential of calcite taken from Somasundaran and Agar (1967) Carbonation The UK s nascent plans for a cement based GDF increase the potential importance of calcite beyond the host rock environment. Carbonation takes place due to dissolution of portlandite by water (1.1). Ca(OH) 2 2 Ca 2OH (1.1) The released alkalinity increases the ph and the capacity and rate of CO 2(g) dissolution (1.2). Initially, the CO 2 dissolves to give carbonic acid (1.2) CO 2 H2O H2CO3 (1.2) The H 2 CO 3 then reacts with OH - to produce bicarbonate and then carbonate (1.3, 1.4). - H2CO3 OH HCO3 H2O - 2- HCO3 OH CO3 H2O (1.3) (1.4) 28

34 The carbonate reacts with dissolved calcium to give calcite (1.5). CO Ca CaCO 3(s) (1.5) Thus, over time much of the available portlandite component of the cement in a repository will be converted to calcite Uranium Relevance Uranium forms a key component of nuclear fuel and is typically the most abundant radionuclide by mass in radioactive wastes. The two most common isotopes of uranium are 235 U (t ½ = 703 Ma; 4.67 MeV) and 238 U (t ½ = 4.46 Ga; 4.26 MeV). They both undergo decay via emission of an alpha particle to 231 Th and 234 Th, respectively. Uranium constitutes the bulk of spent nuclear fuel when measured by mass, although this is not the case in terms of radioactivity. Given uranium's role in the nuclear industry, it could be present in a GDF environment in large quantities (> 80,000 m 3, Table 1.01), and therefore its mobility under these conditions needs to be determined Speciation and environmental behaviour Under natural environmental conditions, uranium is stable in two oxidation states, U(IV) and U(VI). U(IV) has a high charge density and undergoes rapid hydrolysis and thus, typically exhibits low solubility and often precipitates as uraninite (UO 2 ). However, in oxidising conditions U(VI) rapidly forms the linear uranyl moiety UO 2+ 2 (Kaplan et al., 1998). The stability of the UO 2+ 2 ion significantly increases the solubility of U(VI) in comparison to U(IV). Thermodynamic calculations can be used to determine U(VI) speciation across a range of environmental conditions (Figure 1.06). Free UO 2+ 2 dominates the speciation in acidic conditions (< ph 5), however, in more neutral conditions (ph 5-8) and in the absence of 29

35 carbonate, the speciation tends to be dominated by hydroxide species, such as UO 2 (OH) + and (UO 2 ) 3 (OH) + 5. The UO ion forms very stable complexes with the CO 3 ligand (Runde, 2000). The concentration of dissolved carbonate increases with ph for systems in equilibrium with the atmosphere, thus U(VI) speciation is very sensitive to ph and carbonate concentration. If thermodynamic calculations are carried out allowing the systems to equilibrate with atmospheric CO 2, the presence of carbonate species, such as UO 2 (CO 3 ) 2-2 and 4- UO 2 (CO 3 ) 3, will dominate from approximately ph 7 upwards. The stability of the U(VI)- carbonate complexes drastically increases the solubility and mobility of U(VI) in comparison to carbonate free conditions (Barnett et al., 2000; Runde, 2000). In the absence of dissolved - carbonate, hydroxide species, such as UO 2 (OH) 2(aq), UO 2 (OH) 3, and UO 2 (OH) 2-4 dominate speciation through to hyperalkaline conditions (Choppin, 2006). U(VI) sorption by mineral phases is strongly controlled by surface charge and speciation. In acidic waters, mineral surfaces tend to become positively charged. The uranium speciation also tends to become cationic. For oxides, H + will also compete for surface binding sites. Therefore, in acidic conditions, sorption is reduced because of the smaller net attraction between the positively charged surface and cationic uranyl species (Davis et al., 2006). In addition, the formation of the UO 2+ 2 ion enhances the effective charge of the central ion to +3.3 (Choppin, 2006; Runde, 2000). As conditions become more alkaline, the mineral surfaces become more negatively charged, while the uranyl species remains cationic. This increases the net attraction between uranyl and mineral surfaces, and corresponds to the optimal sorption shown in Figure Typically, this strong sorption behaviour would be expected to carry on to high ph, with increasing sorption. However, carbonate can compete with the surface binding sites for U(VI) (Baik et al., 2003). 30

36 Figure Speciation of 1 ppm U(VI) and log total dissolved carbonate (broken line) as a function of ph. Log P CO2 = -3.5 in 0.1 M NaNO 3. Taken from Barnett et al. (2000). Figure Adsorption as a function of ph to subsurface sediments, where different symbols represent sediment collected from three different U.S. Department of Energy sites. Solid/solution ratio = 3.33 g/l in 0.1 M NaNO 3 with log P CO2 = Taken from Barnett et al. (2000). 31

37 1.7.3 U(VI) interaction with calcite Trace amounts of uranium are known to occur in natural calcite. Sturicho et al. (1998) report a highly uniform uranium concentration of 5-35 ppm in a range of calcite samples from the natural environment. X-ray Absorption Near-Edge Structure (XANES) spectroscopy showed that the uranium was predominantly in the tetravalent oxidation state, and Extended X-ray Absorption Fine Structure (EXAFS) spectroscopy showed that the uranium occupied a regular and well defined crystallographic site in natural calcite (Sturchio et al., 1998). They also reported that the main difference between the local calcium and uranium coordination was that the equatorial oxygen shell was split. For calcium, the structure of calcite requires a single equatorial U-O distance of 2.36 Å (N = 6), whereas, the EXAFS U(IV) data were modelled as two equatorial U-O shells at 2.21 Å (N = 3.4 ± 0.8) and 2.78 (N = 2.8 ± 0.9) Å. Tetravalent uranium has an ionic radius of 0.89 Å, a value in the same range of divalent ions (e.g. Zn 2+, Mn 2+ and Cd 2+ ) that can be accommodated in calcite ( Å; Sturchio et al., 1998). The substitution of U(IV) into the calcite structure would cause a charge imbalance. Sturchio et al. (1998) suggested that this imbalance could be resolved by invoking a non-local coupled substitution with sodium, a common impurity in calcite (Ishikawa and Ichikuni, 1984; 1.6); 3Ca 2 U 4 2Na (1.6) Kelly et al. (2003) studied the incorporation of U(VI) in calcite and reported that their sample contained 80 to 500 ppm U(VI). They used EXAFS spectroscopy to show that the U(VI) occupies a well defined site within the calcite lattice (Figure 1.08). This site was modelled with 2 O atoms at 1.80 Å (axial) and 3.8 ± 0.4 equatorial O atoms at 2.41 Å. Additionally, Kelley et al. (2003) fitted 4.3 ± 2.7 C atoms at 3.51 Å which are part of the carbonate ligand. Finally, the authors modelled the EXAFS data with two Ca shells (N = 2.3 ± 0.4; 3.7 ± 0.4) at 32

38 3.78 and 4.01 Å, respectively. The results of Kelly et al. (2003) show that UO 2+ 2 can substitute for a Ca 2+ and two CO 2-3 as shown in Figure This results in a net charge around the uranyl site of +4. In common with Sturchio et al. (1998), the authors relied on the coupled, non-local substitution of sodium to address the local charge balance (1.7). The authors hypothesised any structural defects caused by the absence of the two carbonate ligands could be accommodated by substitution of water. 5Ca 2 UO 2 2 4Na (1.7) Figure A) Local coordination of Ca 2+ in the calcite structure. B) The local structure of UO 2 2+ incorporated in calcite. Taken from (Kelly et al., 2003). 33

39 There is very little information regarding the sorption of U(IV) to calcite mineral surfaces available in the published literature, due to the very low solubility of uraninite (UO 2 ) (Evans et al., 2011). However, U(VI) sorption and coprecipitation with calcite is well documented at circumneutral ph (Dong et al., 2005; Elzinga et al., 2004; Geipel et al., 1997; Reeder et al., 2004b; Reeder et al., 2000; Reeder et al., 2001; Rihs et al., 2004; Wang et al., 2005). The work of Elzinga et al. (2004) studied the reaction of U(VI) with calcite at ph 7.4 and 8.3 and over the concentration range 5 to 5000 µm U(VI) using batch sorption experiments, EXAFS and luminescence spectroscopy. EXAFS analysis of their batch experiments with < 500 µm U(VI) produced data consistent with the formation of uranyl triscarbonate surface complexes and suggested that sorption in their systems was taking place via inner sphere complexation. Unfortunately, their EXAFS analysis could not identify any calcium backscatter to corroborate the proposed inner sphere bonding, and thus the exact coordination environment of U(VI) in this system remained uncertain. The authors then used luminescence spectroscopy to explore whether different U(VI) surface species were present in their experiments. At < 100 µm U(VI), a calcium uranyl triscarbonate complex (in a liebigite-like structure) dominated, whilst for systems in the > 100 µm < 500 µm range, there was evidence that a different, undetermined uranyl carbonate species became significant. The authors suggested that the species was likely to be intermediate between the liebigite-like surface complex and U(VI) incorporated into calcite. X-ray absorption fine structure spectroscopy and luminescence techniques confirm that when UO 2+ 2 is coprecipitated with calcite, it can exist in multiple coordination environments, dependent on the method of precipitation (Reeder et al., 2000b; Reeder et al., 2001b). 34

40 When calcite grows, non-equivalent steps are formed along the <481> and <441> planes (Figure 1.09). In the calcite structure, all carbonate groups are co-planar, exposed at the + steps and orientated so that they are tilted down. However, on the steps the orientation is opposite with exposed carbonate tilted upwards (Figure 1.09; Reeder et al., 2004b). Different metal ions have different preferences for sorption onto either + or step regions on carbonate minerals. In a Ca 2 UO 2 (CO 3 ) 3 complex, the axial oxygens are orientated perpendicular to the carbonate ligands within the equatorial plane. Therefore, at the + steps, the lower axial oxygen interferes with atoms in the lower layer, resulting in unrealistically short O-O and Ca- O distances, and thus sorption of uranyl is sterically hindered (Reeder et al., 2004b). However, at the step, the lower axial oxygen is displaced allowing sorption (Figure 1.09). Elzinga et al. (2004) and Rihs et al. (2004) also suggest that uranyl sorption to calcite is best explained by surface complexation of a calcium-uranyl-carbonate species, via inner-sphere bonding with a surface carbonate. Reeder et al. (2004) carried out fluorescence imaging on U(VI) doped calcite crystals in an attempt to get a better understanding of the sorption behaviour. They proposed that, because the flat portions of the crystal have no structural differences, the differences in uptake are expected to relate to different structure and coordination in the + and - steps. The uranyl can only be accommodated on the calcite surface where it is adjacent to a Ca vacancy or small etch pit (Rihs et al., 2004), which is equivalent to a - step. Rihs et al. (2005) suggested that if this is the case, it implies that one Ca 2+ cation would be released in solution for each uranyl sorbed to the surface, which should give a linear relationship between U(VI) sorption and dissolved Ca 2+. Recent computational molecular dynamics simulations (Doudou et al., 2012) supported preferential sorption of the Ca 2 UO 2 (CO 3 ) 3(aq) species onto the '-' steps, but also suggested that outer sphere complexation on the calcite terrace should be thermodynamically favourable. In addition to surface complexation and incorporation, calcite has been shown to promote the formation of U(VI) 35

41 precipitates, such as schoepite (Carroll et al., 1992; Elzinga et al., 2004; Geipel et al., 1997; Schindler and Putnis, 2004). Figure (A) Ball and stick representation of positive vicinal region of calcite. (B) geometries of and + step regions along the <481> and <441> directions. (C) Ball and stick representation of negative vicinal region of calcite with uranyl complexed with the surface. Taken from Rihs et al. (2004). 36

42 1.8. Neptunium Relevance 237 Neptunium has a half life of approximately 2.14 Ma and primarily decays to 233 Pa via the emission of an alpha particle (4.95 MeV). In the nuclear fuel cycle, 237 Np forms from various neutron reactions, such as successive neutron capture by 235 U and the decay of 241 Am. 237 Np accounts for a significant proportion of the radioactivity present in spent fuels and exhibits complex environmental behaviour. Is seen as one of the most mobile actinides (Kaszuba and Runde, 1999; Runde et al., 1996). This, coupled with the long half-life of 237 Np, means that it is significant in any GDF safety case for a prolonged period of time (Figure 1.10; Morris et al., 2010). Figure Activity of radionuclides in HLW from reprocessing of Spent Nuclear Fuel (SNF) with time after processing. Taken from (Ojovan and Lee, 2005) 37

43 Speciation and environmental behaviour Neptunium has two environmentally important oxidation states, Np(IV) and Np(V). Under reducing conditions, Np will exist in the Np(IV) oxidation state, and behave much like U(IV), i.e., it will undergo rapid hydrolysis resulting in relatively low solubility. However, in oxidising conditions, Np(V) will dominate and the Np will exist as the neptunyl(v) ion (NpO + 2 ), with an effective charge of +2.3 (Choppin, 2006; Runde, 2000). Due to its reduced charge to ionic radius ratio, NpO + 2 is less prone to hydrolysis than UO 2+ 2 and will typically + exist as free NpO 2 (aq) below ph 10. At ph >10, the speciation is dominated by NpO 2 OH (aq) (Figure 1.11). However, much like UO 2+ 2, NpO + 2 forms stable complexes with carbonate, such as Np(V)O 2 CO - 3, and thus the speciation is sensitive to carbonate concentration. In equilibrium with atmospheric CO 2, at ph 8, thermodynamic modelling predicts that 76 % of the neptunyl exists as the Np(V)O 2 CO species, with 24 % present as a free NpO 2 (aq) (Figure 1.11). By contrast, at ph 10.2 under the same conditions, the carbonate species represents 90 % of all Np(V) (Figure 1.11; Heberling et al., 2008c). NpO + 2 exhibits similar environmental 2+ behaviour to UO 2, with sorption increasing with ph (Figure 1.12), however, it is generally more mobile in comparison to UO 2 2+ due to the significantly lower effective charge.. Likewise, the stability of neptunyl(v) carbonate complexes tends to reduce sorption in systems at equilibrium with atmospheric CO 2 (Figure 1.12). Thus Np, and specifically its stability as Np(V), is considered the most mobile of the transuranic elements (Kaszuba and Runde, 1999; Runde et al., 1996). 38

44 Figure Np(V) speciation in solution (Np(V) = 1 x 10-5 M with I = 0.01 M) in a nitrogen atmosphere (top) and with log pco 2 = -3.5 log (bottom). Taken from Schmeide and Bernhard, (2010). 39

45 Figure Neptunium(V) absorption to montmorillonite as a function of ph, with and without atmospheric carbonate (Turner et al., 1998) Interaction with calcite There is little published work on Np(IV) sorption to mineral phases and nothing for calcite. However, it is reported that Np(IV), like other tetravalent actinides, has very low solubility (in absence of competing ligands) with a high affinity (as a coprecipitate) for mineral surfaces compared to Np(V) (Schemeide and Bernhard, 2010). In addition, there are only a handful of papers studying Np(V) behaviour in calcite systems. Heberling et al. (2008a) carried out a NpO + 2 calcite coprecipitation study under ph range , and studied the coordination of Np using EXAFS spectroscopy, similar to that of the uranyl study of Reeder et al. (2001). Heberling et al. (2008a) reported a Np-O eq distance of 2.4 Å. This is longer than the Ca-O eq distance in calcite, but it is indicative of 4-fold coordination around the neptunyl ion, and was 40

46 interpreted as four monodentate bound carbonate ions. Low Debye-Waller factors for the axial and equatorial oxygens were interpreted by the authors as an indicator of low structural disorder. However, increased Debye-Waller factors for the additional oxygen and carbon atoms in the carbonate ligand were interpreted by Herberling et al., (2008a) as disorder in the orientation of the carbonate ligands (Figure 1.13). However, the authors concluded that the NpO + 2 was stable in the calcite structure. The proposed substitution by neptunyl of Ca 2+ and two CO 2-3 in the calcite structure results in a charge imbalance of +3. A coupled substitution with sodium to address the charge imbalance has been suggested (Heberling et al., 2008c), which is similar to UO 2+ 2 studies (Kelly et al., 2003; Sturchio et al., 1998). Figure Model of neptunyl ion incorporated into the calcite structure with neptunyl located at the Ca 2+ site with the two axial neptunyl oxygens substituting for two carbonate ions. Taken from (Heberling et al., 2008c). 41

47 Heberling et al. (2008b) also studied the adsorption of neptunyl to the calcite surface (8 mm carbonate). Similarly to uranyl, neptunyl sorption was reported to peak at ~ ph 8.3. They also reported that at low (< 10 µm) neptunyl concentrations, the ph dependence of sorption was much stronger than at higher concentrations (> 10 µm). Kinetic experiments showed that after hours, the adsorption slowed down (Figure 1.14). However, after four weeks a slow sorption process could still be observed, showing the system had not reached equilibrium (Heberling et al., 2008a). The authors interpreted this kinetic behaviour as the slow dissolution/recrystallization of the calcite gradually incorporating Np(V). Figure Surface loading of 2 µm Np versus time with a solid to liquid ratio of 20 g L -1. Taken from (Heberling et al., 2008b). 42

48 1.9. Europium Relevance Europium is part of the lanthanide series and forms the trivalent Eu 3+ (aq) ion in the natural environment. Eu(III) is often used in nuclear research as an analogue for various trivalent minor actinides such as americium and curium because direct research on Cm(III) and Am(III) is very difficult due to the high specific radioactivities involved (Lee et al., 2006). Americium ( 241 Am t ½ = yr; 243 Am ½ = 7,370 yr) and curium isotopes (t ½ = > 4.7 x 10 3 yr) form significant proportions of the initial activity from spent nuclear fuels. Currently, there are significant quantities of various Cm and Am isotopes in the UK waste inventory (Table 1.2). Therefore, the quantities of the radionuclides and the half lives of the Am and Cm isotopes suggest that they will remain a significant source of activity over prolonged periods of time and are of importance to geodisposal (Figure 1.10). Table 1.2. Activity of Cm and Am isotopes in UK HLW and ILW waste inventories, 2007 (NDA and DEFRA, 2008) Activity (TBq) Isotope t ½ HLW ILW 241 Am 4.33 x x x Am 7.36 x x Cm x x Cm x x Cm x x Cm 8.50 x x Cm 4.73 x x Cm 3.40 x x x

49 1.9.2 Europium speciation Typically, europium exists in the +3 oxidation state and is usually present as a free Eu 3+ in acidic conditions (< ph 7). Above ph 7 hydrolysis takes place and there is a gradual shift of the speciation with increasing ph, from EuOH 2+ (ph 8.7) though to Eu(OH) + 2 (ph 9.0), and Eu(OH) 3(aq) (ph 10.6; Tertre et al., 2006). Above ph 10, the Eu(III) speciation is dominated by the Eu(OH) - 4 species (Figure 1.15). However, in the presence of carbonate, Eu(III) can form stable complexes (Choppin, 2006). The influence of carbonate becomes important above approximately ph 7, and increases with ph, as the total dissolved carbonate concentration increases (assuming equilibrium with the atmospheric CO 2 ; Figure 1.16). Am and Cm generally exist in the +3 oxidation state and exhibit similar chemistry to Eu(III) (Table 1.3), allowing it to be used, within limitations, as an analogue (Tits et al., 2005; Lee et al., 2006; Guillaumont et al., 2003). Table 1.3. Ionic radii for six and eight coordinated trivalent Eu, Am and Cm (Shannon, 1976). CN Eu(III) Am(III) Cm(III)

50 Figure Speciation of 1 x 10-5 M Eu 3+ in 0.5 M NaClO 4 as a function of ph in carbonate free conditions. Taken from Tertre et al. (2006). Figure Speciation of Eu(III) as a function of ph in equilibrium with 0.03% CO 2(g) (thick lines) or 1% CO 2(g) (thin lines). Taken from Zavarin et al. (2005). 45

51 Interaction with cement Calcite constitutes a significant proportion of NIREX Reference Vault Backfill (NRVB) cement. There are several studies in the literature of Eu(III) interaction with calcite. Lakshtanov and Stipp (2004) studied the coprecipitation of Eu(III) with calcite and demonstrated strong Eu(III) partitioning into the solid phase. The authors successfully modelled their data assuming the formation of a Eu 2 (CO 3 ) 3 -CaCO 3 solid solution, which was compatible with the geometry required for Eu(III) incorporation into the calcite lattice (Lakshtanov and Stipp, 2004). This is in contrast to previous work which proposed the formation of a EuOHCO 3 -CaCO 3 solid solution (Curti et al., 2005; Stipp et al., 2003). Subsequently, Eu(III) coprecipitation with calcite has been studied using luminescence spectroscopy (Marques Fernandes et al., 2008). The authors identified three separate Eu(III) sites, two of which were identified as Eu(III) substituted for Ca 2+ in the calcite lattice. The third site the authors considered a Eu(III)-calcite surface complex (Marques Fernandes et al., 2008). Interestingly, the authors compared experiments carried out in Na + and K + systems, where the latter showed strong distortion of the calcite lattice. Therefore, the authors proposed a coupled substitution of Na + and Eu 3+ for two Ca 2+ ions. Other studies have focused on the sorption and uptake behaviour of Eu(III) with calcite, rather than coprecipitation. Zavarin et al., (2005) demonstrated that Eu(III) sorption to calcite varies only slightly over a ph range of 6.5 to 10 with atmospheric CO 2 equilibration, though sorption did appear to decrease at the lowest and highest ph values. The authors suggested that this effect was caused by the stability of the Eu 3+ (aq) ion at low ph, and Eu-carbonate complexation at the high ph (Zavarin et al., 2005). Using a surface complexation approach, Zavarin et al. (2005) successfully modelled their data using two Eu(III)-calcite surface complexes, 2- >CaCO 3 EuCO 3 (log K 3.5) and >CaCO 3 Eu(CO 3 ) 2 (log K -2.4; Zavarin et al., 2005; Figure 46

52 1.17). However, it should be noted that this does not imply a specific coordination and instead only reflects the stoichiometry which best matched the experimental data. Figure Eu(III) sorption to calcite, in equilibrium with atmospheric CO 2. The hollow symbols represent sorption after 24 hours while the filled symbols represent after 30 days. The dashed lines represent the sorbed Eu associated with a particular surface complex whereas, the solid line represents total modelled sorption. Taken from (Zavarin et al., 2005). Another key component of hydrated NRVB are Calcium-Silica-Hydrate (CSH) gels, which occur with a range of Ca/Si ratios. Pointeau et al., (2001) used luminescence spectroscopy to study Eu interactions with CSH gel at high ph (> ph 10). Sorption was very strong (Rd > 1.8 x 10 5 L kg -1 ) across all calcium / silica ratios (Ca/Si) in the CSH ( ) studied. In the higher Ca/Si systems ( ), two distinct types of Eu were found, one of which was assigned as a surface complex and the other as Eu incorporated within the CSH structure. The spectra were similar to that of Eu sorbed to tobermorite, which possesses a structure related to CSH gels. Schlegel et al., (2004) also studied Eu coprecipitation and sorption with CSH as a function of Ca/Si ratio using EXAFS spectroscopy. The authors modelled their EXAFS data of both coprecipitated and sorbed systems with three shells, Si, Si/Ca, and a 47

53 final Ca shell with interatomic distances of 3.2, and Å, respectively. This was consistent with models of the CSH structure and suggested that Eu occupies the Ca sites in the CSH structure in both sorbed and coprecipitated systems. For all Ca/SI ratios, Eu(III) removal was rapid with 90% removal within 15 minutes (Schlegel et al., 2004). For the lower Ca/Si systems ( ), the process seems to be initially surface complexation, followed by diffusion into the structure. Batch data suggested that at a high Ca/Si ratio (1.30), Eu is at least partly removed from solution by co-precipitation. Despite the different mechanisms, the EXAFS data suggest that the final coordination was the same (Schlegel et al., 2004) Aims and Objectives The UK has had a nuclear programme for many years for weapons development through to current power generation activities. Furthermore, in very recent years, the Government has signalled the start of a new power reactor construction programme to maintain energy security and meet climate change obligations (Department for Environment Food and Rural Affairs, 2008). The Government has decided that continued surface storage of the by-products from the nuclear fuel cycle cannot continue indefinitely, due to increased proliferation and terrorist risk. Therefore, it is UK Government policy to construct a deep subsurface geological disposal facility that can contain these wastes over geological time, until their activity has reduced to near background levels. Before any GDF can begin construction and operation, a well defined safety case covering all aspects of geological disposal is required from a regulatory and public perception standpoint. The environmental behaviour of radionuclides is relatively well defined under ambient circumneutral conditions. However, the hydration of a cementitious GDF will generate a region of hyperalkalinity that will surround and extend away the facility, known as the hyperalkaline plume. However, in hyperalkaline systems the environmental behaviour of radionuclides is expected to deviate significantly from that 48

54 observed in circumnetural environments. So far, there has been little work in the academic literature to characterise this behaviour. This represents a significant gap in any safety case for a cementitious GDF and needs to be addressed through research programmes before a GDF programme can progress through to operation. In many areas, research has started to characterise radionuclide behaviour. However, there are notable exceptions, such as the behaviour of actinides with calcium carbonate (calcite) solids and specialty high ph grouts such as NRVB. Both of these materials are expected to be present in the GDF environment. It is impossible to create an informed safety case for a cementitious GDF without characterising radionuclide behaviour with these materials. Therefore, it is critically important that these behaviours are characterised sufficiently. It is the aim of this thesis to attempt to characterise radionuclide behaviour in systems of direct relevance to cementitious geological disposal (i.e. high ph; ph > 10). Specifically, this thesis will study UO 2 2+ behaviour (as a model 6+ oxidation state actinide) in contact with calcite solids in solutions designed to be representative of a cementitious GDF (synthetic hyperalkaline cement leachates; Chapter 3). NpO 2 + (as a model 5+ oxidation state actinide) behaviour in contact with calcite solids in solutions designed to be representative of a cementitious GDF (synthetic hyperalkaline cement leachates; Chapter 4). Eu 3+ interaction with a high ph, calcite containing, cement, NRVB. Eu 3+ is studied as an analogue for the actinides in the 3+ oxidation state, such as Am(III), Cm(III), and Pu(III) (Chapter 5). The study of these three systems will be used to provide information on the environmental mobility of actinides in a range of oxidation states (3+, 5+, 6+) expected to be present in a cementitious GDF environment. An obvious absence in these oxidation states is the 4+ state, which will not be studied, as actinides in oxidation state 4+ are very prone to hydrolysis, 49

55 which results in extremely low solubilities in the high ph range expected, in and around a cementitious GDF. A mechanistic understanding of radionuclide behaviour in these systems will be achieved and characterised using a range of geochemical and analytical techniques, such as: inductively coupled plasma mass spectrometry; x-ray absorption spectroscopy / EXAFS; small angle x-ray scattering; liquid scintillation counting; and, mathematical modelling techniques, etc References Bailey, E.H., Mosselmans, J.F.W., Young, S.D., Time-dependent surface reactivity of Cd sorbed on calcite, hydroxylapatite and humic acid. Mineralogical Magazine 69, Baik, M.H., Hyun, S.P., Hahn, P.S., Surface and bulk sorption of uranium(vi) onto granite rock. Journal of Radioanalytical and Nuclear Chemistry 256, Baldwin, T., Chapman, N., Neall, F., Geological disposal options for High-Level Waste and spent fuel. NDA. Barnett, M.O., Jardine, P.M., Brooks, S.C., Selim, H.M., Adsorption and transport of uranium(vi) in subsurface media. Soil Science Society of America 64, Braney, M.C., Haworth, A., Jefferies, N.L., Smith, A.C., A study of the effects of an alkaline plume from a cementitious repository on geological-materials. Journal of Contaminant Hydrology 13, Chambers, A., Gould, L., Harris, A., Pilkington, N., Williams, S., Evolution of the near field of nirex disposal concept. Nirex. Choppin, G.R., Actinide speciation in aquatic systems. Marine Chemistry 99,

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58 Gaona, X., Wieland, E., Tits, J., Scheinost, A.C., Dähn, R., Np(V/VI) redox chemistry in cementitious systems: XAFS investigations on the speciation under anoxic and oxidizing conditions. Applied Geochemistry 28, Geipel, G., Reich, T., Brendler, V., Bernhard, G., Nitsche, H., Laser and X-ray spectroscopic studies of uranium-calcite interface phenomena. Journal of Nuclear Materials 248, Guillaumont, R., Fanghanel, T., Neck, V., Fuger, J., Palmer, D., Grenthe, I., Rand, M., Update on the chemical thermodynamics of uranium, neptunium, plutonium, americium and technetium. Elsevier, Amsterdam, 960 p. Hay, M.B., Workman, R.K., Manne, S., Mechanisms of metal ion sorption on calcite: Composition mapping by lateral force microscopy. Langmuir 19, Heberling, F., Brendebach, B., Bosbach, D., 2008a. Neptunium(V) adsorption to calcite. Journal of Contaminant Hydrology 102, Heberling, F., Brendebach, B., Bosbach, D., 2008b. Neptunium(V) adsorption to calcite. Journal of Contaminant Hydrology 102, Heberling, F., Denecke, M.A., Bosbach, D., 2008c. Neptunium(V) coprecipitation with calcite. Environmental Science & Technology 42, Ishikawa, M., Ichikuni, M., Uptake of sodium and potassium by calcite. Chemical Geology 42, Kaplan, D., Gervais, T., Krupka, K., Uranium(VI) sorption to sediments under high ph and ionic strength conditions. Radiochimica Acta 80,

59 Karnland, O., Olsson, S., Nilsson, U., Sellin, P., Experimentally determined swelling pressures and geochemical interactions of compacted Wyoming bentonite with highly alkaline solutions. Physics and Chemistry of the Earth 32, Kaszuba, J.P., Runde, W.H., The aqueous geochemistry of neptunium: dynamic control of soluble concentrations with applications to nuclear wasted. Environmental Science & Technology 33, Kelly, S.D., Newville, M.G., Cheng, L., Kemner, K.M., Sutton, S.R., Fenter, P., Sturchio, N.C., Spötl, C., Uranyl incorporation in natural calcite. Environmental Science & Technology 37, Lakshtanov, L.Z., Stipp, S.L.S., Experimental study of europium(iii) coprecipitation with calcite. Geochimica et Cosmochimica Acta 68, Lee, S.-G., Lee, K., Cho, S., Yoon, Y., Kim, Y., Sorption properties of 152 Eu and 241 Am in geological materials: Eu as an analogue for monitoring the Am behaviour in heterogeneous geological environments. Geosciences Journal 10, Madsen, F.T., Clay mineralogical investigations related to nuclear waste disposal. Clay Miner 33, Marques Fernandes, M., Schmidt, M., Stumpf, T., Walther, C., Bosbach, D., Klenze, R., Fanghänel, T., Site-selective time-resolved laser fluorescence spectroscopy of Eu 3+ in calcite. Journal of Colloid and Interface Science 321, McCarter, W.J., Crossland, I., Chrisp, T.M., Hydration and drying of Nirex reference vault backfill. Building and Environment 39,

60 Meleshyn, A., Azeroual, M., Reeck, T., Houben, G., Riebe, B., Bunnenberg, C., Influence of (calcium-)uranyl-carbonate complexation on U(VI) sorption on Ca- and Nabentomtes. Environmental Science & Technology 43, Morris, K., Law, G., Bryan, N., Geodisposal of higher activity wastes. Issues in Environmental Science and Technology. 32 ed. Royal Society of Chemistry; p NDA, DEFRA, UK Radioactive Waste Inventory 2007 Main Report. NDA/RWMD/004. NDA, Geological Disposal: Near-field evolution status report. NDA/RWMD/033. NDA, Geological Disposal: Generic post-closure safety assesment. NDA/RWMD/030. Ojovan, M.I., Lee, W.E., 2005, An Introduction to Nuclear Waste Immobilisation. Elsevier, Oxford. Parkman, R.H., Charnock, J.M., Livens, F.R., Vaughan, D.J., A study of the interaction of strontium ions in aqueous solution with the surfaces of calcite and kaolinite. Geochimica et Cosmochimica Acta 62, Pedersen, K., Subterranean microorganisms and radioactive waste disposal in Sweden. Engineering Geology 52, Perdrial, J.N., Warr, L.N., Perdrial, N., Lett, M.C., Elsass, F., Interaction between smectite and bacteria: Implications for bentonite as backfill material in the disposal of nuclear waste. Chemical Geology 264,

61 Pointeau, I., Piriou, B., Fedoroff, M., Barthes, M.G., Marmier, N., Fromage, F., Sorption mechanisms of Eu 3+ on CSH phases of hydrated cements. Journal of Colloid Interface Science 236, Pointeau, I., Coreau, N., Reiller, P.E., Uptake of anionic radionuclides onto degraded cement pastes and competing effect of organic ligands. Radiochim Acta 96, Reeder, R., Elzinga, E., Tait, C.D., Rector, K.D., Donohoe, R., Morris, D., 2004a. Sitespecific incorporation of uranyl carbonate species at the calcite surface. Geochimica et Cosmochimica Acta 68, Reeder, R.J., Crystal chemistry of the rhombohedral carbonates. Reviews in Mineralogy and Geochemistry 11, Reeder, R.J., Elzinga, E.J., Tait, C.D., Rector, K.D., Donohoe, R.J., Morris, D.E., 2004b. Site-specific incorporation of uranyl carbonate species at the calcite surface. Geochimica et Cosmochimica Acta 68, Reeder, R.J., Nugent, M., Lamble, G.M., Tait, C.D., Morris, D.E., Uranyl incorporation into calcite and aragonite: XAFS and luminescence studies. Environmental Science & Technology 34, Reeder, R.J., Nugent, M., Tait, C.D., Morris, D.E., Heald, S.M., Beck, K.M., Hess, W.P., Lanzirotti, A., Coprecipitation of uranium(vi) with calcite: XAFS, micro-xas, and luminescence characterization. Geochimica et Cosmochimica Acta 65, Rihs, S., Sturchio, N.C., Orlandini, K., Cheng, L.W., Teng, H., Fenter, P., Bedzyk, M.J., Interaction of uranyl with calite in the presence of EDTA. Environmental Science & Technology 38,

62 Runde, W., The chemical interactions of actinides in the environment. Los Alamos Science 26, Runde, W., Neu, M.P., Clark, D.L., Neptunium(V) hydrolysis and carbonate complexation: Experimental and predicted neptunyl solubility in concentrated NaCl using the Pitzer approach. Geochimica et Cosmochimica Acta 60, Sawada, K., Abdel-Aal, K., Tan, K., Satoh, K., Adsorption of lanthanoid ions on calcite. Dalton Transactions, Schlegel, M., Pointeau, I., Coreau, N. and Reiller, P Mechanism of europium retention by calcium silicate hydrates: an EXAFS study. Environmental Science & Technology, 38, Schmeide, K., Bernhard, G., Sorption of Np(V) and Np(IV) onto kaolinite: Effects of ph, ionic strength, carbonate and humic acid. Applied Geochemistry 25, Shannon R., Revised effective ionic radii in halides and chalcogenides. Acta Crystallographica 32, Smellie, J.A.T., Karlsson, F., The use of natural analogues to assess radionuclide transport. Engineering Geology 52, Somasundaran, P., Agar, G.E., The zero point of charge of calcite. Journal of Colloid and Interface Science 24, Stipp, S.L.S., Toward a conceptual model of the calcite surface: Hydration, hydrolysis, and surface potential. Geochimica et Cosmochimica Acta 63,

63 Stipp, S.L.S., Lakshtanov, L.Z., Jensen, J.T., Baker, J.A., Eu 3+ uptake by calcite: Preliminary results from coprecipitation experiments and observations with surface-sensitive techniques. Journal of Contaminant Hydrology 61, Sturchio, N.C., Antonio, M.R., Saderholm, L., Sutton, S.R., Brannon, J.C., Tetravalent uranium in calcite. Science 281, Swanton, S., Alexander, W., Berry, J., Review of the behaviour of colloids in the near field of a cementitious repository. NDA RWMD. Tertre, E., Berger, G., Simoni, E., Castet, S., Giffaut, E., Loubet, M., Catalette, H., Europium retention onto clay minerals from 25 to 150 C: Experimental measurements, spectroscopic features and sorption modelling. Geochimica et Cosmochimica Acta 70, Tits, J., Wieland, E., Bradbury, M.H., The effect of isosaccharinic acid and gluconic acid on the retention of Eu(III), Am(III) and Th(IV) by calcite. Applied Geochemistry. 20, Tits, J., Geipel, G., Mace, N., Eilzer, M., Wieland, E., Determination of uranium(vi) sorbed species in calcium silicate hydrate phases: A laser-induced luminescence spectroscopy and batch sorption study. Journal of Colloid and Interface Science 359, Thompson, D., Pownall, P., Surface electrical properties of calcite. Journal of Colloid and Interface Science 131, Turner, D., Pabalan, R., Bertetti, F., Neptunium(V) sorption on montmorillonite: An experimental and surface complexation modelling study. Clays and Clay Minerals 46,

64 Viallis-Terrisse, H., Nonat, A., Petit, J.C., Landesman, C., Richet, C., Specific interaction of cesium with the surface of calcium silicate hydrates. Radiochimica Acta 90, Wang, Z.M., Zachara, J.M., McKinley, J.P., Smith, S.C., Cryogenic laser induced U(VI) fluorescence studies of a U(VI) substituted natural calcite: Implications to U(VI) speciation in contaminated Hanford sediments. Environmental Science & Technology 39, Wieland, E., Spieler, P., Colloids in the mortar backfill of a cementitious repository for radioactive waste. Waste Management 21, Wieland, E., Tits, J., Bradbury, M.H., The potential effect of cementitious colloids on radionuclide mobilisation in a repository for radioactive waste. Applied Geochemistry 19, Wilson., P.D., The nuclear fuel cycle: From ores to waste. Oxford Science Publications, Oxford. Zachara, J.M., Cowan, C.E., Resch, C.T., Sorption of divalent metals on calcite. Geochimica et Cosmochimica Acta 55, Zavarin, M., Roberts, S., Hakem, N., Sawvel, A., Kersting, A., Eu(III), Sm(III), Np(V), Pu(IV) sorption to calcite. Radiochimica Acta 93,

65 2. Instrumentation and methodology This chapter will outline the theory of the various experimental and analytical techniques used throughout the preparation of this thesis and provides a coherent source of reference to all of the approaches used in the original research work described in Chapters 3, 4 and Experimental Approaches Batch Experiments Batch techniques were used throughout the experimental work in this thesis and a generic description of the experimental approaches are described here. Typically, batch experiments were carried out in 50 ml polypropylene centrifuge tubes. In batch sorption experiments, the required solid (i.e. 5 g, 1 g, and 0.1 g for 1 : 10, 1 : 50, and 1 : 500 systems, respectively) was weighed out (± 2 %) and added to the centrifuge tubes. Then the required solution (typically 50 ml) was added, and the solutions were allowed to equilibrate with the solid for two days. A small volume of radioisotope stock which had no measureable effect on solution ph was then added to each experiment to give the desired final concentration. All radioisotopes were dissolved in a dilute acid (0.01 M HNO 3 ). Where appropriate, control samples were included that were made concurrently with the experiment. These control systems had identical solution chemistry to the other samples except no solid phase was included. Additionally, all experiments included acidified standards consisting of a known activity of radionuclide in: 2 M HCl for Liquid Scintillation Counting (LSC; Section 2.4.1) and gamma spectroscopy (Section 2.4.2); or 2% HNO 3 where the method of analysis was Inductively Coupled plasma - Mass Spectrometry (ICP-MS; Section 2.3.1). In most cases the experimental samples were made in triplicate, though in some cases 5 replicas were created. 60

66 To determine radionuclide concentration, typically a 2 ml aliquot from the experiment was taken and added to a 2 ml centrifuge tube. The aliquot was then centrifuged (in a Sigma 1-14 Microfuge) for five minutes at 3500 rpm to achieve a particle size cut off approximately equivalent to a 0.45 µm filter (Section 2.1.2). A second aliquot was then removed from the centrifuge tube for analysis. For 232 U and (high level; > 1 Bq ml -1 ) 237 Np analysis, the aliquot was added to a 30 ml LSC vial with 5 ml scintillation cocktail (Scintisafe 3, Fisher Scientific) and 2 ml 2 M HCl to acidify and prevent precipitation and sorption. The samples were then counted on a low background LSC (Quantulus) with alpha beta discrimination using a pulse shape analyser value of 75. The samples were counted for sufficient time that the counting error was < 5 %, where counting errors were calculated by equation (2.01), σ = Counts total (2.01) where σ is the standard deviation. 238 U and 237 Np concentrations were determined using ICP- MS (Agilent 7500cx; in spectrum analysis mode with a gas flow rate of 0.9 L min -1 ). The aliquots were added to a 2% HNO 3 solution. The sample was then diluted (typically in a 1:50 ratio) with 2% HNO 3 to ensure that the total dissolved solids were < 0.1%. The dilution was also calculated to place the expected radionuclide concentration within the range of the calibration standards which were at 1, 5, 10, 50, and 100 µg ml -1 for 238 U and 1, 5, 10, 50 and 100 ng ml -1 for 237 Np. Additionally, for 237 Np ICP-MS analysis, a 10 ppb 232 Th internal standard was included to correct for instrumental drift. 152 Eu concentrations were determined using gamma ray spectrometry ( kev 152 Eu line). The 152 Eu containing aliquot was added to a (reproducible) known geometry plastic container (30 ml polypropylene container) and placed on the (Ge semiconductor) γ detector in a reproducible position for 20 minutes. The peak at kev was then integrated and converted to decay events. The data was then converted to an absolute 152 Eu concentration by use of an acidified standard in the same geometry and position relative to determine detector efficiency. 61

67 Centrifugation Calculations When removing aliquots of solution for elemental analysis from uptake experiments, it was important to remove all solid phases and any suspended particulate that was not kept suspended due to Brownian motion (> 0.45 µm). It was important to have a consistent technique for the removal of the solid phases. Stokes' law describes the settling motion of particles in a liquid, as shown in equation (2.02), 2 d g(sp S 1) V = 18η (2.02) where: V is the terminal particle velocity; S 1 is the density of the medium; g is the acceleration due to gravity; S p is the particle density; d is the spherical diameter of the particle; and η is the viscosity of the liquid. With a modified and integrated form of Stokes' law (Weiner et al., 1995), it was possible to calculate the centrifugation time (t) required to achieve a specific particle size cut off, as described by equation (2.03), 18ηln( R s S ) t= ω 2 d 2 ρ (2.03) where: R s is the distance from the centre of rotation to the midpoint of the sample; S is the distance between the centre of rotation and the top of the solution; ɷ is the angular velocity; d is the spherical diameter of the particle; and ρ is the difference in density between the solid and the liquid. It was determined that the centrifuge (Sigma 1-14) used in this study could achieve an effective particle size cut off of 0.45 µm by spinning at 3500 rpm for 5 minutes. 62

68 Ultrafiltration Experiments While centrifugation can be used to select a specific particle size cut off, it can be difficult to isolate very small particles (< 100 nm) in a reliable way. Where a small particle size cut off is desirable, ultrafiltration techniques can be used (Dupré et al., 1999; Nakao, 1994, Pitois et al., 2008). Ultrafiltration is regularly used in the biological sciences to characterise large species such as proteins. In this thesis, ultrafiltration was used to characterise colloid populations in Ch 5. Ultrafiltration is a simple technique where hydrostatic pressure is used to force a solution through a semi-permeable membrane with pore holes of desired size, separating large molecules and particles with a high molecular weight from the solution (Figure 2.01). The standard methodology used 3, 10 and 100 kda filters to fractionate colloid populations, where the filters were approximately equivalent to 1.5, 2.5 and 10 nm, respectively (assuming spherical particles). PolyEtherSulfone (PES) ultrafiltration membranes (Millipore) were routinely used in Ch. 5 in lieu of the more common regenerated cellulose membranes which, in contrast to the PES filters, show low stability in high ph environments. The ultrafiltration membranes were soaked in high ph (inactive) Eu solutions in to condition the filters and saturate Eu binding sites (to prevent sorption skewing results) prior to use. 63

69 Figure Schematic of an ultrafiltration cell Experimental Errors All batch experiments in Chapters 3 and 4 were carried out in triplicate (with one exception which was used 5 replicas) in order to assess experimental error. Chapter 5 followed the approach of Pointeau et al. (2004) and obtained uncertainties from repeating selected, representative systems. The results from the replicate experiments were then combined into a mean value and the error expressed as one standard deviation (σ; equation 2.04) from the mean, 2 Σ( x x) σ = (2.04) n 1 where: x is the experimental value; x is the mean; and n is the number of repeats. The standard deviation can be used to determine the probability that the true value of the measured quantity lies within a given range. Assuming that the experimental data were normally distributed, there is a 68.3 % chance that the true values lie within 1σ of the 64

70 calculated mean, whereas the probability is 95.4 % when the range is expanded to 2σ (Figure 2.02). Figure The normal distribution Synthetic Cement (BIGRAD) Leachates In some of the experimental work described in Ch. 3 and 4, synthetic cement leachates were used which were widely adopted as leachate recipes in the BIGRAD Consortium. These leachates were designed to be representative of a cementitious geological disposal facility in the early and mature stages. The recipes are listed in Table 2.1. The Young Cement Leachate (YCL) was representative of a young cementitious GDF (< 2000 years), where high ph conditions are maintained by dissolution of sodium and potassium hydroxides. The YCL solution was prepared by shaking until the apparent dissolution of all the solids and then maintaining the solution temperature at 40 C for two days before being passing through a 0.45 µm filter (Whatman inline filter; Nylon). This was to ensure undersaturation with respect to calcium hydroxide. The final solution ph of the YCL system was 13.3 ± 0.1. The other leachate solution was Old Cement Leachate (OCL), which was a ph 10.5 ± 0.1 calcium hydroxide solution. No additional preparation of the OCL solution was required. 65

71 Table 2.1. Synthetic cement leachate solution composition OCL YCL ph KOH (g L -1 ) NaOH (g L -1 ) Ca(OH) 2 (g L -1 ) Carbonate Exclusion Dissolved carbonate is an important ligand in natural systems that can strongly influence the environmental behaviour of radionuclides. The capacity for carbonate dissolution in water - increases with ph. This increase is caused by a shift in carbonate speciation from the HCO 3 species to the CO 2-3 species. Additionally, the kinetics of carbonate dissolution are slow, and so it takes a long time for high ph systems to reach equilibrium with atmospheric CO 2(g). Therefore, carbonate ingress has the potential to affect experimental results significantly and for these experiments focussed on a deep GDF environment, where contact with atmospheric CO 2 is not expected, it needs to be prevented. To create an oxic, CO 2 free atmosphere (< 1 ppm) a Coy basic polymer glovebox was used. The glovebox was purged continuously using a pump to bubble air through three 500 ml Duran gas scrubbing bottles (Figure 2.03). Each bottle contained 250 ml 17 M NaOH to scrub out CO 2(g). The air flow was also passed through an additional gas scrubbing bottle containing 250 ml water in order to remove any residual NaOH aerosol. The sodium hydroxide in this system requires refreshing once precipitation of a white sodium carbonate solid in the final dreschel occurs. 66

72 Figure Schematic of CO 2 scrubbing system U Radiotracer Preparation 232 U is a radioisotope with a half life of 68.8 years (8.87 x Bq g -1 ) and decays to 228 Th (half life 1.91 years; 3.03x10 16 Bq g -1 ). The high specific activity of 232 U makes it useful as a low level uranium tracer. Unfortunately, both 232 U and 228 Th decay via the emission of an alpha particle and the decay energies of 232 U and 228 Th are similar (5.52 MeV and 5.31 MeV, respectively). Therefore, it is impossible to analyse for 232 U with energy resolved scintillation counting without removing any ingrown 228 Th. The separation of 232 U from its 282 Th daughter was carried out with a method modified from the literature (Ong and Leckie, 1996). A 0.7 x 10 cm column packed with 100 mg AG1-8X anion exchange resin (chloride form, mesh size) was conditioned with 12 M HCl (15 ml). A stock solution was prepared by adding 20 kbq 232 U to 12 M HCl (15 ml) which was then passed through the column. Three aliquots of 12 M HCl (10 ml) were passed through the column (fractions I-III). The 232 U was eluted with 0.1 M HCl (10 ml, fraction IV). The removal of the daughter isotopes was confirmed using ultra low background LSC (Figure 2.04). 67

73 2.2. Geochemical Modelling PHREEQC is a geochemical modelling software package provided by the U.S Geological Survey that can be used to perform a wide variety of aqueous geochemical calculations. Throughout this thesis, PHREEQC was used to calculate solution speciation and saturation with respect to solid phases for a range of systems, however, it can also be used for a variety of other purposes, such as one dimensional transport and surface complexation modelling. The ANDRA SIT database (ThermoChimie v.7.d June 2011) was used for all geochemical calculations, with additional thermodynamic data added where required. An example script that calculates U(VI) speciation and saturation in a ph 10.5 calcite equilibrated system is shown in Figure PHREEQC uses a database of known stability constants and solubility products for various environmentally relevant reactions in order to determine speciation and saturation at equilibrium. 68

74 Figure LSC spectra of the five fractions (effluent) collected from the separation of 232 U and 228 Th recorded using a Quantulus ultra low background LSC using alpha beta discrimination to isolate the alpha activity. Fraction I, I and III contain 228 Th in concentrated HCl while and IV and V are 232 U containing fractions in 0.1 M HCl. SOLUTION 1 temp 25 ph 10.5 Ca mol/kgw U(6) 10 mg/kgw EQUILIBRIUM_PHASES 1 Calcite 0 10 Figure Example PHREEQC input file to determine the speciation and saturation U(VI) (10 mg / kg water, ph 10.5, 25 C, in equilibrium with 10 moles of calcite solid). 69

75 2.3. Elemental Analysis Inductively Coupled Plasma Mass Spectrometry Inductively Coupled Plasma Mass Spectrometry (ICP-MS; Tanner and Holland, 2001; Linge and Jarvis, 2009; Figure 2.06) is an analytical technique that can be used to determine the elemental composition of a sample down to very low concentrations (typically in the parts per trillion range). In the preparation of this thesis, ICP-MS was used extensively to determine low level 238 U and 237 Np concentration in dilute HNO 3 solutions using a Agilent 7500cx instrument. Details on the calibration standards are described above in Section Figure Schematic of an ICP-MS. Initially, the solution is added to the instrument using a peristaltic pump. It is then converted into a spray by pumping the solution through a nozzle at high pressure (nebuliser). The spray from the nebuliser then passes through a spray chamber, which allows the larger droplets to drain away, leaving only a fine mist. The mist then undergoes evaporation, atomisation, and ionisation in an Inductively Coupled Plasma (ICP) torch (Figure 2.07). 70

76 An ICP torch generates plasma from Ar by using a cooled copper coil to generate an oscillating magnetic field that accelerates any electrons. An accelerated electron ionises an Ar atom to generate two more electrons which are accelerated and ionise more Ar in a cascade reaction. A plasma of Ar + ions and electrons is then generated at a temperature of approximately 10,000 K. The ICP torch consists of three nested quartz tubes. Argon gas is introduced through the middle tube in order to control the positioning of the plasma. Argon gas is also introduced via the outer most tube to provide Ar for the plasma and to act as a coolant to prevent the torch from melting. The inner most tube introduces the sample into the plasma, where it is ionised. Ions from the plasma are passed through several cones, where the pressure is reduced from atmospheric (required for the plasma) to the high vacuum required for the detector. The ions then pass through a quadrupole mass spectrometer. The quadrupole is tuned to deflect all ions, except for those of a specific m/z ratio, away from the detector. This allows the ICP-MS to be sensitive, even at very low concentrations. There are, however, a series of caveats with ICP-MS, such as: mass interferences between polyatomic species (e.g. ArN + ); and a lack of sensitivity around atomic masses of 40, due to the Ar + ions from the plasma. A set of standards of known concentration (for the analyte in question) are required in order to calibrate the instrument. A least squares linear regression of the standard concentration (x) against counts (y) can then be used to used to convert the counts from a sample to an actual concentration. In addition, instrumental drift can occur across a run of samples due to a crusting effect on the cones (Figure 2.06). In cases where this is problematic, an internal standard of constant concentration can be added to the samples to correct for this effect. 71

77 Figure Schematic of an ICP torch. 72

78 2.4. Radiometric Techniques Liquid Scintillation Counting Liquid Scintillation Counting (LSC; Passo and Cook, 1994; Horrocks, 1974) is a method of measuring the radioactivity of a sample to a high degree of accuracy and with some instruments down to low levels (< 0.1 Bq) in comparison to the natural background. Additionally, it is possible to identify isotopes and determine the relative proportions of decays from different isotopes and even discriminate between alpha and beta events. LSC techniques were used throughout this thesis to confirm successful purification of 232 U stocks and determine 232 U and 237 Np concentrations. LSC requires the dissolution of a radioactive sample in a solvent containing a scintillant. Any decay events in the dissolved sample result in an energy transfer to the scintillant and an excited state. Photons are then emitted in the deexcitation process. These photons are then detected by a pair of photomultiplier tubes located on either side of the sample vial (Figure 2.08). The signals are passed through a pre-amplifier, an amplifier and finally to a coincidence circuit. The coincidence circuit only allows the event to be registered when the pulses occur in both photomultiplier tubes simultaneously (the pulses are approximately 15 ns). This system reduces the background noise and allows a greater instrumental sensitivity. Each pulse from the coincidence circuit is taken as a genuine decay event. This allows a liquid scintillation counter to count the absolute number of decay events taking place within a sample with a high efficiency ( 100% for some isotopes). When the number of decay events is plotted as a function of photon energy, it becomes possible to detect peaks belonging to different radioisotopes. There are a variety of different scintillation fluids that are suitable for LSC. Therefore, the position that any given isotope will occupy in an uncalibrated spectrum will differ. However, the relative positions can still provide information on the isotopic composition of a sample and provide absolute measures of radioactivity. Some instruments, such as the Quantulus ultra low background LSC used 73

79 throughout this thesis, are able to distinguish between alpha and beta events through use of a Pulse Shape Analyser (PSA) circuit. The pulse from the photomultiplier tubes differ in character, depending on if the decay is an alpha or beta event. The PSA circuit integrates the tail of the scintillation pulse and compares it with the total scintillation pulse and derives an amplitude independent value to describe the character of the pulse (Figure 2.09). This extra information can be used to separate overlapping peaks produced by alpha and beta decays and provide a more accurate measure of radioactivity. Figure Schematic of a LSC instrument (Passo and Cook, 1994; Horrocks, 1974). 74

80 Figure Example pulse height vs. pulse length for alpha and beta emitters. The PSA value defines the gradient that separates the events (Passo and Cook, 1994) Gamma Spectroscopy Isotopes with unstable nuclei undergo radioactive decay. Common methods of decay are alpha emission, where an alpha particle (two protons, two neutrons) is emitted from the nucleus and beta decay which encompasses positron/electron emission/capture. Often these decays leave the nucleus in an excited state and the subsequent de-excitation results in emission of a high energy photon, known as a gamma ray. The de-excitation event takes place between well defined energy levels and, therefore, the energy of the gamma rays is characteristic of the isotope. Thus, it is possible to identify and quantify radionuclides using gamma spectroscopy (Gilmore, 2008). Gamma spectroscopy was used in this thesis to determine 152 Eu concentrations (Pitois et al., 2008) in batch uptake experiments (Ch. 5). Typically, to record a gamma spectrum a sample is placed on top of a semi-conducting detector, using a lead enclosure to minimise background radiation. Once the gamma photon is emitted from the sample it interacts with and dissipates in the intrinsic region of the semiconducting detector. This promotes a number of electrons into the conduction band of the 75

81 detector and leaves 'holes' in the valence bands, where the number of promotions is proportional to the energy of the photon. Appling a voltage to the detector causes the electrons to migrate to the cathode while the holes move to the anode, creating a current (Gilmore, 2008). This current can be amplified and the signal passed to ancillary counting equipment. The equipment amplifies and shapes the pulses to boost the signal and shorten the duration of the pulse. The analogue signal is then passed through a discriminator circuit to remove the small pulses caused by noise in the equipment and digitalised. The energy of the photon is determined by the number of promotions in the semi detector, and thus, the size of the signal from the detector is proportional to the energy of the gamma photon. Radioisotopes can be identified in a gamma spectrum from one or more characteristic gamma lines. The quantity of these isotopes can be determined by careful use of standards of known concentration. Gamma detectors typically only have low efficiencies due to geometric considerations and this means that standards have to be carefully matrix and volume matched to experimental samples in order to determine correctly detector efficiencies to allow conversion from detector counts to decay events (Gilmore, 2008) Spectroscopy Ultraviolet - Visible Spectroscopy UltraViolet Visible (UV-Vis; Perkampus, 1992) spectroscopy is a common laboratory technique that concerns the absorbance of light in the ultraviolet and visible regions of the electromagnetic spectrum by a sample. When a sample is exposed to light, it is possible that absorption will occur and result in a transition to an excited state. A molecule will only absorb a photon when the energy is of the correct energy to facilitate the promotion of an electron. Thus, molecules absorb light at characteristic wavelengths. The degree of absorption 76

82 at a given wavelength is proportional to the number of absorbing molecules. This relationship is described by the Beer-Lambert Law (2.04), A = εbc (2.04) where: A is absorbance (dimensionless); ε is the molar absorptivity (L mol -1 cm -1 ); b is the path length of the sample (cm) and; c is the concentration of the compound (mol L -1 ). By measuring absorbance over a range of wavelengths, it is possible to identify molecules and measuring absorbance at a single wavelength can be used to determine concentration relative to a set of calibration standards of known concentration. In this thesis the technique was used for determination of ligand (EDTA) concentrations in Ch Luminescence Spectroscopy Luminescence spectroscopy (Lakowicz, 2006) is a technique that involves using a light source to excite the electrons within a molecule. Luminescence spectroscopy is concerned with the photons emitted when the molecule de-excites. Luminescence spectroscopy can be used to determine the vibrational structure and speciation of a molecule. In the preparation of this thesis, luminescence spectroscopy (with excitation at 250 nm and no gating) was used to probe U(VI) speciation in batch uptake experiments in Ch. 4. Luminescence occurs when a photon of sufficient energy promotes an electron to an excited state, as illustrated in Figure The promoted species will then rapidly drop to the lowest available vibrational state via internal conversion (Figure 2.10). The excited species will deexcite to the ground state by emission of a photon. The energy of the photon is typically slightly less than the energy of the absorbed photon due to processes such as internal 77

83 conversion (Lakowicz, 2006). The average time taken for the luminescence to occur is known as the lifetime (τ). The lifetime can be determined using equation (2.05), I(t) t/ τ = I 0 e (2.05) where: I(t) is the intensity at time t; I 0 is the intensity at time 0; and τ is the lifetime. The instrumental set up to record luminescence data involves a monochromatic light source, usually generated using a laser or a xenon flashlamp coupled with a monochromator. The incident light is split, with a small proportion of the photons passed through a reference cell for calibration. The remainder of the photons are passed through the sample. A photomultiplier tube is sited next to the sample and perpendicular to the incident photons to minimise interference. Before reaching the photomultiplier tube, the emitted photons are passed through another monochromator to select for emission energy (Lakowicz, 2006). A schematic of a typical fluorometer is shown in Figure An emission spectrum is collected by fixing the excitation energy and scanning the emission monochromator over a range of wavelengths, whereas, an excitation spectrum is collected by fixing the emission energy and scanning over a range of excitation energies. 78

84 Figure Simplified Jablonski diagram illustrating the processes leading to florescence (Lakowicz, 2006). Figure Schematic of a typical fluorometer (Lakowicz, 2006). 79

85 2.6. Diffraction Techniques Bragg Diffraction X-ray diffraction (XRD; Waseda et al., 2011) refers to a range of analytical techniques that involve the elastic scattering of X-rays by a sample. This can yield important information about the internal structure of a crystalline solid. For example, XRD can be used to determine the distances between planes in the unit cell (d spacing), and therefore can be used to determine the internal crystal structure of a sample. In addition, analysis of XRD data can be also used to 'fingerprint' different crystalline phases and determine the relative approximate quantities of different phases (Waseda et al., 2011) present in a sample (when the phase is greater than approximately 5% total mass). In this thesis, diffraction techniques were used to characterise and confirm the mineralogical purity of experimental materials (Ch. 3, 4 and 5) and probe the structure of U(VI)-calcite precipitates (Ch. 4). All XRD data in the thesis was collected on a Bruker D8 Advance (Cu source) and diffraction patterns collected over a time of approximately one hour. When X-rays are passed through a crystalline solid, peaks in reflection intensity are observed as a function of incident X-ray angle, and hence, d spacing. These peaks in intensity are caused by the constructive interference of the reflected X-rays, which can be understood in terms of Bragg diffraction. When the X-ray wavelength is equal to a function of the d spacing and the incident X-ray angle (Figure 2.07), the conditions for Bragg diffraction are met and constructive interference occurs. This relationship is described by the Bragg equation (2.06), nλ = 2dsin θ (2.06) where: n is the order of reflection; λ is the X-ray wavelength; d is the interplanar spacing (d spacing); and θ is the angle of incidence. When this condition is not met, destructive interference occurs. Thus, d spacings can be calculated from incident X-ray angles that 80

86 produce peaks in reflected X-rays. The peaks are subject to sample orientation in single crystal XRD. Figure Illustration of Bragg diffraction (Waseda et al., 2011) Powder X-ray Diffraction A powdered sample contains a large number randomly orientated of microcrystalline particles. Thus, for each incident X-ray angle that matches with a d spacing in the sample structure, there will be some crystals correctly orientated to allow for Bragg diffraction. Therefore, by scanning a powder sample through a range of incident X-ray angles, all d spacings can be determined. Since the diffraction peaks are caused by the underlying structure of the crystalline phase, the peaks are characteristic and can be used to identify an unknown phase by comparison of the diffraction pattern to a database of known phases (Waseda et al., 2011). See Figure 2.13 for a schematic of a powder XRD instrument. 81

87 Figure Schematic of a powder XRD instrument (Waseda et al., 2011) Microscopy Scanning Electron Microscopy The principle of Scanning Electron Microscopy (SEM; Egerton, 2005) involves directing a beam of high energy electrons towards a solid sample. The interaction of the electron beam with the solid can yield useful information about sample topography and morphology and can be used to qualitative elemental composition. Typically, for SEM analysis, the sample will be held under a vacuum to prevent attenuation of the electron beam by a gas atmosphere. However, Environmental Scanning Electron Microscopy (ESEM) can be used to carry out SEM analysis on samples under low vacuum, high humidity environments. Secondary electrons are generated by the inelastic scattering of the incident beam with the solid sample. This produces excited states, resulting in the emission of low energy secondary electrons. The low energy of the secondary electrons means that only electrons from the surface of the sample will be detected. The mapped secondary electron signal from the SEM can be interpreted as an image of the solid sample and thus can determine morphology and topography (Egerton, 2005). In addition to secondary electron production, elastic scattering of the incident electron beam with the solid sample can occur. This results in a portion of the 82

88 incident beam being backscattered. The degree to which the backscattering occurs is a function of the electron density of the sample (Egerton, 2005). Thus, contrast in electron density can be obtained, where the greater the signal from backscattering electrons, the greater the electron density. SEM instruments are routinely coupled with equipment to carry out Energy-Dispersive X-ray (EDX) spectroscopy which can be used to determine the elemental distribution within a sample (see section 2.5.1). In addition to secondary electrons, X-rays can also be produced by the interaction of the electron beam with the solid sample. The incident electron beam interacts with the sample by inelastic scattering, producing excited states. The excited atoms can emit an X-ray of energy characteristic of a particular element and thus can be used to determine the elemental composition of a sample with mapping (Egerton, 2005). An SEM (FEI Nova 200 dual beam SEM) instrument in secondary electron mode was used to confirm the mineralogy of the calcite solids used in chapters 3 and 4. In addition, the instrument (in backscattering electron mode; equipped with a Kleindiek micromanipulator for in-situ lift-out) was used in Chapter 3 to locate U(VI) coated calcite surfaces and to prepare the thin section that was studied further using TEM Transmission Electron Microscopy Transmission Electron Microscopy (TEM; Egerton, 2005) can be used to determine crystallographic, morphological and elemental composition. The technique is similar to SEM in that it involves using a high energy electron beam to study a solid sample. The key difference is that in TEM the electrons are passed through a thin sample and analysed rather than studying the secondary electrons / backscattered electrons. In TEM, when the electron 83

89 beam is passed through a sample, some electrons are scattered as a function of density, the transmitted electrons are then recorded below the sample. These data show the structural characteristics of the sample and enable much higher resolution than is possible using SEM techniques (< 20 nm). In addition to morphological characteristics, TEM instruments are regularly coupled with Energy-Dispersive X-ray (EDX) spectroscopy equipment which can be used to determine the distribution of elements within a sample (see section 2.5.1). A TEM instrument can also be used to collect Selected Area Electron Diffraction (SAED) data from a TEM thin section. Electron diffraction involves the sample principles as X-ray diffraction (Waseda et al., 2011), except that it uses electrons rather than X-rays. SAED patterns are therefore characteristic of the phase present, and can be used to help identify crystalline phases in much the same way as XRD techniques. TEM instruments were used to study the U(VI)-calcite thin section created as part of Chapter 3. In particular, a FEI Tecnai F20 FEGTEM (with a Gatan Orius SC600 CCD camera) was used for high magnification imaging and collection of the SAED data, whereas, a Philips CM200 FEGTEM fitted with an Oxford Instruments 80 mm 2 X-Max SDD EDX detector (using Aztec software) was used for elemental mapping using EDX techniques Synchrotron Techniques Synchrotron Light Synchrotron facilities are used to produce high intensities of high energy photons for use in a variety of experiments, including X-ray absorption spectroscopy, small angle x-ray scattering and X-ray florescence mapping. A synchrotron in principle generates high energy photons through a four part process (Figure 2.14). The specific details, such as electron energies, will vary from synchrotron to synchrotron, however, The Diamond Light Source Ltd. 84

90 (Oxfordshire, UK) will be used as an example where technical specifications are mentioned. An electron gun produces lower energy electrons (90 kev). These are accelerated to higher energies in a linear particle accelerator (100 MeV). The electrons are inserted into a small booster synchrotron. This uses a shaped trajectory with two straight sections joined with semicircular curves with thirty six dipole bending magnets to accelerate the electrons further to relativistic speeds (3 GeV). The electrons are then inserted into the main storage ring which consists of twenty four straight sections formed into a loop. Surrounding the storage ring is a series of beam lines (experimental hutches) set up for specific experiments, such as X-ray absorption spectroscopy and small angle X-ray scattering. The devices used to generate photons from the electrons in the storage ring broadly fall under two categories. Bending magnets are used to steer the electrons between the straight sections of the storage ring, generating photons in the process. Bending magnets are optimised for generating photons with a wide range of energies (infrared to X-ray). Insertion devices are inserted into the straight parts of the storage ring and consist of a series of high field dipole magnets to oscillate the electrons to generate photons. The two types of insertion devices are wigglers and undulators. Wigglers provide a much higher flux than bending magnets across a similar spectrum. Undulators are similar to wigglers, except they use constructive interference to achieve higher fluxes at particular wavelengths. 85

91 Figure Schematic of a typical Synchrotron facility X-ray Absorption Spectroscopy Overview X-ray Absorption Spectroscopy (XAS) is a term encompassing X-ray Absorption Near Edge Structure (XANES) and Extended X-ray Absorption Fine Structure Spectroscopy (EXAFS) (Newville, 2004). XAS techniques were used in this thesis to determine the local coordination of various U(VI)-calcite reacted solids (Chapter 3; collected at the Diamond Light Source ltd.) and Np(V) solid phases (Chapter 4; collected at the ANKA synchrotron INE beamline). The U(VI) XAS data were collected on the L III -edge (in florescence mode) at room temperature using a 9 element solid-state Ge detector, focusing optics (beam size 0.5 mm 2 ), and a Si-111 double crystal monochromator. Whereas, the Np(V) data was obtained from the Np L III -edge in fluorescence mode (cryogenic) using a 5 pixel solid-state detector (LEGe Canberra) detector, and a Ge(422) monochromator. Both sets of data were normalised using an inline Yttrium foil reference standard. 86

92 The XAS technique involves using an X-ray ( kev) to interact with a core electron. As the X-ray energy approaches that required to excite an electron into an unoccupied level, the probability of absorption increases. When absorption occurs the electron is promoted to an unoccupied level or into the continuum (i.e. is ejected from the atom). Once the core electron has been promoted it will leave behind a core hole. A relaxation process takes place (1-2 femtoseconds later) that involves the decay of a higher energy electron into the core hole. This process will result in a release of energy in the form of a photon (fluorescence) or an Auger electron. The increase in absorption is known as the absorption edge (Figure 2.15). The edges are labelled K (1s), L (2s, 2p (j = 1/2), 2p (j = 3/2)) and M (3s, 3p, 3d), according to the electron that is being excited (Parsons et al., 2002). Importantly, the absorption edge is characteristic to the element and therefore it is possible to tune the energy of the X-rays and thus probe a specific element. However, the exact positioning of the edge can vary slightly, depending on the oxidation state of the element. The X-ray absorption coefficient (µ) is measured as a function of X-ray energy (Newville, 2004; Parsons et al., 2002). These data can yield useful information about oxidation state, coordination geometry and even probe the local structure around the target atom. 87

93 Figure Example U(VI) L III edge XAS spectrum with key regions of interest highlighted (Parsons et al., 2002) Experimental A typical XAS experiment (Figure 2.16) involves multiple ionisation chambers in order to determine the X-ray absorption coefficient. Initially, the X-rays are passed through an ionisation chamber to measure directly the incident X-ray (I 0 ) intensity. The beam then passes through the sample and through another ionisation chamber to measure the transmitted X-ray intensity. Finally, the X-ray intensity is measured again after being passed through a known reference (I R ). The absorption coefficient can be calculated from the transmission data using equation (2.07), where t is the sample thickness (Newville, 2004). The absorption coefficient is also proportional to the fluorescence intensity (I f ) using equation (2.08). Only fluorescence data was collected as part of this thesis. I 0 µt = ln (2.07) I T I f µ I (2.08) 0 88

94 Figure Schematic of a typical beam line set up for the collection of XAS data in transmission and fluorescence modes (Parsons et al., 2002) Analysis An XAS spectrum can be broadly split into two key regions: the near edge (XANES) and the extended fine structure (EXAFS). The pre-edge region can also occasionally yield useful information (Figure 2.16). The XANES part of the spectrum is found at approximately -50 ev to 200 ev from the absorption edge. The XANES can be used to interpret the valence of the absorber atom and to determine coordination environment. Additionally, the subtle features in an XANES spectrum can be used as a fingerprint of the speciation, and the relative contributions of multiple species can be calculated using linear combination fitting (Parsons et al., 2002). The EXAFS part of the spectrum extends far beyond the absorption edge and XANES regions and concerns the minor oscillations in the absorption spectrum. These oscillations are caused by the photoelectron emitted in the relaxation of the core-hole interacting with surrounding atoms (Figure 2.17) and the constructive/destructive interference associated with it. 89

95 Figure Example of scattering between an emitted photoelectron and scattering atoms (Parsons et al., 2002). The EXAFS section of the XAS spectrum is initially prepared for analysis by the subtraction of a background function which removes the major oscillations in the XAS spectrum. This reveals the minor oscillations in the EXAFS which are the fine structure. At this point by convention by the X-ray energy is converted to the wave number of the photoelectron (k) and the EXAFS data are expressed as χ(k), as a function of wave number (Figure 2.18; k-space). Using a Fourier transform on the EXAFS data can provide a plot that approximates to a radial distribution function, known as R-space (Figure 2.18). 90

96 Figure An example of Fe EXAFS data in K (bottom) and R (top) space. Blue lines represent experimental data whereas red lines represent model fit. The dashed lines represent the individual contribution to the spectrum from Fe-O and Fe-Fe scatterers. Taken from (Newville, 2004). The spectrum is interpreted in terms of the EXAFS equation (2.09), χ( k) NS F( k) 2kR 2 = 2 0 2kσ 2r / λ( k ) sin(2kr + φ( k)) e e 2 (2.09) where: N is the path degeneracy; S 2 0 is the amplitude reduction factor; F(k) is the scattering magnitude function; (k) is the scattering phase function; the e -2kσ2 term represents the mean square displacement (disorder due to thermal motion); the e -2r/λ term is the mean free path term (to correct for inelastic scattering); and R is the interatomic distance. 91

97 Typically, a multiple scattering software package (such as FEFF) is used to calculate all possible single and multiple scattering paths (Figure 2.19) from the absorber atom using a set of Cartesian atomic coordinates. Thus, FEFF can calculate the F(k) and ϕ(k) terms of the EXAFS equation from multiple scattering theory. This leaves the terms N, S 2 0, σ 2 and R as unknowns in the EXAFS equation. A second package, uses a multivariate fitting algorithm (Levenberg-Marquardt) to adjust the unknown variables to give a best fit for the experimental data (such as Artemis; Ravel and Newville, 2005). Therefore, it is possible to extract fundamental speciation information regarding coordination numbers and interatomic distances from EXAFS data. Figure Illustration of some simple multiple scattering paths (Parsons et al., 2002) Small Angle X-ray Scattering Small Angle X-ray Scattering (SAXS; Hiemenz and Rajagopalan, 1997; Guinier, 1963; Porod, 1951) is concerned with the small angle elastic scattering of X-rays in a material. The principle is that elastic collisions occur between the incoming X-rays and particles causing scattering in all directions, then constructive and destructive interference causes changes in scattering intensity. In contrast to X-ray diffraction techniques (e.g. XRD), SAXS does not provide information about the internal structure of a material. Instead, SAXS can provide information such as particle morphology, size, and concentration. In this thesis, SAXS 92

98 techniques were used to determine morphology and size of intrinsic U(VI) colloids in high ph systems (Ch. 4). The SAXS data were collected on the I22 SAXS/WAXS beamline (Diamond Light Source Ltd) using a Pilatus 2M detector with camera lengths of 4 m (300 frames, 1 s/frame) and 10 m (60 frames, 10 s/frame). The data were generated using with a 12 kev X-ray beam. A typical SAXS experiment will involve a X-ray beam being passed through a sample, small angle scattered X-rays pass/travel through a vacuum tube and are then counted by a detector (Figure 2.20). A beam stop is used to prevent transmitted X-rays from reaching the detector. The longer the camera length the higher the values of q (scattering vector) can be obtained and thus information about larger particles can be gained (Hiemenz and Rajagopalan, 1997; Guinier, 1963; Porod, 1951). SAXS data may be collected on a bench top instrument or at dedicated synchrotron facilities which allow for higher X-ray intensities and thus increased sensitivity which was necessary for the nanoparticles described in Ch. 4 of this thesis. Figure Schematic of SAXS data collection set up (Hiemenz and Rajagopalan, 1997). 93

99 SAXS data are usually presented as intensity against the scattering vector (A -1 ) (Guinier, 1963; Porod, 1951). The scattering vector can be calculated from equation (2.10), 4π q = λ sin θ (2.10) where: q is the scattering vector; λ is the X-ray wavelength; and θ is the scattering angle. Two key regions are apparent in SAXS data, the Guinier and Porod regimes (Figure 2.21). The Guinier region starts at the lowest values of q and extends up to where the Porod region features begin (Figure 2.16; Guinier, 1963; Porod, 1951). Intensity in the Guinier region is dictated by equation (2.11), 2 2 q R g 3 I(q) = I e 0 (2.11) where I 0 is the intensity at q = 0 and R g is the radius of gyration. The radius of gyration represents the weighted average distance between any point within the particle and the centre. I 0 is estimated by the extrapolation of the SAXS pattern to q = 0 and is related to the excess electrons in the scattering volume by equations (2.12) and (2.13), ( ) 2 I = (2.12) 0 n e I 2 2 = ( ρ ) p n (2.13) 0 V where: n e is the excess electrons in the scattering volume; ρ is the difference in electron density between the medium and the particulate; and V p is the particle volume; and n is the particle number density. An additional parameter can be extracted from SAXS data, the invariant (Q) which can be calculated using equation (2.14) (Porod, 1951). The invariant is related to the volume fraction of the scatterers (Φ) (2.15) (Guinier, 1963). It is possible, if the particle density is low, to 94

100 determine the particle volume (V p ) from the invariant using equation (2.16; Guinier, 1963) (2.16). 2 I(q)q dq Q = (2.14) 2π Q 2 = Φ(1 Φ)( ρ) (2.15) I p (2.16) Q V 0 There are several ways to determine R g. Equation (2.14) can be used to determine particle radius from the value for R g, assuming the particles are spherical and homogenous (Hiemenz and Rajagopalan, 1997). 2 2 R g = R (2.14) 5 3 The Porod region can provide information about particle size and the nature of the particulate such as the surface properties (Porod, 1951). The Porod region is very sensitive to particle size. Figure 2.21 shows an example of data from a monodisperse system. As disorder in the particle sizes increases (the system becomes polydisperse) the features in this region will become less defined, and for a system of high polydispersity, the features can disappear entirely. If the system is relatively monodisperse, the first peak in the Porod region after the Guinier region can be used to determine the particle size (Figure 2.21). Additionally, important information regarding the particulate morphology can be obtained from the Porod region by calculating the gradient on a log-log plot (Figure 2.21; Schmidt, 1991). If the gradient has a value of 4, the particles are perfectly smooth spheres. A gradient between 1 and 3 suggests the particle has mass fractal properties, whereas between a value of between 3 and 4, suggests surface fractal properties. 95

101 Figure Idealised SAXS data for a monodisperse population of spheroid particles, showing key features of the pattern. Model data was generated using the Irena software macro for IgorPro 6.2 (Ilavsky and Jemian, 2009). 2.9 References Dupré, B., Viers, J., Dandurand, J.-L., Polve, M., Bénézeth, P., Vervier, P., Braun, J.-J., Major and trace elements associated with colloids in organic-rich river waters: ultrafiltration of natural and spiked solutions. Chemical Geology 160, Egerton, R.F., Physical principles of electron microscopy. Springer. Gilmore, G., Practical gamma-ray spectroscopy. Wiley-Blackwell. Guinier, A., X-ray diffraction in crystals, imperfect crystals, and amorphous bodies. W. H. Freeman, San Francisco. Hiemenz, P.C., Rajagopalan, R., Principles of colloid and surface chemistry, Third Edition. CRC Press. 96

102 Horrocks, D.L., Applications of liquid scintillation counting. Academic Press, New York. Ilavsky, J., Jemian, P., Irena: tool suite for modeling and analysis of small-angle scattering. Journal of Applied Crystallography 42, Lakowicz, J., Introduction to fluorescence, principles of fluorescence spectroscopy. Springer US. Nakao, S., Determination of pore size and pore size distribution: 3. Filtration membranes. Journal of Membrane Science 96, Newville, M., Fundamentals of XAFS, Chicago. Ong, C.G., Leckie, J.O., Anion-exchange preparation of a 232 U radiotracer for α- particle liquid scintillation counting. Talanta 43, Parsons, J.G., Aldrich, M.V., Gardea-Torresdey, J.L., Environmental and biological applications of extended X-ray absorption fine structure (EXAFS) and X-ray Absorption near edge (XANES) spectroscopies. Applied Spectroscopy Reviews 37, Passo, C.J., Cook, G.T., Handbook of environmental liquid scintillation spectrometry. Packard. Perkampus, H., UV-Vis spectroscopy and its applications. Springer Pitois, A., Ivanov, P.I., Abrahamsen, L.G., Bryan, N.D., Taylor, R.J., Sims, H.E., Magnesium hydroxide bulk and colloid-associated 152Eu in an alkaline environment: colloid characterisation and sorption properties in the presence and absence of carbonate. Journal of Environmental Monitoring 10,

103 Pointeau, I., Coreau, N. and Reiller, P.E. (2008) Uptake of anionic radionuclides onto degraded cement pastes and competing effect of organic ligands. Radiochimica Acta 96, Ravel, B., Newville, M., ATHENA, ARTEMIS, HEPHAESTUS: data analysis for X- ray absorption spectroscopy using IFEFFIT. Journal of Synchrotron Radiation 12, Schmidt, P., Small-angle scattering studies of disordered, porous and fractal systems. Journal of Applied Crystallography 24, Tanner, S.D., Holland, J.G., Plasma source mass Spectrometry special publication. The Royal Society of Chemistry. Waseda, Y., Matsubara, E., Shinoda, K., X-Ray diffraction crystallography: Introduction, examples and solved problems. Springer. Weiner, B., Tschamuter, W., Fairhurst, D., Accurate particle sizing of high density materials to 10 nm Using an X-ray disc centrifuge with a moving source/detector. Brookhaven Instruments Corporation. 98

104 The following is a research manuscript has been accepted for publication in Geochimica et Cosmochimica Acta, October 2014 Chapter 3 99

105 U(VI) behaviour in hyperalkaline calcite systems Kurt F. Smith a,b, Nicholas D. Bryan b,c, Adam N. Swinburne c, Pieter Bots a, Samuel Shaw a, Louise S. Natrajan c, J. Frederick W. Mosselmans d, Francis R. Livens a,c, and Katherine Morris* a a Research Centre for Radwaste Disposal and Williamson Research Centre, School of Earth, Atmospheric and Environmental Sciences, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom. b National Nuclear Laboratory, Chadwick House, Risley, WA3 6AE, United Kingdom. c Centre for Radiochemistry Research, School of Chemistry, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom. d Diamond Light Source Ltd., Diamond House, Harwell Science and Innovation Campus, Didcot, Oxfordshire, OX11 0DE. UK *Corresponding author (katherine.morris@manchester.ac.uk) 100

106 Abstract The behaviour of U(VI) in hyperalkaline fluid / calcite systems was studied over a range of U(VI) concentrations (5.27 x 10-5 µm to 42.0 µm) and in two high ph systems, young and old synthetic cement leachate in batch sorption experiments. These systems were selected to be representative of young- (ph 13.3) and old-stage (ph 10.5) leachate evolution within a cementitious geological disposal facility. Batch sorption experiments, modelling, extended X- ray absorption fine structure spectroscopy, electron microscopy, small angle X-ray scattering and luminescence spectroscopy were used to define the speciation of U(VI) across the systems of study. At the lowest concentrations (5.27 x 10-5 µm 232 U(VI)) significant U removal was observed for both old and young cement leachates, and this was successfully modelled using a first order kinetic adsorption modelling approach. At higher concentrations (> 4.20 µm) in the young cement leachate, U(VI) showed no interaction with the calcite surface over an 18 month period. Small angle X-ray scattering techniques indicated that at high U concentrations (42.0 µm) and after 18 months, the U(VI) was present in a colloidal form which had little interaction with the calcite surface and consisted of both primary and aggregated particles with a radius of 7.6 ± 1.1 and 217 ± 24 Å, respectively. In the old cement leachate, luminescence spectroscopy identified two surface binding sites for U(VI) on calcite: in the system with 0.21 µm U(VI), a liebigite-like Ca 2 UO 2 (CO 3 ) 3 surface complex was identified; at higher U(VI) concentrations (0.42 µm), a second binding site of undetermined coordination was identified. At elevated U(VI) concentrations (> 2.10 µm) in old cement leachate, both geochemical data and luminescence spectroscopy suggested that surface mediated precipitation was controlling U(VI) behaviour. A focused ion beam mill was used to create a section across the U(VI) precipitate - calcite interface. Transmission electron microscope images of the section revealed that the calcite surface was coated with a nano crystalline, U containing phase. Selected area electron diffraction images of the U precipitate 101

107 which was formed at a U(VI) concentration of 4.20 µm were consistent with the formation of calcium uranate. XAS spectroscopy at higher concentrations ( 21.0 µm) suggested the formation of a second U(VI) phase, possibly a uranyl oxyhydroxide phase. These results indicated that in the young cement leachate, U(VI) did not react with the calcite surface unless U(VI) concentrations were very low (5.27 x 10-5 µm). At higher concentrations, speciation calculations suggested that U(VI) was significantly oversaturated and experimental observations confirmed it existed in a colloidal form that interacted with the mineral surface only weakly. In the old cement leachate systems at low concentrations batch sorption and luminescence data suggested that U(VI) removal was being driven by a surface complexation mechanism. However, at higher concentrations, spectroscopic methods suggest a combination of both surface complexation and surface mediated precipitation was responsible for the observed removal. Overall, U(VI) behaviour in hyperalkaline calcite systems is distinct from that at circumneutral ph conditions: at high ph and anything but low U(VI) concentrations, a surface mediated precipitation mechanism occurs; this is in contrast to circumneutral ph conditions where U(VI) surface complexation reactions tend to dominate. 102

108 1. Introduction In many countries the long-term management of intermediate level radioactive wastes will be for their final disposal in a deep Geological Disposal Facility (GDF) (DEFRA, 2008; NEA, 2004; Schwyn et al., 2012). The GDF will be designed to isolate radionuclides from the biosphere for sufficient time to allow the majority of the radioactive material to undergo radioactive decay, although for longer lived radionuclides it is clear that transport away from the engineered barrier cannot be ruled out. Once the waste has been packaged, typically by grouting in steel drums, it will be emplaced in the facility and backfilled. Currently it is anticipated that in the UK, the GDF will be backfilled with cementitious material, and engineering cement (e.g. "shotcrete") will be utilised in many designs. This means that consideration of the alkaline conditions that result from cement materials is of wide relevance (Schwyn et al., 2012; Schmidt, 1991). The cementitious environment is expected to remain alkaline (i.e. ph ) over extended periods and, in the saturated sub-surface, is intended to produce conditions that reduce radionuclide mobility due to formation of insoluble hydrolysis products and increased sorption of cations to negatively charged mineral surfaces (Braney et al., 1993). In a cementitious GDF over time, the portlandite (Ca(OH) 2 ) and calcium silicate hydrate (C-S-H) components of the cement will undergo carbonation via reaction with carbonate in groundwaters to form minerals including calcite (CaCO 3 ; Dow and Glasser, 2003). This will increase the quantity of calcite present in any evolved cementitious GDF. Calcite is also ubiquitous in the natural environment and may well be a component of the host rock environment. Calcite has been shown to sequester metals and radionuclides effectively by adsorption and/or incorporation (Dong et al., 2005; Geipel et al., 1997; Hay et al., 2003; Heberling et al., 2008; Parkman et al., 1998; Rihs et al., 2004a; Zachara et al., 1991; Zavarin et al., 2005). Thus, calcite clearly has the potential to be an important reactive mineral phase for radionuclides within the GDF environment. 103

109 U(VI) is stable in oxidising environments, readily forms the linear uranyl moiety (O=U=O; UO 2+ 2 ) and typically has increased mobility compared to U(IV). In addition, under the hyperalkaline to alkaline conditions relevant to this study, U(VI) is expected to remain stable under anoxic and modestly reducing conditions (Gaona et al., 2011). Certain studies concerning U(VI) interactions with calcite have focused on incorporation and coprecipitation (Curti, 1999; Elzinga et al., 2004; Kelly et al., 2003; Meece and Benninger, 1993; Reeder et al., 2000a; Reeder et al., 2001a), whilst others have studied sorption (Carroll et al., 1992; Elzinga et al., 2004; Geipel et al., 1997; Rihs et al., 2004a). In an evolving geological disposal facility, the geochemical environment will be dynamic and both sorption and incorporation/coprecipitation reactions with radionuclides are possible. There is a significant literature on U(VI) calcite interactions at ambient ph, for example, Elzinga, et. al. (2004) studied the reaction of U(VI) with calcite at ph 7.4 and 8.3 and over the concentration range 5 to 5000 µm using batch sorption experiments, Extended X-ray Absorption Fine Structure (EXAFS) and luminescence spectroscopy. EXAFS analysis of their batch sorption experiments with < 500 µm U(VI) produced data consistent with the formation of uranyl triscarbonate surface complexes and suggested that sorption in their systems was taking place via inner sphere complexation, in agreement with Rihs et al. (2004). Unfortunately, their EXAFS analysis could not identify any calcium backscatterers to corroborate the proposed inner sphere bonding, and thus the exact coordination environment of U(VI) in this system remained undetermined. The authors then used luminescence spectroscopy to explore whether different U(VI) surface species were present in their experiments: at < 100 µm U(VI) a calcium uranyl triscarbonate complex (in a liebigite-like structure) dominated, whilst for systems in the µm range, there was evidence that a different, undetermined uranyl carbonate species became significant. The authors suggested 104

110 this undetermined phase was likely to be intermediate between the liebigite-like (Ca 2 UO 2 (CO 3 ) 3 ) surface complex and U(VI) incorporated into calcite. Indeed, EXAFS and luminescence techniques confirm that when UO 2+ 2 is coprecipitated with calcite, it can exist in multiple coordination environments dependent on the method of precipitation (Reeder et al., 2000b; Reeder et al., 2001b). Furthermore, it has been shown that selective incorporation of UO 2+ 2 during calcite coprecipitation takes place on the - steps of the calcite (104) surface (Reeder et al., 2004; Rihs et al., 2004a). Refinement of this concept by Rihs et al. (2004) suggested that the number of binding sites in their experimental system was too low to account for the U(VI) sorption they observed, and that additional sites could be provided by calcium vacancies or "etch pits", which are effectively equivalent to additional - steps. Recent computational molecular dynamics simulations (Doudou et al., 2012) support preferential sorption of the Ca 2 UO 2 (CO 3 ) 3(aq) species onto the '-' steps but also suggest that outer sphere complexation on the calcite terrace should be thermodynamically favourable. In addition to surface complexation and incorporation, calcite has been shown to promote the formation of U(VI) precipitates, such as schoepite ((UO 2 ) 8 O 2 (OH) 12.12H 2 O; Carroll et al., 1992; Elzinga et al., 2004; Geipel et al., 1997; Schindler and Putnis, 2004). Overall, these studies have developed a good understanding of U(VI) behaviour in calcite systems, including incorporation and surface complexation mechanisms. However, most studies have focused on ambient environmental conditions, whilst in any cementitious GDF, alkaline ph conditions will prevail with ph values > 10.5 dominating over years and with elevated ionic strength compared to typical environmental conditions. This gap in knowledge needs to be addressed in order to underpin any performance assessment for a cementitious GDF. Here, we examine U(VI) interactions with calcite under the hyperalkaline conditions representative of an early stage evolution young and late stage evolution old 105

111 GDF using synthetic Young Cement Leachate (YCL) and Old Cement Leachate (OCL). We present data from batch sorption experiments using varying solid to solution ratios and a wide range of U(VI) concentrations, to develop a mechanistic understanding of U(VI) behaviour across relevant alkaline calcite systems. Batch sorption experiments, geochemical modelling, EXAFS, electron microscopy, Small Angle X-ray Scattering (SAXS) and luminescence spectroscopy have been applied to probe the fate and speciation of U(VI) in these experiments. 2. Methodology 2.1. Modelling All thermodynamic modelling was carried out with PHREEQC (Parkhurst and Appelo, 1996) using the ANDRA SIT database (ThermoChimie v.7.d June 2011). The database was modified to include the formation constant for Ca 2 UO 2 (CO 3 ) 3(aq) (Bernhard et al., 2001). All kinetic modelling was carried out in R (R-Core, 2012). The data were modelled with a simple, first order kinetic adsorption model that assumed the sorption site concentration was proportional to the mass of calcite. The chemical equation used in the kinetic modelling was, k 1 M S M (1) k 2 S and the rate equation was, d[m] k 1 [M][S] k 2 [MS] (2) dt where [M] refers to the metal concentration, [S] is the concentration of calcite surface binding sites, [M S ] is the concentration of U(VI) associated with surface binding sites and k 1 and k 2 are the forward and backward rate constants, respectively. The calcite binding site density value was arbitrary as in experiments, saturation of surface binding sites was not 106

112 approached. However, the use of this arbitrary value did not affect the goodness of fit, as the calcite binding site concentration acts as a scaling constant: for example, if the true value of [S] were lower, then k 1 would increase proportionately, but the fit would not change. In other words, the experimental data provided us with only the product, k 1 [S], not the absolute, separate values of the rate constant and surface site concentration. A value of 1 for surface binding site concentration [S] was used for modelling experimental data with a 0.1 kg L -1 solid to solution ratio and then adjusted proportionately for different solid to solution ratios Sorption Experiments U(VI) behaviour was studied in batch sorption experiments where the systems were preequilibrated with calcite prior to U(VI) additions. All chemicals used were analytical reagent grade and all solutions were made using Milli-Q (18.2 M Ω cm) water. In addition, all experiments and sample manipulations were performed in a CO 2 free environment. The calcite used in this study was a purchased high purity crystalline powder prepared by precipitation (SureChem Products Ltd) and was characterised via powder X-ray diffraction (see supporting information; Figure S1; Bruker D8 Advance). Prior to use, the calcite was passed through a < 63 µm sieve, and the surface area of this sieved material was measured as 0.26 ± 0.01 m 2 g -1 using N 2 B.E.T analysis. Depleted uranium as UO 2 (NO 3 ) 2 (BDH), was dissolved in 0.01 M HCl to give a stock solution concentration of 0.02 M. Uranium tracer ( 232 U(VI) in 2 M HNO 3 ; AREVA) was separated from ingrown 228 Th using ion exchange to allow Liquid Scintillation Counting (LSC) of 232 U, without interference from 228 Th and its daughters (Ong and Leckie, 1996). 107

113 For experiments at ultra-trace U(VI) concentrations, the experiments were spiked with 232 U to a concentration of 5.27 x 10-5 µm (10 Bq ml -1 ). For analysis, a 1 ml sample was mixed with scintillant (Scintisafe 3, Fisher Scientific) and 2 ml 1 M HCl, prior to counting on a LSC (Quantulus, PerkinElmer) instrument with appropriate, matrix matched standards. In this conformation, the detection limit was ~ 0.1 Bq ml -1 and the count time was adjusted such that the counting error was 5 % or less. In experiments bracketing the higher U(VI) concentration range ( µm), U(VI) was used and concentrations were determined using ICP-MS (Agilent 7500cx), again with appropriate, matrix matched standards. All experiments were run in triplicate. This work used two different synthetic solutions that reflected the chemical conditions (particularly ph and calcium concentrations) expected during the evolution of cement in contact with flowing water: OCL; and YCL (Table 1), either as made, or equilibrated for four days with calcite and then filtered (< 0.45 µm; Whatman nylon filter). During preparation to avoid oversaturation, the YCL solution was kept in a CO 2 free environment in an oven at 40 C for 48 hours prior to filtration through a 0.45 µm nylon filter. Selected calcite preequilibrated experiments using YCL, OCL and Milli-Q water were also run, which were equilibrated with calcite for four days, prior to filtration through a 0.45 µm nylon filter. Table 1. Synthetic cement leachate solution composition OCL YCL ph KOH (g L -1 ) NaOH (g L -1 ) Ca(OH) 2 (g L -1 )

114 Batch sorption experiments with solid to solution ratios between 1 : 500 and 1 : 2 were carried out. In these experiments, the calcite was mixed with the OCL or YCL and allowed to equilibrate with the solution for four days prior to spiking with 232 U or U solutions. For higher U(VI) concentrations, the solutions were spiked from a 2.5 mm U(VI) stock solution in 0.01 M HCl to yield final U concentrations between µm. For ultra-trace experiments, 232 U in 0.01 M HCl was added to give a final concentration of (5.27 x 10-5 µm / 10 Bq ml -1 ). All experiments were performed in triplicate unless otherwise stated Luminescence Spectroscopy In order to assess U(VI) speciation across the lower range of experimental concentrations, luminescence spectroscopy was used. Once batch data suggested that the experiments had reached equilibrium, U(VI) labelled calcite samples ( ppm on solids) were prepared for luminescence spectroscopy by centrifugation (500 g, 5 minutes) to leave a moist paste which was then mounted in a Suprasil quartz tube. Steady state emission spectra were recorded at 77 K using an Edinburgh Instrument FP920 Phosphorescence Lifetime Spectrometer equipped with a 5 W microsecond pulsed xenon flashlamp (with single 300 mm focal length excitation and emission monochromators in Czerny Turner configuration) and a red sensitive photomultiplier in a Peltier housing (air cooled, Hamamatsu R928P). Where indicated, delay time and gate time were applied during emission measurements. Lifetime data were recorded following excitation with a microsecond xenon flashlamp, using the multi-channel scaling method. Lifetimes were obtained by tail fit on the data, and the quality of fit judged by minimization of reduced χ 2 and residuals squared. Where the decay profiles are reported as mono-exponential, fitting to a double exponential gave no significant improvement in fit. 109

115 2.4. Electron Microscopy U(VI)-calcite precipitates were examined using both Scanning- and Transmission-Electron Microscopy (SEM/TEM). The SEM images were collected on a FEI Nova200 dual beam SEM/Focused Ion Beam (FIB) in backscatter electron imaging. SEM samples were mounted on a carbon pad, carbon coated and imaged. A thin section (approximately 10 µm x 8 µm x 150 nm) was prepared using a FIB fitted utilising Kleindiek micromanipulators. The thin section was then analysed with TEM using a FEI Tecnai F20 FEGTEM instrument, fitted with a Gatan Orius SC600 CCD camera. The thin section was imaged in high magnification mode and in High-Angle Annular Dark-Field (HAADF) mode to obtain Z contrast images. In addition, Selected Area Electron Diffraction (SAED) patterns were collected. Energy Dispersive X-ray (EDX) analysis was carried out on a Philips CM200 FEG-TEM fitted with an Oxford Instruments 80 mm 2 X-Max SDD detector Extended X-ray Absorption Fine Structure Spectroscopy Selected U(VI) reacted calcite samples ( ppm) were prepared for EXAFS spectroscopy by centrifuging (500 g, 5 minutes) before separating the moist paste for analysis. The solid sample (~0.2 g) was then mounted in a CO 2 free environment onto a Perspex sample holder and sealed with Kapton tape. The sample holder was then heat sealed in a CO 2 free argon environment and the sample stored at 4 C prior to analysis. EXAFS spectra were collected for the U L III -edge on B18 at the Diamond Light Source using a 9 element solid-state Ge detector, focusing optics (beam size 0.5 mm 2 ), and a Si-111 double crystal monochromator. Data collection was carried out at room temperature. Between 8 and 20 QEXAFS scans were collected per sample and averaged in order to increase the signal to noise ratio. Scans were normalised in energy space using an inline yttrium metal reference standard. Background processing was performed using Athena, and EXAFS modelling was 110

116 carried out using Artemis (Demeter; ; Ravel and Newville, 2005). All fitting was carried out using k 1, k 2, and k 3 (typical range, 3-12 k), while the best fit was calculated in R space (typical range, Å) by minimisation of the fit index, R. The reduced χ 2 term was used to assess the significance of extra shells in the fit. A forward through absorber Multiple Scattering (MS) path was included for the axial yl oxygen atoms, where it offered an improvement in the fit (Catalano and Brown, 2004). In the fitting of the U-O axial MS path, the parameters for distance and the Debye Waller-like factor (σ 2 ) were set to double the values for the U-O axial single scattering path (Catalano and Brown, 2004). In all cases, the inclusion of the U-O axial MS path did not require any additional parameters, and in all cases parameterisation used no more than two thirds of the total number of available independent data points Small Angle X-ray Scattering Small angle X-ray scattering was used to investigate the U(VI) colloidal particles in calcite equilibrated YCL solutions. The SAXS samples were prepared by spiking a calcite preequilibrated YCL solution to 42.0 µm U(VI) with a 2.5 mm U(VI) stock solution (in 0.01 M HCl). The sample was then aged under CO 2 free conditions for 18 months. Prior to analysis the sample was passed through a 0.45 µm Nylon filter and U(VI) was found to pass through the filter. SAXS data were collected on the I22 SAXS/WAXS beamline at the Diamond Light Source Ltd. using a Pilatus 2M detector (Henrich et al., 2009). Data were collected with a 12 kev X-ray beam, using a camera lengths of 4 m (300 frames, 1 s/frame) and 10 m (60 frames, 10 s/frame). The collected data were averaged, background subtracted using a U(VI)-free calcite equilibrated YCL and merged into one data set. Data analysis was carried out using the Irena software macro for IgorPro 6.2 (Ilavsky and Jemian, 2009). 111

117 3. RESULTS AND DISCUSSION 3.1. U(VI) Saturation and Speciation PHREEQC thermodynamic speciation and saturation calculations were carried out for 4.20 µm U(VI) in YCL and OCL solutions with and without calcite pre-equilibration (Table 2). These showed that both systems were predicted to be supersaturated with respect to several U(VI) phases. In OCL solutions, these phases were, in order of decreasing saturation, becquerelite (Ca(UO 2 ) 6 O 4 (OH) 6 8(H 2 O)), calcium uranate (CaUO 4 ), and calcium diuranate / calciouranoite (CaU 2 O 7 ). By allowing the model to precipitate becquerelite out of solution, a final equilibrium U(VI) solution concentration of 1.7 x 10-2 µm was predicted. However, if calcium uranate was the solubility limiting phase, the final U(VI) concentration was predicted to be 2.07 x 10-5 µm. In YCL solutions, speciation modelling predicted supersaturation of calcium uranate and clarkeite (Na 2 U 2 O 8 ). Allowing calcium uranate to precipitate, the model predicted a final U(VI) concentration of 2.26 x 10-6 µm, at which point clarkeite was undersaturated. Including calcite equilibration in the model did not significantly alter the model predictions (Table 2). Solubility experiments were carried out in calcite preequilibrated OCL and YCL showed that after equilibration, 42.0 µm U(VI) was stable in solution (< 0.45 µm nylon filter). This was in contrast to the PHREEQC calculations, which modelled the systems as significantly oversaturated. 112

118 Table 2. PHREEQC calculated saturation indices (SI*) for supersaturated U(VI) phases at 4.20 µm in YCL and OCL with and without calcite equilibration YCL calcite equilibrated YCL OCL calcite equilibrated OCL Phase SI* UO 4 Ca (cr) Clarkeite U 2 O 7 Na 2 (s) CaU 2 O 7 :3H 2 O (cr) Becquerelite (nat) Becquerelite (syn) ion activity product * where SI is the saturation index ( log ) solubility product The PHREEQC speciation calculations carried out at 4.20 µm U(VI) for solution speciation suggested that uranyl hydroxide species dominate aqueous speciation in both OCL and YCL - systems (Table 3). In OCL, the dominant uranyl hydroxide species was UO 2 (OH) 3 (86 %), whereas, in YCL it was UO 2 (OH) 2-4 (97 %). The Ca 2 UO 2 (CO 3 ) 3(aq) species was predicted to account for only 0.19 % of U(VI) speciation in OCL, although this was a much higher proportion than in YCL (Table 3). Free UO 2+ 2 accounts for a vanishingly small proportion of U(VI) in OCL and YCL (Table 3). Equilibrating OCL with calcite produced only minor changes in U(VI) speciation, with the most significant being the predicted presence of the 2- (UO 2 ) 2 (CO 3 )(OH) 3 species (0.05 %). In the YCL system, calcite equilibration had even less effect (Table 3). The proportions of Ca 2 UO 2 (CO 3 ) 3(aq) in our experimental systems modelled using thermodynamic calculations were 0.19 x % and 7.9 x % for OCL and YCL systems, respectively. These values were orders of magnitude smaller in comparison to other published calcite-u(vi) studies (Elzinga et al., 2004; Geipel et al., 1997; Reeder et al., 2001a), for example, Elzinga et al., (2003) calculated that the Ca 2 UO 2 (CO 3 ) 3 species accounted for > 70% of all UO 2+ 2 in their ph 7.4 and 8.3 systems equilibrated with atmospheric carbon dioxide. The very much reduced abundance of Ca 2 UO 2 (CO 3 ) 3 in our experiments was due to 113

119 both the very much lower carbonate concentrations, and elevated ph of the OCL and YCL systems. This has the potential to influence U(VI) behaviour in our systems, as the Ca 2 UO 2 (CO 3 ) 3(aq) species has been demonstrated as key in the sorption of U(VI) to calcite surfaces (Elzinga et al., 2004; Reeder et al., 2001a; Rihs et al., 2004b). Table 3. PHREEQC calculated U(VI) speciation (4.20 µm) as fractional composition in YCL and OCL with and without calcite equilibration Species 2- UO 2 (OH) 4 UO 2 (OH) 3 - YCL With calcite OCL With calcite UO 2 (OH) x x x x 10-3 UO 2 (CO 3 ) 3 4- (UO 2 ) 3 (OH) x x x x Ca 2 UO 2 (CO 3 ) 3 (aq) 7.9 x x 10-3 (UO2) 2 (CO 3 )(OH) 3 - CaUO 2 (CO 3 ) 3 2- UO 2 (CO 3 ) x x x x x 10-5 UO 2 (OH) x x x x 10-7 UO 2 (CO 3 ) (aq) 2.2 x x 10-7 UO 2 2+ (UO 2 ) 3 (OH) 5 + (UO 2 ) 4 (OH) x x x x x x x x x x x x 10-8 Overall, the high ph of both OCL and YCL systems lead to speciation calculations being dominated by uranyl hydroxides (UO 2 (OH) 2-4 and UO 2 (OH) - 3 ), though the lower [OH - ] in OCL allowed for minor contributions from uranyl carbonates and the calcium uranyl triscarbonate species. In both YCL and OCL, the modelling suggested that U(VI) was supersaturated with respect to various U(VI) phases, although experimentally 42.0 µm U(VI) was stable in solution over extended periods (as defined by filtration through a 0.45 µm nylon filter). 114

120 3.2. Sorption Experiments Batch U(VI) experiments were carried out in both YCL and OCL systems with varying solid to solution ratios and U(VI) concentrations. Initial experiments were carried out at ultra-trace concentrations (5.27 x 10-5 µm; 10 Bq ml -1 ) using a 232 U tracer to ensure undersaturation with respect to U(VI) oxyhydroxide phases. However, even with the use of an ultra-trace U(VI) spike, the systems were still predicted to be supersaturated with respect to calcium uranate (saturation index 0.84 (OCL) and 0.83 (YCL)) and calcium uranate like precipitates have been observed in high ph cement phase (Macé et al., 2013; Tits et al., 2011). Interestingly, no removal of U(VI) from solution ( 0.45 µm nylon filter) in control experiments without calcite present was observed (Bots et al., In review). The changing 232 U(VI) concentrations with respect to time for both YCL and OCL are shown in Figure 1. In both systems, U(VI) removal took place over 675 hours and after this time the system had reached apparent equilibrium. The proportion of U(VI) removed from solution in YCL at equilibrium was 71 ± 4 %; 27 ± 5 %; and 6 ± 4 % for the 1:10, 1:50, and 1:500 solid to solution ratios, respectively. The proportion of U(VI) removed from solution in OCL at equilibrium was 60 ± 0.9 %; 30 ± 5 %; and 1 ± 4 %, for the 1:10, 1:50 and 1:500 systems, respectively (tabulated K d values are available in the supporting information; Table S1). The geochemical modelling predicted that anionic uranyl hydroxide species (e.g. UO 2 (OH) 2-4, UO 2 (OH) ) dominated in these systems and that the concentrations of free UO 2 (aq) and other cationic U(VI) species were vanishingly small. Calcite has a point of zero charge ranging from 8 to 9.5, depending on the exact solution composition and sample history (Somasundaran and Agar, 1967). Therefore, the calcite surface would be carrying a negative charge in both OCL and YCL systems. Given this and the predicted aqueous speciation, it was unlikely that the observed sorption was due to outer sphere complexation in either YCL or OCL and binding was probably dominated by inner sphere interactions. 115

121 Figure 1. Fraction of U(VI) (5.27 x 10-5 µm) remaining in solution versus contact time with calcite as a function of solid to solution ratio. Lines represent model predictions (calibrated with the 1 : 10 data set only). Error bars are 1σ of five replicates. A sorption model was developed to describe U(VI) sorption kinetics in both OCL and YCL ultra-trace systems. The model was parameterised using the data for the 1:10 systems, and it was then used to predict U(VI) removal from solution with respect to time in the 1:50 and 1:500 systems (Figure 1). The OCL and YCL data were fitted separately to give rate constants (Equation 1) of k 1 = 1.78 x 10-6 L mol -1 s -1 and k 2 = 7.50 x 10-7 s -1 in YCL and k 1 = 8.20 x 10-7 L mol -1 and k 2 = 5.12 x 10-7 s -1 in OCL. The predictions for the 1:50 and 1:500 systems showed a good fit to the experimental data in the YCL and OCL systems (Figure 1). The model was then used to predict removal from solution at equilibrium successfully with respect to solid to solution ratio (Figure 2). Additional experiments were carried out with higher solid to solution ratios (1 : 2, 1 : 5) and were used to validate the sorption model by successfully predicting U(VI) concentration at solid to solution ratios much higher than the 1 : 10 system (Figure 2). These higher solid loading systems reflect 116

122 more realistic ratios for transport in the subsurface where the solid will be present in excess. The sorption model was able to describe U(VI) behaviour at 5.27 x10-5 µm across a wide range of solid solution ratios and this suggested that true surface complexation was occurring in these experiments. Unfortunately, there are no adsorption rate data for uranyl sorption to calcite surfaces available in the literature. However, the calculated adsorption rates in this study were slow compared to uranyl sorption on to other mineral phases such as goethite (Missana et al., 2003). In addition, in other uranium studies with calcite the authors typically allowed no more than a few days for the systems to reach a steady state before analysis (Rihs et al., 2004; Elzinga et al., 2003), which suggests that these high ph systems exhibit slower kinetics than would be expected. This is discussed in more detail below. Figure 2. Fraction of U(VI) (5.27 x 10-5 µm) remaining in solution at equilibrium versus calcite concentration. Symbols represent experimental data, lines represent predictions calculated using the sorption model. Only the filled triangle and circle data points (0.1 g ml -1 calcite) were used in fitting; hollow points are for comparison to the model. Error bars are 1σ from five replicates. 117

123 In order to explore uranium behaviour across a range of concentrations relevant to intermediate level waste disposal where U(VI) will be a significant component of wastes, batch sorption experiments at higher concentrations ( µm) were carried out in YCL and OCL. In experiments with 42.0, 21.0 and 4.20 µm U(VI) YCL solutions, no statistically significant removal of uranium (< 0.45 µm) was observed over a period of 18 months and over a range of solid to solution ratios (1 : 10, 1 : 50, and 1 : 500). At these concentrations, thermodynamic speciation modelling predicted that the YCL solution was significantly oversaturated with respect to calcium uranate (SI = 7.27) yet the U(VI) remained < 0.45µm filterable, probably due to colloid formation. In the OCL experiments with 4.20, 2.10, 0.42 µm U(VI) OCL solutions there was significant U(VI) removal (< 0.45 µm) from solution in all cases by 912 hours (Figure 3). At 4.20 µm U(VI), 87.3 ± 0.5, 89.1 ± 0.5 and 99.9 ± 0.7 % removal took place in the 1:10, 1:50 and 1:500 systems, respectively by 912 hours. This trend continued at 2.10 µm U(VI), with 76.8 ± 6.5, 88.3 ± 0.8 and 49.3 ± 17.4 % removed for the 1 : 10, 1 : 50 and 1 : 500 systems by 912 hours. However, at the lowest U(VI) concentration (0.42 µm), removal was reduced for all solid to solution ratios compared to the higher concentrations: removal of 48.5 ± 1.7, 13.4 ± 0.8, and 0.90 ± 0.5% U(VI) was observed, in the 1 : 10, 1 : 50, and 1 : 500 systems, respectively (K d values are available in the supporting information; Table S1), by 912 hours. In the 4.20 and 2.10 µm U(VI) OCL systems, removal did not appear to depend on solid to solution ratio, with consistently high removal. By contrast, in 0.42 µm U(VI) OCL solutions removal from solution appeared to be directly related to the solid to solution ratio (Figure 3). Control samples without calcite present showed no U(VI) removal (< 0.45 µm) at all concentrations (data not shown). The enhanced U(VI) removal from solution observed in the high U(VI) systems (4.20, 2.10 µm) in comparison to the 0.42 µm system, where the sorption model satisfactorily predicted behaviour, does not fit with a surface complexation mechanism which would be limited by 118

124 the number of surface binding sites. Therefore, in the 4.20 and 2.10 µm systems, U(VI) removal most likely involved a surface mediated precipitation mechanism. This precipitation did not take place in the no calcite experiments and clearly required the calcite surface. This is consistent with removal driven by the precipitation of various U(VI) phases observed by past workers, mainly oxyhydroxides and carbonates, although under different ph and carbonate concentrations to these studied here (Carroll et al., 1992; Elzinga et al., 2004; Geipel et al., 1997; Schindler and Putnis, 2004). However, in our experiments, removal occurred at lower U(VI) concentrations than reported elsewhere. This is probably due to the high ph and low carbonate concentrations in our study, which was aimed at deep disposal conditions. The sorption model was applied to the 0.42 µm U(VI) OCL data and was parameterised using the data from the 1 : 10 systems. This was then used to predict U(VI) removal from solution with respect to time in the 1 : 50 and 1 : 500 systems (Figure 3). The OCL data gave rate constants (Equation (1)) of k 1 = 1.78 x 10-6 L mol -1 s -1 and k 2 = 1.23 x 10-6 s -1. The predictions for the 0.42 µm U(VI) 1 : 50 and 1 : 500 systems provided a satisfactory fit for the experimental data (Figure 3) and thus supported a surface complexation mechanism for this lowest U(VI) concentration experiment. The difference in rates measured in the 0.42 µm and the ultra-trace (5.2 x 10-5 µm) systems may suggest that multiple binding sites are available for U(VI) at the calcite surface, which is consistent with the literature (Elzinga et al., 2004). The kinetics of U(VI) removal were slow across all systems. Interestingly, all experiments were predicted to be supersaturated with respect to calcium uranate, even at low concentrations (5.27 x 10-5 µm) and the observed slow kinetics may be explained by U(VI) being speciated as calcium uranate colloids. Any U(VI) that was present in the form of a colloid would be unable to interact with the calcite surface. The small fraction of U(VI) that is truly dissolved in solution would be free to interact and sorb to the calcite surface. In a static system, this would indeed mean that the system rapidly reaches equilibrium. However, 119

125 if the colloidal population were undergoing a constant, slow dissolution / recrystallisation process, through re-equilibration the fraction of free U(VI) truly dissolved in solution would be replenished. This would lead to a gradual increase in the fraction of U(VI) associated with the calcite surface and a decrease in solution concentrations. Figure 3. Fraction of U(VI) ( µm) remaining in OCL solution versus contact time with calcite as a function of solid to solution ratio (1:10 (A), 1:50 (B), 1:500 (C)). Lines represent predictions from the kinetic model for the 0.42 µm data calculated from the 1:10 data set with rate constants k 1 = 1.78 x 10-6 L mol -1 s -1 and k 2 = 1.23 x 10-6 s -1. Error bars are 1σ of three replicates. 120

126 3.3. OCL Luminescence Spectroscopy To investigate the speciation of U(VI) associated with the calcite solid in the OCL systems, a series of luminescence spectra taken at 77 K were collected from calcite samples reacted with 0.21, 0.42, 2.10 and 4.20 µm U(VI) OCL solutions (1 : 50 solid to solution ratio) and these data were compared with literature values and standards. Following excitation at 250 nm, the emission spectrum recorded for the 0.21 µm sample exhibited a vibrationally resolved structure despite a low signal to noise ratio (Figure 4A). The emission maximum was centred at 501 nm and the estimated E 0-0 transition at cm -1. The kinetic traces were fitted with biexponential decay kinetics, indicative of the presence of two distinct emissive species and the radiative lifetimes were calculated as 736 ± 12 µs (55%) and 1290 ± 20 µs (45%) (501 nm detection; Table S2). The fact that the lifetime of the second emissive species was considerably longer that of the first component used in the kinetic fitting model suggests that the uranyl(vi) ion is in an environment where there are fewer non radiative decay processes operating. This is likely to be due to loss of a carbonate / water ligand(s), which are known to quench the excited electronic state of uranyl(vi) (Natrajan, 2012). These data are remarkably similar to the results of Elzinga et al. (2004) for a calcite sample reacted at ph 8.3 with a 5 µm U(VI) solution (Figure 4B) and for liebigite (Reeder et al., 2000a; Figure 4C) although no lifetime data were reported in these studies. This suggests that, in our study, a liebigitelike Ca 2 UO 2 (CO 3 ) 3 surface complex was forming at low U(VI)-concentrations. 121

127 Figure 4. A: An ungated luminescence spectrum collected from the OCL 0.21 µm U(VI)- calcite sample following 250 nm excitation. B: A luminescence spectrum of a U(VI)-calcite surface complex sample (5 M U(VI), ph 7.4) following 420 nm excitation from Elzinga et al. (2004). C: Luminescence spectrum (293 K) of a liebigite (Ca 2 UO 2 (CO 3 ) 3 ) standard following 420 nm excitation (Reeder et al., 2000a). Luminescence data (Figure 5A) were also collected for a 0.42 µm U(VI) reacted calcite sample following 250 nm excitation at 77 K. This spectrum exhibited an emission maximum at 512 nm, an apparent E 0-0 transition of cm -1 and less resolved vibrational fine structure when compared to the 0.21 µm system. A double exponential fit of the luminescence decay profile gave lifetimes of 141 ± 2 µs (32%) and 515 ± 8 µs (68%) (500 nm detection), which were reduced in comparison to the 0.21 µm data. This spectrum resembled the data collected for the U(VI)-calcite coprecipitate standard but the emission maximum values were red shifted by approximately 7 nm (Figure 5C; Table S2). In addition, the 0.42 µm spectrum closely matched a 'short gate' ( ms) 100 µm sample reported by Elzinga et al. (2003) (Figure 5B) and the radiative lifetimes were considerably shorter than those measured for the 0.21 µm sample suggesting the uranyl(vi) ion had undergone some 122

128 chemical transformations. With this information, we propose that the 0.42 µm spectrum collected in this study represents a second U(VI)-calcite surface coordination environment. Given that the 0.42 µm spectrum was collected ungated, it suggests that in our OCL system, this species was more abundant compared to the experiments of Elzinga et al. (2004). Our work was carried out in low carbonate conditions, in contrast to other studies which have been carried out in equilibration with atmospheric carbon dioxide (Carroll et al., 1992; Elzinga et al., 2004; Schindler and Putnis, 2004) or even with co-addition of carbonate (Elzinga et al., 2004). Figure 5. A: An ungated luminescence spectrum collected from the 0.42 µm U(VI) OCL calcite sample following 250 nm excitation. B: A 'short gate' ( ms) spectrum of 100 µm U(VI) sorbed onto calcite following 420 nm excitation from the literature (Elzinga et al., 2004). C: Ungated luminescence spectrum collected from U(VI) incorporated into calcite standard following 250 nm excitation. 123

129 In OCL at 4.20 µm U(VI), the highest concentration luminescence experiment, the spectrum showed the presence of two species, with peaks at 524, 543 and 567 nm (and minor peaks at 482, 505 nm; Figure 6A). The spectrum collected for the 2.10 µm U(VI) sample was similar to the 4.20 µm U(VI) sample, but the emission bands were red shifted by approximately 10 nm (Figure 6B), suggestive of an elongation of the uranyl bond in this species (Redmond et al., 2011). Additionally, the spectra were vibrationally broadened compared to the lower uranyl loading samples. This is indicative of a lower formal bond order of the uranyl unit and a significantly altered electronic (and therefore chemical) environment in comparison with the lower U(VI) concentration systems. The nm region in both 4.20 and 2.10 µm samples was broadly similar to the spectrum collected from a U(VI) phase precipitated by an oversaturation experiment with 84.0 µm U(VI) OCL solution (Figure 6C) and it also corresponds well to the calcium uranate spectrum reported by Tits et al., 2005 (Figure 6D) and various other U(VI) phases, such as schoepite and becquerelite (Wang et al., 2008). A triple exponential fit of the luminescence decay was required for the 2.10 and 4.20 µm samples, suggesting that U(VI) was hosted in several different environments and highlighting the complicated speciation of uranyl in these samples. The lifetimes calculated for the 2.10 µm system were 26 ± 1 µs (6%), 94 ± 3 µs (35%) and 422 ± 11 µs (57%) (512 nm detection), whereas for 4.20 µm they were slightly shorter at 22 ± 1 µs (13%), 105 ± 4 µs (6%), and 319 ± 11 µs (36%) (503 nm detection). The additional short lived species (22 µs) were in broad agreement with the lifetimes calculated for the 84.0 µm U(VI) OCL precipitate (23 ± 1 µs (35%) and 75 ± 1 µs (53%) µs; 542 nm detection). The extra component to the lifetime fits and the emergence of the red shifted feature, strongly suggested that the spectra for the 2.10 and 4.20 µm samples were a combination of different species with both a U(VI) solid phase and a U(VI) surface complex present. The relative intensities of the minor peaks (492, 512 nm), which we assume are indicative of the surface complex, were greater in the 2.10 µm 124

130 U(VI) sample compared to the 4.20 µm U(VI) sample. Qualitatively, these data indicated that the surface complex component of the spectra was more significant at lower U(VI) concentrations. It is interesting to note that there was an apparent red shift in the luminescence data for the solids as the U(VI) concentration increased. This has been reported previously for calcium uranate solids (Blasse, 1976) and is indicative of a reduction in the uranyl bond order as the form of the uranium species changes. The luminescence data suggest a complex, concentration dependent U(VI) behaviour in the OCL system. At the lowest loadings (0.21 µm), a liebigite-like U(VI)-calcite surface complex was identified. At 0.42 µm U(VI), a second as yet unidentified surface complex dominated the luminescence spectrum. At higher concentrations a different species which was red shifted with respect to the surface complexes appeared to grow in with increased concentration, consistent with the formation of a U(VI) solid phase. 125

131 Figure 6. A: An ungated luminescence spectrum from a 4.20 µm U(VI) OCL calcite sample following 250 nm excitation. B: Ungated luminescence spectrum from a 2.10 µm U(VI) OCL calcite sample following 250 nm excitation. C: Ungated luminescence spectrum from an (84.0 µm) OCL U(VI) precipitate following 250 nm excitation. D: Ungated luminescence spectrum a calcium uranate sample following 420 nm excitation from the literature (Tits et al., 2005). 126

132 3.4. OCL Electron Microscopy In order to characterise the OCL U(VI) surface mediated precipitate further and to explore the interface between it and the calcite solid, a sample from the 4.20 µm U(VI) OCL (1:50) calcite batch sorption experiment was analysed using electron microscopy. SEM images showed crystals (2-20 µm) with the rhombohedral morphology typical of calcite. Backscattering electron imaging (Z constant imaging) showed that the majority of crystals did not have a visible secondary phase, but several crystals were coated with a phase with higher average atomic number, which shows up as brighter regions on the image (Figure 7A). A section (10 µm long x 100 nm thick) was then prepared using a FIB mill in order to examine both the secondary phase and the interface between it and the calcite solid using TEM (Figure 7B). Elemental mapping via TEM-EDX analysis was carried out at the interface between the calcite and the secondary phase (Figure 8). The mapping indicated a uranium rich coating on the calcite surface, with a thickness ranging from approximately 10 nm to 100 nm. Within the limits of analysis for EDX, there was no indication of any uranium in the uppermost layer of the calcite structure (approximately 300 nm from interface, see supporting information; Figure S2). Further EDX analysis of the uranium-rich surface coating showed it contained uranium, calcium and oxygen. The peaks for all other elements are related to either the support grid (Cu and C) or the FIB process (Ga, see supporting information; Figure S3). This suggests that the secondary phase is a U/Ca phase. 127

133 Figure 7. A: Backscattering electron image of calcite crystal coated with an electron dense material, lighter areas indicate higher electron density. The white line indicates the area in which the TEM thin section was prepared. B: SEM image of the calcite thin section (approximately 10 µm wide and 100 nm at its thinnest). Figure 8. Energy dispersive X-ray elemental map overlaid onto high angle annular dark-field image (z contrast) of the calcite surface showing a thin uranium rich (green) coating. 128

134 Further TEM analysis of the uranium-calcite interface showed that the U(VI)-bearing precipitate consisted of elongated nanoparticles, nm in length and approximately 20 nm in thickness (Figure 9). High resolution images of the particles show lattice fringes indicating the particles are nano-crystalline. The individual crystals were orientated roughly parallel to the calcite surface, as shown in Figure 10A. Interestingly, the pristine calcite surface with no secondary phase showed little distortion and a clean termination of the calcite lattice (Figure 10B). However, the uranium coated surface showed indications of distortion and an uneven surface, with the lattice terminating in multiple orientations (Figure 10 B). This suggested that the formation of the uranium solid had altered the calcite surface, either during precipitation or though interaction with the solid once it had formed. Figure 9. High magnification image of a calcite surface coated with a uranium containing phase. The selected area electron diffraction image (Figure 11) was collected from site

135 In order to identify the U(VI) rich precipitate, SAED patterns were collected from several sites with a representative example, site 1, shown in Figure 9. This showed the presence of diffraction rings which is indicative of the crystallites being oriented in multiple directions (Figure 11). The PHREEQC saturation calculations suggested that becquerelite (Ca(UO 2 ) 6 O 4 (OH) 6 8(H 2 O)) was a likely candidate for the precipitate, but the crystallographic data for this phase (Piret-Meunier and Piret, 1982) was a poor fit to the SAED data (see supporting information; Figure S4). Likewise, it was a poor match for other candidate phases, such as schoepite (see supporting information; Figure S5). However, the pattern was a good match for calcium uranate (Figure 11; Zachariasen, 1948) and this is consitent with the composition of the phase as determined by EDX. Calcium uranate phases such as vorlanite (cubic CaUO 4 ; Galuskin et al., 2011) and calcium diuranate / calciouranoite (Ca 2 U 2 O 7 ; Rogova et al., 1974) produce very similar diffraction patterns (see supporting information; Figure S6). Therefore, we have putatively identified the solid U(VI) phase present at the mineral interface in this system as calcium uranate. This is in contrast to other studies which have reported the formation of solid phases, such as schoepite (Schindler and Putnis, 2004) and uranyl hydroxide (Geipel et al., 1997) in calcite sorption experiments, although it is noteworthy that these experiments were at lower ph values and in equilibrium with atmospheric CO 2. A SAED pattern from the calcite solid (see supporting information; Figure S7) was confirmed as a single crystal of calcite. 130

136 Figure 10. A: High magnification TEM image of the uranium containing phase and a distorted calcite surface B: High magnification TEM image of a 'pristine' calcite surface. 131

137 Figure 11. Selected area electron diffraction image collected from U(VI) precipitate (Figure 9, site 1) overlain with powder diffraction data calculated from crystallographic data for calcium uranate (Zachariasen, 1948) OCL Extended X-ray Absorption Fine Structure Spectroscopy Due to the low U(VI) concentrations in many experiments, only a select set of samples with low solid to solution ratios, and high U(VI) uptake (21.0 and 42.0 µm; 1 : 500) were suitable for EXAFS analysis. EXAFS spectra and corresponding Fourier transforms, including fits to the data, for these samples and a standard consisting of calcite coprecipitated with U(VI) are shown in Figure

138 Figure 12. k 3 -weighted χ functions with Fourier transforms for: (A) 0.10 g calcite reacted with 50 ml 21.0 µm U(VI) OCL solution; (B) 0.10 g calcite reacted with 50 ml 42.0 µm U(VI) OCL solution; (C) U(VI) coprecipitated with calcite standard. Solid lines represent experimental data while dashed lines are the fits using the parameters listed in Table 4. The best fit to the EXAFS spectra from the coprecipitated U(VI) calcite standard indicated the U was bonded to 2 axial oxygens at 1.81 Å, and 6 equatorial oxygens at distances between 2.20 and 2.38 Å. In addition, there was a shell of 3 C atoms at 2.90 Å (Table 4). This is consistent with previous models for U(VI) incorporation into calcite which demonstrates that U(VI) incorporated into calcite has a similar coordination environment to a U(VI)-calcite surface complex (Elzinga et al., 2003). However the data collected for the U(VI) reacted calcite samples did not resemble the coprecipitated standard. Therefore, given this and previous data suggesting a surface mediated precipitate, the data were interpreted in the context of a U(VI) solid phase. The spectra are similar to those produced by various U(VI) oxyhydroxides in the literature (Catalano and Brown, 2004), however, the TEM data strongly supported the formation of a calcium uranate solid. It was not possible to fit the EXAFS data with a calcium uranate crystal structure (Zachariasen, 1948) without a significantly shorter 133

139 uranyl like O ax distance of 1.81 Å compared to the 1.97 Å found in the structure. In addition, the XANES spectrum for the 42.0 and 21.0 µm U(VI) reacted calcite samples did not resemble calcium uranate (Macé et al., 2013), and instead appeared closer to a uranyl like schoepite reference, which contained a shorter U-O ax distance as demonstrated by the shoulder in the edge region (Figure 13). Given that becquerelite (Ca(UO 2 ) 6 O 4 (OH) 6 8(H 2 O)) was predicted to be supersaturated in calcite equilibrated OCL, the EXAFS data were fitted using scattering paths and fixed coordination numbers taken from becquerelite crystallographic data (Burns and Li, 2002) which has a shorter O ax distance than calcium uranate (Table 4). The U-O ax -U scattering path was refined to a distance of 1.81 ± 0.01 Å and the five equatorial oxygens were fitted as two shells of oxygen atoms at 2.23 and 2.40 Å, with an occupancy of 2.5 each. This was consistent with becquerelite, which has 5 equatorial oxygen atoms at distances varying from 2.10 to 2.60 Å (Burns and Li, 2002). This offered a significant improvement over a single shell fit (> 5% improvement in reduced χ 2, see supporting information; Table S3). A further two shells of uranium at 3.66 Å and 3.86 Å could also be fitted, which is again consistent with the becquerelite structure which has four uranium atoms at distances varying from 3.7 to 3.9 Å (Burns and Li, 2002). Interestingly, no calcium scatters were identified in the EXAFS modelling, most likely because of its relatively low mass and long distance from the absorber atom (4.20 Å). It is noteworthy that, while becquerelite crystallographic data were used in the calculation of the theoretical scattering paths and amplitudes, subtle differences between becquerelite and other uranium phases such as schoepite (Allen et al., 1996; Finch et al., 1996) would be difficult to distinguish with EXAFS spectroscopy. In summary, SAED data from this study suggested the formation of calcium uranate in the calcite systems at a U(VI) concentration of 4.20 µm. By contrast, EXAFS data from 21.0 µm and 42.0 µm U(VI)-calcite samples were modelled with a U(VI) phase with a shorter axial oxygen distance, best modelled as a uranyl oxyhydroxide mineral 134

140 such as becquerelite. This suggested that the speciation changed between the 4.20 µm and the > 21.0 µm systems, with calcium uranate dominating the samples at lower U(VI) concentrations (2.10 and 4.20 µm) and a uranyl oxyhydroxide precipitate dominating at 4.20 µm and highlights the complexity of the system. Figure 13. Normalised XANES spectra for the U L III edge of a 42.0 µm U(VI) reacted calcite sample (1 : 500) and Ca-uranate (Macé et al., 2013) and schoepite standards. 135

141 Table 4. EXAFS best fit parameters for U(VI)-calcite samples 42.0 µm (A) 21.0 µm (B) Coprecipitated (C) N R (Å) σ 2 (Å 2 ) N R (Å) σ 2 (Å 2 ) N R (Å) σ 2 (Å 2 ) U-O ax ± ± ± ± ± ± U-O eq ± ± ± ± ± ± U-O eq ± ± ± ± ± ± U-C ± ± U-U ± ± ± ± U-U ± ± ± ± S ΔE Reduced χ R Coordination numbers for samples A and B were fixed to those found in becquerelite while sample C were fixed to those found in liebigite. A forward through absorber multiple scattering path was included for the axial oxygens in all systems. 136

142 3.6. YCL Small Angle X-ray Scattering YCL batch sorption experiments suggested that U(VI) was stable in solution at elevated concentration (42.0 µm; < 0.45 µm filtered) over an extended time, and had little affinity for the calcite surface. However, PHREEQC modelling suggested that the system was significantly supersaturated with respect to U(VI) oxydroxide and calcium uranate phases. This highlighted that U(VI) may be present as a stable colloidal phase unable to interact with the calcite surface. To explore this further, a 42.0 µm U(VI) calcite equilibrated YCL solution was aged for 18 months and the filtered solution was then analysed using SAXS. The SAXS data showed scattering above background (Figure 14), indicating the U(VI) was present in a colloidal form. Figure 14. SAXS data collected from a 42.0 µm U(VI) calcite equilibrated YCL solution, aged for 18 months. Solid line represents the experimental data, the dashed line the modelled intensity, and the dotted lines indicate the relative contribution of the two particle populations. The gradient of the Porod region was calculated as

143 The best fit to the SAXS data was achieved using a two particle population model: firstly, a spheroid population to represent primary particles; and secondly a population with an algebraic globule form factor (Reidy et al., 2001) which represent aggregates of the primary particles (Figure 14 and Table 5). A best fit of the model against experimental data suggested the primary particles and aggregates had a radius of 7.6 ± 1.1 Å and 217 ± 24 Å, respectively (Table 5). The Porod slope for the larger particles was approximately -3 (Figure 14), which suggests the particulates had surface fractal properties (Schmidt, 1991). The volume ratio of the primary particle to aggregates was The SAXS data suggested that the U(VI) in the YCL systems was present as suspended nanoparticles. It is not possible to determine the chemical composition of the colloid population using SAXS, however, given the U(VI) saturation calculations (Table 2), calcium uranate is clearly a candidate. It is also not possible to determine the proportion of U(VI) present as a colloid using SAXS, however PHREEQC calculations suggest that, if the system were at equilibrium with calcium uranate, > 99% of U(VI) would be present as a solid phase. Table 5. Summary of fitted parameters in SAXS model Volume Ratio Radius (Å) Polydispersity* (Primary/Aggregate) Primary Particles 7.6 ± ± Aggregated Particles 217 ± ± 0.23 *log-normal distribution 3.7 U(VI) Speciation and Uptake Mechanisms with ph: Concentration and Surface Area. The data collected in this study demonstrate the complex behaviour of U(VI), with the mechanisms of interaction changing from surface complexation, to surface precipitation, to formation of a U(VI) colloidal phase depending on the cement leachate ph, amount of calcite, and U(VI) concentration (Table 6). 138

144 Table 6. Summary of mechanisms of U(VI) uptake to calcite solids in calcite-equilibrated OCL systems. U(VI) concentration range 0.42 µm < 0.21 µm 0.42 µm 2.10 µm > 21.0 µm U(VI)-calcite mechanism Surface complexation dominates removal, as demonstrated with batch data and kinetic modelling. Luminescence spectroscopy suggest a liebigite-like Ca 2 UO 2 (CO 3 ) 3 U(VI)-calcite surface complex as the primary binding site at low concentrations. Luminescence spectroscopy suggest a second U(VI)-calcite binding site exists, resembles spectrum from U(VI) coprecipitated with calcite. Surface mediated precipitation denominates removal. SAED data confirms calcium uranate solids dominate over sampling area. Surface complexation still likely to account for a fraction of U(VI) uptake. Surface mediated precipitation continues. XANES and EXAFS spectroscopy suggest U(VI) speciated as uranyl oxyhydroxide such as becquerelite or schoepite At low UVI concentrations in both the YCL ([U(VI)] = 5.27 x 10-5 µm) and OCL (U(VI) = 5.27 x 10-5 to 0.42 µm) systems, U(VI) removal could be directly related to the available surface area, and was successfully modelled assuming a surface complexation mechanism. Luminescence spectroscopy from a 0.21 µm U(VI) OCL reacted calcite sample suggested a liebigite-like Ca 2 UO 2 (CO 3 ) 3 surface complex identical to that reported by Elzinga et al. (2004). In contrast, the spectrum produced by a 0.42 µm U(VI) OCL reacted calcite sample was different, resembling a redshifted U(VI)-calcite incorporated species and with similarities to spectra reported for U(VI) sorbed to the '-' steps of the calcite (104) surface (Elzinga et al., 2004). This species was interpreted as representing a second, surface binding site of lower U(VI) reactivity than the liebigite-like surface complex which was similar to the calcite incorporated species observed here and reported in the literature (Elzinga et al., 2004; Reeder et al., 2000). The identification of U(VI)-calcite surface complexation implies U(VI) 139

145 removal may be reversible in some systems where the concentration is low. However, it is interesting to note the similarity between the second binding site and U(VI) incorporated into calcite is clearly worthy of further consideration. In batch sorption experiments at higher U(VI) concentrations in OCL ( µm) the relationship between U(VI) removal and calcite surface area broke down. This suggested a new regime where formation of a U(VI) precipitate occurred in agreement with other studies (Carroll et al., 1992; Elzinga et al., 2004; Geipel et al., 1997; Schindler and Putnis, 2004). Indeed, control experiments with calcite equilibrated solutions showed no removal of U(VI) upon filtration, suggesting a calcite surface induced reaction for U(VI) removal in the systems with mineral present. In addition, the luminescence data for higher U(VI) concentrations ( µm U(VI) OCL) showed features which fit a variety of U(VI) solid phases (Tits et al., 2005; Wang et al., 2008). TEM analysis from a sample prepared with 4.20 µm U(VI) indicated that the surface phase was a calcium uranate. However, at 21 µm U(VI) loading, XAS suggested that the surface precipitate was a uranyl oxyhydroxide phase. Given the differences between the XAS (42.0 µm) and TEM data (4.20 µm), it is possible that the system is undergoing predominantly calcium uranate precipitation at lower concentrations, while uranyl oxyhydroxide phases dominate at higher concentrations. This is consistent with the fact that the increased U(VI) concentration significantly increases the saturation of becquerelite (uranyl oxyhydroxide) by approximately five units (see Supporting Information; Table S4). In the ph 13.3 YCL system, U(VI) uptake ( long time scales (18 months), and in the OCL system experiments were typically observed to have long equilibration times for metal uptake (> 1 month). This was surprising as metal 140

146 complexation to mineral surfaces normally displays fast kinetics with typical equilibration times between hours (Rihs et al. 2004; Davis et al. 1987; Giammar and Hering, 2001). Interestingly, SAXS analysis indicated that in YCL at high U(VI) concentrations, U(VI) was largely present as a stable colloidal phase over 18 months which had a primary particle radius of 7.6 Å. These data suggest that for the YCL system, stable colloidal phases will contribute the dominant species (> 99 %) of U(VI) present, presumably as one of the oversaturated phases predicted from the thermodynamic modelling and presumably explaining the poor reactivity of U(VI) at all but ultra-trace levels in the YCL system. Also, if a colloidal population were present in the other OCL and low level YCL systems it could explain the slow kinetics of removal observed in experiments. The presence of colloidal U which is stable over year timescales in high ph systems, is a significant finding and clearly warrants further study in the context of uranium mobility under geological disposal conditions (Bots et al., In review). 4 Conclusions At high ph in model cement leachate systems, U(VI) had several interaction mechanisms with calcite. At low concentrations ( 0.42 µm), U(VI) removal was successfully modelled assuming a surface complexation reaction. Luminescence spectroscopy identified two distinct U(VI) surface binding sites at two different U(VI) concentrations (0.21 and 0.42 µm). At higher concentrations ( 2.10 µm) experiments showed U(VI) removal via a surface mediated precipitation mechanism. This was confirmed by luminescence spectroscopy, which identified a U(VI) phase that corresponds well to various U(VI) solid phases, such as calcium uranate and becquerelite. Furthermore, a calcite crystal coated with a U(VI) solid phase in the 4.20 µm system was identified using TEM as calcium uranate. In experiments at high 141

147 U(VI) concentrations, 21.0 and 42.0 µm, EXAFS spectroscopy successfully identified U-U scatters, indicative of a uranyl oxyhydroxide like solid phase. For YCL systems, U(VI) showed no interaction with the calcite surfaces unless the concentration was very low (5.27 x 10-5 µm) and over a period of 18 months. Where U(VI) removal from solution was observed, the kinetics were slow relative to many metal ion uptake studies and experiments took a month and more to reach apparent equilibrium. Interestingly, despite the fact that U(VI) was stable in solution (< 0.45 µm), the systems were predicted to be supersaturated with various U(VI)-bearing phases. Further characterisation using SAXS techniques confirmed that a U(VI) colloidal form was present in these solutions. These colloids consisted of aggregated and primary particles with a diameter of 217 ± 24 and 7.6 ± 1.1 Å, respectively. Overall, these data demonstrate for the first time the complex nature of U(VI) interactions with calcite solids in hyperalkaline systems representative of young and aged cement leachates. They suggest that for the YCL system representative of early evolution within a GDF, calcite is likely to have limited impact on the mobility of U(VI). However, in OCL systems representative of an aged cementitious geological disposal facility, a complex set of mechanisms allows for significant U(VI) uptake by both surface complexation and surface mediated precipitation of U(VI) solid phases. Under these conditions, the capacity for U(VI) uptake by interactions with calcite is likely to increase with increasing U(VI) concentration and in direct contrast to circumnetural environments. 142

148 5 Acknowledgements Kurt Smith was a Nuclear FiRST DTC student, funded by EPSRC (FIRST EP/G037140/1). This work was carried out as part of the Biogeochemical Gradients and Radionuclide Transport (BIGRAD) consortium, funded by NERC (NE/H007768/1) and with assistance from Louise S. Natrajan (an EPSRC Career Acceleration Fellow) and Adam N. Swinburne (an EPSRC funded postdoctoral researcher). We thank Diamond Light Source for access to beamline I22 (SM5975) and beamline B18 (SP7593) that contributed to the results presented here. The authors would like to thank Paul Lythgoe (geochemical analyses), Mike Ward (electron microscopy / FIB milling), John Waters (electron microscopy / BET analysis), Katie Law (radiochemical separations), Gareth Law and Tim Marshall (XAS data acquisition) for assistance in the preparation of this work. Kurt Smith and Nick Bryan would also like to thank NNL for support. 6 References Allen, P.G., Shuh, D.K., Bucher, J.J., Edelstein, N.M., Palmer, C.E.A., Silva, R.J., Nguyen, S.N., Marquez, L.N., Hudson, E.A., Determinations of uranium structures by EXAFS: Schoepite and other U(VI) oxide precipitates. Radiochim. Acta 75, Bernhard, G., Geipel, G., Reich, T., Brendler, V., Amayri, S., Uranyl(VI) carbonate complex formation: Validation of the Ca 2 UO 2 (CO 3 ) 3(aq.) species. Radiochim. Acta 89, Blasse, G., Nature of the luminescent centers in calcium uranate (Ca 3 UO 6 ). Solid State Commun. 19, Braney, M.C., Haworth, A., Jefferies, N.L., Smith, A.C., A study of the effects of an alkaline plume from a cementitious repository on geological materials. J. Contam. Hydrol. 13,

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155 U(VI) behaviour in hyperalkaline calcite systems Supporting Information Kurt F. Smith a,b, Nicholas D. Bryan b,c, Adam N. Swinburne c, Pieter Bots a, Samuel Shaw a, Louise S. Natrajan c, J. Frederick W. Mosselmans d, Francis R. Livens a,c, and Katherine Morris* a a Research Centre for Radwaste Disposal and Williamson Research Centre, School of Earth, Atmospheric and Environmental Sciences, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom. b National Nuclear Laboratory, Chadwick House, Risley, WA3 6AE, United Kingdom. c Centre for Radiochemistry Research, School of Chemistry, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom. d Diamond Light Source Ltd., Diamond House, Harwell Science and Innovation Campus, Didcot, Oxfordshire, OX11 0DE. UK *Corresponding author (katherine.morris@manchester.ac.uk) 150

156 Figure S1. Powder X-ray diffraction pattern collected from calcite used throughout this study. 151

157 Figure S2. Energy dispersive X-ray spectrum collected from calcite crystal just below the coating. 152

158 Figure S3. Energy dispersive X-ray spectrum collected from coating on calcite sample and show peaks indicative of U. 153

159 Figure S4. Selected area electron diffraction data from uranium containing precipitate overlain with becquerelite crystallographic data. 154

160 Figure S5. Selected area electron diffraction data from uranium containing precipitate overlain with schoepite crystallographic data. 155

161 Figure S6. Selected area electron diffraction data from uranium containing precipitate overlain with Ca 2 U 2 O 7 crystallographic data. 156

162 Figure S7. Selected area electron diffraction data collected from a single calcite crystal. Table S1. Kd values (kg L -1 ) for U(VI)-calcite batch experiments Solid : [U(VI)] (µm) System Solution 5.3 x OCL YCL 1: : : : : : : : : :

163 Table S2. Summary of luminescence and lifetime data Sample ex (nm) em (nm) E 0-0 (cm -1 ) a 1 (µs) b 2 (µs) b 3 (µs) b k 1 (10 3 s -1 ) c 0.21 µm U(IV) ± 12 (55%) 1290 ± 22 (45%) µm U(IV) ± 2 (32%) 515 ± 8 (68%) µm U(IV) ± 1.1 (6%) 94.5 ± 2.7 (35%) 422 ± 11 (57%) µm U(IV) ± 1.3 (13%) 105 ± 4 (52%) 319 ± 11 (36%) 3.13 OCL ± ± precipitate (47%) (53%) Improved signal to noise was obtained using 250 nm excitation. a The apparent electronic origin values (E 0-0 ) were determined from the observed energy of the first vibronic band in the emission spectra. The exact location of E 0-0 is often difficult to ascertain in spectra of uranyl compounds especially at low resolution and its exact E 0-0 may lie at higher energies depending on the local symmetry of the species in question. b Lifetimes quoted are subject to the numerical error given after the lifetime and were collected at: 501 nm (0.21 µm sample); 500 nm (0.42 µm sample); 512 nm (2.10 µm sample); and 503 nm (4.20 µm sample). The relative percentage contributions of each exponential to the total intensity at time = 0 are given in brackets. c The radiative rate constant for the longest lived component in the kinetic trace reported. 158

164 Table S3. EXAFS best fit parameters for U(VI)-calcite samples with a single Oeq shell µm (A) 21.0 µm (B) Coprecipitated (C) N R (Å) σ 2 (Å 2 ) N R (Å) σ 2 (Å 2 ) N R (Å) σ 2 (Å 2 ) U-O ax ± ± ± ± ± ± U-O eq ± ± ± ± ± ± U-C ± ± U-U ± ± ± ± U-U ± ± ± ± S ΔE Reduced χ R Coordination numbers for samples A and B were fixed to those found in becquerelite while sample C were fixed to those found in liebigite. A forward through absorber multiple scattering path was included for the axial oxygens in all systems. 159

165 Table S4. Saturation indices of various U(VI) solid phases in calcite equilibrated OCL systems. U(VI) Phase SI 4.20 µm 42.0 µm Becquerelite (nat) Becquerelite (syn) CaU 2 O 7 :3H 2 O (cr) UO 4 Ca (cr) Schoepite (des) UO 2 (OH) 2 (beta)

166 The following is a research manuscript that has been prepared for publication. Chapter 4 161

167 Np(V) sorption and solubility in high ph calcite containing systems Kurt F. Smith a, Nick D. Bryan b, Gareth Law b, and Katherine Morris a a Research Centre for Radwaste Disposal and Williamson Research Centre, School of Earth, Atmospheric and Environmental Sciences, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom. b Centre for Radiochemistry Research, School of Chemistry, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom Abstract The behaviour of Np(V) in hyperalkaline / calcite systems was studied over a range of concentrations (1.62 x µm) and in two synthetic, high ph cement leachates. The cement leachates were selected to be representative of the chemical conditions expected in a young (ph 13.3, Na +, K +, Ca 2+ ) and old (ph 10.5, Ca 2+ ) cementitious geological disposal facility. These systems were studied using a combination of batch sorption and solubility experiments, X-ray absorption near edge structure (XANES) spectroscopy, and modelling to describe their equilibrium, kinetic and speciation behaviour. In calcite equilibrated old and young cement leachates, the solubilities were 9.68 and µm Np(V), respectively. This was consistent with the formation of a Np(V)O 2 OH (am) solid phase in the old leachate. However, this phase could not explain the solubility in the young cement leachate, as it was undersaturated. In ph 13.3 NaOH solutions, Np(V) solubility decreased with increasing calcium concentration, indicating that calcium may be involved in solubility control in the young system. Analysis using XANES on a precipitate from the young cement leachate 162

168 system showed a spectrum consistent with an elongated Np=O ax bond length compared to that of a neptunyl standard. This suggested that a calcium-neptunate-like phase may be significant in this hyperalkaline experiment. The sorption of Np(V) to the calcite was observed across a range of concentrations and solid to solution ratios in the young cement leachate system. Analysis of these data suggested that a combination of surface complexation and precipitation was probably responsible for the observed Np(V) removal. In the old cement leachate system, sorption across a range of concentrations was dependent on solid to solution ratio and thus binding site density. These data were successfully modelled using a surface complexation approach, which assumed the formation of a monodentate Np(V)-calcite surface complex (>CO 3 NpO 2 ) with a log K of 2.42 and a binding site concentration of 0.15 sites / nm 2. The kinetics of Np(V) removal were also studied with all systems showing slow sorption kinetics with reaction times of weeks to reach equilibrium. In the old cement leachate system, these kinetics were successfully modelled assuming the initial fractionation of Np(V) into two parts, only one of which was immediately available for sorption, but with apparent first order transfer from the unavailable fraction to the available. 1 Introduction Many counties are planning for disposal of higher activity radioactive wastes in an underground Geological Disposal Facility (GDF) at a depth of 200 to 1000 m (DEFRA, 2008; NEA, 2004; Schwyn et al., 2012). In the UK much of the intermediate level waste inventory have already been grouted, and it is expected that the GDF may be backfilled with cementitious material. In addition, cement is likely to be intrinsic to any deep subsurface engineering project. Upon resaturation of the sub-surface, the interaction of the cementitious materials with groundwater will create a region of high ph (> ph 10.5), the chemically disturbed zone (CDZ). In a cementitious GDF, the waste environment and CDZ are expected 163

169 to remain alkaline over extended periods (> 100,000 years). This is intended to produce conditions that reduce radionuclide mobility (Braney et al., 1993). Therefore, investigation of the alkaline conditions generated by cementitious materials is of wide relevance (Schmidt, 1991; Schwyn et al., 2012). Within a cement containing GDF, the portlandite (Ca(OH) 2 ) and calcium silicate hydrate (C-S-H) components of the cement will react with carbonate in groundwaters to form minerals, including calcite (Dow and Glasser, 2003). Calcite has been shown to sequester actinides effectively, by adsorption and/or incorporation, including: + NpO 2, UO 2+ 2, PuO 2+ 2 Am 3+, Th 4+ (Dong et al., 2005; Geipel et al., 1997; Hay et al., 2003; Heberling et al., 2008a; Parkman et al., 1998; Rihs et al., 2004; Zachara et al., 1991; Zavarin et al., 2005). Thus, calcite is likely to be an important mineral phase in many GDF designs, and clearly has the potential to be an important reactive mineral phase for radionuclides within the GDF environment. Several studies have examined Np(V) solubility over a range of environmental conditions, such as ph (5-13), concentration, ionic strength and solution composition (i.e. K + and Na + ) (Kaszuba and Runde, 1999; Neck et al., 1992; Novak et al., 1997; Petrov et al., 2011). In carbonate free environments over ph 10-12, NpO 2 OH (am) and Np 2 O 5 are expected to precipitate (Kaszuba and Runde, 1999; Petrov et al., 2011). Over time, Np(V) solubility is expected to reduce as the initial NpO 2 OH (am) undergoes a slow transition to a less soluble NpO 2 OH (am, aged) over several months (Neck et al., 1992). The phases MeNpO 2 CO 3.xH 2 O and Me 3 NpO 2 (CO 3 ) 2 where Me is Na + or K +, can also precipitate at sufficiently high carbonate concentration ( M) and ionic strength (Novak et al., 1997). There are only a handful of papers studying Np(V) behaviour in calcite systems. In the NpO 2 + calcite coprecipitation study of Heberling et al. 2008b, the coordination of Np upon 164

170 coprecipitation into calcite at ph was examined using EXAFS spectroscopy (Heberling et al., 2008a). The authors fitted their EXAFS data using four shells: 2.1 oxygen atoms at 1.86 Å to account the axial oxygens in the neptunyl unit; 3.9 oxygen atoms at 2.40 Å to represent the closest oxygens in the coordinated carbonate; and, 4.9 carbon and 2.1 oxygen atoms at 3.10 and 3.40 Å, respectively, to account for the remaining atoms in the coordinated carbonate. The Np-O eq distance was longer than the Ca-O eq distance in calcite and was interpreted as Np(V) coordination by four monodentate carbonate ions. Furthermore, low Debye-Waller factors for the axial and equatorial oxygens indicated low structural disorder (Heberling 2008a). The authors concluded that NpO + 2 was stable in the calcite structure. They proposed substitution of Ca 2+ by Np(V) in the calcite structure, where the axial neptunyl oxygens substitute for two adjacent CO 3 2- ions. The authors suggested that the resultant charge imbalance (+3) could be addressed by coupled substitution of Na + for Ca 2+, or the creation of a Ca 2+ vacancy (Heberling et al., 2008a). A similar mechanism has been suggested for UO 2+ 2 incorporation into calcite (Kelly et al., 2003, Sturchio et al., 1998). In addition, neptunyl sorption to calcite was also studied by the same group and was found to reach a maximum at ph 8.3 and at low (< 10 µm) neptunyl concentrations. The ph dependence of sorption was much stronger than at higher concentrations (> 10 µm) (Heberling et. al. 2008b). Kinetic experiments at ph 8.3 showed that after hours, the adsorption slowed down, and, after 4 weeks the system was not at equilibrium. This was ascribed to the structural incorporation of Np(V) in to the calcite, while the solid undergoes gradual dissolution / recrystallization (Heberling et al., 2008b). There are several studies that examine Np(V) solubility in the literature, but, only a few of these studies have focused on hyperalkaline systems, with no work examining the role of Ca 2+. Additionally, there are only a few studies of the interaction on Np(V) with calcite 165

171 surfaces and, to the best of our knowledge, there are no studies that examine Np(V)-calcite interactions in hyperalkaline systems of relevance to cementitious geological disposal. This study examined Np(V) solubility and Np(V) sorption to calcite in high ph cementitious environments. Two synthetic cement leachates representative of the hyperalkaline conditions present in the "young" and "old" stage evolution of a geological disposal facility, termed Young Cement Leachate (YCL) and Old Cement Leachate (OCL), respectively, were studied. Here, we present data from solubility and sorption experiments over a range of Np(V) concentrations and solid to solution ratios and analyse these systems using X-ray Absorption Near Edge Structure (XANES) spectroscopy and geochemical modelling in order to gain a deeper understanding of Np(V) behaviour under these conditions. 2 Experimental 2.1 Materials A high purity chemically precipitated calcite powder (SureChem Products Ltd) was used throughout and was characterised using powder X-ray diffraction (Figure S1; Bruker D8 Advance). Prior to use, the calcite was passed through a 63 µm sieve. For all experiments, 237 Np(V) (purchased from AREVA) was prepared by chemical precipitation (Law et al., 2010). The oxidation state of the Np spike was confirmed as Np(V) immediately prior to dilution into 0.01 M HCl. The oxidation state was confirmed by using UV-vis (UV-1800 Shimadzu) to confirm the presence of peaks at 618 and 980 nm associated with Np(V) and the lack of peaks associated with Np(IV) at 723 and 960 nm. The chemicals used throughout this study were analytical reagent grade unless otherwise stated and all solutions were made using Milli-Q (18.2 M Ω cm) water. 166

172 2.2 Synthetic Cement Leachates In order to reflect the chemical conditions (specifically ph, [Ca 2+ ], [Na + ], and [K + ]) expected during the evolution of cement in contact with flowing water, two solutions were used in experiments: Old Cement Leachate (OCL); and Young Cement Leachate (YCL). Their compositions are given in Table 1. To avoid supersaturation of calcium containing phases, during preparation of the YCL solution, the solution was kept in a CO 2 free environment in an oven at 40 C for 48 hours prior to filtration (< 0.45 µm nylon). Selected experiments required calcite pre-equilibrated YCL and OCL solutions. Here, calcite (1: 100 solid solution) was allowed to pre-equilibrate with the solutions for 4 days before filtration (< 0.45 µm, nylon). Table 1. Synthetic cement leachate solution composition OCL YCL ph KOH (g L -1 ) NaOH (g L -1 ) Ca(OH) 2 (g L -1 ) Np concentration determination In low concentration experiments (< 1 Bq ml -1 ; 0.16 µm), total 237 Np concentrations were determined using Inductively Coupled Plasma Mass Spectrometry (ICP-MS) on an Agilent 7500cx instrument. The ICP-MS samples were prepared by dilution in 2% HNO 3 solution (AnalaR NORMAPUR). The data were calibrated using a six point calibration over the concentration range (0.1 ng L -1 to 100 ng L -1 ). All ICP-MS measurements were corrected for instrumental drift with a 10 µg L Th internal standard. For higher activity 237 Np analysis ( 1 Bq ml -1 ), concentrations were determined using Liquid Scintillation Counting (LSC) on a QUANTULUS instrument (PerkinElmer). LSC samples were routinely prepared by 167

173 combination of 1 ml supernatant with 5 ml scintillant (Scintisafe 3, Fisher Scientific) and 2 ml 1 M HCl. 2.4 Batch and solubility experiments All sample preparation and manipulation was carried out in a CO 2 free (< 1 ppm), aerobic environment (Coy basic polymer glovebox). Sorption experiments were spiked to a final Np(V) concentration between 1.62 x10-3 µm (0.01 Bq ml -1 ) and 1.62 µm (10 Bq ml -1 ) with a Np(V) in 0.01 M HCl stock solution. In the solubility experiments the final concentrations were either 4.22 µm (26 Bq ml -1 ) or 42.2 µm (260 Bq ml -1 ). All sorption and solubility experiments were carried out in triplicate. In some solubility experiments, the calcium concentration was adjusted by the addition of a CaCl 2 solution. 2.5 XAS Measurements In order to understand the speciation of Np in these experiments, an XAS spectrum was collected for a Np(V) precipitate from a calcite equilibrated YCL solubility experiment (42.2 µm). More generally, Np(V) reacted calcite samples were too dilute to allow for XAS analysis. The precipitate was collected by centrifugation and removal of the supernatant. The solid sample was mounted in a CO 2 free environment onto a Perspex sample holder and sealed with Kapton tape. The sample holder was then heat sealed in a CO 2 free argon environment and the sample stored at 4 C prior to analysis. XAS spectra were collected for the Np L III -edge in fluorescence mode on the INE Beamline for actinide research at the ANKA synchrotron light source (Germany), using a cryostat (OptistatDN, Oxford Instruments) and a 5 pixel solid-state detector (LEGe Canberra) detector, and a Ge(422) monochromator. Additional EXAFS spectrum for schoepite was collected for the U L III -edge on B18 at the Diamond Light Source using a 9 element solid-state Ge detector, focusing optics 168

174 (beam size 0.5 mm 2 ), and a Si-111 double crystal monochromator. Background processing was performed using Athena (Demeter; ; Ravel and Newville, 2005) and the data was calibrated against a Y standard. 2.6 Thermodynamic modelling All speciation, saturation and surface complexation thermodynamic calculations were performed using the United States Geological Survey thermodynamic speciation code PHREEQC Interactive (3.0.0) using the ANDRA SIT database (ThermoChimie v.7.d June 2011). 3 Results and Discussion 3.1 Np(V) Solubility A series of calcite equilibrated OCL and YCL 42.2 µm NpO + 2 solutions were allowed to equilibrate in CO 2 free conditions for 550 hours. Thermodynamic calculations using PHREEQC predicted that these systems (at 42.0 µm) should be supersaturated with respect to several Np(V) solid phases, specifically: Np 2 O 5 ; NpO 2 OH (am, aged) ; and NpO 2 OH (am, fresh) (Table 2). Subsequent analysis of the centrifuged supernatant (approximately < 0.22 µm) + from these experiments indicated that removal of NpO 2 from solution had taken place + (Figure 1). The equilibrium concentrations of NpO 2 were determined to be 9.68 ± 0.58 µm and ± µm for the OCL and YCL systems, respectively. A comparison of the experimental data with the predicted solubilities of these solid phases, suggested that NpO 2 OH (am, fresh) was precipitating in the OCL systems (Figure 1). At the low solubility observed in the calcite equilibrated YCL, the system was predicted to be significantly undersaturated with respect to NpO 2 OH (am) (fresh and aged) and it is unlikely that a Np 2 O 5 solid would form in the timeframe of this experiment (Figure 1). Further, the observed 169

175 solubility was significantly higher than that expected for equilibrium with Np 2 O 5 (Figure 1). Therefore, it seemed likely that a phase which was not in the ANDRA SIT database was forming. Table 2. PHREEQC calculated saturation indices for supersaturated Np(V) phases at 42.0 µm in YCL and OCL (Table 1) with calcite equilibration Phase YCL OCL Np 2 O NpO 2 OH (am, aged) NpO 2 OH (am, fresh) Figure 1. Solubility of Np 2 O 5 (red) and, aged (blue) and fresh (green) Np(V)O 2 (OH) (am) solid phases as a function of ph in NaOH systems (PHREEQC predictions). The points at ph 10.5 and 13.3 represent observed Np(V) solubility in the calcite equilibrated YCL and OCL systems. 170

176 Under similar hyperalkaline conditions, calcium has an influence on UO 2+ 2 solubility with U(VI) precipitating in calcium containing phases, such as becquerelite (Ca(UO 2 ) 6 O 4 (OH) 6 8(H 2 O)) and calcium uranate (CaUO 4 ) (Moroni and Glasser, 1995). Therefore, Np(V) solubility as a function of calcium concentrations was also studied in conditions relevant to the YCL system. Here, a series of carbonate free, ph 13.3 NaOH solutions were created with a range of Ca 2+ concentrations (2 x 10-2 M to 2 x 10-6 M) and either 42.2 or 4.22 µm NpO + 2. Analysis of the supernatant ( < 0.21 µm) after 24 hours using + ICP-MS indicated NpO 2 removal across all systems (Figure 2). The Np(V) solubilities at 2 x 10-2 M Ca 2+ after 24 hours were determined as 0.18 ± 0.02 µm and 1.09 ± 0.01 µm for the 4.22 and 42.2 µm systems, respectively. This was in contrast to calcium free experiments in which Np(V) concentrations were determined to be 3.49 ± 0.28 µm and 5.60 ± 0.24 µm for 42.2 and 4.22 µm systems, respectively. The relationship between calcium concentration and observed Np(V) solubility indicated that a calcium containing Np(V) solid was precipitating which was not predicted from thermodynamic modelling. A Log-Log plot of observed Np(V) solubility as a function of calcium concentration in the 42.2 and 4.22 µm systems is shown in Figure 2 A and B. These data demonstrate a clear linear relationship between Np(V) solubility and Ca 2+ concentration as expected if a neptunium-calcium phase was controlling solubility. Although the data indicate that Ca 2+ is involved in the phase(s) they cannot show how other species are involved, which could include O 2- /OH -, H 2 O, or Na +. However, the concentrations of these species will be essentially constant throughout all of the solubility experiments. We can postulate a formula for a phase that would form under these conditions: Ca x Np y O a OH b Na c.eh 2 O. The solubility of such a phase can be expressed in terms of the solubility product (1), where: K sp is the solubility product; A is a constant that includes the activity coefficients for Ca 2+, and NpO + 2, and also the activities of any other species involved in the precipitation of the phase (OH -, Na + etc) since these are constant in the 171

177 experiments (Figure 2); and x and y represent the stoichiometric quantities of Ca 2+ and Np(V), respectively. 2 x y Ksp A [Ca ] [NpO2 ] (1) This can then be rearranged in a linear form (2) where β is a constant (A/K sp ). x log[npo 2 ] log log[ Ca y y 2 ] (2) Therefore, the Ca 2+ : Np(V) ratio in the precipitate can be determined from the gradient of a plot of log [NpO + 2 ] and log [Ca + ] (Figure 2). For reasons of radiological safety (a limit on the activity that could be present in samples analysed using ICP-MS), the calcium concentration could not be determined experimentally. Therefore, an iterative modelling procedure was used to correct for Ca removal in the precipitate. An initial gradient was obtained from the data using the total Ca concentration. This gradient was used to calculate the amount of Ca removed from solution, and the data were replotted. A new gradient was then calculated which was used to recalculate the calcium concentration. This procedure was repeated until the Ca values used in the plot where consistent with the gradient. Because for most data points, the calcium was present in great excess compared to the Np, there was relatively little difference between the initial (dashed lines) and final (solid lines). The final gradients were calculated as ± 0.03 (R ) and ± 0.03 (R ) for the 4.22 µm and 42.2 µm data, respectively. These values suggest that there was an average 0.42 ± 0.03 and 0.23 ± 0.03 calcium atoms per neptunium atom in the 4.22 and 42.2 µm Np(V) systems, respectively. The significant decrease in the ratio of Ca 2+ : Np(V) ions between the two systems suggests that multiple phases may have precipitated. For example the NpO 2 (OH) (am, fresh) phase postulated in the OCL systems seems credible along with a calcium-neptunium containing phase in variable proportion. In high ph cement systems several calcium containing U(VI) minerals have been recorded in the literature, such as 172

178 becquerelite (Ca(UO 2 ) 6 O 4 (OH) 6 8(H 2 O); Moroni and Glasser, 1995) and calcium uranate (CaUO 4 ; Tits et al., 2011). The modelled Ca 2+ : Np(V) ratios in the calcite equilibrated YCL Np(V) precipitate were higher than the stoichiometry of becquerelite which is 0.16 calcium atoms per uranium atom. Therefore, it seems unlikely that the precipitate was an Np(V) analogue of becquerelite. However, calcium uranate has a 1:1 calcium to uranium ratio (Zachariasen, 1948) and variants of the structure contain lower ratios, such as calcium diuranate / calciouranoite (CaU 2 O 7.5H 2 O; (Rogova et al., 1974)). Therefore, it seems plausible that analogous calcium neptunate phase(s) which has not previously been identified may be controlling solubility in these systems. Figure 2. Observed Np(V) solubility after 24 hours as a function of calcium concentration in ph 13.3 NaOH systems with an initial Np(V) concentrations of 4.20 µm (A) and 42.0 µm (B). Triangles/dotted lines represent Np(V) solubility against total Ca concentration. Circles/solid lines represent Np(V) solubility plotted versus Ca concentration corrected for Ca precipitation. The horizontal line represents observed solubility in ph 13.3 NaOH solutions in the absence of added Ca

179 A Np(V) solid phase was recovered from a 42.2 µm calcite equilibrated YCL solubility experiment and analysed using XAS. The resultant XANES spectrum collected from the + sample is presented in Figure 3A with a Np(V)O 2 (aq) reference standard. The post-edge resonance feature (Figure 3) in the XAS spectra is attributable to the axial oxygens in the neptunyl and uranyl units (Denecke et al., 2005; Farges, 1992). The spectrum collected for the calcite equilibrated YCL Np(V) precipitate (Figure 3A) exhibits this feature, demonstrating that reduction of the Np(V) on the beamline had not occurred. However, the + feature appeared in the spectrum at a reduced energy in comparison to the Np(V)O 2 (aq) reference standard and had become less pronounced. The energy of this feature is inversely + proportional to the axial oxygen bond length (Denecke et al., 2005). The Np(V)O 2 (aq) species has typical Np=O ax bond length of 1.81 Å (Ikeda-Ohno et al., 2008). Therefore, this suggests there was a significant increase in the length of the Np=O ax bond in the experimental sample compared to the Np(V)O + 2 standard. Very few Np reference spectra are available, but there are many more U(VI) spectra that contain the uranyl unit that is analogous to neptunyl. Figure 3B also shows XANES reference spectra collected from two U(VI) mineral phases, schoepite (UO 2 ) 8 O 2 (OH) 12 12(H 2 O), and calcium uranate (Rossberg et al., 2014; Macé et al., 2013). Schoepite represents a typical UO 2+ 2 ion with a short U=O ax distance of 1.75 Å (Finch et al., 1996) and calcium uranate is an example of a uranate type structure with a longer U(VI)=O ax distance (1.96 Å; Zachariasen, 1948, Marshall et al., 2014). The increase in the U=O ax bond length between the schoepite and calcium uranate minerals is reflected in the XAS spectra (Figure 3B), in which the resonance feature reduces in energy (Farges et al., 1992). Although, the NpO + 2 and UO 2+ 2 units have different charges, similar changes in the XANES spectra for analogous changes in axial bond length would be expected. Unfortunately, the data were too low in quality to allow for EXAFS analysis. Given the elongated Np=O ax bond in the Np(V) precipitate recovered from the calcite equilibrated YCL 174

180 system, and the similarity to the XAS spectrum produced by a calcium uranate solid phase, the local coordination environment around the Np(V) atom appears similar to that of U(VI) in calcium uranate, suggesting that a calcium neptunate phase may have formed. Figure 3. XANES spectra recorded from (A): a Np(V) precipitate recovered from a 42.2 µm NpO + 2 calcite equilibrated YCL solubility experiment (red line) and reference spectrum of Np(V) in HNO 3 (blue line); (B) XANES reference spectra recorded from two UO 2+ 2 mineral phases, schoepite and calcium uranate (blue and red lines, respectively). The arrows indicate the O ax resonance feature in each spectrum. 3.2 Np(V) sorption to calcite The sorption of Np(V) from solution to calcite surfaces was investigated over a range of concentrations (1.62 x µm), solid to solution ratios (1 : 10, 1 : 50, 1 : 500), and in both OCL and YCL solutions using batch techniques. PHREEQC was used to explore Np(V) speciation in OCL and YCL (Table 1) systems equilibrated with calcite (with no atmospheric CO 2 ). These calculations suggested that in the OCL systems there was sufficient calcite dissolution to allow the NpO 2 (CO 3 ) - species to dominate the Np(V) speciation (Table 3). In addition to the neptunyl carbonate species, free NpO 2 + (aq) and the neutral NpO 2 (OH) (aq) 175

181 species essentially accounted for the remainder (38.1 and 6.05 %, respectively). In contrast, in the calcite equilibrated YCL system, the speciation was predicted to be dominated by hydroxide species with the NpO 2 (OH) 2 - (88.51 %) and the neutral NpO 2 (OH) (aq) (6.42 %) species dominating (Table 3). Interestingly, the predicted carbonate concentrations are sufficiently high for the formation of the neptunyl-carbonate-hydroxide species, NpO 2 (CO 3 ) 2 OH 4-, which accounts for a minor but significant proportion of the speciation (4.5 %). In this system, the free Np(V)O 2 + (aq) ion is predicted to account for only 0.09% of Np(V) aqueous speciation (Table 3). Thus, thermodynamic modelling suggests that in YCL, anionic forms of Np(V) will be dominant. This is likely to be highly unfavourable with respect to outer sphere complexation to the calcite surface, which will also carry a negative charge due to the high ph. By contrast, the predicted speciation in the calcite equilibrated OCL suggests a significant contribution of cationic aqueous species (NpO 2 + (aq) ) and thus outer sphere complexation is possible for the OCL system in addition to inner sphere. Table 3. PHREEQC calculated Np(V) speciation (42.0 µm) as percentage composition in YCL and OCL with calcite equilibration Species OCL YCL NpO 2 (CO 3 ) NpO NpO 2 (OH) (aq) NpO 2 (OH) NpO 2 (CO3) NpO 2 (CO 3 ) 2 OH In all experiments, Np(V) was removed from solution. In the YCL systems, removal was observed over the whole Np(V) concentration range studied (1.62 x µm) and all solid to solution ratios (1 : 10, 1 : 50, 1 : 500). Of the concentrations of study, only the lowest 176

182 concentration was considered undersaturated, which was confirmed in control experiments (no solid phase) with limited precipitation taking place at higher concentrations. At the lowest Np(V) concentration, 1.62 x 10-3 µm, the removal from solution showed a clear dependence on surface area with removal from solution at steady state of: 57 ± 2.1 % (1 : 10); 33 ± 1.1 % (1 : 50); and 15 ± 1.3 % (1 : 500). At the higher Np(V) concentration of 0.16 µm, removal at apparent equilibrium was: 93 ± 1 % (1 : 10); 86 ± 1 % (1 : 50); and 59 ± 1 % (1 : 500). Again, sorption depended on surface area and suggested that the surface plays a role in the removal of Np from solution. However, the percentage of Np(V) removed increased between the lowest and highest concentration systems. The formation of a Np(V)-calcite surface complex would be limited by the number of binding sites available for complexation, therefore the relative sorption of a solution species cannot increase with concentration where a surface complexation mechanism is controlling removal. The experimental data from the YCL calcite sorption experiments was inconsistent with the formation of a Np(V)-calcite surface complex alone. To explain the increase in removal with concentration, at least some precipitation had to take place in these systems. This was confirmed in control experiments with calcite equilibrated YCL solution (no solid), which showed a decrease in total Np solution concentration when the initial Np(V) concentration was 0.16 µm. These experiments indicated that the Np(V) concentration in solution had decreased by 60 ± 3.0 % and 84 ± 1.6 % with initial Np(V) concentrations of 0.16 and 1.62 µm, respectively which confirms the YCL was significantly oversaturated. The OCL systems again show significant removal of Np(V) from solution. At the lowest concentration (1.62 x 10-3 µm) sorption at steady state was measured as 94 ± 1.2 % (1 : 10), 87 ± 1.0 % (1 : 50) and 50 ± 3.0 % (1 : 500). Once again, removal increased with surface area, which suggested surface mediated processes were involved in removal. At the higher 177

183 concentration of 1.62 µm, Np(V) sorption at equilibrium was 92 ± 1.9 %, 69 ± 1.7 % and 15 ± 1.1 % in the 1:10, 1:50 and 1:500 systems, respectively. In contrast to the YCL systems, there was no increase in removal with increasing Np(V) concentrations. Such a result is consistent with a surface complexation mechanism dominating. Sorption data at apparent equilibrium for both YCL and OCL systems are shown in Figure 4. Both YCL and OCL systems show a clear linear relationship with gradients of 0.90 ± 0.12 and 1.69 ± 0.20 (L Kg -1 ) for the OCL and YCL systems, respectively. The gradient is indicative of the number of Np atoms per surface complex (for a surface complexation mechanism). Therefore in the OCL systems, the data are consistent with monodentate complexes with one Np atom per surface complex. This proposed formation of an innersphere complex is consistent with EXAFS spectroscopic data presented by Heberling et al. (2008b) which indicated a surface complexation mechanism was responsible for sorption in Np(V)-calcite systems at ph The same treatment of the YCL data suggests that would be 1.69 Np atoms per surface complex. This is inconsistent with a surface complexation mechanism. Therefore, it seems that precipitation processes dominate in the YCL systems. Given that the removal still depends on the mass of calcite and hence surface area, the surface must play some role in that precipitation, either providing nucleation centres for Np(V)-solid growth or by formation of a discrete Np phase bonded to the surface (Smith et al., In review). A surface precipitation mechanism would also be consistent with the increase in the relative amount of sorption with increasing Np(V) concentration observed in the batch experiments and also the gradient observed in the YCL data in Figure 4. Surface mediated precipitation reactions, including the formation of calcium uranate, have been reported for calcite surfaces with other actinides 178

184 such as U(VI) (Carroll et al., 1992; Elzinga et al., 2004; Geipel et al., 1997; Schindler and Putnis, 2004). Similar reactions could be taking palce with Np(V) in this system. Figure 4. Sorption isotherm from OCL and YCL batch sorption experiments. Lines represent linear fits to the data with gradients of 0.90 ± 0.12 and 1.69 ± 0.20 L Kg -1 for OCL and YCL systems, respectively. It was not possible to fit the YCL systems with a surface complexation approach as the data were inconsistent with the formation of a simple surface complex, however, a surface complexation model was created to explore the data collected from the OCL systems further. Surface speciation was calculated in PHREEQC using the constants outlined in Table 4 (van Cappellen et al., 1993). In the calcite equilibrated OCL solution, the modelling suggested that >Ca surface sites were dominated by the >CaCO - 3 (68.3 %) and >CaOH + 2 (30.3 %) surface species with a minor contribution from >CaOH (1.4 %). By contrast, the >CO 3 surface sites were dominated by >CO - 3 (60.4 %) and >CO 3 Ca + (39.6 %). The equilibrium data from the 1.62 x 10-3 µm and 1.62 µm Np(V) concentrations were fitted using this surface complexation model in order to determine the equilibrium constant for the surface complex. 179

185 In the OCL data, there was a reduction in Np(V) sorption between the lowest (1.62 x 10-3 µm) and highest (1.62 µm) Np(V) systems implying that binding site saturation was approached. Given the reduction in sorption between the lowest and highest concentration systems, in addition to the equilibrium constant, the surface binding site concentration was allowed to vary. It was assumed that only monodentate complexation to >CO 3 binding sites was taking place through the surface reaction (3) (Zavarin et al., 2005). CO 2 3 H NpO2 CO3NpO H (3) A fit of the experimental data yielded a stability constant (Log K) of 2.43 and a surface binding site concentration of 0.15 sites nm -1 (calcite surface area: 0.26 ± 0.01 m 2 g -1 ; BET derived). These parameters clearly reflect the experimental data in the 1.62 x 10-3 µm and 1.62 µm systems, and successfully predict sorption in the 1.62 x 10-2 µm and 0.16 µm systems (Figure 5). Table 4. Summary of calcite-solution surface reactions and constants used in surface complexation model. Reaction Log K >CO 3 H = >CO H >CO 3 H + Ca 2+ = >CO 3 Ca + + H >CaOH 2+ = >CaOH + H >CaOH = >CaO - + H >CaOH + CO 2 = >CaHCO 3 6 >CaOH + CO 2 = >CaCO H

186 Figure 5. Predicted Np(V) sorption at equilibrium against observed Np(V) removal from solution at apparent equilibrium in OCL systems. Circles show data that was used in calculation of model constants, triangles show points that were blind predictions of experimental data. The kinetics of Np(V) sorption to the calcite surface were also examined. In both YCL (Figure 6, Figure S2-4) and OCL (Figure 7) systems, several weeks were required for apparent equilibrium to be reached. Generally, there was an initial rapid increase in sorption within the first 24 hours, followed by further slow and gradual removal from solution. Slow removal kinetics are unusual for surface complexation to mineral surfaces for labile actinide ions such as NpO + 2, and removal is normally expected to occur within hours (Rihs et al. 2004; Davis et al. 1987; Giammar and Hering, 2001). However, Heberling et al. (2008b) observed slow Np(V) removal over 600 hours in their circumneutral ph experiments. The authors suggested this could be caused by a slow dissolution and reprecipitation process at the calcite surface. In addition, a recent study from our laboratories has observed slow removal kinetics over several weeks for U(VI) removal in hyperalkaline calcite systems 181

187 (Smith et. al, In review). Furthermore, the U(VI) in those YCL experiments was colloidal in nature. It is interesting to consider the impact that Np(V) colloids could have on the kinetics of Np removal from solution as Np(V) in a colloidal form would be unable to form a complex with the calcite surface. The slow kinetics observed in this study may involve the gradual dissolution/formation of a colloidal population to form Np(V) that is truly dissolved in solution which may then react with the calcite solid. Slow turnover of the calcite could result in slow incorporation of the Np, which would increase the apparent binding strength of the system and hence a reduction in the free Np concentration. Surface complexation would be expected to be a precursor to incorporation in such a mechanism. Figure µm Np(V) sorption to calcite solid phase in YCL solution with solid to solution ratios of 1 : 10 (circles), 1 : 50 (diamonds) and 1 : 500 (triangles) with an additional calcite free control (square). 182

188 Figure 7. Np(V) sorption to calcite in OCL with 1.62 x 10-3 (A), 1.62 x 10-2 (B), 0.16 (C) and 1.62 (D) µm Np(V) concentrations and solid to solution ratios of 1 : 500 (triangles), 1 : 50 (diamonds) and 1 : 10 (circles) with additional calcite free controls (squares). Lines are model calculations: A and B were used in the fittings while B and C were blind predictions. To explain the kinetics of Np(V) sorption in the OCL systems a coupled kinetic / thermodynamic model was developed, as illustrated in Figure 8, using a bespoke C/PHREEQC program. This model assumed that upon addition of Np(V) to the experiment it 183

189 was instantly fractionated into two components, Np(V) present as instantaneously available for complexation and surface reactions (Np avail ) and a second component that was initially unavailable for such reactions (Np unavail ), which could represent a colloid population. A kinetic reaction was used to represent slow transfer of Np unavail to Np avail (5), assuming that Np avail eventually accounted for all Np(V) and in line with experiments where equilibrium was reached after several weeks. d[np dt avail ] k[np ] (5) Where t is time and k is the rate constant. At the start of the calculation the Np avail was allowed to equilibrate between the solution and calcite surface with the binding site density and equilibrium constant fixed to that calculated for the >CO 3 NpO 2 surface species in the thermodynamic modelling. The solution and surface speciation was then calculated using PHREEQC. The rate constant, k, and the initial fractionation of Np(V) between Np avail and Np unavail was fitted to the experimental data from the 1.62 x 10-3 µm and 1.62 µm data sets (Figure 7). The best fit to these data was achieved with s -1 and 0.51 for the rate constant and initial fraction as Np unavail, respectively (Figure 7). These parameters were then used to predict sorption in the 1.62 x 10-2 µm and 0.16 µm Np(V) systems (Figure 7 B/C). These predictions are a reasonable match for the experimental data given that a single rate and surface complexation constants have been used. In the 0.16 µm system, the prediction was only moderately successful. The model was relatively good at predicting equilibrium and the gradient, however, it was poor at predicting the absolute remaining in solution towards the start of the experiment. In this case, the rate constant seems appropriate, and the problem lies in the initial fractionation of Np(V) into the available and unavailable fractions. The prediction for the 1.62 x 10-2 µm system was a good match for the kinetics and 1:10 and 1:50 equilibrium components of the model. The calculations did not include a term for the reverse unavail 184

190 component of the kinetic reaction (i.e. transfer of Np available to Np unavailable ). While such a reaction must exist, the data suggest that it does not occur in the experiments in Figure 7, and so it is not possible to obtain a value for the reverse rate constant. The value of K required to fit the data here (3.2 x 10-6 s -1 ) is very similar to equivalent rates of removal from solution observed for uranium in OCL and YCL systems (1.8 x 10-6 and 8.2 x 10-7 s -1, respectively (Smith et al., In review). In the case of uranium, there is some evidence of colloidal fractionation and the similarity in the rates of removal from solution may be significant. Figure 8. Schematic diagram of processes used to describe the kinetics of Np(V) removal in OCL systems. 185

191 4 Summary Np(V) solubility in calcite equilibrated YCL and OCL solutions was investigated using a combination of solubility experiments, thermodynamic calculations and XANES spectroscopy. Batch experiments indicated that after 550 hours, the concentrations of Np(V) were 9.68 µm and µm for the calcite equilibrated OCL and YCL systems, respectively. The former was consistent with thermodynamic equilibrium with NpO 2 (OH) (am, fresh) solid phase, however, the observed solubility in the calcite equilibrated YCL system was much lower than predicted, if NpO 2 (OH) (am, fresh) were the limiting phase. In carbonate free, ph 13.3 NaOH systems, there was a linear relationship between the observed Np(V) solubility and calcium concentration. These experiments indicated that the precipitate had an average of 0.42 ± 0.03 and 0.23 ± 0.03 Ca atoms per Np when the initial Np concentration was 4.22 or 42.2 µm Np(V), respectively. XANES data collected from the solid phase formed in the calcite equilibrated YCL system were consistent with an elongated Np=O ax bond length, and it is possible that a calcium neptunate-like phase is the solubility limiting phase at very high ph in equilibrium with calcite. Np(V) sorption to calcite in YCL and OCL solutions was studied using batch experiments. Where possible these data were described using kinetic and thermodynamic modelling approaches. Significant removal from solution was observed across all Np(V) concentrations (1.62 x µm) and solid to solution ratios (1 : 10, 1 : 50, and 1 : 500) in both the YCL and OCL systems. In both systems this removal increased with calcite surface area, suggesting that the surface plays a role in sorption. In the YCL systems the absorption isotherm was inconsistent with simple surface complexation, but could be explained if surface induced precipitation played a role. In the OCL systems, the absorption isotherm was consistent with the formation of a surface complex, and the data were successfully modelled 186

192 using a thermodynamic surface complexation approach. The model assumed the formation of a >CO 3 NpO 2 surface complex with a log K of 2.42 and a surface binding site density of 0.15 sites / nm 2. The kinetics of removal in both OCL and YCL systems was slow, taking several weeks to reach equilibrium. The kinetic behaviour was modelled assuming that, upon addition to the hyperalkaline leachates, the Np(V) fractionates into two components, one of which was available for solution and surface complexation instantaneously and the remainder that was unavailable. A first order kinetic reaction simulated the transfer from the unavailable fraction, and hence the slow sorption observed in the systems: the first order rate constant was fitted as s -1, while the amount initially unavailable was fitted at 51%. The kinetic modelling used the same site densities, surface speciation and equilibrium constant as the thermodynamic modelling. This model was moderately successfully in predicting the kinetics of Np(V) sorption across a range of solid : solution ratios and Np(V) concentrations. Such behaviour would be consistent with a colloidal population. U(VI) colloids have been found in other high ph systems (Smith et. al, in review; Bots et. al, Submitted) and there are some similarities between the rates of complexation observed for U (Smith et. al, 2014; In review) and these for Np data. Overall, this study has examined Np(V) mobility in several hyperalkaline systems of direct relevance to cementitious geological disposal. In the young GDF, our data show that Np(V) solubility is lower than predicted by the precipitation of a simple Np(V) hydroxide phase and an unidentified calcium containing phase must be precipitating. This suggests that Np(V) solubility in a cementitious GDF will be lower than predicted by the current thermodynamic databases and highlights the lack of data available for neptunium solid phases. In contact with 187

193 the calcite surface, Np(V) removal appeared to be at least partially by precipitation, which is perhaps mediated by the calcite surface. The data show that in the old GDF, the primary Np(V) removal mechanism to calcite surfaces would be surface complexation. 5. Acknowledgements Kurt Smith is a Nuclear FiRST DTC student, funded by EPSRC (EP/G037140/1). This work was carried out as part of the Biogeochemical Gradients and Radionuclide Transport (BIGRAD) consortium, funded by NERC (NE/H007768/1). The authors would like to thank Paul Lythgoe for geochemical analyses. We thank the INE beamline, ANKA synchrotron radiation source, Karlsruhe Institute of Technology (Actinet Awards C3-15 and C5-06 ) and Diamond Light Source for access to the B18 beamline (SP7593) that contributed to the results presented here. 6 References Braney, M.C., Haworth, A., Jefferies, N.L., Smith, A.C., A study of the effects of an alkaline plume from a cementitious repository on geological materials. J. Contam. Hydrol. 13, Bots, P., Morris, K., Rosemary, H., Law, G., Mosselmans, F., Brown, A., Doutch, J., Smith, A., Shaw, S., Formation of stable uranium(vi) colloidal nanoparticles in conditions relevant to radioactive waste disposal, Langmuir. Submitted July Carroll, S.A., Bruno, J., Petit, J.C., Dran, J.C., Interactions of U(VI), Nd and Th(VI) at the calcite-solution interface. Radiochim. Acta, 58-9,

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195 Giammar, D.E., Hering, J.G., Time scales for sorption-desorption and surface precipitation of uranyl on goethite. Environ. Sci. & Technol. 35, Hay, M.B., Workman, R.K., Manne, S., Mechanisms of metal ion sorption on calcite: composition mapping by lateral force microscopy. Langmuir, 19, Heberling, F., Denecke, M.A., Bosbach, D., 2008a. Neptunium(V) Coprecipitation with Calcite. Environ. Sci. & Technol. 42, Heberling, F., Brendebach, B., Bosbach, D., 2008b. Neptunium(V) adsorption to calcite. J. Contam. Hydrol. 102, Ikeda-Ohno, A., Hennig, C., Rossberg, A., Funke, H., Scheinost, A.C., Bernhard, G., Yaita, T., Electrochemical and complexation behavior of neptunium in aqueous perchlorate and nitrate solutions. Inorg. Chem. 47, Kaszuba, J.P., Runde, W.H., The Aqueous Geochemistry of Neptunium: Dynamic Control of Soluble Concentrations with Applications to Nuclear Waste Disposal. Environ. Sci. & Technol. 33, Kelly, S.D., Newville, M.G., Cheng, L., Kemner, K.M., Sutton, S.R., Fenter, P., Sturchio, N.C., Spötl, C., Uranyl incorporation in natural calcite. Environ. Sci. & Technol. 37, Marshall, T.A., Morris, K., Law, G.T.W., Livens, F.R., Mosselmans, J.F.W., Bots, P., Shaw, S., Incorporation of Uranium into Hematite during Crystallization from Ferrihydrite. Environ. Sci. & Technol. 48,

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197 Rogova, V.P., Belova, L.N., Kiziyarov, G.N., Kuznetsova, N.N., Bauranoite and metacalciouranoite, new minerals of the hydrous uranium oxides group. Int. Geol. Rev. 16, Rossberg A., Scheinost A.C., Schmeisser N., Rothe J., Kaden P., Schild D., Wiss T., Daehn R "AcReDaS, an Actinide Reference Database for XAS, EELS, IR, Raman and NMR Spectroscopy", Schindler, M., Putnis, A., Crystal growth of schopiete on the (104) surface of calcite. Can. Mineral. 42, Schmidt, P., Small-angle scattering studies of disordered, porous and fractal systems. J. Appl. Cryst. 24, Schwyn, B., Wersin, P., Rüedi, J., Schneider, J., Altmann, S., Missana, T., Noseck, U., FUNMIG Integrated Project results and conclusions from a safety case perspective. Appl. Geochem. 27, Smith, K., Bryan, D., Swinburne, A., Pieter, B., Shaw, S., Natrajan, L., Mosselmans, F., Morris, K., 2014, U(VI) behaviour in hyperalkaline calcite systems. Geochim. Cosmochim. Acta. In review. Tits, J., Wieland, E., Bradbury, M.H., The effect of isosaccharinic acid and gluconic acid on the retention of Eu(III), Am(III) and Th(IV) by calcite. Appl. Geochem. 20, van Cappellen, P., Charlet, L., Stumm, W., Wersin, P., A surface complexation model of the carbonate mineral-aqueous solution interface. Geochim. Cosmochim. Acta. 55,

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199 Np(V) sorption and solubility in calcite containing high ph systems Supporting Information Kurt F. Smith a, Nick D. Bryan b, Gareth Law b, and Katherine Morris a a Research Centre for Radwaste and Decommissioning and Williamson Research Centre, School of Earth, Atmospheric and Environmental Sciences, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom. b Centre for Radiochemistry Research, School of Chemistry, The University of Manchester, Oxford Road, Manchester, M13 9PL, United Kingdom 194

200 Figure S1. Powder X-ray diffraction pattern collected from calcite used throughout this study Figure S x 10-3 µm Np(V) sorption to calcite solid phase in YCL solution with solid to solution ratios of 1 : 10 (circles), 1 : 50 (diamonds) and 1 : 500 (triangles) with an additional calcite free control (square). 195

201 Figure S x 10-2 µm Np(V) sorption to calcite solid phase in YCL solution with solid to solution ratios of 1 : 10 (circles), 1 : 50 (diamonds) and 1 : 500 (triangles) with an additional calcite free control (square). Figure S µm Np(V) sorption to calcite solid phase in YCL solution with solid to solution ratios of 1 : 10 (circles), 1 : 50 (diamonds) and 1 : 500 (triangles) with an additional calcite free control (square). 196

202 The following is a research manuscript that has been published in Mineralogical Magazine, (8) p Chapter 5 197

203 Mineralogical Magazine, December 2012, Vol. 76(8), pp Europium interaction with a vault backfill at high ph R. TELCHADDER, K. SMITH AND N. D. BRYAN* Centre for Radiochemistry Research, School of Chemistry, University of Manchester, Oxford Road, Manchester M13 9PL, UK [Received 20 December 2011; Accepted 26 April 2012; Associate Editor: Nicholas Evans] ABSTRACT Batch experiments have been used to assess the sorption properties of a potential cementitious repository backfill, NRVB (Nirex reference vault backfill), using Eu 3+ as a model trivalent radionuclide and ethylenediaminetetraacetic acid (EDTA) as a competing ligand. The NRVB is an effective scavenger of Eu from solution, with most sorbed within minutes onto the crushed material and less than 1.5% remaining in solution after one day (R d values in the range lkg 1 ). Ultrafiltration showed that nearly all of this remaining Eu (>94%) is attached to NRVB derived colloids or particulates that are mainly retained by a 100 kda ultrafilter. High concentrations of EDTA (>0.01 M) reduced the extent of sorption at apparent equilibrium. The addition of EDTA to a pre-equilibrated system of Eu 3+ andnrvbresultedinatemporarysuspensionofsomeeu,butthisveryquickly returned to the solid phase. There is some irreversibility in these systems, with EDTA able to prevent removal of Eu(III) from solution, but unable to bring it back into solution under the same conditions. KEYWORDS: europium speciation, geological disposal, NRVB, modelling vault backfill. Introduction GEOLOGICAL disposal is UK Government policy for the long-term management of higher activity radioactive wastes (Department for Environment Fisheries and Rural Affairs et al., 2008). The UK currently has no site identified for a geological disposal facility (GDF) and is considering a number of illustrative disposal concepts (Nuclear Decommissioning Authority, 2010). For intermediate-level and some low-level wastes, these include multi-barrier concepts, where waste packages in underground vaults will be backfilled with a cementitious material as part of the engineered barrier system (EBS). One potential backfill is NRVB (Nirex reference vault backfill; Nuclear Decommissioning Authority, 2010; McCarter et al., 2004). Degradation of the backfill will generate a region of high alkalinity (Braney, 1993; Nuclear Decommissioning Authority, 2010) in and around a GDF, which will promote metal * nick.bryan@manchester.ac.uk DOI: /minmag ion hydrolysis and sorption. The NRVB has been designed to provide these properties. It is based on ordinary Portland cement (OPC) with aggregate material containing crushed limestone and hydrated lime. The composition of NRVB (weight fraction) is OPC 26%; limestone 29%; hydrated lime 10 %; water 35%. The backfill, when cured, has a porosity of 0.5, a density of 1730 kg m 3 and a compressive strength of 6 MPa (Francis et al., 1997). The aim of this work is to study the sorption of Eu(III) by the NRVB. A number of authors have studied the interactions of trivalent lanthanide and actinide ions with cements (e.g. Dario et al., 2006), concretes (e.g. Baston et al., 1995), calcium silicate hydrates (C-S-H) (e.g. Pointeau et al., 2001) and calcite (e.g. Zavarin et al., 2005). Pointeau et al. (2004) have suggested that C-S-H is the most important phase in cements (although they studied uranyl ions). For Nd(III), the sorption seems to be dominated by C-S-H, and portlandite and silica interactions are expected to be much less important if they are present (Mandaliev et al., 2010). Pointeau et al. (2001) used lumines- # 2012 The Mineralogical Society

204 R. TELCHADDER ET AL. cence spectroscopy to study Eu interactions with C-S-H. Sorption was very strong (R d > l kg 1 ) for all values of calcium to silicon ratio, C/S, in the C-S-H ( ). For higher C/S systems ( ), two distinct types of Eu were found, which were assigned as surface complex and incorporated within the structure. The spectra were similar to that of Eu sorbed to tobermorite. Although portlandite was present in the C/S = 1.65 sample, there was no evidence for Eu sorption to it. Later, Tits et al. (2003) also used luminescence to identify two types of Cm(III) interactions with C-S-H, one where the Cm retains 1.4 coordinated waters and another with none. Both types are assumed to represent Cm incorporated in a tobermorite-like structure at the Ca sites: the first type is consistent with a labile 7-coordinate Ca site with 2 waters and/or a site with coordination by six structural oxygen atoms and one water; the second type is consistent with a 7-coordinate Ca site, where all coordinating oxygen atoms are structural. With time, transfer of Cm was observed from the type 1 to type 2 sites. Tits et al. (2003) reinterpreted the data of Pointeau et al. (2001), and showed that the Eu data indicated one site with two waters and another with none, and so Cm and Eu behaviour are very similar. Schlegel et al. (2004) studied Eu interactions with C-S-H as a function of C/S. For all samples, removal from solution is rapid, and the net result is incorporation of Eu within a C-S-H structure (at Ca sites). For the lower C/S systems ( ), the process seems to be initial surface complexation, followed by diffusion into the structure. However, batch data suggested that at high C/S (1.3), the Eu is at least partly removed from solution by co-precipitation. Despite the different mechanisms, the final coordination environment was the same. Mandaliev et al. (2010) studied Nd(III) interactions with C-S-H using EXAFS. Changes in Nd Ca/Si bond lengths are consistent with initial inner sphere complexation at the surface, followed by progressive incorporation of Nd at Ca sites within the C-S-H structure. There is some incorporation after 1 day, and there is a significant increase over the following 45 days. Sorption of Eu(III) to calcite is also strong (R d in the range to 10 5 lkg 1,) and rapid at high ph (13.3) (Tits et al., 2005). There is spectroscopic evidence that trivalent radionuclides [Cm(III)] form inner sphere surface complexes on contact with calcite, but that incorporation within the structure takes place with time (Stumpf and Fanghanel, 2002). For Eu in particular, there is evidence that incorporation is possible (Curti et al., 2005). Although there are a number of studies in the literature on europium interactions with calcite, cements and hydrated calcium silicates, there are none describing the interactions with NRVB. Therefore, the primary aim of this work is to study the interactions of Eu 3+ (as a surrogate trivalent actinide) with NRVB and to test whether any Eu associated particulate material is produced in solution. The effect of ethylenediaminetetraacetic acid (EDTA) on sorption is also studied, because as a strongly complexing ligand, it allows us to assess the strength of the NRVB as a sorber and also to test sorption reversibility, which is rarely studied in the literature. Some radioactive wastes may have EDTA present due to its use as a decontamination agent at nuclear facilities (e.g. Toste et al., 1995) and organic ligands can affect the transport of radionuclides by competing with sorption and hydrolysis (Keith-Roach, 2008). Ethylenediaminetetraacetic acid has the ability to chelate trivalent and tetravalent actinides and is known to be persistent in the environment (Means et al., 1980). However, the amounts that may be present in UK wastes disposed to a GDF are relatively small compared to the amounts of cellulosic materials that may degrade to form complexants (estimated 2005 UK inventory, 850 kg; Heath and Williams, 2005). Experiments were performed in NRVB water, which is water that has been allowed to equilibrate with NRVB. The solutions are in equilibrium with portlandite and represent the likely background solution in the engineered barrier system during its evolution. The effect of added NaClO 4 has been studied to determine the effect of changes in the background electrolyte, which might take place during the life of a GDF. Experimental For all experiments, a controlled atmosphere Perspex glove box filled with N 2 was used to exclude CO 2 and prevent carbonation (pco 2 <5 ppm). Before it entered the box, the N 2 gas (supplied as CO 2 free) passed through two Dreschel bottles, with sintered frits, full of saturated NaOH solution to remove any remaining traces of CO 2, followed by a further water filled Dreschel bottle to remove any NaOH aerosols. Even though the samples were kept in the glovebox, they were kept in sealed, gas tight containers, apart from the short time during sampling. 3084

205 SORPTION PROPERTIES OF NRVB The three components of NRVB (ordinary Portland cement, limestone flour and hydrated lime) were mixed in the ratio 265:291:100 for 20 min. Water was then added in a ratio of 1.8:1 (solid:water). The mixture was stirred for 3 h, and then left to set and cure in the glove box for 28 days. The cured cement samples were crushed using a tungsten carbide grinder and then sieved. Only the mm fraction was used in the experiments described here. The Ca and Si contents of the OPC (as SiO 2 and CaO) are 19.5 and 64.4%, respectively. Taking into account the limestone and hydrated lime, the C/S ratio is 8.5. Batch experiments were prepared in polypropylene tubes (Fisher Scientific). The 152 Eu radiotracer was added to each system, so that the total Eu concentration (stable plus radiotracer), [Eu 3+ T ], was M; NRVB was added to each system in multiples of 0.05 g. Experiments were performed with and without NaClO 4 added to the NRVB water. In some systems, EDTA was added to give final, total concentrations of 0.1, 0.01 or M. Where EDTA was added, the final ph was the same as the equivalent experiment in the absence of EDTA to allow comparison. The total solution volume was 10 ml. The solutions were sampled following centrifugation (3000 rpm; 5 min). Based on the density of the NRVB, the rotational velocity, sampling depth, rotor arm length and spinning time, assuming spherical particle shapes, the centrifugation is approximately equivalent to filtration with a 0.5 mm pore size filter. The Eu concentrations were obtained using a Ge semiconductor g-ray detector ( kev g-ray line); EDTA concentrations were measured using UV visible spectroscopy. The UV Vis spectra were recorded on a double beam Cary Varian 500 scan UV vis nir spectrophotometer (range nm, typical scan rate 600 nm min 1 ). The ph measurements were made with a Sentex combination ph electrode and a Seven Easy, Mettler Toledo meter. The electrode was calibrated with ph 7 and 9 buffer capsules (Whatman), and its response tested against a ph 13 buffer solution. Following the approach of Pointeau et al. (2004), uncertainties were estimated by repeating a representative system five times. In one experiment, the solution obtained following centrifugation was analysed by ultrafiltration using a Millipore device attached to a nitrogen cylinder, according to the technique of Pitois et al. (2008). The membranes had pore sizes of 100, 10 and 3 kda (approximately equivalent to 3.1, 1.4 and 0.9 nm, respectively). Before use, the cell and the membranes were washed repeatedly with deionized water and pretreated once with 10 4 M Eu(NO 3 ) 3 solution to saturate the membrane sorption sites: a previous study (Pitois et al., 2008) had shown this was necessary to prevent sorption of 152 Eu. Note: the solutions were not routinely ultrafiltered. Speciation calculations were performed using the United States Geological Survey thermodynamic speciation code PHREEQC Interactive (version , released September 2010) and using the Specific Interaction Theory Database ThermoChimie v.7.b, developed by Amphos 21, BRGM and HydrAsa for ANDRA. The thermodynamic datum for Eu 3+ and EDTA interaction was obtained from Martell and Smith (1974) and corrected to zero ionic strength using the Davies equation. Results Figure 1 shows the percentage of Eu remaining in solution after centrifugation (in absence of NaClO 4 ) plotted vs. the mass of NRVB after 24 h equilibration time. Time series experiments (data not shown) showed that apparent equilibrium was achieved after only a few hours. Only small quantities (<1.5%) of Eu could be detected in solution after 24 h. Control experiments and speciation calculations confirmed that, at the concentrations reported in this work, Eu 3+ is not subject to precipitation as a simple hydroxide. The ph of the NRVB water in contact with and 0.02 g ml 1 NRVB was 12.4Ô0.1 and 12.5Ô0.1, respectively. Europium was equilibrated with NRVB (0.005 g ml 1 ). The supernatant was then ultrafiltered through 100 kda, 10 kda and 3 kda membranes to determine Eu size fractionation (Fig. 2). Ninety-four percent of all Eu in the solution phase was present as clusters or colloidal material (i.e. it appeared in the >3 kda fractions) and 6% was present as free ion or simple complexes in solution. In fact, most Eu (92%) was in the largest (>100 kda) fraction. These colloids/large species must be NRVB derived, because in the absence of NRVB, but at the same ph, the Eu is in the <3 kda fraction. Therefore, the sorption data suggest that the Eu is largely NRVB associated in these systems, either bound directly to the bulk NRVB or to NRVB derived colloids/particulates, and very little of the solution phase loading can be thought of as true solution (e.g. Eu(OH) 3(aq), Eu(OH) 4 (aq)). 3085

206 R. TELCHADDER ET AL. FIG. 1. Plot of Eu 3+ remaining in solution relative to NRVB mass; equilibrated for 24 h; no added NaClO 4 ; solution volume = 10 ml; [Eu T ] = M. Figure 3 shows the the effect of added NaClO 4 (0.1 M) oneu 3+ sorption. For the g ml 1 system, more Eu (approximately 4%) remains in solution compared to a system with no added NaClO 4 (compare the data in Figs 1 and 3, final time point), but for the 0.01 and g ml 1 systems, the increase is modest (<1%), and for the 0.02 g ml 1 system, there is no significant difference. As for NRVB water systems, the process is NRVB mass dependent, and a g ml 1 NRVB system has 6% Eu remaining in solution, compared to approximately 2% for g ml 1 systems and 0.2% for a 0.02 gml 1 system. For solid:solution ratios greater than 0.02 g ml 1, no Eu remained in solution at equilibrium. The ph of the solutions in contact with and 0.02 g ml 1 NRVB was 13.1Ô0.1 and 12.9Ô0.1, respectively. No sorption of EDTA to NRVB was detected at any of the solid:solution ratios studied here. Figure 4 shows the change in solution phase Eu concentration with time as a function of EDTA concentration (with added NaClO 4, 0.1 M) at a solid:solution ratio of 0.03 g ml 1 NRVB, which was chosen because Eu could not be detected in solution at this solid:solution ratio in the absence of EDTA, and so the effect of EDTA could easily be detected. In this experiment, the Eu and EDTA were allowed to equilibrate first before the introduction of the solid NRVB. The amount of Eu remaining in solution increases with EDTA concentration, with up to 30% for 0.1 M EDTA, whereas 0.01 M allows approximately 20% in the solution phase (Fig. 4). A concentration of M EDTA is too small to suspend much Eu indefinitely, however, it does increase the time FIG. 2. Pie chart showing Eu size fractionation in NRVB water determined by ultrafiltration (no added NaClO 4 ). 3086