MineralâWater Interface Reactions of Actinides - ACS Publications

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Mineral−Water Interface Reactions of Actinides Horst Geckeis,* Johannes Lützenkirchen, Robert Polly, Thomas Rabung, and Moritz Schmidt Karlsruhe Institute of Technology (KIT), Institute for Nuclear Waste Disposal (INE), Karlsruhe, P.O.Box 3640, D-76021 Karlsruhe, Germany 3.3.7. Sorption of Uranyl at the Corundum (001) Surface 3.4. Sorption of Trivalent Curium 3.4.1. Sorption of Trivalent Curium Ions at the Corundum (001) Surface 3.4.2. Sorption of Trivalent Curium Ions at the Corundum (110) Surface 3.5. Classical Molecular Dynamics (MD) and Classical Monte Carlo (MC) Simulations 3.5.1. Sorption of Uranyl Carbonate in Feldspar Nanosized Fractures 3.5.2. Sorption of Uranyl on Quartz 3.5.3. Sorption of Uranyl on Goethite 3.5.4. Sorption of Uranyl on Calcite 3.5.5. Sorption of Uranyl on 2:1 Dioctahedral Clays 4. Geochemical Modeling of Actinide Sorption 4.1. Approaches to Ion-Exchange Modeling 4.2. Approaches to Surface Complexation Modeling 4.2.1. Treatment of Heterogeneity and Definition of Surface Sites 4.2.2. Protonation Mechanism and Electrostatics 4.2.3. Actinide Adsorption Model 4.2.4. Input Data for Model Development and Parameter Estimation 4.2.5. Generalization of Results 4.3. Approaches to Modeling Incorporation of Actinides in Matrices 5. Summary Author Information Corresponding Author Notes Biographies Acknowledgments References

CONTENTS 1. Introduction 2. Actinide Sorption Mechanisms 2.1. Outer-Sphere Sorption 2.1.1. Trivalent Actinide Ions 2.1.2. Tetravalent Actinide Ions 2.1.3. Pentavalent Actinide Ions 2.1.4. Hexavalent Actinide Ions 2.2. Inner-Sphere Surface Complexation 2.2.1. Trivalent Actinide Ions 2.2.2. Tetravalent Actinide Ions 2.2.3. Pentavalent Actinide Ions 2.2.4. Hexavalent Actinide Ions 2.2.5. Competitive Sorption 2.3. Actinide Incorporation Processes 2.3.1. Trivalent Actinide Ions 2.3.2. Tetravalent Actinide Ions 2.3.3. Penta/Hexavalent Actinide Ions 2.4. Surface-Induced Actinide Redox Reactions 2.5. Sorption of Intrinsic Actinide Colloids 3. Theoretical Studies of Actinide Sorption 3.1. Background of Actinide Quantum Chemistry 3.2. Quantum Chemical Approaches to Describe Sorption on Mineral Surfaces 3.2.1. Plane Wave DFT Employing Periodic Boundary Conditions (PBC) 3.2.2. Orbital-Based DFT Employing Finite Cluster Models 3.3. UO22+ Sorption 3.3.1. Sorption of Uranyl at Rutile (110) Surface 3.3.2. Sorption of Uranyl at Gibbsite (001) and Edge Faces of Gibbsite 3.3.3. Sorption of the Uranyl Ion onto the Nickel (111) Surface 3.3.4. Sorption of the Uranyl Ion at the Kaolinite (001) Surface 3.3.5. Sorption of the Uranyl Ion onto the Kaolinite (010) Edge Surface 3.3.6. Sorption of the Uranyl Ion onto Goethite © 2013 American Chemical Society

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1. INTRODUCTION Reactions of the light actinide ions (Th to Cm) at the mineral− aqueous solution interface have attracted much attention during recent years. Initially research was mainly driven by the need to obtain solid/liquid distribution constants (KD-values) as input data for performance assessment modeling related to nuclear waste repository projects or to the development of remediation

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dx.doi.org/10.1021/cr300370h | Chem. Rev. 2013, 113, 1016−1062

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Figure 1. Actinide sorption reaction mechanisms.

actinide ions to a specific binding mode can be made depending on the cation distance to the mineral surface. In contrast to bulk spectroscopic methods, XPS, surface scattering techniques and grazing-incidence EXAFS (GIEXAFS) provide structural information specific to the interface. GI-EXAFS operates in a total-reflection mode, where the incoming X-ray beam impinges on the investigated (planar) surface under angles generally near 100 mrad. Taking advantage of the linear polarization of synchrotron radiation and anisotropy of a given system (either the adsorbate, adsorbent or both), this technique allows to examine the system’s spatial orientation, e.g. of adsorbed dipolar actinyl cations such as UO22+ on the surface and thus to identify individual surface sorption sites.20,21 Most methods mentioned above can be applied to hydrated surfaces and thus are considered in situ techniques. XPS, while being surface sensitive, has the disadvantage that measurements are performed in vacuum, which possibly alters the sample. The implementation of quantum chemical approaches supplements spectroscopic tools notably in those cases, where experimental insight cannot or can only insufficiently be achieved4 (see section 3). Sorption reactions may involve pure electrostatic attraction of actinide ions or polymeric actinide species to surfaces with no or little impact on the first metal ion hydration shell. Innersphere surface complexation involves chemical binding of aqueous species and direct contact with mineral surface functional groups. However, depending on geochemical conditions and the type of mineral, reactions other than mere sorption, such as surface induced redox reactions and actinide incorporation into the solid matrix, are possible (see Figure 1). Master parameters controlling sorption mechanisms and determining the extent of sorption are pH, redox conditions, total actinide ion concentration, and the type of mineral surface functional groups, as well as the dynamic properties of the mineral in contact with the aqueous solution. Further parameters, such as salinity or ionic strength, the presence of competing cations and actinide ligating anions exert further influence on surface sorption reactions and actinide surface speciation.

technologies for contaminated sites. Very soon it became apparent that profound understanding of reaction mechanisms is required to develop appropriate geochemical sorption models for a reliable description and prediction of actinide environmental behavior. Sorption reaction mechanisms and actinide surface speciation nowadays can be investigated in great detail with increasing availability of spectroscopic techniques, such as EXAFS (extended X-ray absorption fine structure), TRLFS (time-resolved laser fluorescence spectroscopy), Raman or infrared spectroscopy, XPS (X-ray photoelectron spectroscopy), and X-ray reflectivity techniques (see, e.g., ref 1). EXAFS certainly belongs to the most universal spectroscopic techniques applied to actinide surface speciation studies.2−4 Photoelectron waves excited from an atom by incident X-rays interfere with waves scattered on neighboring atoms and thus create a fine structure in the energy dependent absorption coefficient. Analysis of EXAFS oscillations provides metrical parameters describing short-range order surrounding the absorbing atom, in this case the molecular structure of the surface sorbed actinide species. TRLFS takes advantage of characteristic luminescence properties arising from f−f electron transitions of some actinide ions mainly Cm(III) and UO22+ and represents an efficient tool for geochemical speciation studies. In particular the extremely high sensitivity of Cm(III) TRLFS allows for molecular-level surface speciation at trace concentrations well below one monolayer.5−10 In brief, TRLFS can identify and characterize Cm(III) species in a given system by the bathochromic shift and splitting pattern of fluorescence signals induced by variable ligand fields. Fluorescence quenching in natural systems is controlled by energy transfer to high-energy vibrational modes of H2O/OH−. This can be used to determine the hydration state of Cm(III) by measuring the fluorescence lifetime, which is empirically correlated with the number of quenching entities in the first Cm(III) coordination sphere.8,11−13 Similarly UO22+ TRLFS can be applied for surface speciation. However photophysics in this case is more complex and spectroscopic data, therefore, more challenging to interpret.4,14−18 X-ray reflectivity techniques have been applied to reconstruct the arrangement of molecules and ions at the solid/liquid interface.19 Assignment of adsorbed 1017


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To quantify actinide sorption, batch sorption experiments are usually carried out by varying the chemical conditions. Sorption isotherms are determined by measuring the solid−liquid distribution of an actinide ion at variable metal ion concentration but at fixed pH. The pH dependent actinide sorption is studied at constant metal concentration and provides the so-called “sorption edges”.22−24 Both sorption isotherms and sorption edges provide imperative input data for development of mechanistic geochemical/thermodynamic sorption models, able to account for variable geochemical conditions. Combination of results with complementary spectroscopic and theoretical methods allows for elucidation and verification of sorption mechanisms and actinide surface species. Even though this approach is elaborate, it is much more appropriate for predicting contaminant mobility under natural conditions than using a single distribution coefficient (KD value) determined for a specific geochemical environment (see also discussion in section 4). Such KD values are, in general, only applicable at trace concentration ranges, for reversible sorption reactions and at invariable geochemical conditions. While single distribution coefficients are still used in migration calculations, the development of accurate mechanistic models is obviously reaching limits when the complexity of real-world surfaces becomes involved. The complexity of the light actinide aquatic chemistry has a strong impact on sorption. Under naturally relevant conditions actinide ions can exist in various redox states (An(III), An(IV), An(V), An(VI)). The actinide redox state and speciation of the actinides change due to surface reactions and can form a broad variety of complexes with groundwater constituents, mainly OH− and CO32−. While the tri- and tetravalent actinide species form more or less spherical aquo-ions with 9−10 water molecules in the first coordination sphere, penta- and hexavalent cations form actinyl cations with covalently bound axial oxygen atoms and 4- to 6-fold coordination in the equatorial plane (e.g., ref 25). Steric constraints are thus to be expected for the interaction of actinyl cations with surfaces. Notably the tetravalent actinide but also the hexavalent actinyl cations tend to form oligomeric or even polymeric species which may interact with surfaces by either electrostatic forces or chemical bonds. This chemical complexity has to date prohibited development of any unified model for dealing with actinide sorption under varying geochemical conditions. A further inherent difficulty is related to the unique properties of different mineral surfaces. Even actinide sorption onto pure and well-defined crystal planes is not entirely understood, demonstrating the need for further study. A recent review by Tan et al.4 summarizes available studies on actinide/lanthanide interaction with mineral surfaces, including detailed descriptions of spectroscopic methods. Tan et al. discuss examples of actinide/lanthanide interface reactions in order to elaborate mechanistic insight provided by individual spectroscopic techniques, such as TRLFS and EXAFS. Another review on actinide speciation is ref 26, providing a nice overview of actinide reactions in environmental compartments including complexation with major relevant natural ligands abundant in groundwater and naturally occurring biological (microbial) and mineral surfaces. Another focus is placed on actinide environmental behavior at contaminated sites. More detailed information on aquatic actinide chemistry and notably the related thermodynamics is available from another contribution to this issue.27

The present review deals comprehensively with the various mechanisms of actinide interaction with mineral surfaces, including surface sorption, actinide incorporation into the solid matrix, surface induced redox reactions, and colloid adsorption. As quantum chemical approaches are increasingly applied to elucidate actinide surface structures and tentative binding modes, we dedicate a separate section to this topic, to provide an overview on the methods used and their benefits. The quantitative assessment of actinide behavior in a given environment, such as contaminated sites or nuclear waste disposal systems requires geochemical models to describe, for example, sorption reactions. In the final section, we give an outline on available modeling approaches, their underlying concepts and required parameters, and discuss exemplary applications to actinide interaction with minerals. Actinide reactions with microbial surfaces are quite comparable to interactions with mineral surfaces. Bacterial cell walls consist of exopolymers, proteins and lipids and, thus, contain various surface functional groups potentially capable of interacting with actinide cations. While hydroxyl groups dominate mineral surfaces, bacteria contain in addition carboxylate, phosphate, thiol, and amino moieties, where carboxylate and phosphate groups generally exhibit high affinities for actinide cations (see, for example, ref 28). The range of possible surface sorption mechanisms, such as outersphere binding, inner-sphere surface complexation, surface precipitation, and surface induced redox reactions, have also been observed for actinide−microbe interactions. Geochemical modeling approaches for quantitative description are virtually the same for metal cation sorption to mineral surfaces and biosorption. Under certain conditions, biomass can dominate actinide solid−liquid interface reactions, for example, in soil systems. Unlike mineral surfaces, living cells can actively take up metal ions and transfer them to the cell interior and sequester them in the intracellular medium. Many of these processes have been investigated in recent years by various types of spectroscopies. It is outside the scope of this review to treat this topic comprehensively, but a wealth of studies is available in the open literature; some can be found cited in more comprehensive reviews (e.g., refs 29−34).

2. ACTINIDE SORPTION MECHANISMS The term “sorption” encompasses a variety of possible processes and does not make any distinction with regard to the underlying mechanism. In this section, we deal with actinide reactions at mineral surfaces by various surface phenomena such as outer-sphere attachment and inner-sphere surface complexation of ionic and colloidal actinide species. Beyond pure surface attachment and complexation reactions, actinides can react with minerals by incorporation (mineralization) and surface induced redox reactions (see Figure 1). 2.1. Outer-Sphere Sorption

Cations can interact with negatively charged mineral surfaces by purely electrostatic attraction. Such reactions are well-known for cation interactions with permanently charged clay mineral surfaces (i.e., the typical “ion-exchange” sites). The origin of the relevant surface charge for this type of interaction in clay minerals is isomorphic substitution of, for example, Al in AlO6 octahedral layers by divalent cations (Mg/Fe(II)) and/or Si in tetrahedral SiO4 layers by trivalent cations (Al/Fe(III)). Permanent charge densities can be experimentally determined by quantifying their cation exchange capacity (CEC), which 1018


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ranges, for example, for montmorillonites from ∼0.2 to 0.4 equiv mol−1.35 The usually weak electrostatic nature of cation attachment to permanently charged surfaces renders this interaction readily reversible and strong competition with other cations takes place. Relative sorption strength is quantified by selectivity coefficients defining the exchange of actinide ions with other cations (e.g., Na+, Ca2+) bound to the clay mineral planar sites (for the definition of selectivity coefficients and for respective data see section 4.1 and Table 6). The mostly outer-sphere binding character of actinide ions sorbed to cation exchange sites has been evidenced by applying various spectroscopies36−39 but can also be derived from the interpretation of results obtained from batch experiments performed at different ionic strengths. Solid/liquid distribution ratios are almost pH independent in the low pH region and decrease strongly if ionic strength increases. A general observation is obviously that actinide outer-sphere sorption to clay-type minerals becomes prevalent at low pH and low ionic strength as is described for An(III), An(V), and An(VI).40−42 2.1.1. Trivalent Actinide Ions. For trivalent actinide/ lanthanide cations, TRLFS has been applied to study sorption onto illite and smectite type minerals such as montmorillonite and allows identification of outer-sphere species by their characteristic spectroscopic pattern. Spectroscopic features typical to sorbed Cm(III) species bound to minerals by different interaction modes are given in Figure 2, Table 1, and in more detail elsewhere.43,6

Table 1. Cm TRLFS Data for Species Sorbed onto and Incorporated into Solids, Where Transitory and Mixed Surface/Incorporated Species Are Possible

fractions.44 There are, however, indications for other hitherto unknown active acceptor sites besides structural Fe in the clay. EXAFS investigations suggest arrangement of the entire trivalent lanthanide/actinide aquo complex in the hexagonal cavities of the montmorillonite basal plane (Figure 3).44 Investigations using X-ray reflectivity (XR) and resonant Figure 2. Fluorescence emission spectra of various Cm(III) species and typical emission lifetimes; excitation wavelength λex = 396.6 nm (see also data in Table 1).

Despite close to 100% Cm(III) uptake onto smectite minerals at low ionic strength and at pH up to 4−5 observed in batch experiments, the fluorescence emission band maximum for the 6D7/2 → 8S7/2 transition of sorbed Cm(III) at 593.8 nm remains unchanged compared to that of the aquo ion.37 This is explained by the unaffected first hydration sphere and thus an unchanged ligand field. Slight shortening of the Cm fluorescence lifetime is explained by fluorescence quenching due to dipole−dipole interaction over a distance of ∼8 Å with iron in the clay structure. A correlation of the Eu/Cm fluorescence lifetime with the Fe content of the clay was established and allows quantifying Cm(III) outer-sphere surface

Figure 3. Proposed arrangement of an actinide(III)(H2O)9 aquo ion in hexagonal cavities of the montmorillonite basal plane;44 the structure is based on EXAFS analysis; Sm(III) is taken as a chemical analogue to Am(III). Reprinted with permission from ref 44. Copyright 2008 American Chemical Society. 1019


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while at low surface coverage an inner-sphere bidentate surface complex was observed. Chang et al.53 concluded from deconvolution analysis of their cryogenic TRLFS spectra obtained for UO 2 2+ adsorbed onto gibbsite a minor contribution from outer-sphere surface species. According to that study the outer-sphere species exists only at the lowest ionic strength investigated (0.001 mol dm−3) and disappears at higher electrolyte concentrations.

anomalous X-ray reflectivity (RAXR) on Pu(III) sorption onto the muscovite (001) plane, which has a structure comparable to that of montmorillonite45 (pH 3, 0.1 mol dm−3 NaClO4) revealed Pu to be located at a distance about 18 Ǻ above the surface. This finding is neither consistent with the existence of outer-sphere nor with inner-sphere bound species. The authors interpret their experimental results with tentative Pu complexation to an adventitious carbon layer at the muscovite surface or redox reactions leading to the formation of Pu(IV) nanoparticles. 2.1.2. Tetravalent Actinide Ions. In general, the affinity of tetravalent cations toward hydrolysis and chemical binding to surface hydroxyl groups is considered strong even at low pH so that purely electrostatic interactions with mineral surfaces are considered irrelevant.41 This is reflected by the onset of experimentally determined sorption edges at very low pH values ( titanium > silicium oxides. This finding roughly corresponds to the valence coordination number ratio (VCNR) defined by Pauling, which is a measure for Lewis base properties and thus the electron density of the respective surface hydroxyl groups.55,57 The presence of strong complexing organic and inorganic ligands effectively competes with surface complexation. Especially carbonate and carboxylic acids can significantly suppress actinide uptake by solution complexation or competitive sorption, but also favor formation of ternary surface complexes. A detailed discussion of the interplay of actinide solution chemistry and reaction with mineral surfaces follows further below. 2.2.1. Trivalent Actinide Ions. In the literature a great number of studies on inner-sphere sorption of trivalent actinide ions onto pure and composite minerals is available: aluminum and iron oxides and hydroxides, clay minerals, igneous rock minerals, titanium oxides, calcite, etc. Am(III) sorption was studied mainly by batch sorption experiments,41,56,58−61 while only a few spectroscopic studies have been published.62−64 In a combined EXAFS and TRLFS study of Am(III) sorption onto smectite and kaolinite inner-sphere surface compexation is studied at pH 6.62 The trivalent actinide ions coordinate with 4 oxygen atoms from the surface and 5 remaining hydration sphere water molecules. Am(III) sorption onto ferrihydrite results in a bidentate corner-sharing surface species based on EXAFS data.63 Two different inner-sphere surface complexes have been identified also on quartz up to pH 9.4.65 In this case the sorption data have been interpreted as bidentate Am/ Cm(III) coordination to the surface. Very often the trivalent lanthanides are used as chemical homologues to simulate the behavior of the trivalent actinides Am(III) and Cm(III) with similar ionic radii. Analogous behavior could be clearly verified for e.g. Eu(III), Gd(III), Am(III), and Cm(III) sorption onto γAl2O3.58 Surprisingly, Am(III) binding to Na-illite was reported 1020


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to be significantly stronger than Eu(III) binding.42 This finding is currently not understood, but might be explained by the differing sorption properties of different illite batches used for the experiments. Cm(III) TRLFS sorption studies are available for a wide range of solids encompassing γ-Al2O3,7,66,67 α-Al2O3,68,69 gibbsite,69,70 bayerite,69 illite,71,72 kaolinite,37,555 smectite,7,37,72 quartz,7,65 silica,5 calcite,73 feldspars,74 and several phosphate minerals.75 A compilation of the spectroscopic findings can be found elsewhere.4,6 Tan et al.4 reviewed TRLFS (and EXAFS) results for Cm(III) (but also Eu(III) and U(VI)) sorption onto a variety of mineral surfaces and provide a nice comparison of spectroscopic findings such as peak positions, fluorescence lifetimes, number of water molecules in the first coordination sphere of the respective actinide ions and assignments to surface complex structures. Inner-sphere surface complexation induces characteristic peak shifts in Cm(III) emission spectra from 593.8 nm for the Cm(III) aquo ion to ∼598 up to 607 nm for the respective surface complexes. The associated fluorescence lifetimes extend from ∼65 μs (Cm(III) aquo ion) to 110−220 μs in the various inner-sphere surface complexes (see Table 1 and Figure 2), which corresponds to about 2−5 H2O equivalents remaining in the first coordination sphere. These investigations are (partly) compared with additional batch sorption studies using Eu(III) or Gd(III). To some extent spectroscopic data have been used as a basis for validating or establishing geochemical sorption models. Unfortunately, the iron (hydr)oxide minerals, which are considered as strong sorbents attenuating metal ion migration in the environment, cannot be studied using TRLFS with Cm(III) as a fluorescence probe. The strong luminescence quenching of Fe caused by energy transfer from excited Cm(III) states to electronic Fe levels leads to the complete extinction of the fluorescence signal. Isomorphous aluminum (hydr)oxides have frequently been used as model minerals instead.38,68 Surface X-ray scattering experiments have demonstrated similar structures of the hematite- and corundum-water interface,76,77 however, subtle differences in surface water ordering and reactivity were also identified. Variations in adsorbed metal ion surface speciation have been discovered as well.78 The solubility of aluminum oxides is different to iron oxides; their solubility is higher compared to that of the respective hydroxides. Therefore, surface alteration reactions take place and surface aluminum hydroxide (bayerite, gibbsite) formation may occur. Sample pretreatment and purification procedures can, thus, influence surface properties significantly.79−82 In all TRLFS experiments using aluminum (hydr)oxides inner-sphere surface complexes of Cm(III) are found. It is remarkable, that for all aluminum oxides/ hydroxides, for aluminum containing clay minerals as well, three different inner-sphere surface complexes with very similar spectroscopic features are reported. This finding points to a comparable chemical environment of Cm(III) at the different mineral surfaces with similar surface ligands. Fluorescence lifetime results indicate that about half of the hydration sphere of 9 H2O/OH− ligands is removed in the course of surface complexation.62,69,72,83 Combined EXAFS and TRLFS studies on Am(III)/Gd(III)/Lu(III) and TRLFS study of Cm(III) sorption onto γ-Al2O3, smectite and kaolinite, show that Am/ Cm/Gd/Lu all form inner-sphere surface complexes with a coordination to ∼9 (or 8) oxygen atoms (comprised of both surface and ligating aqua molecules). This confirms the assumption that a coordination very similar to that of the

aquo ion is retained in the surface complex. TRLFS with Cm(III) suggests mostly coordination to 4 ± 1 oxygen atoms at the surface and to 5 ± 1 water molecules. At increasing pH, a slight decrease of the coordination number was observed in EXAFS results, which was attributed to destructive interference because of slightly different metal−oxygen distances notably for the ternary hydroxo surface species.62 Spectroscopically identified species forming in the absence of complexing ligands other than OH− and H2O in different pH regions were assigned to the following stoichiometry: SOH + Cm 3 + + 5H 2O ⇔ SOCm(OH)x 2 − x (H 2O)5 − x + (1 + x)H+

(1)

with x = 0, 1, and 2 and SOH being (generic) surface hydroxyl groups. Overall, the experimental findings suggest that the pH dependent surface speciation will be controlled by the deprotonation of one surface hydroxyl group. However, because of the surface structure of the minerals considered, further coordination to about 3−4 neighboring surface oxygen atoms takes place simultaneously. Broad fluorescence emission bands in TRLFS and high Debye−Waller-factors in EXAFS point to the existence of slightly differing chemical environments with asymmetry in the metal−O pair distribution arising from variations in the first coordination sphere (number of coordinating water molecules, bonding to different groups, and possibly variable distances of metal binding to ligating hydroxyl groups).62,83 As will be discussed later, these assumptions can be confirmed to some extent by other spectroscopic studies and by quantum chemical calculations. The existence of different binding sites at the surface of sapphire (α-Al2O3) was spectroscopically identified at different crystal planes of single crystals. The (001), (110), (012), (104), and (018) planes of α-Al2O3 showed different affinities toward Cm(III) sorption under otherwise identical experimental conditions as evidenced by α-spectroscopy and audioradiography.68,69 TRLFS results for Cm(III) sorbed onto the (001) surface show distinct differences with regard to peak position and fluorescence lifetime compared to the other four crystal planes and are interpreted as indicators for different binding strengths. This is likely due to the fact that this plane shows the largest relaxation from its ideal cut when contacted with water. Similar conclusions come from XPS analysis of Cm(III) 4f7/2 electron binding energies. EXAFS on Gd sorption on gibbsite (α-Al(OH)3, platelets) and bayerite (β-Al(OH)3, rods), which possess clearly distinguishable planar and edge faces with different binding sites (number, orientation, and nature of aluminol sites) revealed a single metal−oxygen shell for both systems with similar metal−oxygen distances but two metal−Al distances.69 The relative large Debye−Waller factors obtained again indicate either slightly distorted structures or a mixture of different species. Findings obtained from TRLFS studies with Cm(III) and EXAFS data for Gd(III) could only be explained by assuming a mixture of different Gd(III) surface complexes existing at the mineral surface: two basal and three edge sites at the gibbsite surface (see Figure 4) and two basal and four edge sites at the bayerite surface are in agreement with EXAFS findings with respect to coordination numbers and atomic distances. In these coordination environments, Gd(III) is bound either in bi- or tridentate fashion by Al−OH surface groups. TRLFS data are, however, less compatible with bidentate coordination and rather point to higher coordination 1021


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Figure 4. Normal view of the gibbsite (001) surface showing the surface unit cell and possible sorption sites.69 Doubly coordinated aluminol groups are depicted in red, singly coordinated aluminol groups in blue. On the basis of EXAFS measurements, the “suitable” binding sites are B2, B5, E1, E4, and E5 (B, basal sites; E, edge sites).

by surface oxygen atoms.69 The discussion of suitable binding sites was based on the assumption that ideal surfaces derived from crystallographic data of bayerite and gibbsite exist in the experiments in contact with solution. Impurities, defects, crystallographic steps at the surface, surface relaxation in response to solution contact, or the presence of amorphous phases, etc., may lead to more variability in binding sites. Analysis and assignment of metal−oxygen distances determined by EXAFS is as well not always straightforward (see, e.g., ref 84). One can not exclude that longer metal−oxygen distances usually assigned to coordinated hydration water might stem from binding to surface oxygen atoms. Another example for the variety of possible binding sites at a mineral surface is provided by studies of Cm(III) sorption onto γ-Al2O3 and kaolinite. Experiments in the presence of carbonate have been performed by direct excitation of the Cm(III) metal ion at 20 K,85 where Marques et al. did not excite Cm(III) ions adsorbed on γ-Al2O3 at λex = 396.6 nm from the 8S7/2 ground state to the F-state but rather scanned the excitation light wavelength from 596−622 nm thus populating lower 6D7/2 states. The resulting emission spectra are reproduced in Figure 5. The sharp line in the emission spectra reflects the fluorescence line resonant with excitation (RFL). The broad emission band at higher wavelengths represents emission from lower lying levels of the excited states. Such an observation can be interpreted by the presence of a variety of surface bound Cm-species with slightly different chemical environment and thus different ligand field splitting of Cm-f-orbitals, an interpretation in accord with EXAFS results above. One outcome of all these studies is that metal ion binding to mineral surfaces (often) cannot be assigned to a single, and well-defined surface species with a fixed structure as in crystal lattices. A similar conclusion is derived from the observation that emission spectra of surface bound Cm recorded at very low temperatures (approximately 10 K) are broad, even though energy levels thermally populated at room temperature are suppressed. Such is the case for Cm(III) sorbed onto calcite,86 α-Al2O3 single crystals, and clay minerals, respectively. Only if Cm(III) is incorporated into the substrate structure and located in well-defined positions within a crystal lattice are narrow fluorescence peaks obtained in cryogenic TRLFS.86 Quantum chemical calculations are in agreement with this spectroscopybased picture, as they revealed local energy minima with only

Figure 5. Fluorescence emission spectra of the6D7/2 → 8S7/2 transitions measured at T < 20 K for the Cm(III)-carbonate/γAl2O3 system as a function of excitation wavelength (596−622 nm). The sharp bands represent the fluorescence emission resonant with the excitation; the broad sidebands correspond to the fluorescence emission from the lower lying 6D7/2 excited states86 (RFL = resonant fluorescence line). The fluorescence spectrum excited at 396.6 nm (representing the absorption maximum of the Cm(III) aquo ion) is added for comparison. Reprinted with permission from ref 85. Copyright 2010 American Chemical Society.

slight differences in energies for different surface sorbed trivalent element species87,88 (see discussion below). All these findings obviously suggest that even well-defined surfaces offer different sites for metal ion sorption, where the trivalent actinide ion exhibits variability in its chemical environment. Even though variability in surface bound actinide structures exist on a molecular scale, spectroscopic characterization of actinide surface speciation provides complementary and consistent information suitable to validate surface complexation models. This is demonstrated in Figure 6 for a TRLFS study of Cm(III)/Eu(III) sorption onto Ca montmorillonite. The Cm(III) species distribution derived from peak deconvolution of TRLFS spectra is compared with the prediction of a surface complexation model (2SPNE SC/CE).71 There is good agreement between model calculations/predictions and TRLFS spectroscopic findings which supports model assumptions with respect to surface speciation.71,72 Where exactly Cm/ Am/Eu(III) are located at the montmorillonite surface is not entirely clear from this data, however. On the basis of TRLFS and EXAFS data,62,89 location of the metal ion at Al octahedral at the smectite (010) edge plane has been proposed.90 This appears to have been validated in a recent EXAFS study on Am interaction with a Mg-based smectite mineral (hectorite). The first oxygen coordination shell is found at d(Am−O1) = 2.42(1) Å. Mg and Si shells containing 10 dissolved silica increases due to enhanced dissolution of the alumosilicate. This is proposed to cause the appearance of silicato-Cm(III) surface species, identified by additional TRLFS bands. The fraction of these species is enhanced if the silicate concentration increases. Similar spectral features have been recently reported in a study of Cm(III) sorption onto illite and montmorillonite at pH > 10 and attributed to the same mechanism.102 Both studies confirm an earlier Cm(III) sorption study on quartz that suggested formation of ternary silicato surface species in sufficient excess of dissolved silica.65 2.2.2. Tetravalent Actinide Ions. The tetravalent actinides Th(IV), U(IV), Np(IV), and Pu(IV) have very low solubilities and a strong tendency toward hydrolysis under relevant natural groundwater conditions.103 This leads to a strong interaction (sorption) with any kind of surfaces, even at low pH.104−106 Precipitation or polymer/colloid formation due to oversaturation46,107,108 have to be expected as side reactions in sorption studies of tetravalent actinides. Among the actinides only Th(IV) is exclusively stable in the tetravalent oxidation state. The other actinides U−Pu are redox active and form different oxidation states depending on electro-chemical boundary conditions (for surface induced redox reactions, see below). To avoid associated experimental complications, many sorption studies reported in the literature have been performed with thorium. Typically, An(IV) sorption is modeled with monodentate surface complexes, while spectroscopic investigations suggest rather that bidentate surface complexes form. In the following selected examples are given. Thorium sorption onto silica was modeled assuming monodentate complexation at silanol surface sites and considering the different Th(IV) hydrolysis species (Th(OH)x4−x with x = 0−4).109 The same surface speciation was used in a surface complexation model to describe Th(IV) sorption onto illite and montmorillonite.41,42,110 In a polarized EXAFS study, however, the authors concluded that Th(IV) binds to the montmorillonite surface by sharing double corners with Si tetrahedra at clay platelet edge sites.46 An EXAFS study on Th(IV) sorption onto iron corrosion products (magnetite and ferrihydrite) points to a bidentate surface coordination to the iron phases as well.111 A bidentate corner sharing Th(IV) complex, where the Th(IV) atom shares one O atom with each of two [FeO6] octahedra was also found for ferrihydrite. In case of magnetite, two different surface complexes were identified: bidentate corner sharing arrangements to [FeO6] octahedra and to [FeO4] tetrahedra.111 In the presence of strongly complexing ligands such as carbonate, phosphate or carboxylic acids, ternary An(IV) surface complexes have been reported. Uptake of Th(IV) and Np(IV) has been studied at high pH (pH = 11.0−13.6), under conditions where γ-Al2O3 and SiO2 surfaces are negatively charged and at elevated carbonate concentrations where tetravalent actinide ions form anionic carbonato complexes.

Figure 6. Top: Cm(III) species distribution derived from evaluation of TRLFS spectra in presence of 0.25 g dm−3 Ca-montmorillonite and 2 × 10−7 mol dm−3 Cm(III) in 0.066 mol dm−3 CaCl2 solutions. The Cm3+ aquo ion and three Cm(III) surface complexes identified. The gray areas visualize experimental uncertainty ranges. Bottom: Calculation of Cm(III) species distribution using the 2SPNE SC/CE model and parameters taken from a previous Eu(III) sorption modeling study on the same material.71,72 Reprinted with permission from ref 71. Copyright 2005 Elsevier.

influence surface speciation of actinide ions. In TRLFS experiments of Cm(III) sorption onto calcite, formation of ternary carbonato surface complexes was observed.91 Cm(III) sorption experiments on kaolinite and on γ-Al2O3 in the presence of CO32− show, that at least two different Cm(III) carbonato-surface species occur.85 In a similar sorption study on montmorillonite Eu(III) surface species containing carbonate were also detected by TRLFS and the batch sorption data could be modeled with the 2SPNE SC/CE model71 including two additional surface species, namely, SOCmCO3, and SOCmOHCO3−.92 The same model approach including ternary surface carbonato complexes was successfully used to describe experimental data for Eu(III) and Cm(III) sorption onto natural clay rock (Callovo−Oxfordian and Opalinus clay).93 Because of the strength of An(III) carbonato surface complexes, the presence of dissolved An(III) carbonato complexes only moderately attenuates An(III) sorption onto clay minerals at circumneutral pH. Naturally occurring organic carboxylic acids such as humic and fulvic acids (or respective model ligands) also have strong complexing properties toward trivalent actinides and as a result influence sorption when present.60,61,94−98 Depending on the ratio/amount of organic matter sorbed at the mineral surface and in solution, its overall effect on metal ion sorption strongly depends on pH. At low pH, where significant amounts of 1023


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carbonate groups at steps of the calcite (104) face. Two additional carbon backscatterers at 2.94 ± 0.02 Å are conclusively interpreted as belonging to two bidentate coordinated carbonate groups, pointing away from the surface, thus forming a ternary surface complex (see Figure 7).136

The observed decrease in distribution coefficient with increasing carbonate concentration is explained by the simple assumption that negatively charged complexes will not (or barely) sorb.112 Declining Pu(IV) sorption onto calcite with increasing pH was also explained by the competitive formation of aqueous plutonium−carbonate species.113 Nevertheless, for the description of Pu(IV) sorption the postulation of two different Pu(IV)-hydroxo-carbonato surface species was required. In the presence of citrate, causing formation of a dissolved 1:2 Pu(IV)-citrate complex (Pu(cit)22−), Pu(IV) sorption onto Mn oxide in the pH range between 3 and 7 was explained by dissociation of Pu(IV) from dissolved Pu(cit)22− and subsequent sorption onto the solid.114 In this case, ternary surface complexes of Pu(IV) involving citrate were not observed. The presence of natural organic matter (humic and fulvic acid) has a strong impact on thorium sorption, is affected by the extent of organic interaction with mineral surfaces and thereby depends on pH.56,96,106,109,115 General trends of actinide(IV) partitioning between solid and dissolved organic matter are inherently similar to observed behavior of trivalent actinide ions. The effect of phosphate was investigated for Th(IV) sorption onto gibbsite116 and TiO2.117 In both cases, increased Th(IV) uptake was observed in the presence of phosphate. 2.2.3. Pentavalent Actinide Ions. Neptunium(V) is the most relevant pentavalent actinide in the context of nuclear waste disposal. It is quite stable under oxic conditions, and because of its low effective charge, it also has a lower tendency toward sorption reactions than the tri- and tetravalent actinide ions.118 Under reducing conditions Np(V) can be reduced at mineral surfaces to Np(IV) (see below). Reduction due to interaction with high-intensity X-ray beams in EXAFS studies has also been observed.119 Generally, Np(V) undergoes innersphere sorption at higher pH. This was demonstrated in NpO2+ sorption studies on different iron minerals120,121 using XPS122 or EXAFS,123 as well as on kaolinite,124,125 gibbsite,126,127 silica,56,112,128 α-and γ-Al2O3,56,112,124,125,129 clay minerals,41,42,130,131 and feldspar.132 Upon sorption to goethite, for instance, an inner-sphere surface complex was detected in an EXAFS study with only subtle differences in spectra obtained for the aquo ion, where Np(V) is coordinated by 4 oxygen atoms at a distance of 2.51 ± 0.03 Å in the equatorial plane.123 By using attenuated total reflection (ATR) infrared spectroscopy, inner-sphere complexes were also found for Np(V) sorption onto SiO2, TiO2, and ZnO and similar bidentate surface complexes have been postulated for all three minerals.133 This was concluded from the spectroscopic fingerprint observed for the stretching frequency of Np(V)− O bond. Due to strong competition with carbonate complexation, Np(V) sorption onto kaolinite124 and calcite113 at ambient CO2 partial pressure decreases at pH > 8, down to zero at pH 10. The CO2 influence on Np(V) speciation and sorption onto hematite was investigated by EXAFS. Both a bis-carbonato Np(V) inner-sphere surface complex coordinated via a bidentate Np(V)−O2−Fe linkage and a tris-carbonato outersphere complex were found.134 Np(V) sorption onto calcite using EXAFS also points to inner-sphere complexation, which was concluded from the observed decrease in Np(V)−O-yl distance and the presence of 3−6 carbon backscattering atoms for the sorbed species.135 In a recent publication dealing with the same system, results suggest that Np(V) is linked to

Figure 7. Most likely neptunyl−calcite adsorption complex based on EXAFS results and geometrical considerations: neptunyl adsorbs at step edges on the calcite (104) face as a bidentate inner-sphere biscarbonato complex. Green lines indicate bonds to the six equatorial oxygen atoms. (Np, green; Ca, blue; O, red; C, gray).136 Reprinted with permission from ref 136. Copyright 2011 Elsevier.

Np(V) sorption on soil predominantly occurs via surface complexation on clay minerals.137,138 The presence of strong complexing organic molecules influences Np(V) uptake: the extent of sorption is determined primarily by the affinity of the organics for the mineral surface.56,139−141 In the presence of humic and fulvic acids, enhanced sorption is generally observed at lower pH compared to the actinide-mineral binary system. Because of competition effects with sorbed and dissolved organic ligands, reduced sorption is detected at higher pH. Potential reduction of Np(V) to Np(IV) has also been discussed.105 The 2SPNE SC/CE model, already mentioned for trivalent actinides, was adapted to describe Np(V)−illite and montmorillonite sorption data. Surface complexation constants describing Np(V) interaction with these two clay minerals have been evaluated for two aqueous Np(V) species: NpO2+ and NpO2(OH)0.41,42 A compilation of Np(V) surface complexation constants on a multitude of mineral phases142 and complexation constants for the sorption of different Np(V)− carbonato complexes on clay buffer material have been reported.143 Only few papers on Pu(V) sorption have been published and results reported compare well with Np(V) results.113,144 However, Pu(V) reduction at surfaces is much more pronounced than it is for Np(V).47 2.2.4. Hexavalent Actinide Ions. As uranium can be handled easily in many laboratories and is of interest as a pollutant at many contaminated sites worldwide,145,146 a large variety of sorption studies onto a manifold of different solid phases has been published. This large amount of literature cannot be discussed entirely in the present review. While U(VI) is most stable under oxidizing conditions, only a few studies are published on Pu(VI) due to fast reduction.147−149 As a consequence of the dioxocation structure the effective charge in the equatorial plane of the AnO22+ ion is +3.3.150 Strong 1024


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applied to elucidate the structure of U(VI) surface species and the underlying sorption mechanism.156,158,171,153,172−176 In agreement with the previously discussed TRLFS studies, usually mononuclear bidentate (or “edge sharing” at, e.g., [FeO6] octahedra) surface complexes51,153,156,173,177 and, depending on the metal ion concentration, surface precipitates and/or surface bound oligomeric species were identified.158,178 U(VI) interaction with polycrystalline and monocrystalline (110) TiO2 results in two different bidendate surface complexes as evidenced by EXAFS: one complex with shared edges and one with shared apical oxygen atoms of the TiO6 octahedron.51 As already discussed before, predominantly outer-sphere U(VI) species were surprisingly detected at the (001) surface. In a comprehensive study using GI-EXAFS, CTR diffraction, and XPS the binding sites and surface complexation of U(VI) adsorbed on the (11̅02) surfaces of α-Al2O3 and α-Fe2O3 have been analyzed.153 Samples were prepared under ambient conditions at pH 5 and 1 mmol dm−3 U(VI) for α-Al2O3 and pH 7 and 85 μmol dm−3 U(VI) for α-Fe2O3. The available binding sites on the (11̅02) surfaces were constrained through analysis of bond valences and steric conditions. It could be shown that U(VI) forms ternary uranyl−carbonato surface complexes with surface oxygen atoms that are singly coordinated to aluminum or iron (e.g., Fe−OH groups). On the α-Al2O3 (11̅02) surface, a monodentate complex results, whereas on the α-Fe2O3 (110̅ 2) surface, the binding is bidentate to adjacent singly coordinated oxygen sites (i.e., binuclear with regard to the Fe−OH groups involved or, as an equivalent term, “corner sharing” of two adjacent [FeO6]octahedra). Differences in protonation of the singly coordinated oxygen atoms, surface charging, U(VI) aqueous speciation (because of different experimental conditions), substrate structure, or the electronic structure of surface functional groups are assumed to be responsible for these differences in adsorption geometry. For the isostructural surfaces, the observed surface complexes are different from the bidentate, mononuclear (edge sharing) surface complexes of U(VI) typically found for U(VI) sorption on powdered aluminum- and iron-(oxyhydr)oxides and clay minerals using EXAFS spectroscopy. It was concluded that the presence of monodentate-mononuclear and bidentate-binuclear complexes may have been overlooked in past EXAFS studies on such substrates, as these complexes have U−Al or U−Fe interatomic distances too large to be easily detected by EXAFS spectroscopy.153 Very recently the influence of solution composition and, therefore, the uranium speciation on sorption onto the magnetite (111) surface was investigated. U(VI) sorption takes place exclusively in the presence of Ca and carbonate with no reduction to U(IV).179 Apparently, the high thermodynamic stability of the Ca2UO2(CO3)3 complex stabilizes uranium in oxidation state VI and shifts the stability field for U(VI) to lower Eh values. This finding once again clearly points to the relevance of aqueous speciation for the nature of surface reactions and surface speciation. Uranium(VI) forms very stable carbonato complexes in solution and as a consequence uranium sorption in the presence of dissolved CO2 is strongly suppressed in comparison to the carbonate free system.155,180−182,26,151,183 The effect for U(VI) is much more pronounced than, e.g., for the trivalent actinides.183 Description of uranium sorption on natural rocks and sediments in the presence of carbonate is only possible by postulating formation of various ternary uranium carbonato

hydrolysis and carbonate complexation are known reactions in hexavalent actinide chemistry, as well as the formation of U(VI) silicato species and formation of polynuclear (dimeric, trimeric etc.) species.26 In the presence of Ca2+, ternary Ca−U− carbonato complexes, such as Ca2UO2(CO3)3, are formed.145 As uranium aqueous speciation is quite complex, interactions with mineral surfaces are difficult to correctly describe.151 Spectroscopic and theoretical studies on the nature of U(VI) inner-sphere surface complexation usually suggest coordination in a bidentate fashion.152,153 As in case of Cm(III), the fluorescence of the UO22+ cation has been utilized to TRLFS experiments for in situ surface speciation often combined with EXAFS or XPS, for example, on silica,154,155 albite,156 alumina,155,157 calcite,158 gibbsite,159 smectites,155,160,161 kaolinite,162 muscovite,163 natural sediments,146 and phosphates.164−170 In all cases primarily bidentate inner-sphere surface complexes for UO22+ were identified. At increasing pH and metal ion concentration, ternary hydroxo and carbonato surface complexes and polynuclear complexes form and finally U(VI) hydroxide/carbonate precipitates. SiOUO2 type complexes form at surfaces of the minerals silica154 and albite.156 Similarly, a TRLFS study of U(VI) interaction with gibbsite was interpreted as indicating formation of a bidentate mononuclear inner-sphere surface complex, in which the uranyl ion is bound to two reactive OH− groups at the broken edge, linked to one Al atom. In addition, polynuclear uranyl(VI) surface species were observed.159 Montmorillonite was considered in a conceptual model as a composite of Al(OH)3 and silica. Sorption and spectroscopic data were interpreted as a combination of U(VI) binding to AlOH and SiOH surface groups.155,161 Likewise, inner-sphere U(VI) surface complexes were identified at kaolinite162 and muscovite surfaces.163 U(VI) adsorption complexes forming at the calcite surface were assigned to triscarbonato-like complexes by EXAFS and luminescence spectroscopy at low metal ion concentrations.158 Cryogenic laser induced fluorescence spectroscopy allows to obtain spectra with much better resolution and intensities compared to measurements at room temperature.17,146 This technique was used in an extended study on uranium sorption onto fine grained sediment materials from the Hanford 300 area in simulated Ca-saturated groundwater by comparing their behavior with a series of reference minerals.146 By measuring the surface area-normalized KD values the following affinity series for UO22+ sorption was obtained: 6-line-ferrihydrite > North Carolina-chlorite ≈ California-clinochlore > quartz ≈ Michigan-chlorite > illite > montmorillonite. In the selected Hanford sediments two major types of adsorbed U(VI) species were identified: U(VI) adsorbed onto quartz (or amorphous silica) and onto phyllosilicates. Comparison with literature spectroscopic data suggested that U(VI) exists primarily as inner-sphere complexes, bound to surface silanol groups on quartz and as surface U(VI) triscarbonato complex on phyllosilicates. For completeness one must note that for different depths below the surface of the Hanford 300A vadose zone and for different sediment size fractions, U(VI) can be found in different chemical forms.146 In an additional cryogenic laser induced fluorescence spectroscopy study on U(VI) contaminated Hanford sediments from the BX Tank Farm U(VI) was found as a uranophane-type solid or soddyite.17 Unlike TRLFS, which mostly provides a sort of fingerprint information for U(VI) surface species, EXAFS can be directly 1025


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Figure 8. Various modes of UO22+ interaction with goethite: (a) in absence of PO43−/bidentate edge sharing and bidentate corner sharing surface complexes and (b) in presence of PO43−/ternary phosphate surface complexes with UO22+ directly bound to goethite surface hydroxyl groups or (c) to surface adsorbed PO43− ions.195 Reprinted with permission from ref 195. Copyright 2012 American Chemical Society.

surface complexes.151,184−186 Different stoichiometries have been proposed in the literature for these species, a portion of these in modeling studies: SOUO2(CO3HCO3)2−,151,184,185 SOUO2(OH)(CO3)2−,151,154 SO(UO2) Cax(CO3)y,158,179 S(CO3)UO2(CO3)23−,120,187 SOUO2HCO3,151 SOUO2CO3−,151 SOUO2(CO3)23−,151  SOUO2(HCO3)2−,151 and SOUO2OH(CO3)24−.151 In addition to sorption of pure carbonato and bicarbonato complexes, mixed hydroxo-carbonato and carbonato-bicarbonato surface complexes have been proposed. Sorption of Cauranium−carbonato species (such as Ca2UO2(CO3)3), which dominate aqueous U(VI) speciation in Ca containing systems at neutral/slightly alkaline pH conditions145 have not yet been experimentally verified but had to be considered in modeling to successfully describe U(VI) uptake data.141,188The existence of bicarbonato surface complexes may be questionable, as bicarbonate is known to be a weak ligand for actinide ions in solution. The presence of ternary uranyl-carbonato surface complexes has also been verified in various EXAFS studies, for example, refs 172 and 176. Despite the richness of possible surface U(VI) carbonato species, spectroscopic, and theoretical studies on uranyl sorption behavior in carbon dioxide bearing systems indicate these to show a generic pattern146,189,190 as similar surface complexes are frequently reported. Uranyl uptake curves show characteristic cation behavior at low pH and in the presence of carbonate an adsorption envelope is observed at higher pH that is reminiscent of anion adsorption.191,192 This again points to the importance of solution speciation, in this case for uranyl sorption, which is particularly clear in the case of the abovementioned desorption edge that is caused by carbonate complexation. If full congruence between the adsorption patterns occurs on different (but composite) sorbents, it is often explained by the presence of a dominant solid phase. However, as discussed above, a recent comparison between a number of different solids actually shows that the distribution coefficients of uranyl

may strongly vary for comparable conditions146 suggesting that not one dominant solid phase is present in all the solids compared or if individual oxides are compared (see also section 4.2.3). The presence of PO43− can have a significant impact on uranyl speciation as well.193−195 Description of U(VI) adsorption to different amorphous and crystalline ironoxide/ hydroxide containing solids in presence of phosphate by a surface complexation modeling approach assuming formation of a ternary FePO4(UO2) surface complex and hydrous ferrous oxide being the dominant reactive sorbent turns out to be inappropriate. While the approach works for modeling U(VI) sorption in absence of phosphate, a consistent model for phosphate adsorption to all ironoxides/hydroxides in a reported study failed.193 One reason for the failure might be the formation of surface precipitates in addition to the assumed surface complexes, as has been experimentally evidenced for U(VI) sorption onto alumina by TRLFS at a total PO43‑ concentrations >10−4 mol dm−3.194 A recent EXAFS study reveals the complexity of U(VI) interactions with phosphates in the presence of a mineral surface. Beside precipitation of meta autunite phases, two types of ternary phosphate surface complexes at the goethite surface were spectroscopically identified: UO22+ bridging the surface and the PO43‑ ion and surface adsorbed PO43− coordinating the UO22+ cation195 (Figure 8). As for carbonate and phosphate anions, ternary U(VI) surface complexes have been found in the presence of arsenate. In the context of remediation of uranium contaminated sites enhanced uranium sorption on aluminum oxide pretreated with arsenate was observed and interpreted as being due to the formation of ternary surface UO22+-arsenato complexes or surface precipitation.196 The influence of humic substances on U(VI) sorption was also studied and was again observed to be affected by the distribution of the organic between the solid and aqueous phase.197,198 An EXAFS study of the ternary U(VI)/kaolinite/ 1026


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2.3. Actinide Incorporation Processes

humic acid system showed U(VI) prefers direct binding to kaolinite over sorption to (pre)adsorbed HA.199 Additional TRLFS experiments prove U(VI) coordinates directly to humic acid under certain conditions,162 with U(VI) directly bound to the kaolinite surface, acting as a bridge to coadsorbed HA.162 The existence of ternary humate surface complexes also turned out to be necessary for describing uranyl sorption onto hematite106 and onto other solids. U(VI) sorption was also studied on phosphate containing minerals, which were discussed as potential engineered barriers in nuclear waste repositories or contaminated sites. Bidentate inner-sphere U(VI) surface complexes form o n Th4P2O7(PO4)4, ZrP2O7 and Zr2O(PO4)2 surfaces; depending on the chemical conditions combined XPS, EXAFS, TRLFS and surface complexation modeling studies suggest (S− O)2UO2, (S−O)2UO2(NO3)−, and (S−O)2UO2(OH)− type surface species.164−166 Similar species were postulated for the LaPO4 system as well, where slightly different species were identified for low and high surface coverage.167 The different species were interpreted as uranyl interacting with “strong” and “weak” sites, respectively. At the surface of ZrP2O7 and Zr2O(PO4)2, ZrOH and phosphate groups are responsible for metal ion sorption. A slight increase in sorption is found with increasing temperature from 25−90 °C,168,170 where mainly the surface phosphate groups react upon a temperature increase. Ternary UO22+ surface complexes form in presence of citric and oxalic acid at the ZrP2O7 surface.169 Monodentate surface complexes of UO22+ and ternary (hydroxo, carbonato etc.) complexes are considered appropriate to describe experimental data for uranium interaction with mineral surfaces in many surface complexation modeling exercises.42,151,142 In contrast, spectroscopic findings overwhelmingly point to bidentate uranium surface species, as discussed in detail above. As for the trivalent actinide cations, this apparent contradiction can be easily resolved, if we postulate monodentate surface complexes are required to explain the pH dependent sorption behavior (i.e., only one oxygen atom must be deprotonated to account for the pHdependence of sorption using surface complexation models). On a molecular scale, however, the UO22+ cation obviously coordinates to two neighboring oxygen atoms on mineral surfaces. 2.2.5. Competitive Sorption. Geochemical modeling of actinide sorption in multicomponent systems currently relies on the assumption of competitive metal ion interaction with hydroxyl groups at mineral surfaces. The extent of individual metal ion sorption is determined by the respective complexation constants. Recently, this concept has been put into question by a study on competitive sorption of actinide/ lanthanide and other metal ions onto montmorillonite and illite.110 Interestingly, only metal ions with similar chemistry (e.g., valence state) were observed to compete with one another, whereas metal ions with dissimilar chemical properties were not so that competition seems to be selective.110 Eu(III), Nd(III), and Am(III) were reported to show competitive behavior. In contrast, competition between Th(IV) and U(VI) was not observed. From these findings, the authors conclude that multiple sets of surface sites exist at the clay mineral surfaces, which are selective with regard to the sorption of certain types of metal ions. Other types of metal ions are not able to compete for sorption on those selective sites.

When describing actinide interactions at the mineral/water interface, the solid phase itself is generally considered invariant: only ion, electron, or proton exchange processes at surface hydroxyl groups are taken into account. However, mineral dissolution, alteration, and secondary phase formation, as a consequence of geochemical variations, can occur as well, thereby modifying sorption mechanisms in such a way that trace contaminants ultimately are incorporated into the solid matrix. In general, inclusion of radionuclides into the bulk structure of host minerals represents a potentially efficient pathway for actinide immobilization. Coprecipitation and mineralization processes are discussed as being relevant for actinide immobilization during corrosion of used UO2 fuel, high-level vitrified waste, cementitious waste, steel containers, and cast iron inlets.200 Formation of crystalline actinide bearing solids is furthermore of interest for the development of specific waste matrices for actinide disposal (e.g., ref 201) and of dedicated fuel for nuclear fission reactors or targets for transmutation techniques. Even under overall chemical equilibrium conditions, minerals cannot be considered as inert. Ionic minerals, such as calcite or barite, show fast surface dynamics in contact with electrolyte solutions, where more than one monolayer continuously exchanges with solutes. Dissolution rates can be taken as a measure for the surface dynamics and range from 6.6 × 10−7 to 0.89 mol m−2 s−1 (25 °C202). Under such conditions, it is likely that dissolved metal ions participate in dynamic crystal surface rearrangement processes, especially if they are present in trace concentrations. For clay and oxide minerals, surface dynamics are much slower and dissolution rates are orders of magnitude smaller (e.g., smectites k = 6.0 × 10−12−1.7 × 10−11 mol m−2 s−1; 35 °C, 0.1−4 mol dm−3 KOH203). Incorporation of trace elements into oxide minerals under equilibrium conditions are therefore difficult to observe in laboratory experiments but can nevertheless occur in nature for long contact times. Because of the anticipated reducing conditions in the far field of radioactive waste repositories in deep geological formations, actinide ions are usually present in their low oxidation states. Americium and curium, which contribute to the radiotoxicity of high level waste, occur exclusively in their trivalent state in aqueous solution. Even plutonium is expected to exist in the trivalent state in a strongly reducing environment. Because of these considerations, we place specific emphasis on actinides in their reduced state in this part of the review. 2.3.1. Trivalent Actinide Ions. It is not surprising that actinide interaction with dynamic solids, such as calcite, involves various sorption steps beyond pure surface adsorption. TRLFS studies with Cm as fluorescent probe added to a variety of solids revealed no simple classification in Cm(III) outersphere and inner-sphere surface complexes. The following systems were studied (maxima of fluorescence emission bands and fluorescence lifetimes are given in brackets): Cm/calcite (λem = 607.5 nm, τ = 314 μs), Cm/Th4P6O23 (λem = 606.7 nm, τ = 290 μs), Cm/Zr2O(PO4)2 (λem = 605.0 nm, τ = 260 μs), and Cm/amorphous aluminosilicate (λem = 606.8 nm, τ = 518 μs) (see also Table 1). All samples were prepared by adding Cm(III) to the respective solid phase. The Cm-species of interest were present after relative short equilibration times (i.e., within days). Due to the fast formation kinetics of these species they probably represent adsorbed species. Fluorescence emission spectra for the respective species, however, exhibit a significant red shift and clearly longer life times. Luminescence 1027


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lifetimes of 260−518 μs correspond to a loss of seven to eight water molecules in the actinide ion’s first coordination sphere, which is not the case for the inner-sphere surface complexes discussed above. However, the spectra of these Cm species are different from those typically found for Cm(III) incorporated into mineral phases without water molecules in its first hydration sphere (strong red shift to emission wavelengths >609 nm and emission lifetimes in the range of milliseconds), Therefore this sorption species can be considered as intermediate between surface sorption and incorporation into the mineral structure (solid solutions). This simplified view, however, must be viewed with care. Lifetimes may, for example, be influenced by type and arrangement of anions in the host phase ligand system. Curium incorporated into solid hydroxides such as brucite (Mg(OH)2) is assumed to be located at Mgsites and thus surrounded by 6 OH− ions in a distorted octahedral environment. Assuming that a single O−H oscillator in the first coordination sphere contributes half the measured energy transfer of a water molecule, 204 experimental fluorescence lifetimes measured for Mg(Cm)(OH)2 (τ = 165 μs; see Table 1) would correspond to exactly six OH− ions. Varying M−O distances and distorted crystal structures may complicate an exact quantification of energy transfer.205,209 Nevertheless, quenching entities, notably H2O and OH−, in the host lattice will certainly shorten fluorescence lifetimes of incorporated Cm(III) ions. Parenthetically, energy transfer mechanisms to MoO42− entities are also assumed to be responsible for the relatively short lifetime of Cm(III) fluorescence in powellite.206 Transition from inner-sphere surface complexation to a species considered as incorporated has been observed in sorption studies with gibbsite,70 Ca-montmorillonite,72 and silica.5,65 In these systems, the pH-dependent formation of incorporated Cm-species is observed to be a consequence of dissolution/reprecipitation events. Respective spectra exhibit long lifetimes corresponding to a complete loss of hydration sphere and, in the case of gibbsite, a very strong red shift (610 nm) typical of incorporated species. Curium incorporation in this case is most likely a consequence of exceeding solubility limits due to pH variation: precipitating gibbsite subsequently covers surface sorbed Cm with subsequent displacement of coordinated water. A similar process has been observed for Cm adsorbed onto Ca-montmorillonite at pH above 12. At high pH considered relevant for nuclear waste disposal conditions in the presence of corroding cementitious phases, formation of CSH phases is favored, which might incorporate or cover the adsorbed Cm inner-sphere complex. In case of silica, incorporation has also been concluded from TRLFS studies.5 However, long fluorescence lifetimes might as well be the consequence of the formation of ternary silicate surface complexes instead of mineralization65 (see discussion on inner-sphere surface complexation of trivalent actinide ions on clay minerals and quartz). Panak and co-workers207 performed a TRLFS study on Cm/ Am interaction with amorphous hydrous alumosilicates, which exist in natural groundwater as precursor to clay mineral formation. Two species were identified, which exhibit moderate spectroscopic shifts to 598.5 and 601.9 nm and short lifetimes of ∼85 μs, indicating substantial hydration of Cm(III) in the colloid structure. However, with increasing pH a third Cmspecies becomes predominant and is characterized by a fluorescence emission at 606.8 nm and τ = 518 ± 25 μs, pointing to the transition of a hydrated actinide species into an

incorporated colloidal actinide species, accompanied by complete loss of hydration sphere. The Ca(Sr)−Cm−CO3−SO4 system has been studied in depth, in order to understand interaction mechanisms on the molecular level.211−213,217,218 With a large set of sparsely soluble mineral phases this system allows systematic studies of solubility effects, ligand coordination strength, charge compensation, host phase structure, etc., on the actinide incorporation process. In coprecipitation experiments under carefully controlled oversaturation conditions Cm(III) was observed to form three different species upon interaction with calcite, two incorporated and one surface sorption species. Analysis of fine structures in the Eu(III) and Cm(III) luminescence emission spectra after selective excitation of individual sites reveals low symmetries for the sorption and for one incorporated species, but a well-defined trigonal symmetry (C3) for the second incorporated Eu/Cm species similar to the undisturbed lattice site of Ca in calcite (C3i). Because of the similar ionic radii substitution of Ca2+ ions by Cm3+ ions in the crystal lattice is possible.211,217 This is confirmed by EXAFS analysis of Am incorporated into calcite.219 Spectroscopic evidence for charge compensation by coupled substitution of two Ca2+ ions by one Cm3+ and one Na+ ion is reported211. Earlier studies on analogous trivalent lanthanide systems proposed a vacancy model for charge compensation in the solid solution with Eu2(CO3)3 and CaCO3 as endmembers,220 and another but less consistent model solid solution with EuOHCO3 and CaCO3 endmembers was also discussed. It is remarkable that incorporation occurs also when Cm is added to a calcite suspension under overall equilibrium conditions, reflecting the dynamics of the solid/liquid system and the thermodynamic stability of the respective solid solution.73 The effects of ligand strength and local symmetry on actinide incorporation were investigated in a comprehensive comparative study of coprecipitation experiments with the CaCO3 polymorph aragonite and gypsum (CaSO4 2H2O)212 and an analogous study comparing the isostructural Sr minerals strontianite (SrCO3) and celestite (SrSO4)218 as host matrices. The studies show that the strongly complexing carbonate anion favors the incorporation process. Cm(III) is only weakly associating with gypsum and celestite but attaching strongly to the carbonate phases. This behavior correlates with the solubility of the respective Cm(III) endmembers, which should be significantly higher for the sulfate solid than for the respective carbonate phase. In addition, the investigation of isostructural Sr mineral solid solutions yields clear evidence for a strong effect of the host lattice symmetry on the incorporation process.218 While association of Eu/Cm to gypsum occurs mainly by surface sorption, incorporation predominates for celestite. The different uptake mechanism is believed to be the consequence of different symmetries in both solids with different abilities to accommodate Eu/Cm in the lattice. A similar observation can be made based on the comparison of incorporation into three CaCO3 polymorphs calcite, aragonite, and vaterite, which show a dependence of the guest ion speciation on the host lattice symmetry.211−213 Lower lattice symmetry appears to favor the incorporation of Eu/Cm. Only one clearly defined incorporation species is observed in vaterite and aragonite, while calcite shows sorption species as well. Upon phase conversion of Cm(III) doped metastable vaterite to calcite Cm(III) remains incorporated. A calcite solid solution forms which is nearly identical to that formed by 1028


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coprecipitation. This finding again demonstrates the strong thermodynamic driving force toward formation of a Eu/Am/ Cm−calcite solid solution. The analysis of rare earth distributions in natural calcite samples leads to the same conclusion.221 Ca2+ containing minerals like calcite or calcium silicate hydrate (CSH) phases are ubiquitous in the geosphere or as secondary phase components of cementitious engineered barrier systems for the long-term storage of nuclear waste. Investigations with Cm(III) in Ca2+ minerals reveal an extraordinary red shift in the emission spectra in comparison to those typical for Cm(III) inner-sphere sorption species. Peak maxima are found between 609 and 624.4 nm. Emission lifetimes of these species are longer than 700 μs, indicating that the actinide ion lost its complete hydration shell and is incorporated into the mineral lattice within short time, with no evidence for surface sorbed species. An EXAFS study on the chemically analogous europium reaction confirms the assumption that Eu/Cm(III) is located at Ca-sites in the crystal lattices.222 Cm(III) interaction with a number of phosphates in aqueous suspensions revealed that significant incorporation is only observed for fluoroapatite.75 Cm(III) TRLFS revealed in this case narrow emission bands having spectra similar to synthesized samples with Cm(III) doped Ca 10 (PO 4 ) 6 F 2 samples and the solid separated from suspension after reaction with dissolved Cm(III). The authors conclude that Cm(III) ions exchange for Ca2+ ions in channels of the apatite structure. In summary, the findings discussed above emphasize Cacontaining minerals being favorable hosts for trivalent actinide/ lanthanide ion incorporation. Studies on Cm(III) coprecipitated with other phosphates revealed more than one species. In Th4P6O23, two major and one minor site (regarded as interstitial Cm(III) required for charge compensation) have been identified.75 Isomorphic substitution of La(III) by Cm(III) in LaPO4 leads to Cm(III) occupancy in four different sites, presumably because of disturbances in the lattice causing a deviation from the ideal structure. Also in this case a minor species is found, which is attributed to an interstitial monohydroxo-complex.216 In the context of natural analogue studies for nuclear waste repositories, iron minerals have long been discussed as possible host phases for actinide ions.223 In the vicinity of a nuclear waste repository, coupled redox/coprecipitation reactions are expected to occur. Because of the considerable masses of iron in container materials Fe(II/III) oxi/hydroxides are relevant chemical retention barriers.224 Even though compatibility of ionic radii of trivalent f-elements and Fe(II/III) is poor, spectroscopic evidence for the possible incorporation of trivalent actinides upon ferrihydrite alteration to goethite has been reported.63 Secondary phases, such as molybdates, clay minerals, etc., forming during the corrosion of high-level vitrified waste in contact with groundwater have been invoked as potential host phases for actinide and lanthanide incorporation or sorption.225 The formation of Cm containing solid solutions with powellite (CaMoO4) has been spectroscopically confirmed.226−228 Benchmark studies demonstrate that actinide ions in principle can be incorporated into clay minerals, such as the Mg-based smectite hectorite as well.210 The authors followed a synthesis route where the smectite structure is built up from Cm(III) coprecipitated in brucite, which is reacted with silica in order to form the typical TOT layered smectite structure. A recent

EXAFS study provided detailed information on the structure of Am(III) incorporated into hectorite showing it to be located in a slightly distorted MgO6 octahedral layer environment.64 In all EXAFS studies of dilute solid solutions actinide ions exist as minor components (FeOCO2UO2(H2O)3 or (>FeOH2)UO2CO3)) are assumed to form upon uranyl sorption onto the goethite surface in the presence of inorganic carbon solution. In calculations of these complexes, there were only four ligands in the equatorial plane with U−Owater bond distances from 2.32 to 2.54 Å. When two water molecules are replaced by one CO32− ion as ligand in the equatorial plane U− Osurface distances became significantly longer than the distances to the oxygen atom of the CO32− anion (2.65/2.32 Å). 3.3.7. Sorption of Uranyl at the Corundum (001) Surface. Both plane wave DFT317,320 as well as cluster calculations321 were used to study uranyl sorption onto the corundum (001) surface. At the corundum (001) surface, reactive sites are doubly coordinated aluminol groups or their deprotonated counterparts. Moskaleva et al.320 studied uranyl inner- and outer-sphere complexes at the corundum (001) surface (see Figure 17 and Table 4). A rough estimate of the

Figure 15. (a) Structure of an uranyl inner-sphere complex at the kaolinite (010) surface at the S0 termination. (b) Structure of an uranyl inner-sphere complex at the kaolinite (010) surface at the S1 termination. U, yellow; O, red; Al, blue; Si, green; H, white. This figure is only schematic and does not show the actual converged optimized structure.

different uranyl complexes at the surface with different U− Owater distances were found.375 3.3.6. Sorption of the Uranyl Ion onto Goethite. Sherman et al.319 investigated sorption of uranyl onto goethite in a joint theoretical/experimental approach. For the theoretical part of their study they used a minimal Fe2(OH)4(H2O)6 cluster model. At this cluster either edge- (E2) or corner- (C2) sharing sorption of uranyl is possible (see Figure 16). Edge-

Figure 17. (a) Structural arrangement of the uranyl ion in an innersphere complex at the corundum (001) surface and (b) in an outersphere complex at the corundum (001) surface. U, yellow; O, red; Al, blue; H, white. This figure is only schematic and does not show the actual converged optimized structure.

reaction enthalpy showed that formation of an inner-sphere complex is an endothermic reaction (ΔE = 78 kcal/mol), whereas a reaction energy of ΔE = −6 kcal/mol was found for the outer-sphere complex with its weak binding. The main uncertainty in this energy estimate is the theoretical determination of the protonation energy of a water molecule H2O → H3O+. In the outer-sphere complex UO22+(H2O)3(OH)2 the uranyl was calculated to be almost linear with an OylUOyl angle of 176° (inner-sphere surface complex: 149°). The water ligand pointing toward the surface has a U−Owater bond length of 2.30 Å, whereas the other two water molecules have U−Owater bond lengths of 2.59 and 2.68 Å. Glezakou and de Jong321 carried out a thorough investigation of inner- and outer-sphere complexes at this surface using a cluster model. They used stoichiometric clusters (Al2O3)n with n = 8, 11, 14, and 18, based on the detailed DFT investigation with PBC and plane-wave basis sets of Hinneman and

Figure 16. (a) Structure of a corner-sharing uranyl inner-sphere complex at the goethite (101) surface. (b) Structure of an uranyl edgesharing inner-sphere complex at the goethite (201) and (010) surface. U, yellow; O, red; Fe, violet; H, white. This figure is only schematic and does not show the actual converged optimized structure.

sharing was only observed at the (210) and (010) surface, whereas C2 type sorption occurred at the dominant (101) surface. Note that we kept the notation for the Miller indices used in the original papers. For both sorption sites, bidentate inner-sphere uranyl complexes retain its 5-fold coordination. Both U−Owater and U−Osurface distances were nearly the same for both the E2 and the C2 case (Table 4). When H2O molecules were replaced by OH− in the UO22+(H2 O)3 1040


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+ 12H2O system, where six water molecules comprise the first hydration shell and six further water molecules were placed on the top of the cluster in order to simulate surface hydration. Different DFT functionals were benchmarked with MP2 calculations, because DFT lacks the description of dispersion interactions379 and respective benchmark studies are not available in the literature. The outcome of this investigation was that structures optimized by DFT calculations using the BP86 functional and the cc-pVDZ basis set380 were very close to those obtained from MP2. This was taken as a validation of DFT as an appropriate theoretical tool for this type of studies. Trivalent actinide/lanthanide ions were found to form stable tridentate inner-sphere complexes at the corundum (001) surface with six water molecules remaining in the first coordination shell (see Figure 18a). The most stable complex was formed when one or two surface protons were removed. Structural information is summarized in Table 5. Distances of the three Cm−O bonds to the surface varied considerably for all investigated complexes. Hence the complexes were always tilted on the surface. The overall average distance R between the metal ion and all nine surrounding oxygen atoms was at 2.64 Å, which is somewhat longer than EXAFS analysis results for the respective lanthanide complexes.67 Energy calculations resulted in many local minima for surface complex structures with only marginal differences, all populated in thermal equilibrium.88 In this respect, theoretical calculations support the interpretation of relative high Debye−Waller factors obtained in EXAFS and broad emission bands in TRLFS for surface complexes (see discussion in section 2.2.1). Whether covalency plays a role in oxygen-actinide bonds is currently discussed in the literature.372 A simple attempt was made to estimate the role of covalency in the Cm-O bonds of the inner-sphere complex of Cm at the corundum (001) surface by localizing the molecular orbitals and calculating the occupation numbers of all the orbitals. The orbitals of all nine Cm−O bonds were found to have an occupation number of ∼1.8 originating from the oxygen atom and ∼0.2 from the curium ion. Hence the binding character is not purely ionic; there is a small covalent contribution. If this is a significant sign of a covalent binding must be confirmed by further theoretical calculations, as put forward by Kaltsoyannis,372 as well as experimentally. 3.4.2. Sorption of Trivalent Curium Ions at the Corundum (110) Surface. For investigating Cm(III) sorption onto the corundum (110) surface322,323 an Al27O75H67 cluster was used. Reactive sites are either singly, doubly or triply coordinated aluminol groups or their deprotonated counterparts. The system finally selected to study inner-sphere complexes at the hydrated corundum (110) surface was Al27O75H67 + Cm3+ + 12H2O. At the (110) surface tetradentate inner-sphere complexes with five water molecules in the first coordination shell were found (see Figure 18b). The Cm−Os (surface oxygen) distances obtained were 2.37, 2.50, 2.59, and 2.92 Å, where the shortest distance is that between the metal ion and the deprotonated aluminol group. Cm−Ow distances lie between 2.50 and 2.75 Å. The average values obtained are listed in Table 5. For inner-sphere complexes at the corundum (001) and (110) surfaces, variations of the individual Cm−Ow distances are ±0.18 and ±0.25 Å, respectively. Both theoretical studies and TRLFS studies performed by Rabung et al.68 observed a similar trend for the binding strength of Ln/An(III) corundum (001) and (110) surface complexes. Calculations show that the

Carter.378 To model a fully hydroxylated surface, a layer of protons was added on top of the surface. With the cluster model they were able to examine both charged and neutral systems. The authors obtained structures of free and sorbed (UO2(OH)m(H2O)n)q and studied their dependence on the charge q and the number of water molecules n in the first and second coordination shell. The largest deviation from linearity of the uranyl moiety was found for the sorption of the bare uranyl cation (without water ligands) at the fully hydroxylated surface, OylUOyl angle of 166° (Table 4). Adding three water molecules to the first coordination shell increased the OylUOyl angle to 176° (nearly linear). In this complex the U−Owater bond distances were 2.48−2.51 Å. The most stable inner-sphere complex was found with three water molecules in the first coordination shell and two in the second coordination shell, an OylUOyl angle of 168°, and U−Owater bond distances of 2.53−2.61 Å. Only minor changes were observed compared with gas phase calculations. In the most stable outersphere complex, the uranium atom is located at a distance of 4.64 and 4.90 Å above the closest surface oxygen atom. A consideration of binding energies by the authors321 also revealed that inner-sphere complexes are energetically favored over the outer-sphere complexes at the corundum (001) surface. Both the DFT with PBC and cluster ansatz model strategies to describe sorption of uranyl ions onto the corundum (001) surface yield conflicting results concerning (1) which interatomic distance is longer, the distance to the water ligands U−Owater or the distance to the surface U−Osurface and (2) the energetically most stable species. Concerning the bond distances, one possible reason for the conflicting results may be the different theoretical methods applied. The reason for the differences in the energy estimates may simply lie in the larger uncertainties associated with energy predictions as mentioned above. 3.4. Sorption of Trivalent Curium

3.4.1. Sorption of Trivalent Curium Ions at the Corundum (001) Surface. Sorption of trivalent lanthanides and actinides onto the corundum (001) surface was studied by Polly et al.87,88 using an Al31O60H21 cluster as model for the hydroxylated corundum (001) surface (see Figure 18a). Based

Figure 18. (a) Structure of a tridentate curium inner-sphere complex at the corundum (001) surface and (b) of a tetraentate curium innersphere complex at the corundum (110) surface. Cm, yellowl; O, red, Al, blue, H, white.

on earlier spectroscopic studies of Rabung et al.68 it was assumed that 5 ± 1 H2O molecules remain in the first coordination sphere of the trivalent lanthanide/actinide cation upon sorption to corundum. The final model to describe the formation of actinide inner-sphere complexes at the hydrated corundum (001) surface was therefore the Al31O60H21 + Cm3+ 1041


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Table 5. Optimized Bond Distances of the Inner-Sphere Cm3+ Complexes at the Corundum (001) and (110) Surface in Åa surface (001) (110)

system 3+

Al31O60H20 + Cm + 12H2O Al27O75H66 + Cm3+ + 12H2O

complex

Cm−Os

rs

Cm3+−Ow

rw

R

tri-IS tetra-IS

2.42, 2.64, 2.96 2.37, 2.50, 2.59, 2.92

2.67 2.60

2.53, 2.59, 2.60, 2.64, 2.69, 2.71 2.50, 2.51, 2.52, 2.61, 2.75

2.63 2.58

2.64 2.59

a

tri-IS, tetra-IS = tridentate, tetradentate inner-sphere complex; OS = outer-sphere complex; rs,w = average binding distances between curium and surface and water O atoms, respectively; R = average overall binding distance for all Cm−O bonds.

Ca2UO2(CO3)3 at the calcite (101̅4) (natural cleavage plane) surface and at the acute or obtuse steps having (314̅8) and (312̅ 1̅ 6) surfaces, respectively. Several different inner-sphere species were found at the (101̅4) and (314̅8) calcite faces. At the obtuse (31̅2̅16) surface they found possible outer-sphere complex formation. Using free energy considerations, they found that inner-sphere complexes at this surface are energetically less favorable. Formation of an inner-sphere complex requires surmounting a free energy barrier associated with displacement of water molecules to form new interactions to the carbonate ions of the surface. In summary, the authors found a general preference of outer-sphere complexes over inner-sphere complexes at the investigated calcite surfaces. 3.5.5. Sorption of Uranyl on 2:1 Dioctahedral Clays. Greathouse et al. reported a number of classical MD and MC studies of uranyl sorption onto 2:1 dioctahedral clay minerals, focusing their investigations on montmorillonite,386,386−388 but also studied beidellite and pyrophyllite.389 With classical MC and MD, they determined that uranyl sorbed in the interlayer of montmorillonite with negative charge sites in the octahedral sheet.386 In these outer-sphere complexes, uranyl retains five water molecules in its equatorial plane and the uranyl molecular axis is oriented normal to the interlayer surfaces. In a further study combining MC simulations with experiment and considering montmorillonite with charged sites in both the octahedral and tetrahedral sheets, they found outer-sphere complexes with the OylUOyl molecular axis tilted ∼45° to the surface.386,387 In another study388 focused on montmorrillonite in the presence of sodium counterions and carbonato ligands, the main result was a significant decrease of uranyl sorption upon increase of UO2CO3 species, while sodium sorption remained constant. In a systematic MD simulation study of aqueous uranyl adsorption onto montmorillonite, pyrophyllite, and beidellite, the same authors389 focused on atomistic details to help explaining experimental findings. They found that at low uranyl concentrations, the pentaaquouranyl complex dominates in solution, which readily adsorbs to the clay basal plane. At higher uranyl and carbonato concentrations, the mono(carbonato) complex forms in solution and the amount of uranyl sorption decreases. The uranyl surface complexes are dominated by the pentaaquo cation.

number of bonds between the metal ion and the surface increases when changing from the (001) to the (110) surface. Rabung et al.68 reported n(001) = 5 ± 1 water molecules in the Cm(III) first coordination sphere at the (001) and only n110 = 3 ± 1 at the (110) surface, thus indirectly pointing to a larger number of binding interactions to the (110) plane. Even though the quantum chemical calculations are not in full agreement with spectroscopic measurements, both studies show a consistent trend. 3.5. Classical Molecular Dynamics (MD) and Classical Monte Carlo (MC) Simulations

Both classical molecular dynamics (MD) and classical Monte Carlo (MC) simulations rely on various classical force fields. Various studies of actinide sorption in this field are summarized here in brief, focusing on the sorption of uranyl onto different minerals, nanosized fractures in feldspar,381 quartz,382,383 goethite, 384 calcite, 385 and various 2:1 dioctahedral clays.386,386−389 In these studies, the uranyl force field of Guilbaud and Wipff390 was used. Only Steele et al.384 developed their own interaction potential for uranyl. 3.5.1. Sorption of Uranyl Carbonate in Feldspar Nanosized Fractures. Kerisit and Liu381 presented MD simulations of diffusion and sorption of Ca2UO2(CO3)3, UO22+, CO32−, and UO2CO3 in nanosized fractures of feldspar, which is of great relevance at the Hanford site in Washington state. They showed in their calculation that the presence of feldspar considerably reduces the diffusion coefficient of these species. 3.5.2. Sorption of Uranyl on Quartz. Two different groups investigated this problem, Greathouse et al.382 and Boily and Rosso.383 Greathouse et al.382 studied the structure and dynamics of the uranyl ion and its aquo, hydroxyl and carbonato complexes near the hydrated quartz (010) surface. For uranyl, they found outer-sphere complexes formed at the singly protonated surface, whereas inner-sphere complexes formed at the deprotonated surface. In all cases, the uranyl maintained its 5-fold coordination. Boily and Rosso383 studied sorption of the neutral UO2(OH)2 species at the (001), (010), and (101) surfaces of quartz. At the (001) surface they identified mononuclear monodentate complexes. At both the other surfaces binuclear bidentate inner-sphere complexes were observed. Furthermore, they found four water molecules coordinated to uranyl in the equatorial plane at the (001) and (010) surfaces and a 5-fold coordination at the (101) surface. The authors also reported that the (010) surface can stabilize more species than the other two surfaces. 3.5.3. Sorption of Uranyl on Goethite. Steele et al.384 used atomistic modeling methods to predict uranyl species at different goethite surfaces. On the basis of agreement of calculations with the experimentally determined EXAFS Fe−U distance of 3.5 Å, they suggest that uranyl species are likely to be bound to the hydrated (110) and (001) faces of goethite. 3.5.4. Sorption of Uranyl on Calcite. Doudou et al.385 performed classical MD simulations at 10 K to determine the structures of inner- and outer-sphere complexes of

4. GEOCHEMICAL MODELING OF ACTINIDE SORPTION Traditionally distribution coefficients have been used to describe sorption of actinides and their retention on mineral surfaces in migration processes, as well as to assess their mobility and propagation in environmental systems from a given source. Distribution coefficients correspond to the overall uptake of solutes in contact with a given mass of minerals, soil or rock of interest, that is, KD = q/c 1042

(2) dx.doi.org/10.1021/cr300370h | Chem. Rev. 2013, 113, 1016−1062

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Table 6. Averaged Logarithmic Values for (Gaines−Thomas and Vanselov) Selectivity Coefficients for Na−Illite42,399 and Na− Montmorillonite41,400 exchange pair

illite (Gaines−Thomas, average values for a range of ionic strength)

Na+/M+ Na+/M2+

0.0 (H+, NpO2+), 1.1 (K+) 0.7 (UO2+), 1.0 (Mg2+, Ca2+), 1.1 (Ni2+), 1.3 (Co2+) 1.0 (Al3+), 1.9 (Eu3+, Am3+)

Na+/M3+

montmorillonite41 (Gaines−Thomas, average values for a range of ionic strength) −0.1 (NpO2+) 0.1 (UO2+), 0.4 (Cd2+), 0.5 (Ni2+), 0.6 (Zn2+, Co2+, Pb2+) 2.2 (Eu3+, Am3+)

where KD is the distribution coefficient (in dm−3 kg), q is the amount of solute adsorbed per mass of mineral (in mol kg−1), and c is the total dissolved equilibrium concentration of the solute (in mol dm −3 ). These distribution coefficients correspond to linear and reversible adsorption isotherms and are rather easily implemented in codes that describe the migration of a selected solute through a porous medium for example. They are neither linked to a retention mechanism, nor do they necessarily consider solution and surface speciation explicitly. Both these aspects can be resolved mechanistically, as shown in another contribution to this issue (solution speciation) and the previous sections (surface speciation). Furthermore, many other (nonlinear) isotherms exist391 that rule out the application of a constant distribution coefficient in migration calculations. We later show in an illustrative way how the use of single distribution coefficients may lead to significant error in performance assessment calculations. In principle, the drawbacks of the distribution coefficient concept are overcome by using appropriate adsorption models (and here we will focus on ion exchange and surface complexation/precipitation models). As discussed in previous sections, various general phenomena contribute to sorption from aqueous solutions. For all these phenomena, that is, ion-exchange, surface complexation, and absorption/incorporation (including surface precipitation, coprecipitation) models are required that go beyond simple distribution coefficients. Such model approaches will be addressed in the following sections.

montmorillonite400 (Vanselov, reduced to zero ionic strength) 0.0 (Li+), 0.1 (H+), 0.3 (K+) 0.1 (Cu2+), 0.2 (Mg2+, Ca2+, Sr2+, Zn2+, Cd2+, Co2+), 0.3 (Ni2+) NA

exchange will interfere and has to be separated from other contributions. Common ways to summarize ion-exchange equilibria are via Vanselow coefficients395 or using the Gaines−Thomas formalism.396 Vanselow coefficients are calculated according to the following equation K v = xM × γNa 2[Na +]2 × x Na −2 × γM −1[M2 +]−1

(7)

where xi = {X i} × ({X 2M} + {XNa})−1

(8)

Thus from the formalism based on the simple mass law equation the corresponding Vanselow coefficient can be obtained for M2+−Na+ exchange by the relation K v = K 2 × ({X 2M} + {XNa})

(9)

The Gaines−Thomas formalism, which has been used in the model by Bradbury and Baeyens,71,397,398 involves formulation of an exchange reaction for a trivalent actinide (An3+) in the same way 3NaX + An 3 + = AnX3 + 3Na +

(10)

K GT,An/Na = fAn × γNa 3[Na +]3 × fNa −3 × γAn−1[An 3 +]−1 (11)

Here, f j are the equivalent fractions defined as the equivalents of j sorbed per unit mass divided by the cation exchange capacity (CEC), that is, fAn = 3{X3An} × (CEC)−1 and fNa = {XNa} × (CEC)−1

4.1. Approaches to Ion-Exchange Modeling

(12)

Ion exchange models have long been in use in various disciplines. With respect to actinides they have been important in describing the ionic strength dependent uptake behavior of clays and clay minerals. Models that must cover a broad range of pH, need to include ion-exchange and surface complexation mechanisms. Probably the first such model was designed for divalent metal adsorption onto kaolinite and montmorillonite.392,393 The models were implemented in the speciation formalism394 considering balances and writing chemical equilibria and mass law equations: XNa + H+ = XH + Na +

(3)

K1 = {XH} × γNa[Na +] × {XNa}−1 × γH−1[H+]−1

(4)

2XNa + M2 + = X 2M + 2Na +

(5)

For illite and montmorillonite, both in sodium form, Table 6 lists a number of published selectivity coefficients. Consistent data sets for ion exchange parameters appear to be scarce. Theoretically the selectivity coefficients should primarily depend on charge, but in reality this is not the case. For example the relatively small selectivity coefficients for UO22+ as compared to other divalent cations might be explained by steric hindrances of the linear dioxo cation. The values listed in Table 6 show trends with ion charge, which is expected for a largely electrostatically controlled process. However, there appear to be different trends for the two montmorillonite samples. The ion exchange approach has also been used to model radionuclide interaction with oxides401 and carbonates.402 However, for these two classes of mineral surface complexation models are more widely used.142,187,403−408 These models are the most prominent ones and will be discussed in detail in the following section.

K 2 = {X 2M} × γNa 2[Na +]2 × {XNa}−2 × γM −1[M2 +]−1

4.2. Approaches to Surface Complexation Modeling (6)

Surface complexation (thermodynamic sorption) models have been extensively reviewed and described.409−418 While they were originally designed by inorganic and environmental chemists, nowadays they are largely implemented in geochemical disciplines (the same is actually true for solution speciation). Surface complexation models treat adsorption

X is an exchanger site and curly brackets denote mass specific concentrations of surface species in mmol kg−1. γi are aqueous solution activity coefficients, and [i] are concentrations in mol dm−3of aqueous species. There is no surface charge development due to ion exchange. However, in a titration the proton 1043


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Figure 19. (A) Consecutive steps in the development of a surface complexation model. Model decision, input data, and model parameters. Various protonation schemes and electrostatic models are available: (E)CCM, (extended) constant capacitance model; DLM, diffuse layer model; BSM, basic Stern model; (E)TLM, (extended) triple layer model; TPM, three plane model; FLM, four layer model. Omission of electrostatic factors yields the nonelectrostatic model (NEM). Charge Distribution (CD) requires multiplane adsorption models (see panel B): In panel A, the charge of adsorbed Na+ would be distributed, in panel B it would be placed as a point charge in the 1-plane. PZNPC, point of zero net proton charge; IEP, isoelectric point; log Ki°, equilibrium constants; Ci, capactitances (see panel B); lζ, slip plane distance (see panel B). (B) Illustrative sketch of a typical electrostatic interfacial model (three-plane model, TPM). Ψo, potential in the o-plane (often called surface potential); σo, absolute surface charge density (net charge) in the o-plane; Ψ1 and σ1, potential and absolute charge density in the 1-plane; Ψd and σ2, potential and absolute charge density in the 2-plane; σd, diffuse layer charge density; C1 and C2, capactitances; ζ, zeta-potential; lζ, slip plane distance; ΨI, potential at a distance “l” from the 2-plane. R, T, F, and IC are gas constant, absolute termperature, Faraday’s constant, and ionic strength in molar units, respectively, ε = εo × εr (εo = permittivity of vacuum and εr = dielectric constant of solution).

processes similar to aqueous speciation schemes. By defining surface functional groups as surface ligands that interact with solutes to form surface complexes and taking into consideration, where necessary, variable surface charge, the usual mass law equation and mole balance formalism can be applied.394 Gibbs energy minimization approaches also exist419 but are not as widely used. The speciation based models allow prediction of changes in the distribution coefficient of a certain element in contact with a given sorbent as a function of pH, ionic strength, concentration of other complexing solutes (in solution or at the

surface) or competing solutes for relevant adsorption sites, and other variables. The most advanced models achieve this at a level of sophistication that includes details of the surface (in terms of structure), takes into account the structure of the electrical interfacial layer, and constrains structure and bonding of the surface complexes by information from spectroscopic studies or theoretical calculations that have been reviewed in much detail in the previous sections. Experimental studies illustrate the complexity of even well-defined single crystal surface structures, which may actually consist of two domains 1044


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(e.g., of hematite (001)420). Some observations concerning sorption on single crystal samples like extended outer-sphere complexes that differ from sorption complexes on powders,153 as mentioned before, may result from physical properties inherent to the single crystals (such as smoothness, contrasting rough particle surfaces).421 Inclusion of detailed mechanistic knowledge into adsorption models increases confidence in their reliability, for example, when used in predicting migration of actinides from a waste disposal site to the biosphere. Unlike solution speciation, where substantial efforts are made to obtain sound thermodynamic data for actinides and other elements relevant to nuclear waste disposal401 in aqueous solution, the lack of consistent surface complexation model parameters for multicomponent systems presently hampers comprehensive implementation of sorption phenomena in performance assessment. Various studies have been initiated by the NEA/OECD to establish a basis for sorption modeling. 401,422,423 While the respective reports401,422,423 illustrate the state-of-the-art of the modeling and the progress made from the use of distribution coefficients to surface complexation models, consistent application of the latter is by far not as advanced as is the case for solution speciation and solubility. A standard sorption model has yet to be agreed upon, which would be highly beneficial and help establish consistent and comprehensive databases, as is common practice for aqueous solutions and solubility. The situation is complicated by the fact that the variety of different surface complexation models is overwhelming. In the following we will give a short overview on the many possible available options for surface complexation modeling. We will divide the surface complexation approach into its major components and discuss these one by one. Surface complexation models involve writing chemical reactions for the formation of surface species in the same way as for aqueous solutions. An example for such a surface chemical reaction representing in a generic form the adsorption of a solute (Ann+) to a surface functional group (SOH) is given by the following equation, which is obtained from eq 1 for x = 0 and neglecting hydration waters SOH + An n + ⇔ SOAn−1 + n + H+

Among these, it is further possible to distinguish two different types of triply coordinated sites, for example, on the (110) face. A full model version would include all these sites. The multisite-complexation (MUSIC) model426−428 permits estimation of protonation constants for all of these different oxygen atoms exposed to solution.429 A possible simplification of a full model results from the outcome of these estimates, showing in this case that only one pH-dependent protonation reaction occurs in the usual pH-range. Addition of a pH-silent site, bearing a charge, is required to describe the overall experimental surface charge.430 The general concept of the MUSIC approach has been experimentally verified. For example, distinct differences between crystal planes of the same mineral have been observed experimentally68,76 and experimental results currently distinguish between different reactive sites on colloidal particles.431 Such results can elegantly be characterized by advanced theoretical methods432 illustrated in the previous section. Even effects that may become relevant for very small particles 433 can in principle be included. Figure 19A summarizes early decisions within surface complexation models. Once the types of sites to be considered in the model are decided on, the respective site densities have to be determined. For mechanistic sites, this can be accomplished from surface structure. For generic sites this is necessarily arbitrary and introduces the first severe complication in relation to solution studies, where analytical determination of all total component concentrations is usually possible. While for surfaces this has often been done by saturation experiments, it was shown that the potentiometric titrations used to this end434−436 are not suitable.437 Since the distinction between generic and mechanistic models for laboratory scale particles has been discussed in some detail, it is interesting to see that on the regional scale a similar distinction in modeling approaches exists.438 The analogue for the generic model is the so-called generalized composite approach (GCA), where a sediment or a mineral assemblage are represented by generic sites. In this kind of modeling, the site densities have to be estimated, which is usually done in various ways (all are necessarily rather arbitrary though they may rely on previous work). Acid−base properties also are usually not determined because of experimental problems.151 The model then more or less consists of actinide adsorption reactions (to be fitted to experimental adsorption data) and aqueous speciation (from tabulated thermodynamic data) but often is rather successful in describing experimental observations and predicting migration experiments. Since the acid−base properties are not always included in GCA approaches, transport of protons or hydroxide ions in the concomitant migration experiments may not be well-described, but in many environmental scenarios the systems are to some extent buffered. The mechanistic (MUSIC) analogue would be the so-called component additivity approach (CAA). In this approach the mineral composition of interest is determined and known relevant surface complexation parameters for the mineral components are used in a predictive way in simulations. Comparisons of the approaches to various systems exist.151,190,438−440 The option of using the expected dominant sorbent in a complex sorbent has been successfully applied.183,441−443 This short overview shows that there are various, often equally successful, applications of surface complexation models to handle complex situations. 4.2.2. Protonation Mechanism and Electrostatics. The second model component is related to surface charge. As

(13)

where SOAn−1+n is the actinide surface complex formed. The use of such chemical reactions and associated mass law equations for reactions on oxide minerals started in the second half of the previous century.424 In the next sections an overview will be given on how to design a surface complexation model following Figure 19A. 4.2.1. Treatment of Heterogeneity and Definition of Surface Sites. Use of an adsorption equation like eq 13 within the speciation formalism requires definition of the surface site in terms of type and number. Surface sites can be treated as generic surface sites, when either nothing is known about the type of sites on a given sorbent (e.g., when the surfaces heterogeneity cannot be mechanistically resolved) or the use of “mechanistic” sites is not possible for some other reason. Mechanistic sites in turn might be defined based on knowledge of surface structure. Both the presence of various sites on a given crystal plane and the occurrence of various crystal planes on a given particle introduce site heterogeneity even on welldefined particles. For goethite, for example, structural considerations allow to distinguish as many as three different hydroxyl groups in general terms. They differ in coordination of the surface oxygen to either one, two or three iron atoms.425 1045


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fundamental process on mineral surfaces. It is experimentally quantified by (i) the proton related surface charge density (typically via potentiometric surface titrations444,445) or (ii) the surface, diffuse layer, or zeta-potentials measured via singlecrystal electrodes, force distance curves, or electrokinetic methods, respectively.462 For minerals like calcite, constituent ions also affect isoelectric points (where the zeta-potential is zero).463 The action of charge determining ions and the resulting surface charge and potentials affect all surface reactions and a model for the electrostatics at the interface is required. Mineral surface titration curves cannot be described by a single surface site with one or two pK values445 unlike weak-acids, such as acetate. This was realized early on and in principle at least two limiting solutions to the problem exist: (i) Inclusion of many (i.e., a sufficient number of different) sites with their distinct pK values397 or distribution of values450,464 and (ii) use of one site with one or two discrete pK values with some “electrostatic” correction term.434,449 The first approach is hardly ever followed and the second one unfortunately has been followed in many different ways. Inclusion of electrostatic terms causes shifts of the 50:50 species distribution at the pKa value if (i) several sites with various protonation properties are involved in the speciation scheme or if (ii) for a 1-site model fractional charges are asymmetric. The surface complexation formalism involves the definition of so-called “intrinsic” stability (int) constants, which are valid for zero surface potential and operational, “apparent” stability (app) constants, which change, for example, with pH. The latter are defined as the product of the intrinsic constants (that are real constants) and an electrostatic factor (P) that accounts for Kapp change (e.g., with pH).

shown in Figure 19A, the basic charge component of the models has two subcomponents, protonation mechanism and electrostatics. The basic charging is due to interactions of the surface hydroxyls with the charge-determining ions (proton and hydroxide ion for oxide minerals). The various surface species create a net charge that can be directly observed by electrokinetics and is usually experimentally characterized in a simple monovalent electrolyte solution.444 It depends on pH and ionic strength and is described by surface chemical reactions, such as SOH + H+ ⇔ SOH 2+, K+

(14)

SOH ⇔ SO− + H+, K −

(15)

and concomitant mass law equations. The above two reactions generically symbolize protonation and deprotonation of mineral surfaces. They are part of the very first surface complexation models,434,445−447 which would nowadays be termed 1-site-2-pK models. Because early on no mechanistic sites were considered, a generic site was involved that could occur in three protonation states. Based on the above two reactions or their sum (see Figure 19) there is a specific pH value at which the surface is uncharged, the so-called point of zero net proton charge (pznpc). In the 2-pK-model, the pznpc is given by pH pznpc =

1 (log K+ − log K −) 2

(16)

Sometimes the more general term point of zero charge is used but a precise definition is required for the various notions of points of zero charge to avoid confusion. Since the occurrence of three protonation states on one oxygen atom within a narrow pH-range is not realistic, the socalled 1-pK concept was introduced.448−450 In its original form it might be referred to as the generic form of the MUSIC model developed later426,428 SOH−1/2 + H+ ⇔ ≡ SOH 2+1/2 , 10−pHpznpc

K+,app = [SOH 2+] × [SOH]−1 × [H+]−1

(18)

K+,app = K+,int × P

(19)

If the electrostatic factor is unity the apparent and intrinsic constants are identical. This (i.e., P = 1) occurs at the point of zero charge in the most common cases. Introduction of an electrostatic factor accounts for the change of the apparent constants with variable charge. A very simple model would be the Nernst model (that considers the difference from the pznpc to obtain a surface potential),465 or the so-called constant capacitance model,434 which involves a linear relationship between (measurable) proton related surface charge density and the (model-dependent) surface potential

(17)

For the 1-pK approach, the logarithm of the equilibrium constant is equal to the pznpc. Being able to estimate451−454 or measure the pznpc has the considerable advantage of avoiding determination of two equilibrium constants required in the 2pK analogue. The situation is more complex as soon as electrolyte binding to the charged sites is involved455 or whenever multisite models are used.429 Nowadays both 1-pK and 2-pK approaches are used. Although the 1-pK mechanism is often thought to be related to the fractional charge of +0.5 and −0.5, within the MUSIC approach any fractional charge can be introduced. This was initially accounted for by Pauling bond valence57 but is nowadays done by the Brown−Altermatt approach.428,456 Within the same framework, protonation constants of individual surface functional groups can be estimated.426,428,457−459 Although the approach has been criticized,460 it is usually very successful in a predictive sense and a clear improvement over the previously used generic models. Both 1-pK and 2-pK reactions occur in the general MUSIC framework.426,428 However, the 2-pK variant within the MUSIC model will result in huge differences in pK values (about 10 pK units), while in the classical generic 2-pK models these are fitting parameters and may be quite close to each other.455,461 The basic charging behavior at the solid-solution interface is part of a multicomponent equilibrium system and a

(20)

σ=C×Ψ

Surface charge density σ (in C m−2) can be computed from molar concentrations of the charged surface species, that is, [ SOH2+] − [SO−] for the case discussed above, multiplied by a factor that relates molar concentration to surface specific units and Faraday’s constant. With a known capacitance C (in F m−2), the surface potential Ψ is retrieved and the electrostatic factor (P) P = exp( −Δz × F × Ψ × R−1 × T −1)

(21)

is obtained that links an apparent operational constant to the corresponding intrinsic one. In eq 21, F is Faraday’s constant, R is the ideal gas constant, and T is absolute temperature. In the calculation of the electrostatic factor, Δz is the amount of charge transferred in a given reaction, that is, Δz = 1 for the formation of SOH2+ from SOH. In this case, Δz equals 1046


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basic Stern model, BSM;479 the triple layer model and extensions thereof, TLM,480 and ETLM;481,482, the three plane model,369 TPM (shown as an example in Figure 19B in great detail), which is similar to but in its numerical implementation not equal to the TLM; and the four layer model, FLM.483 Other variants of variable charge models exist,484−486 and comparisons between the models are available.473,479,487 The different models may include varying numbers of planes of adsorption and require an appropriate number of charge − potential - relationships to solve the resulting mathematical problem (see Figure 19B for the TPM). Charges in the planes of adsorption are calculated from balances taking into account the absolute charge of the various surface complexes. When required, overall neutrality of the interfacial region is included and finally a set of equations is obtained that corresponds in number to the unknowns and the system can be numerically solved.394 Figure 20 shows an example for the surface charging of hydrous ferric oxide (HFO) obtained from a generic 2-site diffuse layer model 22 and from a recent CD-MUSIC approach.488,489

the charge of the surface species formed. The capacitance value C is unknown and usually obtained from fitting the model to experimental data. The electrostatic factor causes K+,app to decrease with increasing positive potential (i.e., increasing  SOH2+ concentration causes an increase in σ and a decrease in P). In 1-pK models the treatment is different since charge transferred and net charge of the surface species formed differ, that is, for SOH2+1/2 arising from SOH−1/2 Δz = 1, but the net surface charge is calculated from 1/2 × ([SOH2+1/2] − [SOH−1/2]).411,466−468 The constant capacitance approach suits the constant medium approach, in which parameters are valid for a given electrolyte composition and concentration.469,470 The introduction of additional interfacial planes (Figure 19B, cf below) allows distinction between proton adsorption (plane of the surface hydroxyls, o-plane), and medium ions (assumed to form ion pairs with charged surface hydroxyls). The model shown in Figure 19B also permits description of the measurable zetapotential, which pertains to the slip plane, that is, the potential at the stagnant layer of water molecules and ions that moves with the particles in an applied electric field. Modeling the zetapotential typically involves either assuming that it occurs at the onset of the diffuse layer (i.e., ζ = Ψd) or adding an adjustable parameter (the slip plane distance lζ) that relates the surface potential at a distance l (Ψl) to Ψd by the equation given on Figure 19B. This latter is more successful in describing experimental data463,471,472 but yields excessive lζ at lower ionic strength. Figure 19B shows a sketch of the electrical interfacial layer (EIL) in terms of a three plane model (TPM).473 In electrokinetics, a stagnant layer exists that moves with a particle in an applied electrostatic field.474,475 Charge of adsorbed entities can be placed in the plane of the surface hydroxyls (o-plane, subject to the surface potential Ψo) and contributes to the surface charge density σo. Bound counterions are placed further away from the surface either in terms of point charge, (as in Figure 19B, where Na+ in the 1plane is subject to the potential Ψ1, and contributes to the net charge density σ1). Cl− is fixed in the 2-plane that coincides with the beginning of the diffuse layer. Two parallel capacitors (C1 and C2) are introduced to obtain required charge (σi) potential (Ψi) relationships to solve the system of equations. The electroneutrality condition involves the sum of the actual charges in all planes of adsorption plus the diffuse layer charge that is calculated from the Gouy−Chapman equation for z/z electrolytes. Water dipole effects can be included and the geometry of the particles can be taken into account. In many codes, plane geometry is typically implemented, that is, the Gouy−Chapman equation (given in Figure 19B). It is valid for z/z background electrolytes, the concentration of which defines the overall ionic strength and is not affected by other components in the system (such as protons, hydroxide ions, or other solution components). Since the introduction of surface complexation models many different electrostatic models have been proposed and used, the acronyms of which appear in Figure 19A. They encompass the nonelectrostatic model (NEM), which is within the so-called 2SPNE SC/CE mentioned earlier successfully applied to clays,41,42,131 but not for oxide-minerals;476 the Nernst model;465,477 the constant capacitance model and its extension used in the constant medium approach, CCM434 and ECCM;478 the purely diffuse layer model, DLM;446,447 the

Figure 20. Calculated HFO surface charge density (σo) as a function of pH at 1 mmol dm−3 (full line) and 100 mmol dm−3 (dashed line) salt concentration obtained with a 2-site diffuse layer model (DLM, red)22 and a recent CD-MUSIC approach (black).488.

While both models agree well at low salt content and within 3 pH-units of the pznpc, there are strong differences beyond these conditions. These differences have repercussions on modeling actinide adsorption to HFO in that, for example, the resulting electrostatic factors from the two models will be different. 4.2.3. Actinide Adsorption Model. Once a model for basic charging (as in Figure 20) is obtained, the next step is to analyze adsorption of the solute of interest (Figure 19). This requires writing appropriate, balanced chemical reactions in terms of the surface sites involved (resulting in mono- or multidentate surface complexes, as in a recent modeling example490), using the correct coefficient for the actinide (i.e., >1 for polynuclear surface complexes) and the appropriate proton stoichiometry. If the CD-concept is used, CD-factors are required. These correlate with stability constants and affect 1047


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proton stoichiometry. Unfortunately, specific experimental proton stoichiometry435 data are collected by very few groups. The CD (charge distribution) concept427 permits distribution of the charge (of, e.g., actinide ions) among different planes (e.g., in Figure 19B the charge transferred to the interface by Am3+ could be distributed among the o- and 1planes). The decision on how to distribute the charge could be based on structural information, spectroscopic results, or goodness-of-fit to the adsorption data. The CD factors greatly influence the model inherent proton stoichiometry491 and attempts have been undertaken to estimate them from advanced theoretical calculations.492 At trace actinide concentrations, basic charging alone influences adsorption. Higher actinide surface concentration will contribute more strongly to the overall charge and may ultimately lead to charge reversal (observed for cation adsorption to negatively charged silica493). For trace concentrations of contaminants, the basic charging remains the most important factor determining the overall surface charge of a mineral. In the environment it is potentially affected by adsorbed ubiquitous anions like carbonate or silicate. Also if the trace contaminant of interest is already present on the sorbent, the mole balances are no longer correct.417 However, this is probably not a typical case for actinides migrating from a waste repository or in a contaminated site environment. The diversity of the possible model variants that can be constructed (see Figure 19) results in various complications. Depending on the nature of the model the intrinsic constants for adsorption of a given actinide to a given mineral will be different. This is similar to the situation for aqueous solutions, where one may obtain different stability constants at infinite dilution depending on the model used for calculating activity coefficients. A different debate concerns itself with treatment of multidentate surface complexes.494,495 All these aspects have created the unfortunate situation that any systematic selfconsistent compilation of surface complexation stability constants based on literature references alone is next to impossible. Basic charging affects the actinide adsorption model and choices made for the basic charging model propagate to this part of the modeling. Actinide adsorption introduces new adjustable parameters (log Ki° and CD-factors, see Figure 19A) beyond the previously defined basic charging related parameters. In Figure 19, Am3+ is considered to be adsorbed as an inner-sphere bidentate complex, releasing two H+ from SOH groups. The overall charge transferred in the formation of the surface is therefore Δz = 3 − 2 = +1. The charge of the adsorbed Am3+ can be distributed over the o (zo,Am) and 1 (z1,Am) planes (zo,Am + z1,Am = 3). Assuming 9-coordination of Am, zo,Am = 2/9 × 3 and z1,Am = 3 − zo,Am. Figure 21A depicts a comparison of model calculations for uranyl on HFO (building on the acid−base models shown in Figure 20) and Figure 21B depicts the model on quartz. In Figure 21, the predicted overall uptake curves do not differ much for the different models, while the surface speciation in the models does. In the HFO case (Figure 21A), the DLM calculation is a blind prediction from the linear free energy relationship (LFER) for HFO22 in Figure 22 (monodentate surface complexes on generic sites, the properties of which were obtained by fitting a wide range of experimental data including surface charging and metal adsorption data, but not uranyl), while the CD-MUSIC model 488,489 was calibrated on

Figure 21. Calculated uranyl uptake as a function of pH at 100 m mol dm−3 salt concentration, 0.1 μmol dm−3 overall uranyl concentration and 600 m2 dm−3 solid concentration (A) on HFO obtained with a 2site diffuse layer model (DLM, black)22 and a recent CD-MUSIC (TPM) approach (red)488,489and (B) on quartz calculated with a 3-site nonelectrostatic model (NEM)496 and a 1-site CD-MUSIC (BSM) approach.497 OS means outer-sphere and IS-1 and IS-2 are two innersphere surface complexes as discussed in the text.

experimental uranyl uptake data involving spectroscopic information about the structure of the surface complexes (i.e., bidentate binding to singly coordinated hydroxyl groups). The uptake of uranyl on HFO (Figure 21A) in conjunction with Figure 20 highlights that adsorption of cations via surface complexation occurs despite electrostatic repulsion. The calculated slopes of the two overall uptake curves differ somewhat, but the pH for the predicted onset of adsorption is in the same pH range for both generic and mechanistic models. The surface speciation however is very different. The DLM model involves monodentate adsorption of the uranyl ion onto strong and weak sites, with the release of one proton, resulting in SiOUO2+ surface complexes. The CD-MUSIC approach involves a bidentate surface complex, (FeOH)2−0.1UO2+1.1, which hydrolyses with increasing pH. For quartz (Figure 21B), the calculated overall adsorption curves obtained from the two models differ by no more than 1048


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Figure 22. Linear free energy relationships (LFERs) for the adsorption of cationic species on a variety of sorbents. The LFERs relate (intrinsic) stability constants for the equation SOH + Men+ ⇔ SOMe−1+n + H+ to the corresponding first hydrolysis reaction OH− + Men+ = MeOH−1+n, where n covers values from 1 to 4. (for detailed discussion see text).

zero net proton charge, pznpc) and zeta-potential data (yielding the isoelectric point, IEP) can be used. The basic charging model requires stability constants (log Ki°) for the previously defined sites within the chosen protonation scheme and if involved in the model, CD factors, capacitance values (Ci), and slip plane distance (lζ). Finally, the actual actinide adsorption model is designed while the basic charging model must not be changed any more. The nature of the surface complexes needs to be defined in terms of stoichiometry and charge distribution. For this, spectroscopic data and computational chemistry can be used as a priori constraints. The model is then fitted to actinide adsorption data. Further information required for the overall model is a scheme for aquatic actinide speciation. Numerical estimation of unknown parameters is handled by fitting a defined model to relevant experimental data, that is, for obtaining acid−base properties (or primary charging) to model fits to titration and electrokinetic data are performed and for actinide interaction parameter fits to adsorption data (see Figure 19). Various options exist to this end, including directly coupled speciation/optimization codes,501 or shell options.65,502,503 Ideally, direct speciation results are used for this, such as those obtained from TRLFS, for example, for the Cm/Am-quartz system.65 The shell options have the advantage that an arbitrary number of data set or various sources of data can be combined; furthermore, experimental error estimates can be more easily treated in a consistent way, which is required for objective fitting.504 4.2.5. Generalization of Results. Since model parameters pertaining to actinide adsorption in multicomponent systems are linked to all subsystem parameters (i.e., aqueous speciation and surface acid−base properties), it is imperative to compile self-consistent parameters within a given model framework. Such attempts can be found for the generic site approach within a DLM, the simplest model that allows a variation of ionic strength. Databases exist for various solids, including HFO,22 hydrous manganese oxide,505 goethite,506 and gibbsite.507 Whenever such approaches result in linear relationships

0.2 pH units. For this example, the nonelectrostatic model (NEM) was calibrated on experimental column data, which required three different (generic) surface sites with a total of 4 uranyl surface complexes to obtain a good fit to breakthrough and elution curves.496 The CD-MUSIC approach497 was calibrated on experimental batch data with a single surface site and 3 uranyl surface complexes. The latter were constrained from a theoretical study suggesting the presence of an outersphere surface complex498 and a number of spectroscopic studies suggesting inner-sphere bidentate surface complexes.49,499 The inner-sphere surface complex (IS-1) in the CD-MUSIC approach has the same structure as the one for the HFO model analogue, while the CD-factors differ. The second inner-sphere surface complexes within the CD-MUSIC approach (IS-2) is a hydrolysis product of IS-1, as for the HFO. Unlike for HFO, for quartz a further hydrolysis step of the IS-2 was not required. For the nonelectrostatic model contributions from the weak, intermediate (S2), and strong sites differ from the generic HFO model (where the strong site dominates). This has been discussed in some detail elsewhere.500 While for both HFO and quartz the overall sorption is similar, surface speciation, which is related to mechanistic understanding of the adsorption process, differs strongly. This suggests that comparison of available surface speciation schemes with concomitant stability constants and charge distribution factors is not meaningful and comprehensive modeling of data is required within a self-consistent model framework. This would be a substantial effort requiring refitting available data. 4.2.4. Input Data for Model Development and Parameter Estimation. As shown in Figure 19A, input (experimental data and constraining information) for model development encompasses particle morphology for the sorbent, structural data either from bulk crystallography or surface diffraction to obtain information related to mechanistic surface sites in terms of type and number (site densities). For the basic charging model acid−base data (yielding ideally the point of 1049


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will also interfere. Remarkably for Pu the LFER in the inset covers various oxidation states. Note also, that the two sets of LFERs are parallel (those for Pu and those obtained for goethite and silica for different mostly divalent transition metal ions), i.e. same slopes for goethite and silica in both evaluations. Figure 22 provides a rough idea about the affinity of cations in general toward the different surfaces with silica being the least efficient sorbent and HFO (ferrihydrite) the most efficient one. Figure 21 is in agreement in that for identical total uranyl concentration, surface area concentration, and same ionic strength, the adsorption edge on quartz (which is considered to be similar to silica) is at clearly higher pH than on HFO. Results were all obtained within simple models having highly simplified electrostatic parts and rather generic treatment of sites (note that for the clays the resulting surface speciation for Cm(III) and the multisite approach have been verified spectroscopically518). These simple models are probably best suited for performance assessment. They have, e.g., been used for modeling simple519 and complex520 mineral assemblages, and detailed roadmaps for their application in real-world situations have been presented.406,423,521,522 While simple (more generic) approaches do not necessarily include the available mechanistic knowledge, the framework of the MUSIC model allows this to a significant extent.458,488−491,523,524 As stated above, the definition of surface sites, protonation mechanism and proton affinity constants are based on surface structural considerations and can be predicted.427,428 Other concepts allow the estimation of points of zero charge,451−453 acid−base constants, ion-pair formation constants, and capacitance values for generic site models.454,482,525 Current models allow the location of point charges of inner-sphere (charge of the actinide placed in the surface plane) and outer-sphere surface complexes (charge placed in the plane of background electrolyte adsorption, which creates the concomitant competition).481,526 Ionic strength effects in macroscopic adsorption data do not necessarily distinguish outer- from inner-sphere surface complexes.527 Beyond viewing adsorbed ions in terms of point charges, charge distribution over the available planes of adsorption appears more realistic.427 Potential contributions from oriented water dipoles at surfaces, are nowadays considered.369 The structure and stoichiometry of the surface complexes are best evaluated from available structural investigations (spectroscopy, theory) combined with numerical fitting of the macroscopic data. Comprehensive studies on selected systems applying some variant of the CD-MUSIC approach have treated the adsorption of uranyl to HFO (ferrihydrite),489,528,529 to goethite,319 to SiO2,497 of trivalent actinides or trivalent analogues of actinides to TiO2,490,530 and to SiO265 and are also being applied to nonoxide minerals like calcite.531 Compared to the high number of citations of MUSIC papers, the actual application of the model remains still comparatively limited. In particular, a comprehensive, self-consistent database within the CD-MUSIC model is lacking that would allow more surface chemistry input compared to the generic models. Bouby et al.490 based their model analogue for Am(III) sorption onto TiO2 on two different structural studies, assuming either monodentate490 or multidentate532 surface complexes. Both model variants were finally applied and equally successful, another incidence that clearly shows that fitting a model to macroscopic data alone cannot prove an adsorption mechanism.

between intrinsic adsorption constants and corresponding aqueous solution constants (e.g., hydrolysis constants for metal ions) predictions of adsorption constants are possible for ions that have not been experimentally studied. In the above cases, this would permit obtaining parameters for actinides on HFO, goethite, and gibbsite. However, this is currently not possible whenever ternary complexes with ligands beyond hydroxide ions are involved. Other approaches exist along similar lines, though with different surface−site and electrostatic model combinations for pure acid−base properties454,508−510 or adsorption constants for many different systems.511 One prominent example concerns actinide adsorption to illite and montmorillonite, which are major components in some radioactive waste disposal concepts.512 The NEM-approach involves, as discussed previously, both surface complexation and an ion exchange mechanism. This 2SPNE SC/CE model is quite successful in blind predictions of adsorption data513 and LFERs were obtained for the surface complexation part of this model. Figure 22 shows a series of LFERs. The concept appears to generally hold, despite lack of mechanistic input and different assumptions made. The data in Figure 22 for the “oxidic” sorbents, HFO (also referred to as 2-line ferrihydrite, Fe(OH)3), goethite (α-FeOOH), and gibbsite (α-Al(OH)3) were all evaluated by Dzombak and co-workers22,506,507 with a purely diffuse layer model. The data for Illite and montmorillonite were evaluated by Bradbury and Baeyens within a nonelectrostatic model.41,42,131 For HFO (ferrihydrite) and clays, multisite models were used, while for goethite and gibbsite single-site models were deemed sufficient. The HFO LFERs were discussed by Tiffreau et al.514 in the sense that the lower slope for the strong site LFER will cause inconsistency insofar the adsorption of strongly hydrolyzing ions will be favored on weak sites. The data for alumina from Huang and Stumm447 have been previously evaluated by Schindler515 and pertain to the DLM, while the data on silica were evaluated within a Nernst model. The electrostatic model for the anatase data which are included in the compilation by Schindler and Sposito,414 is not specified in available literature. Since the concomitant acid−base data were evaluated with a CCM,434 it can be assumed that that this electrostatic model was also used for the adsorption data. Data for adsorption of rare earth elements on basaltic rock,516 evaluated within a DLM illustrates the applicability of the approach to composite minerals. Although both the electrostatic models and the evaluation of parameters like surface site density are not consistent among the different LFERs, the analogy between adsorption and hydrolysis in solution appears to be similar for a wide range of surfaces based in Figure 22. Such LFERs allow prediction of the respective constants for solutes that have not been experimentally studied or are, as might often be the case, for example, for extremely redox-sensitive actinides, difficult to study. Recently surface complexation parameters within the DLM have been analyzed for Pu and Th517 and LFERs are available. Two LFERs for Pu (one for goethite and another for silica) are shown in the inset to Figure 22. They clearly do not coincide with the LFERs that had been independently obtained for the two solids.465,506 The difference may be due to differences in the underlying acid−base models. Thus the site density for silica used by Schindler et al.465 was about 5.5 sites nm−2, while that for the Pu-study was 2.3 sites nm−2. For goethite the difference in site density is less spectacular (2.3 vs 2.2 sites nm−2), but other parameters like acid−base constants 1050


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Ternary surface complexes (i.e., surface complexes involving, for example, an actinide and an anion, like carbonate or silicate addressed in previous sections) as well as competition between two solutes110 or physical mixtures of sorbents189can be handled by surface complexation models.193,489,533 Multisite models can also account for the ion-specific competition that was addressed in an earlier section. The approach was also extended to include the adsorption of natural organic matter and all the charge regulating aspects involved.534 Consequently the model framework is sufficiently flexible to allow in principle the description of all the discussed adsorption phenomena though at the cost of simplicity. With increasing complexity of the models there is an increase in the number of adjustable parameters. The greater the number of parameters that can be constrained from available knowledge, the more reliable the estimation of the remaining ones. Insensitivity of model parametrization to experimental titration and adsorption data has long been realized.450,479,487 Therefore, mechanistic modeling requires support in terms of process understanding from spectroscopy and advanced computational chemistry. This has also been recognized by waste management organizations in relation to safety assessment. These organizations focus on process understanding and distribution coefficients as the most valuable outcome from sorption studies.535 A difficult issue is how to transfer a maximum of process understanding to a final retention model.522 While this is in principle possible with the existing models (like the MUSIC model), the situation in the real world is probably far too complex for detailed models and reactive transport codes are not necessarily capable of implementing such advanced approaches.536 Adsorption treated by surface complexation models is one aspect of sorption or retention processes involved in mobilization/immobilization of acitinides, redox-reactions on surfaces is another (cf., section 2.4). The chemistry related to surface catalyzed redox-reactions is not typically included in adsorption models. In principle, the formalism of surface complexation, however, allows such phenomena to be described and coupling between kinetic processes and surface speciation has been successfully applied to other aspects of surface chemistry, including dissolution processes coupled to MUSIC type models.537 Furthermore, actinides might interact with solids via incorporation into the solid matrix, a process that can occur through various pathways introduced previously (see section 2.3) including coprecipitation, diffusion, or precipitation processes at surfaces. Approaches to model the latter are addressed in the following section.

xM M(s) + xAn An(s) = M(xM)An(xAn)(s)

(22)

where xM and xAn are mole fractions such that xM + xAn= 1. As an example, the free energy of an isostructural binary solid solution is then related to the sum of the standard Gibbs energies of the end-members multiplied by their mole fractions plus a mixing term. The mixing term requires a mixing model, which can be considered a kind of “activity” model. At low mole fractions of the incorporated actinide the treatment may be simply via the dilute solid solution formalism. The more similar the sizes of M and An and the greater the similarity of the crystal lattice of the end-members, the closer the solid solution will be to an ideal one. For nonideal solid solutions, the required complexity of the model to describe nonideality is debatable, similar to surface complexation modeling. Various models exist539 and implementation of solid solution models in mass-law/mass-balance codes is possible. Gibbs energy minimization has been successfully applied to several systems relevant to actinides.540,541 One general concern in the equilibrium treatment is the assumption of “homogeneous” solid solutions, an assumption that will be hardly ever met for the current context. In nearly all cases, the incorporation of actinides into some matrix will be proceeded by an adsorption step, which can be described by a surface complexation model. For dynamic matrix systems, like calcite or brucite, the adsorbed species will begin to be incorporated.234,542,543 For bulky matrix particles, it is hardly perceivable that this will ultimately lead to homogeneous solid solutions, however. Therefore, sometimes adsorption and coprecipitation processes are combined.135,249,544,545 Farley, Dzombak, and Morel546 designed a model that can describe the transition from adsorption to surface precipitation to homogeneous solid solution formation (c.f., Figure 23).

4.3. Approaches to Modeling Incorporation of Actinides in Matrices

The incorporation of actinides into matrix compounds falls under the general term absorption.538 Absorption encompasses processes like coprecipitation or surface precipitation, which are usually treated in terms of solid solutions. A detailed account on this in the context of actinides is given by Bruno et al. in a very general way with respect to actinides539 and more recently by Bruno and Montoya.1 The heterogeneous precipitation/incorporation regime of an isotherm is related to the formation of solid solutions. For the general formation of a binary solid solution, one may write the reaction between the two end-members (M(s) for a matrix component and An(s) for an actinide) in equilibrium with a homogeneous solid solution of intermediate composition as

Figure 23. Illustrative adsorption isotherm for an actinide on an oxidic sorbent at constant pH and ionic strength (blue line). The stars show various possible measurement results that may either be on the equilibrium isotherm (blue, green and red) or not.

While the model in its usual application involves the assumption of an ideal solid solution, it has been shown that the approach can be generalized toward nonideal solidsolutions and also to solid solutions limited to the surface region (i.e., only part of the matrix is involved in solid-solution formation).547 The BET (Brunauer, Emett, and Teller) equation (a standard for gas adsorption on surfaces548 used 1051


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but also at a mechanistic level,369 taking significant details of surface properties and sorbed species into consideration, including temperature variation.458,550 Unlike for solutions, no peer-reviewed comprehensive sorption databases for actinides using such models exist. The available databases22,505−507 and comprehensive studies41,42,131 are very helpful but not sufficient. Databases that accumulate surface complexation parameters within any model551 can be seen as cost-efficient alternatives, with the increased risk of predictive failure due to parameter inconsistencies.552,553 Hopefully in the future reevaluation of sorption data within a model that is consistent with aqueous speciation data from the NEA-TDB401 will be initiated so that a peer-reviewed reliable database will be available for performance assessment.

to evaluate specific surface area) is ultimately obtained for multilayer precipitation (condensation) of the adsorbing solute on the matrix surface. It can be simulated by decreasing the amount of the matrix element that is involved in solid solution formation to very small values. The approach is capable of covering the full range of adsorption, heterogeneous precipitation, and solid-solution formation processes, but this is only possible with a very simple electrostatic model. The interfacial model should not involve individual surface properties pertaining to the two end-members like capacitances. The model originally proposed by Farley, Dzombak, and Morel546 was finally implemented with a purely diffuse layer model in the surface complexation database for hydrous ferric oxide.22 It has since then been extended for trace metal sorption to calcite.543 Though frequently cited the surface precipitation model is rarely applied. In the context of actinide sorption Zhu549 has estimated parameters for the uranyl ion within this model for HFO and for calcite. The formalism of the surface precipitation model is straightforward for ideal solid solutions, but not all available speciation codes can translate that formalism. When such processes are included in performance assessment, distribution coefficients are usually defined for the adsorption process. One major problem that has been realized early on is that the various mechanisms of adsorption, absorption or homogeneous precipitation can hardly be distinguished based on macroscopic data alone.538 More importantly, the use of experimental distribution coefficients within a performance assessment can lead to grossly erroneous results, if an inappropriate mechanism is considered. This is shown schematically in Figure 23, where the blue line represents the surface precipitation isotherm,546 with pure adsorption at low concentrations (which ideally may be linear, as shown, and justify in the range of linearity the use of a constant distribution coefficient), followed by site saturation at higher concentrations (which may be apparent, that is, caused by electrostatic effects and/or competition) and finally surface precipitation, which results in a (surface) solid solution. Homogeneous precipitation is also shown and usually occurs at higher dissolved concentrations than the heterogeneous processes. It was common that a site specific experimental study reported no more than one distribution coefficient for a given system (sorbent, solution, radionuclide). This single point may originate from any (but only one) of the four points shown in Figure 23. For performance assessment one would obtain any of the four resulting distribution coefficients (the slopes of the linear lines from the origin to the points and beyond). Only one of them qualifies as a constant distribution coefficient (the one in blue and only in the linear part of the full isotherm). The red and violet points, if used in terms of a distribution coefficient, would predict too strong retention of the respective actinide (resulting in under-estimation of a final dose from the actinide). A distribution coefficient from the green data point would yield a “conservative” value, not on purpose though, but rather by accident. Figure 23 illustrates the requirement for a thorough experimental, spectroscopic and theoretical study to learn as much as possible about the sorption mechanisms along the isotherm. It also shows that a simple distribution coefficient can hardly account for a broad concentration range and that more comprehensive models are required instead. In principle tools to treat sorption phenomena within thermodynamic concepts are available. The processes can be described on a more generic level capturing influences of environmental parameters in a consistent and appropriate way,

5. SUMMARY Clear progress has been achieved over recent years in fundamental understanding of actinide sorption mechanisms and actinide-mineral surface speciation. Individual sorption mechanisms, such as outer-sphere sorption, inner-sphere surface complexation, incorporation reactions, as well as surface induced redox reactions, have been identified and the different processes can be distinguished from each other. Application of spectroscopic techniques and quantum chemical calculations mostly draw remarkably coherent pictures and provide detailed insights into the molecular structure of respective surface actinide species. Application of cluster models and plane wave DFT with periodic boundary conditions appear to provide consistent results and allow validation of DFT approaches. Their applicability together with the development of pseudopotentials paves the way to quantum chemical treatment of larger systems. This further enhances the relation to experimental mineral/aqueous solution interface studies. Both, spectroscopies and theory, suggest that actinide surface complexes cannot be considered as species with distinct uniform structures. Apparently, even at well-defined single crystal surfaces a mixture of different species with slightly variable structures exist. Formation of ternary and quaternary surface complexes and their importance is now unequivocally established for actinide ions in spectroscopic studies. The richness of reported structures and stoichiometries for, for example, surface actinide carbonato and hydroxo-carbonato uranyl species, however, points to remaining uncertainties in interpreting experimental and spectroscopic data. Actinide “surface solid-solution” formation certainly requires more investigation. Actinide incorporation into a variety of mineral phases has been experimentally demonstrated, and appears to be notably dominant, for example, for trivalent actinide interaction with Ca-bearing minerals. Studies of actinide incorporation into iron oxide or silicate minerals when in contact with aqueous solutions close to equilibrium and with incompatible ionic radii are more challenging due to slow reaction rates and respective experimental difficulties. Related investigations are subject of a current European collaborative project.554 Redox reactivity of mineral surfaces attracts increasing attention, as it may severely impact the mobility of the light actinides uranium, neptunium, and plutonium. It appears likely, though not entirely verified, that electron transfer from Fe(II) solid phases is kinetically preferred for surface sorbed actinide ions over reduction of dissolved actinide species by Fe(II) in solution. In this context, natural Fe(II) containing minerals, (co)adsorbed Fe(II) in deep groundwater systems, or magnet1052


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Biographies

ite as a major anaerobic corrosion product of steel canisters in waste storage facilities represent huge electron donor reservoirs. Actinides species will therefore most likely be present in deep geological repositories in oxidation states III or IV, with low solubility and strong sorption as a consequence. For most of the studies discussed in the present review establishment of thermodynamic equilibrium is assumed. Redox and mineralization reactions, however, are partly characterized by slow kinetics and time frames in the order of years. Data on actinide reaction kinetics are hardly available and in part difficult to acquire, although they are essential for the reliable prediction of actinide fate and transport in natural systems over longer time scales. Most existing data for actinide interaction with mineral surfaces are available for uranium because of its low specific radioactivity and the ease of use. For transuranium elements, such as plutonium, however, safety requirements usually are so strict that they only can be handled in dedicated laboratories. This is notably true for synchrotron based X-ray spectroscopic studies, since only a few beamlines dedicated to actinide research are available worldwide. The situation becomes even worse if chemical reactions of actinides, such as redox transitions, are going to be examined spectroscopically in situ. The necessity for establishing and keeping such experimental facilities is thus evident. The observation of “selective competition” of metal ions for sorption sites is certainly remarkable and also worrying. It is obvious that geochemical calculations applying surface complexation constants for individual metal ions without knowledge of sorption competition bear large uncertainties. The idea that metal ion selective surface sites exist at mineral surfaces definitely calls for further research. Defining the exact nature of mineral surface sites and their charges still represents a central challenge for quantifying actinide sorption by geochemical sorption modeling. Multisite complexation models derive surface sites and their acid/base properties from the known structure of ideal crystal planes and are extremely successful in describing well-defined systems. They are, however, difficult to apply to real-world, natural systems although some progress is being made.556 The latter are heterogeneous on various scales. Therefore, pragmatic approaches are used for natural sediments, rocks and soils where generic surface sites need to be defined. Further simplifications have been proposed for clay rock systems, where nonelectrostatic models have been successfully applied to describe metal ion sorption. As a consequence, various types of sorption models exist, which limits the use of the model specific (thermodynamic) data that are available.

Horst Geckeis studied chemistry at the university in Saarbrücken, Germany and completed his Ph.D. in nuclear chemistry. Since 2008, he has been head of the institute for nuclear waste disposal (INE) and holds a professorship for radiochemistry at the Karlsruhe Institute of technology (KIT). His research interests relate to all aspects of nuclear waste disposal but focus specifically on the aquatic chemistry of longlived fission products and actinides and their interaction with mineral and colloid surfaces. He is currently head of the nuclear chemistry division of the German Chemical Society and member of an advisory committee on nuclear waste disposal nominated by the German Federal Ministry for the Environment, Nature Conservation and Nuclear Safety (BMU).

Johannes Lützenkirchen is a staff scientist at KIT-INE. Johannes started as chemical engineer and completed his diploma at the University of Karlsruhe (TU), now KIT, with a focus on water chemistry and transport processes. From Karlsruhe, he moved to Strasbourg, France, where he received a Ph.D. (1996) from the Louis Pasteur University (ULP), in Environmental Physical Chemistry on Surface Complexation and Precipitation Models. During his Ph.D. studies, he spent time with Freddy Dumont at the Free University of Brussels (ULB), Belgium, for experimental studies. From 1996 to 1999 he was employed as a Post-doc at Umeå University (UmU) in the group of Staffan Sjöberg with a focus on experimental work, mainly related to iron oxides. From Northern Sweden, he moved to Baden, Switzerland, where he was employed as a geochemist with Colenco Power Engineering and involved in long-term safety assessment of radioactive waste repositories. Since 2001, he has been employed at KIT-INE, where he has been predominantly working within applied projects until 2007. In the last 5 years, he has become more involved in fundamental research, involving sorption studies, interfacial processes in general, geochemical equilibrium modeling, and reactive transport.

AUTHOR INFORMATION Corresponding Author

*E-mail: [email protected]. Notes

The authors declare no competing financial interest. 1053


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Moritz Schmidt, born in Herdecke, Germany, studied Chemistry at the Rheinisch-Westfälische Technische Hochschule Aachen (RWTH) and the Ruprecht-Karls-Universität Heidelberg. He received his Ph.D. in 2009 under supervision of Prof. Thomas Fanghänel. After a two-year post-doctoral fellowship with P. Fenter and L. Soderholm at Argonne National Laboratory, he became a post-doctoral researcher at the Institute for Nuclear Waste Disposal of the Karlsruhe Institute of Technology. His research interests include the geochemistry of the actinides, particularly with respect to nuclear waste storage applications, and the application of modern analytical tools (TRLFS, CTR, RAXR) for the elucidation of molecular scale processes at the mineral/aqueous interface.

Robert Polly studied physics at the Technical University Graz, Austria, and completed his Ph.D. in 1998 on a combined experimental and quantum chemical topic. He worked for two years, 1998−2000, at the University of California, San Diego at the Supercomputer Center on the development of explicit correlated multiconfigurational ab initio methods. Coming to Germany, he joined the institute for theoretical chemistry at the University of Stuttgart, and since 2005, he has been a staff scientist at the Institute for Nuclear Disposal at KIT. His speciality is theoretical chemistry approaches to the investigation of

ACKNOWLEDGMENTS The authors wish to thank M.A. Denecke for her extremely helpful comments and linguistic corrections. We also thank Th. Stumpf and B. Schimmelpfennig for valuable discussions on many aspects of actinide-mineral surface interactions. The thorough reviews of five anonymous reviewers are gratefully acknowledged. They all helped to improve the present paper significantly. This research has received partial funding from the German Federal Ministries of Economics and Technology (BMWi) and Education and Research (BMBF) under contract nos. 02 E 10961 and 02NUK019A as well as from the European Community’s Seventh Framework Programme (FP7/2007− 2013) under Grant Agreement no. 232631 (ACTINET I3).

actinide reactions and incorporations at the mineral/water interface using ab initio and density functional theory methods.

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Thomas Rabung studied chemistry at the University Saarbrücken (Germany) and performed his doctoral thesis at the former Research Center Karlsruhe, now Karlsruhe Institute of Technology (KIT, Germany). He completed his Ph.D. in 1998 and now he is a staff scientist at the Institute of Nuclear Waste Disposal at KIT. He has long-standing experience in the field of radionuclide interactions at the mineral−water interface including the influence of organic matter. A main tool in these studies is the use of laser fluorescence spectroscopy (TRLFS). He is involved in many national and international research programmes and projects and has coordinated two research projects financed by the European Commission. 1054


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