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Is the thermodynamic approach appropriate to describe natural dynamic systems? (Status and limitations)
Nuclear Engineering and Design 202 (2000) 143 – 155 www.elsevier.com/locate/nucengdes
Is the thermodynamic approach appropriate to describe natural dynamic systems? (Status and limitations) J.I. Kim * Institut fu¨r Nukleare Entsorgungstechnik (INE), Forschungszentrum Karlsruhe GmbH, P.O. Box 3640, 76021 Karlsruhe, Germany Accepted 20 March 2000
Abstract The long-term performance assessment of a nuclear waste repository entails sound knowledge on the migration behaviour of radionuclides in a given aquifer system, which can in principle be appraised by a well-founded thermodynamic approach. For various reasons, in such an endeavour, a number of uncertainties are involved, which cause difficulties in the application of a thermodynamic approach to natural dynamic systems. This paper summarises the present state of the thermodynamic approach being applied to the assessment of various geochemical reactions of radionuclides and discusses the limitations involved in application to laboratory systems as well as dynamic natural systems. © 2000 Elsevier Science B.V. All rights reserved.
1. Introduction The geochemical behaviour of radionuclides in aquifer systems has been described, to a certain extent, by a thermodynamic approach. The thermodynamic data appraised for this purpose are, for obvious reasons, derived from individual equilibrium reactions of laboratory system, namely from experiments of a ‘closed’ system. The real natural aquifer system is dynamic and ‘open’, per definition in opposition to a laboratory system, for emerging geochemical reactions of individual radionuclides. Along the migration pathway, a given radionuclide in natural aquifer systems in-
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teracts with a variety of geochemical surroundings. In this, although pH and Eh values may remain largely invariant for certain cases, the water-constituent components change their concentrations, and the specific or gross reactivity of geo-matrix surfaces varies with space and time. Hence, the radionuclide concerned undergoes a multiplicity of nanoscopic reactions along its migration, in which individual processes may be not equilibrated or irreversible. For the appraisal of such complexities, a challenging question indubitably arises: ‘is the thermodynamic approach appropriate to describe natural dynamic systems?’ The question inevitably triggers off another: ‘is there any alternative scientific tool available then?’. This paper summarises the present state of available thermodynamic approaches that can be
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applicable for the performance assessment of a multi-barrier system, which is presumed to function as a safe confinement for a given nuclear repository. Limitations of the approaches, as perceived for the moment, are to come into view simply as repercussions of the present discussion. To provide a simplified overview on a multiplicity of convoluted chemical reactions in natural systems, the present discussion is separated into the following three areas: the laboratory systems (closed systems), the natural systems (open systems), and the applicability of laboratory knowledge to natural systems.
2. What do we need to know? All long-lived radionuclides to be disposed of are immobilised in different solid phases, called engineered barriers, that are contained by backfills known as geo-engineered barriers and then by natural geological barriers. Whatever the perceived scenarios are to come after, the long-term performance assessment of such a multi-barrier system entails, upon water intrusion, the quantification of the geochemical behaviour of individual radionuclides in the environment of each barrier. There are two gross geochemical reactions, namely immobilisation and mobilisation of radionuclides in each specific barrier. They are reciprocated in nature. The balance of the two reactions for a given radionuclide in each barrier can be appraised by the ratio of immobilisation rate to mobilisation rate. How can the two reaction rates be quantified with high certainty? Can they be delivered by a thermodynamic approach?
Fig. 1. Chemical states of radionuclides controlling their mobility in natural aquatic systems.
Whatever the engineered barriers are made of, they are thermodynamically unstable in the nearfield environment once water comes into contact with them, and undergo a variety of geochemical reactions. Solid phases of waste are first subject to corrosion, leaching and dissolution of individual components, the processes of which generate the ‘new’ secondary solid phases that appear to be stable thermodynamically as compared with the primary solid phases in the near-field environment. Upon the phase conversion of engineered barriers, the radionuclides dissolved are submitted to hydrolysis, complexation, redox reaction and colloid generation, and largely immobilised by incorporation into the secondary solid phases that are partially equilibrated in the new environment. Under these conditions, the chemistry of radionuclides in the aqueous phase is submerged into the geochemical condition evolved from water interactions with waste, backfill and surrounding geological matrices. The chemical states of all metallic radionuclides to be disposed of can be classified on the basis of their oxidation states, from mono-valent to heptavalent, namely M(I), M(II), M(III), M(IV), M(V), M(VI) and M(VII). To this category belong, as illustrated in Fig. 1, all radioactive elements of long half-life. The two characteristics are recognised for all elements in aquifer systems: the oxidation state specific chemical behaviour, on the one hand, and the chemical reactions governed by the effective charge of individual ions, on the other. These characteristics can be followed as a primary guide for understanding of their general chemical behaviour in both laboratory and natural systems. For M(I) and M(II), much of their geochemical behaviour is known, since they belong to the natural element groups of major component both in aqueous and solid phases, and undergo reversible chemical reactions for the most part. The important nuclides belonging to these categories are 237Cs(I) and 90Sr(II), as fission products, and 226 Ra(II), as one of the decay series of 238U. As these elements are naturally present, their thermodynamic behaviour can be correlated to the chemical systematic of alkali metal ions and alkaline earth metal ions. For actinides, which are present
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under natural conditions as M(III), M(IV), M(V) and M(VI), their chemical behaviour is much more difficult to describe, since they are submitted to a multiplicity of geochemical reactions. However, for M(VI), the chemistry of U(VI) (Grenthe et al., 1992) observable in nature can be closely followed, to which transuranic actinides of the same oxidation state can be compared to a certain extend. To M(VII) belongs only 99Tc and, if present as TcO− 4 , it is least interactive with geochemical surroundings. Under reducing conditions, which are mostly the case for deep aquifers, the actinides of M(V) and M(VI), as well as Tc(VII), become reduced to M(III) or M(IV). Any chemical compound of M(IV) is sparingly soluble in water at neutral pH and the particular chemical properties lead to its strong tendency toward the colloid formation.
3. Applicability of a thermodynamic approach to laboratory systems (closed systems) All thermodynamic constants relevant to the appraisal of the geochemical behaviour of radionuclides are determined by laboratory experiments, besides those evaluated by indirect approaches. The application of the data to describe the geochemical behaviour of radionuclides in laboratory systems appears, therefore, accountable. However, the geochemical study in the laboratory is largely made with geological samples under simulated natural conditions. In the laboratory, such conditions are kept in a closed state but the radionuclides introduced are exposed to a variety of geochemical reactions. Therefore, the so-called laboratory system deals with pure samples as well as natural geological samples, for which the speciation is performed in order to evaluate the applicability of a thermodynamic approach. In the present paper, the laboratory system is discussed accordingly. To facilitate discussion, the subject is divided into three areas of chemical speciation: aqueous phase, solid phase, and solid – water interface. Although the chemical reactions of radionuclides are convoluted within the three phases, their reactions can be discussed separately in individual phases.
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3.1. Speciation in the aqueous phase In natural aquifers, the dominant chemical reactions of metal ions, in particular actinide ions, are hydrolysis and carbonate complexation, followed by many minor complexations with waterconstituent ligands depending on their concentrations, such as chloride, sulphate, phosphate, etc. (Kim, 1993). Humic acid is also one of the important aquatic constituents, with its concentration varying widely in different groundwater (Kim, 1993). Hydrolysis is the most important primary geochemical reaction for actinides and its tendency follows the effective charge as given by (Chopin, 1983): An4 + \ AnO22 + ] An3 + \ AnO+ 2 The effective charges of AnO22 + and AnO+ 2 are known as 3.29 0.1 and 2.39 0.2, respectively (Chopin, 1983), because their linear oxygen bindings expose the equatorial charge larger than the nominal charge. This tendency is also followed by other complexation, depending on the ligand concerned. Hydrolysis facilitates the colloid generation of actinides through a formation of oxygen bridges, either real or pseudo colloids (Kim, 1994). The speciation can be reasonably made for aqueous actinide ions based on hitherto known thermodynamic constants for binary reactions of hydrolysis and complexation with some important water-constituent ligands. However, for the ternary complexation, the experimental constants are poorly known. The relevant thermodynamic data for An4 + are largely missing, because this ion endures strong hydrolysis that leads to colloid formation in aqueous phase. The present state of knowledge for the thermodynamic speciation of actinides and some important fission products in aqueous phase is summarised as follows. For binary reactions: M(I), no M(III), sufficomplexation ciently known
M(V), sufficiently known
M(II), well known
M(VI), sufficiently known
M(IV), poorly known
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Fig. 2. Stability regions of aqueous hydroxo and carbonato species of Am(III) and Cm(III) in 1 M NaCl at 25°C; (1yz) = An(OH)y (CO3)3z − y − 2z; hatched area, possible stability region of mixed species (Neck et al., 1997).
The ternary complexation is poorly known for all of M(III), M(IV), M(V) and M(VI). Because of the high electric charge, the ternary complexation of various kinds is expected particularly for M(IV), and to a somewhat less extent for M(III) and M(VI). They are presumed to be predominantly hydroxo complexes, which are difficult to speciate experimentally. However, their formation constants may be approximated for some particular cases on the basis of each corresponding binary complexation. A typical example for the speciation of aqueous species of M(III) is illustrated in Fig. 2, which is calculated based on the hitherto known thermodynamic constants (Neck et al., 1997). This figure demonstrates the applicability of the thermodynamic speciation in the aqueous phase under various geochemical conditions, i.e. at different pH and PCO2 values, for the formation of hydroxides (1Y0), carbonates (10Z) and hydroxo carbonates (1YZ). In laboratory systems, the thermodynamic speciation is for the moment practicable for M(III), M(V) and M(VI), at different ionic strength, either in aqueous phase under controlled conditions or in natural groundwater. A further extension of
the thermodynamic database for aqueous actinide ions of various species, particularly for M(IV), is still necessary. In all groundwater, colloids are always present, which are composed of inorganic, organic or a combination of both components (Kim, 1994). They are produced either by heterogeneous nucleation of water-constituent metal ions through oxygen bridges or by dispersion of the surface composites of geo-matrices. They are small in size (B 100 nm) and chemically reactive for polyvalent radionuclides. Such radionuclides are readily sorbed onto groundwater colloids and thus present in groundwater as so-called pseudo colloids. The speciation of these colloidal actinide species is difficult to realise but if given groundwater colloids are well characterised, the formation of pseudo colloids can be modelled. Nonetheless, a new model approach is yet to be developed to describe their formation process, structural conformation and stability.
3.2. Speciation in the solid phase There are two solid phases to be dealt with: pure solid phases of actinides, which are composed of actinides themselves as major components, and solid solutions that contain actinides in a bulk of other solid phases. On the former species, the extent of our present knowledge is somewhat favourable. Even so, there is a large number of various solid species of actinides that remain not apprehended for their formation under aquatic conditions. The knowledge on the latter species is virtually non-existent, besides on uranium in natural minerals (Grenthe et al., 1992). However, for actinide solid solutions, the oxidation state specific natural analogues may be taken for the comparative purpose, although their applicability is not always straightforward, e.g. rare earth elements for M(III), Th or Zr for M(IV), U for M(VI), and Re for Tc(VII). The present knowledge on thermodynamics of pure solid phases for the oxidation state specific actinides and some fission products is summarised as: M(I), well known
M(III), sufficiently known
M(V), sufficiently known
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M(II), well known
M(IV), poorly known
M(VI), sufficiently known
The pure solid-phase formation is directly related to the speciation in the aqueous phase, e.g. the conversion of the solid phase is accountable for the change in the solubility of a given element. A typical example (Neck et al., 1998) is shown in Fig. 3. The solubility experiment of Am(III) as a function of pH at PCO2 =1 and PCO2 = 10 − 2 demonstrates a continuous correlation, as shown in the upper part of Fig. 3. Once the thermodynamic speciation of solid phases is applied, as
Fig. 3. Upper part: Solubility of Am(III) in 0.1 M NaClO4 (Neck et al., 1998); experimental data from Meinrath (1991). Lower part: Stability regions of Am(III) solid phases at I = 0.1 M and 25°C (Neck et al., 1998).
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shown in the lower part of Fig. 3, it becomes apparent that, during the experiment, the solid phase is converted from the original amorphous Am2(CO3)3 at near-acidic pH to a ternary phase of AmOHCO3 at neutral pH, and further to NaAm(CO3)2 at near-alkaline pH. A thermodynamic approach is thus applicable to laboratory systems for the prediction of solid-phase conversion as a function of both pH and PCO2. The estimation of the Am(III) solubility under dynamic geochemical conditions is also possible, once such a conversion becomes predictable. The speciation in pure solid phases is also possible, based on the present knowledge, for the actinides of M(III), M(V) and M(VI). For M(IV), more pertinent thermodynamic data are yet to be known for this purpose. In dissolution equilibrium, the solubility of a given solid phase of M(IV) is difficult to determine, because M(IV) tends to form colloids. This fact disturbs the speciation of M(IV) solid phases equilibrated in water and, as a consequence, the solubility of any M(IV) compound is known poorly. Although the solubility of MO2 has been determined by a large number of experiments, the reported solubility product varies considerably for all actinides of M(IV). The experimental solubility product of PuO2, for example, varies from log Ksp = −51.8 to log Ksp = − 60.2, whereas its calculated value appears to be log Ksp = − 63.8 (Kim and Kanellakopulos, 1989). Such a large range of difference is mainly attributable to the colloid generation, which is sensitively affected by individual experimental conditions. The speciation of actinides in solid solution is, as mentioned already, hardly possible, because the relevant thermodynamic data are not available for the moment. Re-immobilisation of long-lived radionuclides, after dissolution of the primary solid waste packages, depends directly on the subsequent formation of solid solution. Therefore, knowledge on the secondary solid-phase formation under near-field geochemical conditions is an essential prerequisite of the source term quantification for the long-term performance assessment. For the moment, only laboratory simulation experiments can provide such knowledge.
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3.3. Solid–water interface reactions The migration behaviour of individual radionuclides in geo-engineered and geological barriers depends on their solid – water interface reactions at surfaces of available geo-matrices, which are composites of different minerals. At near-neutral pH, the surface of geo-matrices is in generally covered with the surface-bound anionic oxygen or neutralised hydroxide. Cationic radionuclides interact on the surface to form either ion association (outer sphere complex) or complexation with surface oxygen ligands (inner sphere complex). The former process is valid largely for M(I) and M(II), while M(III), M(IV), M(V) and M(VI) attend the latter reaction. This concludes that aqueous actinide ions undergo mainly the stable surface complexation. In addition, another possible interface reaction is the surface precipitation (thereafter mineralisation), to which metal ions of very low solubility are subjected, i.e. M(III), M(IV) and M(VI). The surface sorption behaviour of a given radionuclide is often quantified simply by its concentration distribution between solid geo-matrix and water. The quotient is commonly known as Kd, which is an operational parameter borrowed from the conventional ion exchange separation. Whereas a laboratory ion exchanger bears reversible chemical reactions, the sorption process of metal ions onto geo-matrices carries on both reversible and irreversible reactions, which are convoluted each other with time. The so-called ‘Kd concept’ is often used to calculate the migration behaviour of individual radionuclides in a given aquifer system, although the data are mostly derived from sort-term laboratory experiments. Since Kd is a simple operational parameter, which depends on various unknown geochemical effects involved, it accordingly has no solid thermodynamic foundation; its applicability to the long-term performance assessment has become the subject of controversial debate. An alternative to the Kd concept is to evaluate thermodynamic parameters for the complexation of individual radionuclides onto the surface of geo-matrices. This can be realised by the study of
solid–water interface reactions for a given radionuclide on a number of single minerals and thereafter by a combination of the results proportionally to a composite model. The development of such a surface complexation model is in progress in a number of laboratories worldwide (Davis et al., 1998). A simplified reaction process on each mineral surface is: − lp n(.SOH) + MLm p − p + n/l − n l (.SO)n MLm + (n/l)L+ nH+ p − n/l
Ksc
where n(SOH) represents the number (n) of surface functional sites to be interacted with a given − lp metal ion species, MLm , complexed with the p p l− number of ligands L . In laboratory systems, the pKa values of n(SOH) for individual minerals and the complexation constants of the metal ion − lp species concerned MLm can be determined p and, hence, the quantification of the presented reaction becomes possible as a function of the pH and L concentration. However, the surface sorbed species is to be known for this purpose either by direct or by indirect speciation. Once the surface complexation constants, Ksc, of different radionuclides are known for various minerals, a composite model can be produced for a geo-matrix of mineral mixture. An integrated modelling procedure is still to be developed for a broad application, although a number of different surface complexation modellings for particular laboratory examples are available in the literature (Stumm, 1987; Dzombak and Morel, 1990). A typical example of the surface complexation study is shown in Fig. 4, in which the laser-spectroscopic speciation of the Cm(III) sorption onto silica surfaces of colloidal suspension is compared with thermodynamic modelling (Chung et al., 1998). The two kinds of surface sorbed species are recognised by spectroscopy with pH variation, from which their formation constants can be derived as corroborated by thermodynamic calculation (dark lines). Another example is shown in Fig. 5 for a comparison of the surface specific sorption (Ka) of U(VI) onto different minerals and mineral composites as a function of pH (Waite et al., 2000).
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Contrary to the Kd concept, the surface complexation modelling is a thermodynamic approach that provides a deep insight into the underlying chemical reactions at solid–water interfaces. While this approach facilitates the quantification of the primary solid–water interface reactions, its applicability to the appraisal of the long-term geochemical behaviour of radionuclides in a given aquifer system is still to be corroborated by a wide range of experiments. The present state of knowledge on solid–water interface reactions of pertinent elements is as follows: Fig. 4. Laser spectroscopic speciation of Cm(III) sorbed on 0.5 g/l silica. Dark lines, fit to a surface complexation model (Chung et al., 1998).
M(I), sufficiently known M(II), sufficiently known
M(III), poorly known M(IV), not known
M(V), somewhat known M(VI), somewhat known
As M(IV) forms colloids (either real or pseudo colloids) at neutral pH, its solid–water interface reactions on geo-matrices are difficult to quantify.
4. Applicability of a thermodynamic approach to natural systems (open systems)
4.1. Source-term quantification/near-field chemistry (closed/open systems)
Fig. 5. Surface-specific sorption of U(VI) on various natural minerals as a function of pH (Waite et al., 2000).
Fig. 6. Geochemical reactions that govern the generation of mobile radionuclide species and their migration.
Practical examples can be taken from the source-term quantification, which is made up within the near-field geochemistry that can be considered as either closed or open system. When the near field is intruded by fresh water and the corrosion of waste packages proceeds, the geochemistry of the environment changes above all pH, Eh and PCO2 values. This change can be modelled by a thermodynamic approach once the corrosion and leaching products can be assessed with time. However, most important is to appraise the formation of secondary solid phases produced from dissolution of the original solid waste and its packages. The solubility of each new solid phase then results in the aqueous concentrations of radionuclides, which are accountable for the source terms of individual radionuclides. Fig. 6 is a schematic illustration for the solid-phase
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conversion and the generation of mobile aqueous species, which depicts the near-field chemistry. The source-term assessment practised up to now has been relied solely on the solubility of each primary solid waste, e.g. spent fuel, HAW glass or cement, without taking into account the solid-phase conversion. The formation of each secondary solid phase is a thermodynamic process but is kinetically controlled with respect to the dissolution rate of the primary solid phase and advection of groundwater. Therefore, the condition can be considered as either a closed or an open system, since the radioactive elements concerned can be present in either the equilibrium or non-equilibrium state. The secondary solid phases are always found in the laboratory dissolution experiment of solid waste. As a typical laboratory example, Fig. 7 demonstrates the formation of various secondary mineral phases from the HAW glass dissolution, i.e. smectite, powellite and baryte, that may incorporate M(III) and M(IV) as solid solutions in clay minerals, M(III) in powellite and M(II) in baryte (Grambow et al., 1997). The predictive modelling of such secondary mineral phase formation under near-field natural conditions may be practicable by a thermodynamic
approach, when sufficient knowledge on such chemistry is attained from laboratory simulations. Along with the secondary solid-phase formation, the new aqueous species are produced in the near field, which accompanies the colloid generation. In this, two different kinds of colloids can be distinguished: ‘real’ colloids that consists of radioactive elements themselves, e.g. M(IV); and ‘pseudo’ colloids that are composed by the sorption of radionuclides onto aquatic colloids present in the system. In the near field, the generation of real colloids is less likely because the pseudo colloid formation prevails. The generation of pseudo colloids in a laboratory system can be described by a thermodynamic approach. However, a direct application of such results to dynamic natural systems is not possible, but the laboratory experience provides an insight into the possible geochemical processes for the colloid generation. The extent of the pseudo colloid generation in the near field is difficult to predict but a simulation experiment may facilitate its approximation, e.g. the maximum possible concentration. As a simplified example, the time-dependent dissolution of actinides under near-field conditions is illustrated in Fig. 8. The saturation concentration
Fig. 7. Formation of secondary phases from HAW glass dissolution (after 1 year of corrosion of Cogema-type glass in Mg-rich salt solution at 190°C) (Grambow et al., 1997). Comparison of analysed mineral composition with geochemical modelling.
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Fig. 8. Schematics of controls for the dissolution behaviour of actinides (Kim and Grambow, 1999).
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the elements of other oxidation states display a varying degree of colloid formation. A significant amount of colloids is generated by M(III) and M(IV), to a lesser extent by M(VI), followed by M(V). Since M(III), M(V) and M(VI) are not subject to the real colloid generation, the results, as evidenced in Fig. 9, are certainly attributable to their pseudo colloids. The thermodynamic approach is not capable of describing precisely the radionuclide behaviour in dynamic geochemical systems, such as the solubility, the migration behaviour, etc. For the nearfield chemistry, thermodynamic modelling is, nonetheless, an indispensable tool for the approximate assessment of given geochemical conditions and also for the semi-quantitative appraisal of the possible geochemical behaviour of individual radionuclides.
4.2. Migration in far-field systems (open systems)
Fig. 9. Dissolved fraction (%) of radionuclides from powdered, high burn-up, spent fuel samples after 200 days of corrosion in granitic water at 25°C (Geckeis et al., 1998).
of a given actinide element comprises, then, the thermodynamic solubility and the maximal possible colloid concentration. Therefore, the sourceterm quantification of actinides is difficult to attain by a thermodynamic approach alone but can be supported by an empirical approach based on a variety of simulation experiments. Fig. 9 illustrates the dissolution of radionuclides from the corrosion of spent fuel in granite groundwater at 25°C, as given by relative release rates (%) after 200 days of water contact (Geckeis et al., 1998). The colloid concentration of individual radioactive elements is quantified by ultrafiltration at approximately 2 nm pore size. M(I) and M(VII) do not show the colloid formation, while
The migration from the near field to the far field of a given repository results in a dilution of individual radionuclide concentrations to a level much less than their solubility, i.e. in the nanomole range for actinides. Therefore, in the far-field aquifer system, the chemical properties of radionuclides are expected to submerge in the chemical behaviour of aquatic constituent elements, particularly chemical homologues, present in much higher concentrations. If actinides are present as complexed species in aquatic phase, like carbonate or humate, etc., their speciation can be still made by a thermodynamic approach, but only qualitatively. Once the system becomes dynamic by aquatic flow, the kinetics of individual reactions become more important, since many reactions are driven to a non-equilibrium state in the open system. The geochemical behaviour of radionuclides in such dynamic systems calls for an appropriate scientific mean to describe. For a lack of substantiated solutions, the migration behaviour of radionuclides in the far-field environment has been parameterised by the socalled ‘Kd concept’, which takes into account only the radionuclide concentration distribution in a given aquifer system between water and solid
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matrices. The simplicity of Kd facilitates, obviously, a numerical modelling of the radionuclide migration. However, its applicability to the longterm performance assessment of a given geological barrier, the far field, has been attended with controversial debates, because its ‘black-box’-like concept does not provide an insight into the underlying geochemical reactions. Fig. 10 illustrates a schematic presentation of solid–water interactions of M(III) in one of the deep Gorleben aquifer systems (Gohy 573). The naturally occurring cerium, Ce(III), is determined in groundwater, colloids and geo-matrix present in equilibrium for many thousand years. Concentrations of Ce(III) found in the three different phases enable the calculation of their distribution ratios (R) among three given phases. As is apparent from this figure, an over 95% of Ce(III) is present in a colloidal state, which can hardly be exchanged by Ce(III) incorporated in the geo-matrix. The distribution of Ce(III) between groundwater, including colloid-borne, and geo-matrix gives a concentration ratio of R =483. According to the Kd concept, this value leaves us to perceive a considerable retardation in the migration of Ce(III) in the aquifer system. If groundwater colloids migrate without retardation, the colloidborne Ce(III) is possibly accompanied by the same fate. As a result, the Kd concept misrepresents the real situation in nature.
As an alternative to the Kd concept, surface complexation, solid–water interface reactions of individual radionuclides onto a composite geomatrix, has been put into modelling either to replace Kd or possibly to substantiate it. The development of the surface complexation model (SCM) is a tedious undertake, since it entails a wide range of experiments for individual radionuclides with respect to a variety of natural mineral components under different geochemical conditions. The hitherto endeavoured SCM concept is based faithfully on a thermodynamic approach (Dzombak and Morel, 1990); nonetheless, its modelling capability is not sufficiently developed enough for the application to natural systems. This concept can describe the sorption behaviour of some radionuclides in laboratory system onto mineral surfaces (cf. Fig. 5) and also onto aquatic colloids of homogeneous nature (cf. Fig. 4). Shall it be, one day, applicable to natural dynamic systems? In the far field, as shown in Fig. 10, the primary migration process of radionuclides, particularly actinides, may be attributable to the mobilisation of colloid-borne species. Pseudo colloids may facilitate the actinide migration without retardation as fast as the water flow velocity in porous aquifer systems. The migration behaviour of colloidborne actinides depends on their chemical binding state onto colloids: whether they are bound reversibly or irreversibly. The reversible binding
Fig. 10. Solid –water interactions of M(III) in a natural aquifer system: distribution of Ce(III) in solution, solid and colloidal phases.
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Fig. 11. Desorption kinetics of Eu(III) from humic colloids: comparison between Eu(III) sorbed in the laboratory and Eu(III) bound naturally to humic colloids (Geckeis et al., 1999).
Fig. 12. Retardation and migration of different radionuclide species relative to water transport velocity.
results in a competitive reaction of a given actinide between the surfaces of colloid and geo-matrix, and hence leads to its retardation of migration, whereas the irreversible binding promotes colloids as a carrier for the actinide migration along with the water flow. Fig. 11 shows a laboratory desorption experiment of Eu(III), which is sorbed onto humic colloids at different reaction times in laboratory (Geckeis et al., 1999). For the purpose of comparison, the same desorption experiment is performed also for Eu(III) that is already present as sorbed on humic colloids in two of the Gorleben groundwaters. To accelerate the reaction, a chelate cation exchanger as solid form is introduced in each groundwater sample. As appears in
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this figure, the desorption of Eu(III) bound onto colloids in natural groundwater is much slower than that to form colloids onto which Eu(III) is sorbed in laboratory. The sorption and desorption behaviour of Eu(III) onto and from humic colloids in laboratory system can be described by the surface complexation model, whereas, for the natural system, the reaction behaviour of Eu(III) cannot be appraised by the same approach, because its irreversible sorption appears to prevail in natural colloids. Fig. 12 shows a comparison of two different migration processes. Radionuclides, which undergo reversible sorption reactions with geo-matrices, may be treated for their migration by an ion exchange process with its retardation parameter Rf. To this category belong M(I) and M(II) and, to a certain extent, also M(V) and M(VI). The colloid-borne radionuclides may migrate as fast as the water flow velocity if they are irreversibly bound onto mobile colloids; nonetheless, if they are bound irreversibly onto colloids of unstable nature, then filtration or precipitation accelerates their retardation. To this category belong M(III) and M(IV). A thermodynamic approach is not available to describe such natural dynamic reactions. A combination of Figs. 10–12 explains the following facts. (1) A thermodynamic approach is not yet directly applicable to dynamic natural systems; for example, the quantitative appraisal of the radionuclide migration in the far field. However, it is an indispensable tool for understanding the underlying geochemical reactions as well as for the assessment of key uncertainties involved in the available migration modelling. (2) The thermodynamic approach fails to describe the migration behaviour of colloid-borne radionuclides. Still, then, the colloid generation either by homogeneous nucleation (real colloid formation) or by heterogeneous nucleation (pseudo colloid formation) can be approximated by an extra thermodynamic approach. The geochemical modelling of the colloid-borne radionuclide migration is yet to be developed, for which empirical experience on the underlying kinetic and thermodynamic processes is an essential prerequisite.
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5. Perspectives A large number of thermodynamic data are available for the appraisal of geochemical reactions of long-lived radionuclides, actinides and some fission products, which have been integrated in various geochemical codes or in coupled codes. However, the hitherto available data do not satisfy the necessary modelling of all significant geochemical reactions of essential radionuclides, particularly actinides. The present situation is summarised in a simplified form as follows. What do we know at present? 1. Aquatic chemical reactions of radionuclides in laboratory systems (closed systems). Thermodynamics of solubility and complexation (mostly binary complexes). Speciation of solid and solution phases (mostly binary complexes). 2. Thermodynamic speciation modelling by various geochemical codes (closed system). Capability is limited largely to binary complexes. Speciation of M(IV) is limited to a large extent by a lack of its relevant thermodynamic data. 3. Solid–water interface interaction of radionuclides in trace concentrations for reversible processes (for radionuclides of Z E3+ , the modelling is still in the developing state). 4. Phenomenological understanding of radionuclide migration processes. What do we need to know? 1. Aquatic chemical reactions of radionuclides in natural multi-component systems (open systems). 2. Surface complexation models (SCM) for the broad application: for composite geo-matrices in natural dynamic systems (to replace the Kd concept!). 3. Irreversible geochemical reactions of radionuclides. Secondary solid-phase formation in the near field: formation of solid solution of radionuclides. Colloid generation: both homogeneous and heterogeneous nucleation.
4. Key uncertainties involved in the application of thermodynamics to natural dynamic systems (open and closed systems). 5. Applicability of the laboratory results to natural systems, for which the following prevalent conditions are to be taken into account. Laboratory system Closed system Short-term reactions Equilibrium reactions Reversible processes
Natural systems Open system Long-term reactions Non-equilibrium reactions Irreversible processes
The presently available knowledge on the aquatic thermodynamics of actinides and fission products facilitates largely the understanding of their prevalent geochemical behaviour, which hence makes it possible to distinguish the essential reactions from a bulk of convoluted geochemical reactions. For laboratory multi-complex systems, the thermodynamic speciation of underlying geochemical reactions is largely possible, whenever the relevant thermodynamic data are available. For various reasons, as discussed in this paper, the applicability of a thermodynamic approach to dynamic natural systems is considerably limited. Nonetheless, the thermodynamic knowledge is of cardinal importance for understanding a variety of far-reaching geochemical reactions in the near and far fields, and allows the quantification of, if not altogether at once, individual isolated geochemical reactions separately. For the analysis of key uncertainties involved in the long-term performance assessment of a given repository, a thermodynamic approach can play a significant role for the corroboration of parameters describing the given geochemical conditions, like pH, Eh etc., the probable source terms of individual radionuclides and their migration behaviour. For this purpose, well-founded thermodynamic data for a wide range of geochemical reactions of actinides are absolutely necessary.
References Chopin, G.R., 1983. Solution Chemistry of the Actinides. Radiochim. Acta 32, 43.
J.I. Kim / Nuclear Engineering and Design 202 (2000) 143–155 Chung, K.H., Klenze, R., Park, K.K., Paviet-Hartmann, P., Kim, J.I., 1998. A study of the surface sorption process of Cm(III) on silica by time-resolved laser fluorescenec spectroscopy (I). Radiochim. Acta 82, 215. Davis, J.A., Coston, J.A., Kent, D.B., Fuller, C.C., 1998. Application of the surface complexation concept to complex mineral assemblages. Environ. Sci. Technol. 32, 2820. Dzombak, D., Morel, F.M.M., 1990. Surface Complexation Modeling Hydrous Ferric Oxide. Wiley, New York. Geckeis, H., Grambow, B., Loida, A., Luckscheiter, B., Smailos, E., Quinones, J., 1998. Formation and stability of colloids under simulated near field conditions. Radiochim. Acta 82, 123. Geckeis, H., Rabung, Th., Kim, J.I., 1999. Kinetic aspects of metal ion binding to humic substances, In: Buckau, G. (Ed.), Effects of Humic Substances on the Migration of Radionuclides: Complexation and Transport of Actinides, Second Technical Progress Report FZKA 6324, Karlsruhe Research Centre (FZK), p. 47. Grambow, B., Loida, A., Dressler, P., Geckeis, H., Gago, J., Casas, I., de Pablo, J., Gime´nez, J., Torrero, M.E., 1997. Chemical Reaction of Fabricated and High-Burnup Spent UO2 Fuel with Saline brine, EUR 17111 EN, European Commission, Brussels. Grenthe, I., Fuger, J., Konings, R.J.M., Lemire, R.J., Muller, A.B., Nguyen-Trung, C., Wanner, H., 1992. Chemical Thermodynamics of Uranium: NEA-TDB, OECD. North-Holland, Amsterdam.
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Kim, J.I., 1993. The chemical behavior of transuranium elements and barrier functions in natural aquifer systems. In: Interante, C.G., Pablan, R.T. (Eds.), Scientific Basis for Nuclear Waste Management XVI, Vol. 294. Materials Research Society Symposium Proceedings, Pittsburgh, PA, p. 3. Kim, J.I., 1994. Actinide colloids in natural aquifer systems. Mater. Res. Soc. Bull. 19, 47. Kim, J.I., Grambow, B., 1999. Geochemical assessment of actinide isolation in a German salt repository environment. Eng. Geol. 52, 221. Kim, J.I., Kanellakopulos, B., 1989. Solubility products of plutonium(IV) oxide and hydroxide. Radiochim. Acta 48, 145. Meinrath, G., 1991. Carbonat-komplexierung des dreiwertigen Americiums unter grundwasserbedingungen. Thesis. Technische Universita¨t Mu¨nchen. Neck, V., Fangha¨nel, T., Kim, J.I., 1997. Mixed hydroxo-carbonate complexes of neptunium(V). Radiochim. Acta 77, 167. Neck, V., Fangha¨nel, Th., Kim, J.I., 1998. Aquatische Chemie und Thermodynamische Modellierung von trivalenten Aktiniden, Report FZKA 6110. Forschungszentrum Karlsruhe. Stumm, W., 1987. Aquatic Surface Chemistry: Chemical Processes at the Particle – Water Interface. Wiley, New York. Waite, T.D., Davis, J.A., Fenton, B.R., Payne, T.E., 2000. Approaches to modelling Uranium(VI) adsorption on natural mineral assemblages. Radiochim. Acta (in press).
Is the thermodynamic approach appropriate to describe natural dynamic systems? (Status and limitations) J.I. Kim * Institut fu¨r Nukleare Entsorgungstechnik (INE), Forschungszentrum Karlsruhe GmbH, P.O. Box 3640, 76021 Karlsruhe, Germany Accepted 20 March 2000
Abstract The long-term performance assessment of a nuclear waste repository entails sound knowledge on the migration behaviour of radionuclides in a given aquifer system, which can in principle be appraised by a well-founded thermodynamic approach. For various reasons, in such an endeavour, a number of uncertainties are involved, which cause difficulties in the application of a thermodynamic approach to natural dynamic systems. This paper summarises the present state of the thermodynamic approach being applied to the assessment of various geochemical reactions of radionuclides and discusses the limitations involved in application to laboratory systems as well as dynamic natural systems. © 2000 Elsevier Science B.V. All rights reserved.
1. Introduction The geochemical behaviour of radionuclides in aquifer systems has been described, to a certain extent, by a thermodynamic approach. The thermodynamic data appraised for this purpose are, for obvious reasons, derived from individual equilibrium reactions of laboratory system, namely from experiments of a ‘closed’ system. The real natural aquifer system is dynamic and ‘open’, per definition in opposition to a laboratory system, for emerging geochemical reactions of individual radionuclides. Along the migration pathway, a given radionuclide in natural aquifer systems in-
* Tel.: +49-7247-822231; fax: +49-7247-824308. E-mail address: [email protected] (J.I. Kim).
teracts with a variety of geochemical surroundings. In this, although pH and Eh values may remain largely invariant for certain cases, the water-constituent components change their concentrations, and the specific or gross reactivity of geo-matrix surfaces varies with space and time. Hence, the radionuclide concerned undergoes a multiplicity of nanoscopic reactions along its migration, in which individual processes may be not equilibrated or irreversible. For the appraisal of such complexities, a challenging question indubitably arises: ‘is the thermodynamic approach appropriate to describe natural dynamic systems?’ The question inevitably triggers off another: ‘is there any alternative scientific tool available then?’. This paper summarises the present state of available thermodynamic approaches that can be
0029-5493/00/$ - see front matter © 2000 Elsevier Science B.V. All rights reserved. PII: S 0 0 2 9 - 5 4 9 3 ( 0 0 ) 0 0 3 4 0 - X
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applicable for the performance assessment of a multi-barrier system, which is presumed to function as a safe confinement for a given nuclear repository. Limitations of the approaches, as perceived for the moment, are to come into view simply as repercussions of the present discussion. To provide a simplified overview on a multiplicity of convoluted chemical reactions in natural systems, the present discussion is separated into the following three areas: the laboratory systems (closed systems), the natural systems (open systems), and the applicability of laboratory knowledge to natural systems.
2. What do we need to know? All long-lived radionuclides to be disposed of are immobilised in different solid phases, called engineered barriers, that are contained by backfills known as geo-engineered barriers and then by natural geological barriers. Whatever the perceived scenarios are to come after, the long-term performance assessment of such a multi-barrier system entails, upon water intrusion, the quantification of the geochemical behaviour of individual radionuclides in the environment of each barrier. There are two gross geochemical reactions, namely immobilisation and mobilisation of radionuclides in each specific barrier. They are reciprocated in nature. The balance of the two reactions for a given radionuclide in each barrier can be appraised by the ratio of immobilisation rate to mobilisation rate. How can the two reaction rates be quantified with high certainty? Can they be delivered by a thermodynamic approach?
Fig. 1. Chemical states of radionuclides controlling their mobility in natural aquatic systems.
Whatever the engineered barriers are made of, they are thermodynamically unstable in the nearfield environment once water comes into contact with them, and undergo a variety of geochemical reactions. Solid phases of waste are first subject to corrosion, leaching and dissolution of individual components, the processes of which generate the ‘new’ secondary solid phases that appear to be stable thermodynamically as compared with the primary solid phases in the near-field environment. Upon the phase conversion of engineered barriers, the radionuclides dissolved are submitted to hydrolysis, complexation, redox reaction and colloid generation, and largely immobilised by incorporation into the secondary solid phases that are partially equilibrated in the new environment. Under these conditions, the chemistry of radionuclides in the aqueous phase is submerged into the geochemical condition evolved from water interactions with waste, backfill and surrounding geological matrices. The chemical states of all metallic radionuclides to be disposed of can be classified on the basis of their oxidation states, from mono-valent to heptavalent, namely M(I), M(II), M(III), M(IV), M(V), M(VI) and M(VII). To this category belong, as illustrated in Fig. 1, all radioactive elements of long half-life. The two characteristics are recognised for all elements in aquifer systems: the oxidation state specific chemical behaviour, on the one hand, and the chemical reactions governed by the effective charge of individual ions, on the other. These characteristics can be followed as a primary guide for understanding of their general chemical behaviour in both laboratory and natural systems. For M(I) and M(II), much of their geochemical behaviour is known, since they belong to the natural element groups of major component both in aqueous and solid phases, and undergo reversible chemical reactions for the most part. The important nuclides belonging to these categories are 237Cs(I) and 90Sr(II), as fission products, and 226 Ra(II), as one of the decay series of 238U. As these elements are naturally present, their thermodynamic behaviour can be correlated to the chemical systematic of alkali metal ions and alkaline earth metal ions. For actinides, which are present
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under natural conditions as M(III), M(IV), M(V) and M(VI), their chemical behaviour is much more difficult to describe, since they are submitted to a multiplicity of geochemical reactions. However, for M(VI), the chemistry of U(VI) (Grenthe et al., 1992) observable in nature can be closely followed, to which transuranic actinides of the same oxidation state can be compared to a certain extend. To M(VII) belongs only 99Tc and, if present as TcO− 4 , it is least interactive with geochemical surroundings. Under reducing conditions, which are mostly the case for deep aquifers, the actinides of M(V) and M(VI), as well as Tc(VII), become reduced to M(III) or M(IV). Any chemical compound of M(IV) is sparingly soluble in water at neutral pH and the particular chemical properties lead to its strong tendency toward the colloid formation.
3. Applicability of a thermodynamic approach to laboratory systems (closed systems) All thermodynamic constants relevant to the appraisal of the geochemical behaviour of radionuclides are determined by laboratory experiments, besides those evaluated by indirect approaches. The application of the data to describe the geochemical behaviour of radionuclides in laboratory systems appears, therefore, accountable. However, the geochemical study in the laboratory is largely made with geological samples under simulated natural conditions. In the laboratory, such conditions are kept in a closed state but the radionuclides introduced are exposed to a variety of geochemical reactions. Therefore, the so-called laboratory system deals with pure samples as well as natural geological samples, for which the speciation is performed in order to evaluate the applicability of a thermodynamic approach. In the present paper, the laboratory system is discussed accordingly. To facilitate discussion, the subject is divided into three areas of chemical speciation: aqueous phase, solid phase, and solid – water interface. Although the chemical reactions of radionuclides are convoluted within the three phases, their reactions can be discussed separately in individual phases.
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3.1. Speciation in the aqueous phase In natural aquifers, the dominant chemical reactions of metal ions, in particular actinide ions, are hydrolysis and carbonate complexation, followed by many minor complexations with waterconstituent ligands depending on their concentrations, such as chloride, sulphate, phosphate, etc. (Kim, 1993). Humic acid is also one of the important aquatic constituents, with its concentration varying widely in different groundwater (Kim, 1993). Hydrolysis is the most important primary geochemical reaction for actinides and its tendency follows the effective charge as given by (Chopin, 1983): An4 + \ AnO22 + ] An3 + \ AnO+ 2 The effective charges of AnO22 + and AnO+ 2 are known as 3.29 0.1 and 2.39 0.2, respectively (Chopin, 1983), because their linear oxygen bindings expose the equatorial charge larger than the nominal charge. This tendency is also followed by other complexation, depending on the ligand concerned. Hydrolysis facilitates the colloid generation of actinides through a formation of oxygen bridges, either real or pseudo colloids (Kim, 1994). The speciation can be reasonably made for aqueous actinide ions based on hitherto known thermodynamic constants for binary reactions of hydrolysis and complexation with some important water-constituent ligands. However, for the ternary complexation, the experimental constants are poorly known. The relevant thermodynamic data for An4 + are largely missing, because this ion endures strong hydrolysis that leads to colloid formation in aqueous phase. The present state of knowledge for the thermodynamic speciation of actinides and some important fission products in aqueous phase is summarised as follows. For binary reactions: M(I), no M(III), sufficomplexation ciently known
M(V), sufficiently known
M(II), well known
M(VI), sufficiently known
M(IV), poorly known
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Fig. 2. Stability regions of aqueous hydroxo and carbonato species of Am(III) and Cm(III) in 1 M NaCl at 25°C; (1yz) = An(OH)y (CO3)3z − y − 2z; hatched area, possible stability region of mixed species (Neck et al., 1997).
The ternary complexation is poorly known for all of M(III), M(IV), M(V) and M(VI). Because of the high electric charge, the ternary complexation of various kinds is expected particularly for M(IV), and to a somewhat less extent for M(III) and M(VI). They are presumed to be predominantly hydroxo complexes, which are difficult to speciate experimentally. However, their formation constants may be approximated for some particular cases on the basis of each corresponding binary complexation. A typical example for the speciation of aqueous species of M(III) is illustrated in Fig. 2, which is calculated based on the hitherto known thermodynamic constants (Neck et al., 1997). This figure demonstrates the applicability of the thermodynamic speciation in the aqueous phase under various geochemical conditions, i.e. at different pH and PCO2 values, for the formation of hydroxides (1Y0), carbonates (10Z) and hydroxo carbonates (1YZ). In laboratory systems, the thermodynamic speciation is for the moment practicable for M(III), M(V) and M(VI), at different ionic strength, either in aqueous phase under controlled conditions or in natural groundwater. A further extension of
the thermodynamic database for aqueous actinide ions of various species, particularly for M(IV), is still necessary. In all groundwater, colloids are always present, which are composed of inorganic, organic or a combination of both components (Kim, 1994). They are produced either by heterogeneous nucleation of water-constituent metal ions through oxygen bridges or by dispersion of the surface composites of geo-matrices. They are small in size (B 100 nm) and chemically reactive for polyvalent radionuclides. Such radionuclides are readily sorbed onto groundwater colloids and thus present in groundwater as so-called pseudo colloids. The speciation of these colloidal actinide species is difficult to realise but if given groundwater colloids are well characterised, the formation of pseudo colloids can be modelled. Nonetheless, a new model approach is yet to be developed to describe their formation process, structural conformation and stability.
3.2. Speciation in the solid phase There are two solid phases to be dealt with: pure solid phases of actinides, which are composed of actinides themselves as major components, and solid solutions that contain actinides in a bulk of other solid phases. On the former species, the extent of our present knowledge is somewhat favourable. Even so, there is a large number of various solid species of actinides that remain not apprehended for their formation under aquatic conditions. The knowledge on the latter species is virtually non-existent, besides on uranium in natural minerals (Grenthe et al., 1992). However, for actinide solid solutions, the oxidation state specific natural analogues may be taken for the comparative purpose, although their applicability is not always straightforward, e.g. rare earth elements for M(III), Th or Zr for M(IV), U for M(VI), and Re for Tc(VII). The present knowledge on thermodynamics of pure solid phases for the oxidation state specific actinides and some fission products is summarised as: M(I), well known
M(III), sufficiently known
M(V), sufficiently known
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M(II), well known
M(IV), poorly known
M(VI), sufficiently known
The pure solid-phase formation is directly related to the speciation in the aqueous phase, e.g. the conversion of the solid phase is accountable for the change in the solubility of a given element. A typical example (Neck et al., 1998) is shown in Fig. 3. The solubility experiment of Am(III) as a function of pH at PCO2 =1 and PCO2 = 10 − 2 demonstrates a continuous correlation, as shown in the upper part of Fig. 3. Once the thermodynamic speciation of solid phases is applied, as
Fig. 3. Upper part: Solubility of Am(III) in 0.1 M NaClO4 (Neck et al., 1998); experimental data from Meinrath (1991). Lower part: Stability regions of Am(III) solid phases at I = 0.1 M and 25°C (Neck et al., 1998).
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shown in the lower part of Fig. 3, it becomes apparent that, during the experiment, the solid phase is converted from the original amorphous Am2(CO3)3 at near-acidic pH to a ternary phase of AmOHCO3 at neutral pH, and further to NaAm(CO3)2 at near-alkaline pH. A thermodynamic approach is thus applicable to laboratory systems for the prediction of solid-phase conversion as a function of both pH and PCO2. The estimation of the Am(III) solubility under dynamic geochemical conditions is also possible, once such a conversion becomes predictable. The speciation in pure solid phases is also possible, based on the present knowledge, for the actinides of M(III), M(V) and M(VI). For M(IV), more pertinent thermodynamic data are yet to be known for this purpose. In dissolution equilibrium, the solubility of a given solid phase of M(IV) is difficult to determine, because M(IV) tends to form colloids. This fact disturbs the speciation of M(IV) solid phases equilibrated in water and, as a consequence, the solubility of any M(IV) compound is known poorly. Although the solubility of MO2 has been determined by a large number of experiments, the reported solubility product varies considerably for all actinides of M(IV). The experimental solubility product of PuO2, for example, varies from log Ksp = −51.8 to log Ksp = − 60.2, whereas its calculated value appears to be log Ksp = − 63.8 (Kim and Kanellakopulos, 1989). Such a large range of difference is mainly attributable to the colloid generation, which is sensitively affected by individual experimental conditions. The speciation of actinides in solid solution is, as mentioned already, hardly possible, because the relevant thermodynamic data are not available for the moment. Re-immobilisation of long-lived radionuclides, after dissolution of the primary solid waste packages, depends directly on the subsequent formation of solid solution. Therefore, knowledge on the secondary solid-phase formation under near-field geochemical conditions is an essential prerequisite of the source term quantification for the long-term performance assessment. For the moment, only laboratory simulation experiments can provide such knowledge.
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3.3. Solid–water interface reactions The migration behaviour of individual radionuclides in geo-engineered and geological barriers depends on their solid – water interface reactions at surfaces of available geo-matrices, which are composites of different minerals. At near-neutral pH, the surface of geo-matrices is in generally covered with the surface-bound anionic oxygen or neutralised hydroxide. Cationic radionuclides interact on the surface to form either ion association (outer sphere complex) or complexation with surface oxygen ligands (inner sphere complex). The former process is valid largely for M(I) and M(II), while M(III), M(IV), M(V) and M(VI) attend the latter reaction. This concludes that aqueous actinide ions undergo mainly the stable surface complexation. In addition, another possible interface reaction is the surface precipitation (thereafter mineralisation), to which metal ions of very low solubility are subjected, i.e. M(III), M(IV) and M(VI). The surface sorption behaviour of a given radionuclide is often quantified simply by its concentration distribution between solid geo-matrix and water. The quotient is commonly known as Kd, which is an operational parameter borrowed from the conventional ion exchange separation. Whereas a laboratory ion exchanger bears reversible chemical reactions, the sorption process of metal ions onto geo-matrices carries on both reversible and irreversible reactions, which are convoluted each other with time. The so-called ‘Kd concept’ is often used to calculate the migration behaviour of individual radionuclides in a given aquifer system, although the data are mostly derived from sort-term laboratory experiments. Since Kd is a simple operational parameter, which depends on various unknown geochemical effects involved, it accordingly has no solid thermodynamic foundation; its applicability to the long-term performance assessment has become the subject of controversial debate. An alternative to the Kd concept is to evaluate thermodynamic parameters for the complexation of individual radionuclides onto the surface of geo-matrices. This can be realised by the study of
solid–water interface reactions for a given radionuclide on a number of single minerals and thereafter by a combination of the results proportionally to a composite model. The development of such a surface complexation model is in progress in a number of laboratories worldwide (Davis et al., 1998). A simplified reaction process on each mineral surface is: − lp n(.SOH) + MLm p − p + n/l − n l (.SO)n MLm + (n/l)L+ nH+ p − n/l
Ksc
where n(SOH) represents the number (n) of surface functional sites to be interacted with a given − lp metal ion species, MLm , complexed with the p p l− number of ligands L . In laboratory systems, the pKa values of n(SOH) for individual minerals and the complexation constants of the metal ion − lp species concerned MLm can be determined p and, hence, the quantification of the presented reaction becomes possible as a function of the pH and L concentration. However, the surface sorbed species is to be known for this purpose either by direct or by indirect speciation. Once the surface complexation constants, Ksc, of different radionuclides are known for various minerals, a composite model can be produced for a geo-matrix of mineral mixture. An integrated modelling procedure is still to be developed for a broad application, although a number of different surface complexation modellings for particular laboratory examples are available in the literature (Stumm, 1987; Dzombak and Morel, 1990). A typical example of the surface complexation study is shown in Fig. 4, in which the laser-spectroscopic speciation of the Cm(III) sorption onto silica surfaces of colloidal suspension is compared with thermodynamic modelling (Chung et al., 1998). The two kinds of surface sorbed species are recognised by spectroscopy with pH variation, from which their formation constants can be derived as corroborated by thermodynamic calculation (dark lines). Another example is shown in Fig. 5 for a comparison of the surface specific sorption (Ka) of U(VI) onto different minerals and mineral composites as a function of pH (Waite et al., 2000).
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Contrary to the Kd concept, the surface complexation modelling is a thermodynamic approach that provides a deep insight into the underlying chemical reactions at solid–water interfaces. While this approach facilitates the quantification of the primary solid–water interface reactions, its applicability to the appraisal of the long-term geochemical behaviour of radionuclides in a given aquifer system is still to be corroborated by a wide range of experiments. The present state of knowledge on solid–water interface reactions of pertinent elements is as follows: Fig. 4. Laser spectroscopic speciation of Cm(III) sorbed on 0.5 g/l silica. Dark lines, fit to a surface complexation model (Chung et al., 1998).
M(I), sufficiently known M(II), sufficiently known
M(III), poorly known M(IV), not known
M(V), somewhat known M(VI), somewhat known
As M(IV) forms colloids (either real or pseudo colloids) at neutral pH, its solid–water interface reactions on geo-matrices are difficult to quantify.
4. Applicability of a thermodynamic approach to natural systems (open systems)
4.1. Source-term quantification/near-field chemistry (closed/open systems)
Fig. 5. Surface-specific sorption of U(VI) on various natural minerals as a function of pH (Waite et al., 2000).
Fig. 6. Geochemical reactions that govern the generation of mobile radionuclide species and their migration.
Practical examples can be taken from the source-term quantification, which is made up within the near-field geochemistry that can be considered as either closed or open system. When the near field is intruded by fresh water and the corrosion of waste packages proceeds, the geochemistry of the environment changes above all pH, Eh and PCO2 values. This change can be modelled by a thermodynamic approach once the corrosion and leaching products can be assessed with time. However, most important is to appraise the formation of secondary solid phases produced from dissolution of the original solid waste and its packages. The solubility of each new solid phase then results in the aqueous concentrations of radionuclides, which are accountable for the source terms of individual radionuclides. Fig. 6 is a schematic illustration for the solid-phase
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conversion and the generation of mobile aqueous species, which depicts the near-field chemistry. The source-term assessment practised up to now has been relied solely on the solubility of each primary solid waste, e.g. spent fuel, HAW glass or cement, without taking into account the solid-phase conversion. The formation of each secondary solid phase is a thermodynamic process but is kinetically controlled with respect to the dissolution rate of the primary solid phase and advection of groundwater. Therefore, the condition can be considered as either a closed or an open system, since the radioactive elements concerned can be present in either the equilibrium or non-equilibrium state. The secondary solid phases are always found in the laboratory dissolution experiment of solid waste. As a typical laboratory example, Fig. 7 demonstrates the formation of various secondary mineral phases from the HAW glass dissolution, i.e. smectite, powellite and baryte, that may incorporate M(III) and M(IV) as solid solutions in clay minerals, M(III) in powellite and M(II) in baryte (Grambow et al., 1997). The predictive modelling of such secondary mineral phase formation under near-field natural conditions may be practicable by a thermodynamic
approach, when sufficient knowledge on such chemistry is attained from laboratory simulations. Along with the secondary solid-phase formation, the new aqueous species are produced in the near field, which accompanies the colloid generation. In this, two different kinds of colloids can be distinguished: ‘real’ colloids that consists of radioactive elements themselves, e.g. M(IV); and ‘pseudo’ colloids that are composed by the sorption of radionuclides onto aquatic colloids present in the system. In the near field, the generation of real colloids is less likely because the pseudo colloid formation prevails. The generation of pseudo colloids in a laboratory system can be described by a thermodynamic approach. However, a direct application of such results to dynamic natural systems is not possible, but the laboratory experience provides an insight into the possible geochemical processes for the colloid generation. The extent of the pseudo colloid generation in the near field is difficult to predict but a simulation experiment may facilitate its approximation, e.g. the maximum possible concentration. As a simplified example, the time-dependent dissolution of actinides under near-field conditions is illustrated in Fig. 8. The saturation concentration
Fig. 7. Formation of secondary phases from HAW glass dissolution (after 1 year of corrosion of Cogema-type glass in Mg-rich salt solution at 190°C) (Grambow et al., 1997). Comparison of analysed mineral composition with geochemical modelling.
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Fig. 8. Schematics of controls for the dissolution behaviour of actinides (Kim and Grambow, 1999).
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the elements of other oxidation states display a varying degree of colloid formation. A significant amount of colloids is generated by M(III) and M(IV), to a lesser extent by M(VI), followed by M(V). Since M(III), M(V) and M(VI) are not subject to the real colloid generation, the results, as evidenced in Fig. 9, are certainly attributable to their pseudo colloids. The thermodynamic approach is not capable of describing precisely the radionuclide behaviour in dynamic geochemical systems, such as the solubility, the migration behaviour, etc. For the nearfield chemistry, thermodynamic modelling is, nonetheless, an indispensable tool for the approximate assessment of given geochemical conditions and also for the semi-quantitative appraisal of the possible geochemical behaviour of individual radionuclides.
4.2. Migration in far-field systems (open systems)
Fig. 9. Dissolved fraction (%) of radionuclides from powdered, high burn-up, spent fuel samples after 200 days of corrosion in granitic water at 25°C (Geckeis et al., 1998).
of a given actinide element comprises, then, the thermodynamic solubility and the maximal possible colloid concentration. Therefore, the sourceterm quantification of actinides is difficult to attain by a thermodynamic approach alone but can be supported by an empirical approach based on a variety of simulation experiments. Fig. 9 illustrates the dissolution of radionuclides from the corrosion of spent fuel in granite groundwater at 25°C, as given by relative release rates (%) after 200 days of water contact (Geckeis et al., 1998). The colloid concentration of individual radioactive elements is quantified by ultrafiltration at approximately 2 nm pore size. M(I) and M(VII) do not show the colloid formation, while
The migration from the near field to the far field of a given repository results in a dilution of individual radionuclide concentrations to a level much less than their solubility, i.e. in the nanomole range for actinides. Therefore, in the far-field aquifer system, the chemical properties of radionuclides are expected to submerge in the chemical behaviour of aquatic constituent elements, particularly chemical homologues, present in much higher concentrations. If actinides are present as complexed species in aquatic phase, like carbonate or humate, etc., their speciation can be still made by a thermodynamic approach, but only qualitatively. Once the system becomes dynamic by aquatic flow, the kinetics of individual reactions become more important, since many reactions are driven to a non-equilibrium state in the open system. The geochemical behaviour of radionuclides in such dynamic systems calls for an appropriate scientific mean to describe. For a lack of substantiated solutions, the migration behaviour of radionuclides in the far-field environment has been parameterised by the socalled ‘Kd concept’, which takes into account only the radionuclide concentration distribution in a given aquifer system between water and solid
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matrices. The simplicity of Kd facilitates, obviously, a numerical modelling of the radionuclide migration. However, its applicability to the longterm performance assessment of a given geological barrier, the far field, has been attended with controversial debates, because its ‘black-box’-like concept does not provide an insight into the underlying geochemical reactions. Fig. 10 illustrates a schematic presentation of solid–water interactions of M(III) in one of the deep Gorleben aquifer systems (Gohy 573). The naturally occurring cerium, Ce(III), is determined in groundwater, colloids and geo-matrix present in equilibrium for many thousand years. Concentrations of Ce(III) found in the three different phases enable the calculation of their distribution ratios (R) among three given phases. As is apparent from this figure, an over 95% of Ce(III) is present in a colloidal state, which can hardly be exchanged by Ce(III) incorporated in the geo-matrix. The distribution of Ce(III) between groundwater, including colloid-borne, and geo-matrix gives a concentration ratio of R =483. According to the Kd concept, this value leaves us to perceive a considerable retardation in the migration of Ce(III) in the aquifer system. If groundwater colloids migrate without retardation, the colloidborne Ce(III) is possibly accompanied by the same fate. As a result, the Kd concept misrepresents the real situation in nature.
As an alternative to the Kd concept, surface complexation, solid–water interface reactions of individual radionuclides onto a composite geomatrix, has been put into modelling either to replace Kd or possibly to substantiate it. The development of the surface complexation model (SCM) is a tedious undertake, since it entails a wide range of experiments for individual radionuclides with respect to a variety of natural mineral components under different geochemical conditions. The hitherto endeavoured SCM concept is based faithfully on a thermodynamic approach (Dzombak and Morel, 1990); nonetheless, its modelling capability is not sufficiently developed enough for the application to natural systems. This concept can describe the sorption behaviour of some radionuclides in laboratory system onto mineral surfaces (cf. Fig. 5) and also onto aquatic colloids of homogeneous nature (cf. Fig. 4). Shall it be, one day, applicable to natural dynamic systems? In the far field, as shown in Fig. 10, the primary migration process of radionuclides, particularly actinides, may be attributable to the mobilisation of colloid-borne species. Pseudo colloids may facilitate the actinide migration without retardation as fast as the water flow velocity in porous aquifer systems. The migration behaviour of colloidborne actinides depends on their chemical binding state onto colloids: whether they are bound reversibly or irreversibly. The reversible binding
Fig. 10. Solid –water interactions of M(III) in a natural aquifer system: distribution of Ce(III) in solution, solid and colloidal phases.
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Fig. 11. Desorption kinetics of Eu(III) from humic colloids: comparison between Eu(III) sorbed in the laboratory and Eu(III) bound naturally to humic colloids (Geckeis et al., 1999).
Fig. 12. Retardation and migration of different radionuclide species relative to water transport velocity.
results in a competitive reaction of a given actinide between the surfaces of colloid and geo-matrix, and hence leads to its retardation of migration, whereas the irreversible binding promotes colloids as a carrier for the actinide migration along with the water flow. Fig. 11 shows a laboratory desorption experiment of Eu(III), which is sorbed onto humic colloids at different reaction times in laboratory (Geckeis et al., 1999). For the purpose of comparison, the same desorption experiment is performed also for Eu(III) that is already present as sorbed on humic colloids in two of the Gorleben groundwaters. To accelerate the reaction, a chelate cation exchanger as solid form is introduced in each groundwater sample. As appears in
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this figure, the desorption of Eu(III) bound onto colloids in natural groundwater is much slower than that to form colloids onto which Eu(III) is sorbed in laboratory. The sorption and desorption behaviour of Eu(III) onto and from humic colloids in laboratory system can be described by the surface complexation model, whereas, for the natural system, the reaction behaviour of Eu(III) cannot be appraised by the same approach, because its irreversible sorption appears to prevail in natural colloids. Fig. 12 shows a comparison of two different migration processes. Radionuclides, which undergo reversible sorption reactions with geo-matrices, may be treated for their migration by an ion exchange process with its retardation parameter Rf. To this category belong M(I) and M(II) and, to a certain extent, also M(V) and M(VI). The colloid-borne radionuclides may migrate as fast as the water flow velocity if they are irreversibly bound onto mobile colloids; nonetheless, if they are bound irreversibly onto colloids of unstable nature, then filtration or precipitation accelerates their retardation. To this category belong M(III) and M(IV). A thermodynamic approach is not available to describe such natural dynamic reactions. A combination of Figs. 10–12 explains the following facts. (1) A thermodynamic approach is not yet directly applicable to dynamic natural systems; for example, the quantitative appraisal of the radionuclide migration in the far field. However, it is an indispensable tool for understanding the underlying geochemical reactions as well as for the assessment of key uncertainties involved in the available migration modelling. (2) The thermodynamic approach fails to describe the migration behaviour of colloid-borne radionuclides. Still, then, the colloid generation either by homogeneous nucleation (real colloid formation) or by heterogeneous nucleation (pseudo colloid formation) can be approximated by an extra thermodynamic approach. The geochemical modelling of the colloid-borne radionuclide migration is yet to be developed, for which empirical experience on the underlying kinetic and thermodynamic processes is an essential prerequisite.
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5. Perspectives A large number of thermodynamic data are available for the appraisal of geochemical reactions of long-lived radionuclides, actinides and some fission products, which have been integrated in various geochemical codes or in coupled codes. However, the hitherto available data do not satisfy the necessary modelling of all significant geochemical reactions of essential radionuclides, particularly actinides. The present situation is summarised in a simplified form as follows. What do we know at present? 1. Aquatic chemical reactions of radionuclides in laboratory systems (closed systems). Thermodynamics of solubility and complexation (mostly binary complexes). Speciation of solid and solution phases (mostly binary complexes). 2. Thermodynamic speciation modelling by various geochemical codes (closed system). Capability is limited largely to binary complexes. Speciation of M(IV) is limited to a large extent by a lack of its relevant thermodynamic data. 3. Solid–water interface interaction of radionuclides in trace concentrations for reversible processes (for radionuclides of Z E3+ , the modelling is still in the developing state). 4. Phenomenological understanding of radionuclide migration processes. What do we need to know? 1. Aquatic chemical reactions of radionuclides in natural multi-component systems (open systems). 2. Surface complexation models (SCM) for the broad application: for composite geo-matrices in natural dynamic systems (to replace the Kd concept!). 3. Irreversible geochemical reactions of radionuclides. Secondary solid-phase formation in the near field: formation of solid solution of radionuclides. Colloid generation: both homogeneous and heterogeneous nucleation.
4. Key uncertainties involved in the application of thermodynamics to natural dynamic systems (open and closed systems). 5. Applicability of the laboratory results to natural systems, for which the following prevalent conditions are to be taken into account. Laboratory system Closed system Short-term reactions Equilibrium reactions Reversible processes
Natural systems Open system Long-term reactions Non-equilibrium reactions Irreversible processes
The presently available knowledge on the aquatic thermodynamics of actinides and fission products facilitates largely the understanding of their prevalent geochemical behaviour, which hence makes it possible to distinguish the essential reactions from a bulk of convoluted geochemical reactions. For laboratory multi-complex systems, the thermodynamic speciation of underlying geochemical reactions is largely possible, whenever the relevant thermodynamic data are available. For various reasons, as discussed in this paper, the applicability of a thermodynamic approach to dynamic natural systems is considerably limited. Nonetheless, the thermodynamic knowledge is of cardinal importance for understanding a variety of far-reaching geochemical reactions in the near and far fields, and allows the quantification of, if not altogether at once, individual isolated geochemical reactions separately. For the analysis of key uncertainties involved in the long-term performance assessment of a given repository, a thermodynamic approach can play a significant role for the corroboration of parameters describing the given geochemical conditions, like pH, Eh etc., the probable source terms of individual radionuclides and their migration behaviour. For this purpose, well-founded thermodynamic data for a wide range of geochemical reactions of actinides are absolutely necessary.
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