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Waste form corrosion modeling: comparison with experimental results
Waste Management 21 (2001) 741±752
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Waste form corrosion modeling: comparison with experimental results Bernhard Kienzler *, Berthold Luckscheiter, Stefan Wilhelm 1 Forschungszentrum Karlsruhe, Institut fuÈr Nukleare Entsorgung (INE), Karlsruhe, Germany Received 7 April 2000; received in revised form 18 December 2000; accepted 2 January 2001
Abstract Results of leaching experiments using active and inactively simulated HLW glass in concentrated NaCl solution are described. Measured solution concentrations of glass components, ®ssion products and actinides are compared with computed data. The computed pH value corresponds with the ®ndings from experiments with inactively simulated glass samples. Moreover, the concentrations of silica and strontium can be described adequately by reaction path modelling. The computed U concentration is explained by the precipitation of schoepite or Na2U2O7. The computed Am concentration signi®cantly exceeds the measured data. This may be attributed to sorption processes on corrosion products of the glass, which are relevant also for lanthanide elements under the conditions of the experiments. This hypothesis is tested by solid solution approaches and by computing sorption of Am onto SiO2 precipitates. # 2001 Elsevier Science Ltd. All rights reserved. Keywords: Glass corrosion; Actinides; Geochemical modelling
1. Introduction The associated risk of a radioactive waste disposal system depends on the quantity of radionuclides mobilised in the case of groundwater (brine) intrusion to the waste. In the presence of water or brines, waste forms may be corroded and radioactive elements are leached. New solid reaction products will be formed, which are thermodynamically favoured under the given geochemical conditions. For performance assessment and longterm safety analyses, model calculations predicting the quantity of dissolved radionuclides in the solutions as a function of time are required. Within the time periods under consideration, the natural environment as well as engineered structures will change signi®cantly and an unavoidable uncertainty regarding the future state or evolution of the repository system has to be taken into account. Therefore, the geochemical isolation potential for radionuclides has to be evaluated [1]. * Corresponding author. Tel.: +49-7247-82-4467; fax: +49-724782-4308. E-mail address: [email protected] (B. Kienzler). 1 Present address: Colenco Power Engineering AG, Baden, Switzerland.
In the last years, experiments and related model approaches describing the leaching of waste glasses have been published. The models comprise dierent processes involved, such as kinetically controlled glass network dissolution, precipitation of secondary phases, ion exchange, silica diusion through amorphous surface layers and speci®c interactions between dissolved ions and the glass surface [2±8]. In the present study, a thermodynamic reaction path model [9] is applied and the results are compared with the experimental ®ndings concerning the behaviour of HLW glass in NaCl brines. The comparisons cover the dissolution model of the waste glass, the development of the geochemical environment and the concentrations and speciation of radionuclides. Comparisons of experiments to model predictions can indicate whether the database available for 25 C can be applied for the experimental conditions or not. Experiments with highly radioactive glass samples and with inactive glass (having the same chemical composition as the active glass but without radioactive isotopes and actinide elements) are used for the comparisons. The highly radioactive glass has been prepared by the Commissariat aÁ l'eÂnergie atomique's (CEA) Centre de la ValleÂe du RhoÃne, Marcoule, France. The glass composition is similar to that of the COGEMA
0956-053X/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved. PII: S0956-053X(01)00008-3
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glass R7T7 produced in the vitri®cation plant R7 and T7 at La Hague, France. Concentrations of radioactive elements, except for plutonium, are lower in the CEA glass than in the COGEMA R7T7 glass, because a different waste was used by CEA.
2. Experimental Static corrosion tests were performed in dierent brines relevant for minerals present in German salt domes. The composition of the NaCl brine (`solution 3' [10]) used for the experiments described here has the following composition: 6.233 mol (kg H2O) 1 NaCl, 0.032 mol (kg H2O) 1 CaSO4 (320 g dm 3 NaCl and 3.8 g dm 3 CaSO4). At 55 C pH is 5.9 and the density is 1.203 g cm 3. The composition of the HLW glass is reported elsewhere [9]. The tests with the highly radioactive glass samples were performed in tantalum lined autoclaves at temperatures of 110 and 190 C over periods of 100 and 365 days. By using 3.835 g of glass powder with a grain size of 71±100 mm and 40 ml solution, the ratio of the glass surface area to the solution volume was about 5000 m 1. The atmosphere in the vessels consisted of argon. At the end of the duplicate tests, the autoclaves were cooled to ambient temperatures. Afterwards, pH and Eh were measured by means of a Ross type electrode and a Pt electrode, respectively. Measured pH was corrected for liquid junction potential [11]. Due to the Fe concentration in the solutions, we assume the redox potential is buered by Fe(II)/ Fe(III). Measured Eh values scatter between 0 and 150 mV. The solutions were ®ltered through 0.45 mm ®lters and aliquots of the ®ltrates were ultra®ltered through an 18 AÊ membrane ®lter to detect the presence of colloids. Prior to analyses, radionuclides had to be separated, both from the highly concentrated salt solution and from other elements. Separation was performed either by liquid-liquid extraction or by extraction chromatography procedures. Concentrations of glass constituents, ®ssion products and actinides were analysed by ICP±AES, ICP±MS and nuclear analytical methods, such as a-, g-spectroscopy and liquid scintillation counting (LSC). Details of the experiments are described elsewhere [9]. Sorption behaviour of americium onto the alteration layer of the corroded glass was investigated by speci®c experiments. In addition, the homologue elements neodymium and cerium were studied as simulates for the trivalent actinides. Pre-corroded simulated HLW glass powder was applied as substrate. Sorption isotherms for Am, Nd and Ce were determined, using various initial concentrations of these elements. The tests are performed in brines and diluted aqueous solutions at 80 C in the pH range 2±10.
3. Model The radionuclide concentrations and possible precipitation of solid phases were modelled with the software package EQ3/6 Rel.7.2a [12]. The initial solution composition including all species involved in further reactions was computed. To simulate corrosion of the waste glass by reaction path modelling, small amounts of a `special reactant' having the element composition of the waste glass were `added' numerically to the solutions and the new equilibrium was calculated by EQ6. A thermodynamic equilibrium between the solution and solid phases was assumed for each step. This procedure is implemented as `relative rate law' in the EQ6 code. The in®nitesimal steps of the EQ6 run were considered to represent the progress of waste form dissolution. This reaction path model had been previously used to evaluate the reaction progress of HLW glass in magnesium-rich brines [9]. It computed the evolution of element concentrations over a large range of the reaction progress. The calculated data were compared with experimental ®ndings obtained for various surface to volume ratios. In the reaction path computations, the redox potential could not be treated as an independent variable because oxygen is not considered in the used `special reactant' approach. Therefore, the Eh was ®xed by the O2 (gas) fugacity as an input parameter to EQ3/6. 4. Database Predictive geochemical modelling of thermodynamic properties of radionuclides in complex aquatic systems requires a thermodynamic database as well as an appropriate thermodynamic model. The ion interaction model (Pitzer approach) is applicable to highly concentrated aqueous solutions. The database applied for the present computations is based on the Pitzer `data0' ®le provided with EQ3/6 code including additional data for uranium [13,14] and other actinides [15,16] measured in our institute. Based on available experimental data, comprehensive sets of thermodynamic data (chemical potentials, thermodynamic constants) and model parameters (Pitzer parameter) for the major homogeneous and heterogeneous equilibria of americium/curium(III) and neptunium(V) were predicted [17]. For the high pH range expected to prevail for glass waste form dissolution in NaCl solutions, Pitzer interaction coecients for silicate and alumina species have been incorporated into the database [18]. Reactions and corresponding log K data for the radionuclides are given in Tables 1 and 2. Uranium data are taken from Refs. [19±21], data of plutonium and other tetravalent actinides were evaluated by Neck [22]. Since measured concentrations in the corroding solutions are available only for room temperature, all simulations were performed at 25 C.
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5. Comparison between measured and computed data and discussion This section presents a comparison between the measured and computed HLW glass dissolution progress and the extent of dierences between experimental and numerical results. Furthermore, the reasons for the observed dierences are discussed. It will be shown for which cases model and database describe the experiments suciently well and where the model or database are incomplete or where the reaction mechanisms are not yet understood.
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5.1. Dissolved glass components and pH Computation of the reaction of the waste glass in NaCl brine was started at an initial pH of 7.5. Computed and measured concentrations and pH are presented in Fig. 1 as a function of the progress of glass dissolution. Depending on the progress of the dissolution of the HLW glass the computed pH increased to 8.5. The experiments covered a reaction progress between 0.005 and 0.016 kg (kg H2O) 1 glass in the case of active samples and up to 0.026 kg (kg H2O) 1 glass for inactive samples. The experimental pH for inactive
Table 1 Thermodynamic data of aqueous radionuclide species applied for the model calculations Reaction
Log K
Ref.
Uranium VI VI VI VI VI VI VI VI VI IV IV IV IV
UO2+ 2 UO2OH+ UO2(OH)2(aq) UO2(OH)3 UO2(OH)24 (UO2)2(OH)2+ 2 (UO2)3(OH)2+ 4 (UO2)3(OH)+ 5 (UO2)3(OH)7 3+ UOH U(OH)2+ 2 U(OH)+ 3 U(OH)4(aq)
() () () () () () () () () () () () ()
2 H++U4++H2O+0.5 O2(aq) 1 H++UO2+ 2 +1 H2O 2 H++UO2+ 2 +2 H2O 3 H++UO2+ 2 +3 H2O 4 H++UO2+ 2 +4 H2O 2 H++2 UO2+ 2 +2 H2O 4 H++3 UO2+ 2 +4 H2O 5 H++3 UO2+ 2 +5 H2O 7 H++3 UO2+ 2 +7 H2O 1 H++U4++H2O 2 H++U4++2 H2O 3 H++U4++3 H2O 4 H++U4++4 H2O
33.78 5.70 11.50 19.22 33.03 5.68 11.93 15.59 31.05 0.40 0.40 3.80 9.30
[19] [19] [20,21] [19] [19] [19] [19] [19] [19] [19] [12] [12] [19]
Neptunium V V V IV IV IV IV
NpO+ 2 NpO2OH(aq) NpO2(OH)2 NpOH3+ Np(OH)2+ 2 Np(OH)+ 3 Np(OH)4(aq)
() () () () () () ()
3 H++0.25 O2(aq)+Np4++1.5 H2O 1 H++H2O+NpO+ 2 2 H++2 H2O+NpO+ 2 1 H++Np4++H2O 2 H++Np4++2 H2O 3 H++Np4++3 H2O 4 H++Np4++4 H2O
9.87 11.31 23.54 1.00 2.80 5.80 9.60
[12] [15] [15] [22] [22] [22] [22]
Plutonium VI V IV IV IV IV III III III III III
PuO2+ 2 PuO+ 2 PuOH3+ Pu(OH)2+ 2 Pu(OH)+ 3 Pu(OH)4(aq) PuOH2+ Pu(OH)+ 2 Pu(OH)3(aq) PuCl2+ PuCl+ 2
() 2 H++Pu4++H2O+0.5 O2(aq) () 3 H++0.25 O2(aq)+Pu4++1.5 H2O () 1 H++Pu4++H2O () 2 H++Pu4++2 H2O () 3 H++Pu4++3 H2O () 4 H++Pu4++4 H2O () 1 H++Pu3++1 H2O () 2 H++Pu3++2 H2O () 3 H++Pu3++3 H2O ()Pu3++1 Cl ()Pu3++2 Cl
8.13 2.94 0.60 0.60 2.29 7.98 2.80 15.19 25.69 0.24 +0.74
[12] [12] [22] [22] [22] [22] [16] [16] [16] [16] [16]
Americium III III III III III
AmOH2+ Am(OH)+ 2 Am(OH)3(aq) AmCl2+ AmCl+ 2
() 1 H++Am3++1 H2O () 2 H++Am3++2 H2O () 3 H++Am3++3 H2O ()Am3++1 Cl ()Am3++2 Cl
2.80 15.19 25.69 0.24 +0.74
[16] [16] [16] [16] [16]
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samples corroded at 190 C with a high reaction progress is about 0.3 pH units below the computed pH. Radioactive samples show a signi®cant lower pH in the range of 7.2. In order to test the model approach and to compare the pH at experimental temperatures, the temperature eect of pH was computed for the pure NaCl brine according to the procedure described by Ref. [12]: the quenched (cooled) solution's composition was used to de®ne the system at room temperature. Then, a reaction path for temperature increase was computed by EQ6. It was essential to take into account the complete composition of the solution. Using this procedure, the computed pH decrease of the NaCl brine was about 0.2 units and about 0.6 unit at 110 and 190 C, respectively. Since the thermodynamic data-set for highly concentrated solutions is incomplete, computation of the eect of the temperature quenching on the pH of the experimental solutions is not possible with this procedure. The lithium and boron concentrations are used as a measure of the dissolution progress of the HLW glass matrix. Boron (B) concentrations are used to calculate the reaction progress. However, B could used to scale the glass dissolution up to the dissolution of 0.1 kg (kg H2O) 1 glass. Above this value, concentration of B is
limited by precipitation of borax. The Li concentrations calculated with the relative rate model agree well with the experimental data. According to the calculations, amorphous silicate and calcium bearing solids, such as the clay minerals nontronite and mesolite were predicted to precipitate. This is corroborated by the fact that clay mineral corrosion products of HLW glass in NaCl brine are experimentally detected [23]. Measured Si concentration are in excellent agreement with the computed values. Si solubility is controlled by amorphous silica. Calcium concentrations in the experiments are found to be only two times lower than computed values. Anhydrite precipitation is not computed. 5.2. Fission products Fig. 2 compares experimental and calculated concentrations of the ®ssion products Sr and Mo. Measured and computed strontium concentrations agree well for both active and inactive glasses. The computed increase of the Sr concentration is parallel to that for the Li concentration and lies in the range of the experimental scatter. The experimental Mo concentrations are in the same range of 510 5 to 210 4 mol (kg
Table 2 Thermodynamic data of solid radionuclide phases applied for the model calculations Reaction U
UO3.2H2O Schoepite Na2U2O7 Na-Becquerelite CaU2O7 Becquerelite (UO2)2(SiO4).2H2O Soddyite U(OH)4(am) USiO4
Log Ksp +UO2+ 2 +3
+
() 2 H
H2O
+ () 6 H++UO2+ 2 +2 Na +3 H2O
Refs.
5.37
[13]
22.60
[19]
26.45
[14]
()
2+ 6 H++UO2+ +3 H2O 2 +Ca
()
4 H++2 UO2+ 2 +1 SiO2(aq)+4 H2O
6.03
[27]
() ()
4 H++U4++4 H2O 4 H++U4++1 SiO2(aq)+2 H2O
8.05
[22] [12]
Conite Np(OH)4(am) NpO2(s) Np2O5 NpO2OH(s) NpO2OH(am)
() () () () ()
4 H++Np4++4 H2O 4 H++Np4++2 H2O 2 H++H2O+2 NpO+ 2 1 H++H2O+NpO+ 2 1 H++H2O+NpO+ 2
0.81 7.8 7.98 4.52 5.22
[22] [15] [15] [15] [15]
Pu
Pu(OH)4(am) Pu(OH)3(am) PuO2OH(am)
() () ()
4 H++Pu4++4 H2O 3 H++Pu3++3 H2O 1 H++PuO+ 2 +3 H2O
2.51 14.49 5.40
[22] [16] [12]
Am
Am(OH)3(am) Am(OH)3(cr)
() ()
3 H++Am3++3 H2O 3 H++Am3++3 H2O
17.00 15.02
[16] [16]
Mo
CaMoO4.0.2H2O Powellite MgMoO4.2H2O MgMoO4.0.52H2O NaNdMoO4 Nd(OH)3(s)
() 1 Ca++1 MoO24 +0.2 H2O
7.95
[32]
1.34 2.10 8.94 15.59
[32] [32] [32] [12]
Np
Nd
() 1 Mg++1 MoO24 +2 H2O () 1 Mg++1 MoO24 +0.5 H2O () 0.5 Na++0.5 Nd3++1 MoO24 +0.16 H2O () 3 H++Nd3++3 H2O
B. Kienzler et al. / Waste Management 21 (2001) 741±752
H2O) 1, for both inactive and active experiments at 110 and 190 C. Assuming that Mo concentration is not controlled by solid phases, the model predicts concentrations in the upper range of the experimental scatter. Consideration of powellite (CaMoO4.0.2 H2O) or trace amounts of Na0.5Nd0.5MoO4.0.2 H2O would result in a decrease of the Mo concentration by almost one order of magnitude. Moreover, the evolution of the computed Mo concentration as function of the reaction progress is completely dierent from measured [Mo]. However, the formation of powellite has been detected by SEM and EDX (micro X-ray analyses) in many corrosion experiments with simulated HLW glass in NaCl and in MgCl2 solutions [9,24,25]. The discrepancy between computed and measured Mo concentrations can not yet be explained. Computations show that the solubility of neodymium is controlled by solid Nd(OH)3 at a concentration of 10 6 mol (kg H2O) 1. Measured concentrations are below the detection limit (110 6 mol (kg H2O) 1). Therefore these data are not shown in Fig. 2. 5.3. Uranium Dierent from the studied components of the glass matrix and the ®ssion products under investigations, the
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actinide elements can exist in several oxidation states. Fig. 3 shows the Eh±pH diagram for dissolved uranium species. Ranges of Eh and pH data measured in dierent experiments are given as well. Under the conditions prevailing in the experiments, U exists in the pentavalent or hexavalent state. Results of reaction path calculations for the uranium are shown in Fig. 4. Uranium concentrations in NaCl brine are about 210 6 mol (kg H2O) 1. Computations were performed for redox potentials above 50 mV where uranium is in the hexavalent state. Computations indicated that the rather high U concentration at the computed pH of 8.5 is compatible with schoepite (UO3.2H2O) or Na2U2O7 as solubility-controlling solids. The predominating aqueous U species is computed to be UO2(OH)3 . For the computed Si concentrations the precipitation of uranium silicate phases, such as uranophane (Ca[(UO2)SiO3OH]2.5H2O) or soddyite ((UO2)2SiO4.2H2O) is expected from the model calculations and uranium silicate phases have been observed in the experiments [25]. For the dissolution of soddyite, the EQ3/6 database provides log K=0.39. Application of this value results in [U] 310 9 mol (kg H2O) 1 which is not compatible with measured uranium concentrations. However, Kosp values given in literature dier signi®cantly.
Fig. 1. Computed and measured concentrations and pH of the HLW glass matrix elements as a function of the reaction progess in NaCl brine. Upper abscissa: experimental time scale. Measured pH: +: T=110 C, inactive R7T7 samples; X: T=190 C, inactive R7T7 samples; &: T=190 C, active R7T7 sample.
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Measured solubility products by Nguyen et al. [26] and Moll et al. [27] have resulted in log K=5.74 and log K=6.03 0.53, respectively (see Table 2). Uranium concentrations of 210 6 mol (kg H2O) 1 are computed, if a log K=6.03 is used in EQ3/6. This concentration corresponds to the experimental ®ndings. On the other hand, lower log K values have also been reported in literature. Measured solubility for synthetic soddyite phases have resulted in log K=3.9 0.7 [28]. The solubility has been determined at an initial pH=3. In Ref. [29] log K for a natural soddyite has been determined to be 3.0 2.9 at a pH8.8. For uranophane published solubility data dier signi®cantly, too. In Ref. [26] log K=8.5 1.5 is reported, experiments with synthetic uranophane have shown log K=11.7 0.6 [28] and 7.8 0.8, respectively, for a natural uranophane at a pH between 8.0 and 9.3. 5.4. Plutonium Pu can occur in dierent oxidation states, too. Fig. 5 shows the Eh±pH diagram for dissolved Pu species and the ranges of Eh and pH data measured in the experiments. Under these conditions, Pu exists only in tetravalent state and solid Pu(OH)4(am) controls solubility. The computed concentration of 610 10 mol (kg
H2O) 1 is obtained for the glass experiments. Fig. 4 shows a good agreement between the computed results and the experimental ®ndings for Pu. 5.5. Other actinides In the experiments dissolved Np and Am concentrations were determined, as well. Under the conditions relevant for these experiments, Am exists only in the trivalent state. Np, however, can occur in tetra- or pentavalent states. Fig. 6 shows the Eh±pH diagram for dissolved Np species and the measured Eh and pH data. The Eh±pH range observed in the experiments is just below the phase boundary between Np(V) and Np(IV) in the tetravalent region. In the case of Np(IV), solid Np(OH)4(am) would limit solubility to 110 9 mol (kg H2O) 1. This concentration is one order of magnitude below measured data which scatter between 310 8 and 610 9 mol (kg H2O) 1 (see Fig. 4). For the pentavalent state calculated Np concentrations are in the range of the measured values. For this condition, solid Np(OH)4(am) or Np2O5(cr) may control the dissolved concentration of 210 7 mol (kg H2O) 1 or of 210 6 mol (kg H2O) 1, respectively. Signi®cant sorption of Np(V) is reported in literature at the pH relevant for glass corrosion in NaCl solution
Fig. 2. Computed and measured concentrations of Sr and Mo released from HLW glass in NaCl brine. ^, Sr inactive glass sample (T=110 C); , Mo inactive glass sample (T=110 C); *, Sr inactive glass sample (T=190 C); & Mo inactive glass sample (T=190 C); &, Sr active R7T7 sample; ^, Mo active R7T7 sample.
B. Kienzler et al. / Waste Management 21 (2001) 741±752
Fig. 3. Eh±pH diagram of uranium (chemical activity of U: 10 of aqueous U species.
6
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mol (kg H2O) 1). Dotted lines: stability ®eld of water; dashed lines: stability ®elds
Fig. 4. Computed pH and computed and measured uranium, neptunium, and plutonium concentrations released from the HLW glass in NaCl brine. *, uranium experimental ^, neptunium experimental &, plutonium experimental.
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Fig. 5. Eh±pH diagram of plutonium (chemical activity of Pu: 10 10 mol (kg H2O) 1). Dotted lines: stability ®eld of water, dashed lines: stability ®elds of aqueous Pu species, solid line: stability ®eld of Pu(OH)4(am).
Fig. 6. Eh±pH diagram of neptunium (chemical activity of U: 10 ®elds of aqueous Np species.
7
mol (kg H2O) 1). Dotted lines: stability ®eld of water, dashed lines: stability
B. Kienzler et al. / Waste Management 21 (2001) 741±752
[30]. A shift of Np to the pentavalent state may be caused by a-radiolysis. The total a-decay rate of the glass samples is about 2.5108 Bq/g. Taking a G-value of 2 into account, at the surface of the glass powder about 10 10 mol (kg H2O) 1 oxidising species are formed per second in a 20-mm surface layer of the glass powder. These oxidising species may change the redox conditions locally with respect to the bulk of the solution and transfer Np species into the pentavalent state. In the case of americium, the measured concentrations are in the range of 110 10 mol (kg H2O) 1, whereas the calculated concentrations are by three orders of magnitude higher (Fig. 7). Measured Am concentration is close to the detection limit and do not allow application of speciation techniques such as time resolved laser ¯uorescence (TRLF) spectroscopy. The in¯uence of solid solutions to limit the Am concentration has been investigated. Solid solutions are phases of variable compositions which can be considered as mixtures between two or more `end-member' minerals. EQ3/ 6 provides features to treat solid solutions. The thermodynamic activity of a solid solution component is de®ned by multiplication of the mole fraction of the component with its activity coecient in the solid phase. For the computations, an ideal solid solutions approach is chosen using activity coecients of 1. Two
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types of end members are considered in this study: hydroxides and molybdates. One model considers the hydroxide end members of Am, lanthanides and aluminium, the other approach uses the two amorphous hydroxides Am(OH)3(am) and SiO2(am) (see Fig. 7). Calculations including possible precipitation of solid containing Am are not successful in resolving the discrepancy between the measured and calculated concentrations. Molybdate phases do not signi®cantly control the Am concentration. In addition, consideration of the presence of traces of CO2 gas in the leaching vessels resulting in the formation of Na±Am carbonates would not explain the low Am concentrations. The sorption experiments with Am onto pre-corroded simulated HLW glass powder show high retention factors (Rs) up to 105 ml/g independent on the initial concentration of the element. This ®nding indicates a sorption process. Details of the experiments and the results are reported elsewhere [31]. Measured distribution coecients for Am, Nd and Ce are given in Fig. 8 as a function of pH. These data can be used to interpret the discrepancy between measured and calculated Am concentrations. We assume that Am is sorbed onto precipitated, amorphous SiO2. The amount of precipitated SiO2(am) is computed by the model. The sorption data and the amount of precipitated SiO2(am)
Fig. 7. Americium concentrations released from the HLW glass in NaCl brine: measured [Am] and results of computations applying dierent model approaches. ~, americium experimental.
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Fig. 8. Measured distribution coecients for Ce, Nd and Am on the corrosion products of the corroded HLW glass matrix in NaCl brine. &, Am after 36 days experiment; *, Am after 10 days experiment.
are used to calculate an expected value for the Am concentration. As shown in Fig 7, better agreement with the measured Am concentrations is obtained by this model. Similar behaviour is also expected for Nd. Systematic investigations of the relevant processes, however, are extremely dicult, because the concentrations are close to the detection limit. As shown in Fig. 5, Pu concentration is controlled by solubility of Pu(OH)4(am). Sorption of Pu may be relevant, if Pu is shifted into the trivalent or pentavalent oxidation state. Corresponding experiments with HLW glass in NaCl brines have not been conducted up to now.
6. Summary and conclusions 6.1. Redox potential Dissolved concentrations computed for the actinides and for iron depend signi®cantly on the redox potential. The rather high calculated U concentrations are in agreement with the experiments. However, several U(VI) solids may account for solubility limiting phases. Experimental Eh and pH indicate clearly that Pu is well situated in the stability ®eld of Pu(IV). Based on measured pH and Eh of the bulk solution, one would expect Np to be in the tetravalent state, too. The measured Eh and pH are close to the stability ®eld of Np(V). One can assume that at the surface of the glass powder
radiolytically formed oxidizing species change locally the redox state of the solution and oxidize Np species. 6.2. Temperature eect Concentrations measured after cooling the solutions to ambient temperature can be explained by the computations for 25 C. The pH measured for 110 and 190 C is about 0.3 units lower than computed. The model is not able to describe the pH under radiation by the highly radioactive samples. The computed behaviour of Mo diers considerably from the experimental ®ndings. In the pH range between 3 and 4 the solubility constant Kosp of powellite is almost independent of the temperature [32]. However, at pH values above 6 preliminary experiments show a strong increase of solubility with temperature. We assume that the higher solubility at the experimental temperature may contribute to the dierences between measured and calculated Mo concentrations. High-temperature Pitzer data for important species and the identi®cation of temperature-dependent sorption reactions for the actinide-containing solids are not available. At present the model is not well de®ned for elevated temperatures. Ksp and stability constants of dissolved species can be extrapolated for higher temperatures by means of H and entropy data, such as performed in Ref. [33]. However, it has to be recognised, that for most of the equilibrium constants signi®cant errors persist in highly concentrated aqueous solutions.
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6.3. Thermodynamic data
References
The measured uranium concentrations indicate that Na2U2O7, schoepite or a silicate phase such as soddyite can be the limiting phase. Especially for silicate phases, the solubility products given in literature dier by several orders of magnitude. Under the prevailing conditions, these phases have been detected at the surface of the corroded glass. Applying the latest published log Kosp for soddyite [27], calculations result in a good agreement with measured U concentrations. However, the variation of the solubility products of silicate phases reported in literature is considerable. For the tetravalent state, neptunium concentrations are calculated to be one order of magnitude below the experimental ®ndings. Concentrations in the range of the experiments are computed only for Np(V). For solid Np phases, also rather large discrepancies of log Kosp for oxides and hydroxides are reported in literature. In the case of Am, the experiments show signi®cantly lower dissolved concentrations than predicted by assuming precipitation of an Am(OH)3 phase. Several mechanisms are investigated which may account for this, such as the formation of solid solutions. The discrepancy is resolved when sorption of Am onto SiO2(am) is considered. A simple solid solution approach, such as presented in this study, cannot yet describe the measured data. Additional work is required to qualify such an approach on the basis of the endmembers Am(OH)3(am) and SiO2(am). For plutonium a good agreement between the experimental ®ndings and the computed Pu(OH)4(am) solubility is obtained. It is important to note the in¯uence of the prevailing Eh and pH conditions: if Pu(III) would dominate, a signi®cantly higher solubility would be expected. In this case, sorption processes similar to those modelled for americium have to be considered in order to describe the experimental concentrations. The current study demonstrates the capability and the limits of geochemical modelling to predict dissolution of HLW glass, the development of the geochemical environment, as well as the concentrations and the speciation of mobilized radionuclides. Since thermodynamic data at higher temperatures are not experimentally veri®ed, further experimental work is needed to understand the mechanisms and reactions controlling dissolved actinide and ®ssion product concentrations.
[1] Kim JI, Grambow B. Geochemical assessment of actinide isolation in a German salt repository environment. Engineering Geology 1999;52:221±30. [2] Grambow B. Nuclear waste glass dissolution: mechanism, model and application. JSS Project, Swedish Nuclear Fuel and Waste Management Co 1987;87±02. [3] Bourcier WL, Pfeifer DW, Knauss KG, McKeegan KD, Smith DK. A kinetic model for borosilicate glass dissolution based on the dissolution anity of a surface alteration layer. Mat Res Soc Symp Proc 1990;176:209±16. [4] Bourcier WL. Waste glass corrosion modelling: Comparison with experimental results. Mat Res Soc Symp Proc 1994;333:69±82. [5] Jollivet P, Nicols M, Vernaz E. Estimating the alteration kinetics of the French vitri®ed high-level waste package in a geologic repository. Nuclear Technology 1998;123:67±81. [6] Jantzen CM, Ramsey WG, Prediction of radioactive waste glass durability by the hydration thermodynamic model: Application to saturated repository environments. 13th symposium on the scienti®c basis for nuclear waste management. Boston, MA (USA). 27 Nov±2 Dec 1989. Mat Res Soc Symp Proc, 1990;176:217±28. [7] Advocat T, Vernaz E, Crovisier JL, Clement A, Gerard F. Modeling of solution renewal with the KINDIS code: Example of R7T7 glass dissolution at 90 C. 18th international symposium on the scienti®c basis for nuclear waste management. Kyoto (Japan). 23±27 Oct 1994, Murakami X., Takashi Ewing RC, editors Scienti®c basis for nuclear waste management 18. Part 1. Materials Research Society. pp. 31±38, 1995. [8] Buck EC, Bates JK. Microanalysis of colloids and suspended particles from nuclear waste glass alteration. Applied Geochemistry 14(5):635±53 [9] Grambow B, Lutze W, Kahl L, Geckeis H, Bohnert E, Loida, A, Dressler P. Long-term safety of radioactive waste disposal. Retention of Pu, Am, Np and Tc in the corrosion of COGEMA glass R7T7 in salt solutions. Final Report. Report FZKA 5767, Karlsruhe: Forschungszentrum Karlsruhe, 1996. [10] Arbeitskreis HAW-Produkte, Korrosionsexperimente an verglasten radioaktiven AbfaÈllen. -SalzloÈsungen und S/V-VerhaÈltnis, Stellungnahme Nr. 3, 1986. [11] Grambow B, MuÈller R. Mat Res Soc Symp Proc 1990;176:229±40. [12] ``EQ3/6, A Software Package for Geochemical Modeling Version 7.2a.'' University of California, Lawrence Livermore National Laboratory. Software produced under the Designated Unclassi®ed Subject Area (DUSA) ``Yucca'', 1990±1993. [13] DõÂaz Arocas P, Grambow B. Solid-liquid phase equilibria of U(VI) in NaCl-brines. Geochim Cosmochim Acta 1998; 62(2):236±45. [14] Sandino A, Grambow B. Solubility equilibria in the U(VI)-Ca-KCl-H2O system: Transformation of schoepite into becquerelite and compreignacite. Radiochimica Acta 1994;66/67:37±43. [15] FanghaÈnel Th, Neck V, Kim JI. Thermodynamics of neptunium(V) in concentrated salt solutions: II. Ion interaction (Pitzer) parameters for Np(V) hydrolysis species and carbonate complexes. Radiochimica Acta, 1995;69:169±76. [16] Neck V, FanghaÈnel Th, Kim JI. Aquatische Chemie und thermodynamische Modellierung von trivalenten Actiniden. Report FZKA 6110, 1998. [17] FanghaÈnel T, Weger HT, Koennecke T, Neck V, Paviet-Hartmann P, Steinle E, Kim JI. Thermodynamics of Cm(III) in concentrated electrolyte solutions. Carbonate complexation at constant ionic strength (1 m NaCl), Radiochimica-Acta 1998;82:47±53. [18] Reardon EJ. An ion interaction model for the determination of chemical equilibria in cement/water systems. Cement and Concrete Research 1990;20:157±92.
Acknowledgements The work is supported by the Bundesamt fuÈr Strahlenschutz (BfS) within the project: ``Erstellung eines integrierten Nahfeldmodells von Gebinden hochaktiver AbfaÈlle im Salzstock Gorleben: geochemisch fundierter Quellterm fuÈr HAW-Glas, abgebrannte Brennelemente und Zement''.
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[19] Grenthe I, Fuger J, et al. Chemical thermodynamics of uranium. In: Wanner H, Forest I, editors. OECD/NEA, ISBN 0-444-89381-4. Amsterdam: North-Holland Elsevier Science Publishers B.V, 1992. [20] Silva RJ. Mechanisms for the retardation of uranium (VI) migration. Fall Meeting of the European Materials Research Society. Strasbourg (France). 5±8 Nov 1991. [21] Choppin GR, Mathur JN. Hydrolysis of actinyl(VI) cations. Radiochimica Acta 1991;52/53:25±8. [22] Neck V, Kim JI. Solubility and hydrolysis of tetravalent actinides. Report FZKA 6350, Karlsruhe: Forschungszentrum Karlsruhe, 1999. [23] Abdelouas A., Etude de l'alteÂration de verres rhyolitiques au contact de saumures naturelles (Bolivie) Application aÁ l'eÂtude du comportement aÁ long terme du verre nucleÂaire R7T7. Report FZKA 5777, Karlsruhe: Forschungszentrum Karlsruhe, 1996. [24] Abdelouas A, Crovisier J-L, Lutze W, MuÈller R, Bernotat W. Structure and chemical properties of surfacelayers developed on R7T7 simulated nuclear waste glass altered in brines at 190 C. Eur J Mineral 1995; 7:1101±13. [25] Luckscheiter B, Nesovic M. Langzeitsicherheit der Endlagerung radioaktiver AbfaÈlle: Entwicklung und Charakterisierung eines Glasprodukts fuÈr den HAWC der WAK. Report FZKA 5825, Karlsruhe: Forschungszentrum Karlsruhe, 1996. [26] Nguyen SN, Silva RJ, Weed HC, Andrews JE. Standard Gibbs free energies for formation at 30 C of four uranyl silicates: Soddyite, uranophane, sodium boltwoodite and sodium weeksite, Report LLNL UCRL-JC-106032, 1991.
[27] Moll H, Geipel G, Matz W, Bernhard G, Nitsche H. Solubility and speciation of (UO2)2SiO4x2H2O in aqueous systems. Radiochimica Acta 1996;74:3±7. [28] Casas I, Perez I, Torrero E, Bruno J, Cera E, Duro L. Dissolution studies of synthetic soddyite and uranophane. Report SKB-TR±97-15, Stockholm: Svensk KaÈrnsbraÈnslehantering AB, 1997. [29] Casas I, Bruno J, Cera E, Finch RJ, Ewing RC. Kinetic and thermodynamic studies of uranium minerals. Assessment of the long-term evolution of spent nuclear fuel. Report SKB-TR±9416, Stockholm: Svensk KaÈrnsbraÈnslehantering AB, 1994. [30] Del Nero M, Advocat T, Jollivet P, Bontems G. Sorption of neptunium(V) on silicate gels, Proc. 2nd NUCEF Int. Symp. NUCEF'98; November 16±17, 1998, JAERI-Conf 99-004, 1999:584-595. [31] Luckscheiter B, Kienzler B. Determination of sorption isotherms for Eu, Th, U and Am on the gel layer of corroded HLW glass. submitted to J Nucl Mater. [32] Grambow B, MuÈller R, Rother A. Determination of molybdate mean ionic activity coecients for the assessment of radionuclide mobility. Radiochimica Acta, 1992;58/59:71±7. [33] MeÂnard O, Advocat T, Ambrosi JP, Michard A. Behaviour of actinides (Th, U, Np and Pu) and rare earth (La, Ce, and Nd) during aqueous leaching of a nuclear waste glass under geological disposal conditions. Applied Geochemistry 1998;13:105± 26.
www.elsevier.com/locate/wasman
Waste form corrosion modeling: comparison with experimental results Bernhard Kienzler *, Berthold Luckscheiter, Stefan Wilhelm 1 Forschungszentrum Karlsruhe, Institut fuÈr Nukleare Entsorgung (INE), Karlsruhe, Germany Received 7 April 2000; received in revised form 18 December 2000; accepted 2 January 2001
Abstract Results of leaching experiments using active and inactively simulated HLW glass in concentrated NaCl solution are described. Measured solution concentrations of glass components, ®ssion products and actinides are compared with computed data. The computed pH value corresponds with the ®ndings from experiments with inactively simulated glass samples. Moreover, the concentrations of silica and strontium can be described adequately by reaction path modelling. The computed U concentration is explained by the precipitation of schoepite or Na2U2O7. The computed Am concentration signi®cantly exceeds the measured data. This may be attributed to sorption processes on corrosion products of the glass, which are relevant also for lanthanide elements under the conditions of the experiments. This hypothesis is tested by solid solution approaches and by computing sorption of Am onto SiO2 precipitates. # 2001 Elsevier Science Ltd. All rights reserved. Keywords: Glass corrosion; Actinides; Geochemical modelling
1. Introduction The associated risk of a radioactive waste disposal system depends on the quantity of radionuclides mobilised in the case of groundwater (brine) intrusion to the waste. In the presence of water or brines, waste forms may be corroded and radioactive elements are leached. New solid reaction products will be formed, which are thermodynamically favoured under the given geochemical conditions. For performance assessment and longterm safety analyses, model calculations predicting the quantity of dissolved radionuclides in the solutions as a function of time are required. Within the time periods under consideration, the natural environment as well as engineered structures will change signi®cantly and an unavoidable uncertainty regarding the future state or evolution of the repository system has to be taken into account. Therefore, the geochemical isolation potential for radionuclides has to be evaluated [1]. * Corresponding author. Tel.: +49-7247-82-4467; fax: +49-724782-4308. E-mail address: [email protected] (B. Kienzler). 1 Present address: Colenco Power Engineering AG, Baden, Switzerland.
In the last years, experiments and related model approaches describing the leaching of waste glasses have been published. The models comprise dierent processes involved, such as kinetically controlled glass network dissolution, precipitation of secondary phases, ion exchange, silica diusion through amorphous surface layers and speci®c interactions between dissolved ions and the glass surface [2±8]. In the present study, a thermodynamic reaction path model [9] is applied and the results are compared with the experimental ®ndings concerning the behaviour of HLW glass in NaCl brines. The comparisons cover the dissolution model of the waste glass, the development of the geochemical environment and the concentrations and speciation of radionuclides. Comparisons of experiments to model predictions can indicate whether the database available for 25 C can be applied for the experimental conditions or not. Experiments with highly radioactive glass samples and with inactive glass (having the same chemical composition as the active glass but without radioactive isotopes and actinide elements) are used for the comparisons. The highly radioactive glass has been prepared by the Commissariat aÁ l'eÂnergie atomique's (CEA) Centre de la ValleÂe du RhoÃne, Marcoule, France. The glass composition is similar to that of the COGEMA
0956-053X/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved. PII: S0956-053X(01)00008-3
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glass R7T7 produced in the vitri®cation plant R7 and T7 at La Hague, France. Concentrations of radioactive elements, except for plutonium, are lower in the CEA glass than in the COGEMA R7T7 glass, because a different waste was used by CEA.
2. Experimental Static corrosion tests were performed in dierent brines relevant for minerals present in German salt domes. The composition of the NaCl brine (`solution 3' [10]) used for the experiments described here has the following composition: 6.233 mol (kg H2O) 1 NaCl, 0.032 mol (kg H2O) 1 CaSO4 (320 g dm 3 NaCl and 3.8 g dm 3 CaSO4). At 55 C pH is 5.9 and the density is 1.203 g cm 3. The composition of the HLW glass is reported elsewhere [9]. The tests with the highly radioactive glass samples were performed in tantalum lined autoclaves at temperatures of 110 and 190 C over periods of 100 and 365 days. By using 3.835 g of glass powder with a grain size of 71±100 mm and 40 ml solution, the ratio of the glass surface area to the solution volume was about 5000 m 1. The atmosphere in the vessels consisted of argon. At the end of the duplicate tests, the autoclaves were cooled to ambient temperatures. Afterwards, pH and Eh were measured by means of a Ross type electrode and a Pt electrode, respectively. Measured pH was corrected for liquid junction potential [11]. Due to the Fe concentration in the solutions, we assume the redox potential is buered by Fe(II)/ Fe(III). Measured Eh values scatter between 0 and 150 mV. The solutions were ®ltered through 0.45 mm ®lters and aliquots of the ®ltrates were ultra®ltered through an 18 AÊ membrane ®lter to detect the presence of colloids. Prior to analyses, radionuclides had to be separated, both from the highly concentrated salt solution and from other elements. Separation was performed either by liquid-liquid extraction or by extraction chromatography procedures. Concentrations of glass constituents, ®ssion products and actinides were analysed by ICP±AES, ICP±MS and nuclear analytical methods, such as a-, g-spectroscopy and liquid scintillation counting (LSC). Details of the experiments are described elsewhere [9]. Sorption behaviour of americium onto the alteration layer of the corroded glass was investigated by speci®c experiments. In addition, the homologue elements neodymium and cerium were studied as simulates for the trivalent actinides. Pre-corroded simulated HLW glass powder was applied as substrate. Sorption isotherms for Am, Nd and Ce were determined, using various initial concentrations of these elements. The tests are performed in brines and diluted aqueous solutions at 80 C in the pH range 2±10.
3. Model The radionuclide concentrations and possible precipitation of solid phases were modelled with the software package EQ3/6 Rel.7.2a [12]. The initial solution composition including all species involved in further reactions was computed. To simulate corrosion of the waste glass by reaction path modelling, small amounts of a `special reactant' having the element composition of the waste glass were `added' numerically to the solutions and the new equilibrium was calculated by EQ6. A thermodynamic equilibrium between the solution and solid phases was assumed for each step. This procedure is implemented as `relative rate law' in the EQ6 code. The in®nitesimal steps of the EQ6 run were considered to represent the progress of waste form dissolution. This reaction path model had been previously used to evaluate the reaction progress of HLW glass in magnesium-rich brines [9]. It computed the evolution of element concentrations over a large range of the reaction progress. The calculated data were compared with experimental ®ndings obtained for various surface to volume ratios. In the reaction path computations, the redox potential could not be treated as an independent variable because oxygen is not considered in the used `special reactant' approach. Therefore, the Eh was ®xed by the O2 (gas) fugacity as an input parameter to EQ3/6. 4. Database Predictive geochemical modelling of thermodynamic properties of radionuclides in complex aquatic systems requires a thermodynamic database as well as an appropriate thermodynamic model. The ion interaction model (Pitzer approach) is applicable to highly concentrated aqueous solutions. The database applied for the present computations is based on the Pitzer `data0' ®le provided with EQ3/6 code including additional data for uranium [13,14] and other actinides [15,16] measured in our institute. Based on available experimental data, comprehensive sets of thermodynamic data (chemical potentials, thermodynamic constants) and model parameters (Pitzer parameter) for the major homogeneous and heterogeneous equilibria of americium/curium(III) and neptunium(V) were predicted [17]. For the high pH range expected to prevail for glass waste form dissolution in NaCl solutions, Pitzer interaction coecients for silicate and alumina species have been incorporated into the database [18]. Reactions and corresponding log K data for the radionuclides are given in Tables 1 and 2. Uranium data are taken from Refs. [19±21], data of plutonium and other tetravalent actinides were evaluated by Neck [22]. Since measured concentrations in the corroding solutions are available only for room temperature, all simulations were performed at 25 C.
B. Kienzler et al. / Waste Management 21 (2001) 741±752
5. Comparison between measured and computed data and discussion This section presents a comparison between the measured and computed HLW glass dissolution progress and the extent of dierences between experimental and numerical results. Furthermore, the reasons for the observed dierences are discussed. It will be shown for which cases model and database describe the experiments suciently well and where the model or database are incomplete or where the reaction mechanisms are not yet understood.
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5.1. Dissolved glass components and pH Computation of the reaction of the waste glass in NaCl brine was started at an initial pH of 7.5. Computed and measured concentrations and pH are presented in Fig. 1 as a function of the progress of glass dissolution. Depending on the progress of the dissolution of the HLW glass the computed pH increased to 8.5. The experiments covered a reaction progress between 0.005 and 0.016 kg (kg H2O) 1 glass in the case of active samples and up to 0.026 kg (kg H2O) 1 glass for inactive samples. The experimental pH for inactive
Table 1 Thermodynamic data of aqueous radionuclide species applied for the model calculations Reaction
Log K
Ref.
Uranium VI VI VI VI VI VI VI VI VI IV IV IV IV
UO2+ 2 UO2OH+ UO2(OH)2(aq) UO2(OH)3 UO2(OH)24 (UO2)2(OH)2+ 2 (UO2)3(OH)2+ 4 (UO2)3(OH)+ 5 (UO2)3(OH)7 3+ UOH U(OH)2+ 2 U(OH)+ 3 U(OH)4(aq)
() () () () () () () () () () () () ()
2 H++U4++H2O+0.5 O2(aq) 1 H++UO2+ 2 +1 H2O 2 H++UO2+ 2 +2 H2O 3 H++UO2+ 2 +3 H2O 4 H++UO2+ 2 +4 H2O 2 H++2 UO2+ 2 +2 H2O 4 H++3 UO2+ 2 +4 H2O 5 H++3 UO2+ 2 +5 H2O 7 H++3 UO2+ 2 +7 H2O 1 H++U4++H2O 2 H++U4++2 H2O 3 H++U4++3 H2O 4 H++U4++4 H2O
33.78 5.70 11.50 19.22 33.03 5.68 11.93 15.59 31.05 0.40 0.40 3.80 9.30
[19] [19] [20,21] [19] [19] [19] [19] [19] [19] [19] [12] [12] [19]
Neptunium V V V IV IV IV IV
NpO+ 2 NpO2OH(aq) NpO2(OH)2 NpOH3+ Np(OH)2+ 2 Np(OH)+ 3 Np(OH)4(aq)
() () () () () () ()
3 H++0.25 O2(aq)+Np4++1.5 H2O 1 H++H2O+NpO+ 2 2 H++2 H2O+NpO+ 2 1 H++Np4++H2O 2 H++Np4++2 H2O 3 H++Np4++3 H2O 4 H++Np4++4 H2O
9.87 11.31 23.54 1.00 2.80 5.80 9.60
[12] [15] [15] [22] [22] [22] [22]
Plutonium VI V IV IV IV IV III III III III III
PuO2+ 2 PuO+ 2 PuOH3+ Pu(OH)2+ 2 Pu(OH)+ 3 Pu(OH)4(aq) PuOH2+ Pu(OH)+ 2 Pu(OH)3(aq) PuCl2+ PuCl+ 2
() 2 H++Pu4++H2O+0.5 O2(aq) () 3 H++0.25 O2(aq)+Pu4++1.5 H2O () 1 H++Pu4++H2O () 2 H++Pu4++2 H2O () 3 H++Pu4++3 H2O () 4 H++Pu4++4 H2O () 1 H++Pu3++1 H2O () 2 H++Pu3++2 H2O () 3 H++Pu3++3 H2O ()Pu3++1 Cl ()Pu3++2 Cl
8.13 2.94 0.60 0.60 2.29 7.98 2.80 15.19 25.69 0.24 +0.74
[12] [12] [22] [22] [22] [22] [16] [16] [16] [16] [16]
Americium III III III III III
AmOH2+ Am(OH)+ 2 Am(OH)3(aq) AmCl2+ AmCl+ 2
() 1 H++Am3++1 H2O () 2 H++Am3++2 H2O () 3 H++Am3++3 H2O ()Am3++1 Cl ()Am3++2 Cl
2.80 15.19 25.69 0.24 +0.74
[16] [16] [16] [16] [16]
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samples corroded at 190 C with a high reaction progress is about 0.3 pH units below the computed pH. Radioactive samples show a signi®cant lower pH in the range of 7.2. In order to test the model approach and to compare the pH at experimental temperatures, the temperature eect of pH was computed for the pure NaCl brine according to the procedure described by Ref. [12]: the quenched (cooled) solution's composition was used to de®ne the system at room temperature. Then, a reaction path for temperature increase was computed by EQ6. It was essential to take into account the complete composition of the solution. Using this procedure, the computed pH decrease of the NaCl brine was about 0.2 units and about 0.6 unit at 110 and 190 C, respectively. Since the thermodynamic data-set for highly concentrated solutions is incomplete, computation of the eect of the temperature quenching on the pH of the experimental solutions is not possible with this procedure. The lithium and boron concentrations are used as a measure of the dissolution progress of the HLW glass matrix. Boron (B) concentrations are used to calculate the reaction progress. However, B could used to scale the glass dissolution up to the dissolution of 0.1 kg (kg H2O) 1 glass. Above this value, concentration of B is
limited by precipitation of borax. The Li concentrations calculated with the relative rate model agree well with the experimental data. According to the calculations, amorphous silicate and calcium bearing solids, such as the clay minerals nontronite and mesolite were predicted to precipitate. This is corroborated by the fact that clay mineral corrosion products of HLW glass in NaCl brine are experimentally detected [23]. Measured Si concentration are in excellent agreement with the computed values. Si solubility is controlled by amorphous silica. Calcium concentrations in the experiments are found to be only two times lower than computed values. Anhydrite precipitation is not computed. 5.2. Fission products Fig. 2 compares experimental and calculated concentrations of the ®ssion products Sr and Mo. Measured and computed strontium concentrations agree well for both active and inactive glasses. The computed increase of the Sr concentration is parallel to that for the Li concentration and lies in the range of the experimental scatter. The experimental Mo concentrations are in the same range of 510 5 to 210 4 mol (kg
Table 2 Thermodynamic data of solid radionuclide phases applied for the model calculations Reaction U
UO3.2H2O Schoepite Na2U2O7 Na-Becquerelite CaU2O7 Becquerelite (UO2)2(SiO4).2H2O Soddyite U(OH)4(am) USiO4
Log Ksp +UO2+ 2 +3
+
() 2 H
H2O
+ () 6 H++UO2+ 2 +2 Na +3 H2O
Refs.
5.37
[13]
22.60
[19]
26.45
[14]
()
2+ 6 H++UO2+ +3 H2O 2 +Ca
()
4 H++2 UO2+ 2 +1 SiO2(aq)+4 H2O
6.03
[27]
() ()
4 H++U4++4 H2O 4 H++U4++1 SiO2(aq)+2 H2O
8.05
[22] [12]
Conite Np(OH)4(am) NpO2(s) Np2O5 NpO2OH(s) NpO2OH(am)
() () () () ()
4 H++Np4++4 H2O 4 H++Np4++2 H2O 2 H++H2O+2 NpO+ 2 1 H++H2O+NpO+ 2 1 H++H2O+NpO+ 2
0.81 7.8 7.98 4.52 5.22
[22] [15] [15] [15] [15]
Pu
Pu(OH)4(am) Pu(OH)3(am) PuO2OH(am)
() () ()
4 H++Pu4++4 H2O 3 H++Pu3++3 H2O 1 H++PuO+ 2 +3 H2O
2.51 14.49 5.40
[22] [16] [12]
Am
Am(OH)3(am) Am(OH)3(cr)
() ()
3 H++Am3++3 H2O 3 H++Am3++3 H2O
17.00 15.02
[16] [16]
Mo
CaMoO4.0.2H2O Powellite MgMoO4.2H2O MgMoO4.0.52H2O NaNdMoO4 Nd(OH)3(s)
() 1 Ca++1 MoO24 +0.2 H2O
7.95
[32]
1.34 2.10 8.94 15.59
[32] [32] [32] [12]
Np
Nd
() 1 Mg++1 MoO24 +2 H2O () 1 Mg++1 MoO24 +0.5 H2O () 0.5 Na++0.5 Nd3++1 MoO24 +0.16 H2O () 3 H++Nd3++3 H2O
B. Kienzler et al. / Waste Management 21 (2001) 741±752
H2O) 1, for both inactive and active experiments at 110 and 190 C. Assuming that Mo concentration is not controlled by solid phases, the model predicts concentrations in the upper range of the experimental scatter. Consideration of powellite (CaMoO4.0.2 H2O) or trace amounts of Na0.5Nd0.5MoO4.0.2 H2O would result in a decrease of the Mo concentration by almost one order of magnitude. Moreover, the evolution of the computed Mo concentration as function of the reaction progress is completely dierent from measured [Mo]. However, the formation of powellite has been detected by SEM and EDX (micro X-ray analyses) in many corrosion experiments with simulated HLW glass in NaCl and in MgCl2 solutions [9,24,25]. The discrepancy between computed and measured Mo concentrations can not yet be explained. Computations show that the solubility of neodymium is controlled by solid Nd(OH)3 at a concentration of 10 6 mol (kg H2O) 1. Measured concentrations are below the detection limit (110 6 mol (kg H2O) 1). Therefore these data are not shown in Fig. 2. 5.3. Uranium Dierent from the studied components of the glass matrix and the ®ssion products under investigations, the
745
actinide elements can exist in several oxidation states. Fig. 3 shows the Eh±pH diagram for dissolved uranium species. Ranges of Eh and pH data measured in dierent experiments are given as well. Under the conditions prevailing in the experiments, U exists in the pentavalent or hexavalent state. Results of reaction path calculations for the uranium are shown in Fig. 4. Uranium concentrations in NaCl brine are about 210 6 mol (kg H2O) 1. Computations were performed for redox potentials above 50 mV where uranium is in the hexavalent state. Computations indicated that the rather high U concentration at the computed pH of 8.5 is compatible with schoepite (UO3.2H2O) or Na2U2O7 as solubility-controlling solids. The predominating aqueous U species is computed to be UO2(OH)3 . For the computed Si concentrations the precipitation of uranium silicate phases, such as uranophane (Ca[(UO2)SiO3OH]2.5H2O) or soddyite ((UO2)2SiO4.2H2O) is expected from the model calculations and uranium silicate phases have been observed in the experiments [25]. For the dissolution of soddyite, the EQ3/6 database provides log K=0.39. Application of this value results in [U] 310 9 mol (kg H2O) 1 which is not compatible with measured uranium concentrations. However, Kosp values given in literature dier signi®cantly.
Fig. 1. Computed and measured concentrations and pH of the HLW glass matrix elements as a function of the reaction progess in NaCl brine. Upper abscissa: experimental time scale. Measured pH: +: T=110 C, inactive R7T7 samples; X: T=190 C, inactive R7T7 samples; &: T=190 C, active R7T7 sample.
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Measured solubility products by Nguyen et al. [26] and Moll et al. [27] have resulted in log K=5.74 and log K=6.03 0.53, respectively (see Table 2). Uranium concentrations of 210 6 mol (kg H2O) 1 are computed, if a log K=6.03 is used in EQ3/6. This concentration corresponds to the experimental ®ndings. On the other hand, lower log K values have also been reported in literature. Measured solubility for synthetic soddyite phases have resulted in log K=3.9 0.7 [28]. The solubility has been determined at an initial pH=3. In Ref. [29] log K for a natural soddyite has been determined to be 3.0 2.9 at a pH8.8. For uranophane published solubility data dier signi®cantly, too. In Ref. [26] log K=8.5 1.5 is reported, experiments with synthetic uranophane have shown log K=11.7 0.6 [28] and 7.8 0.8, respectively, for a natural uranophane at a pH between 8.0 and 9.3. 5.4. Plutonium Pu can occur in dierent oxidation states, too. Fig. 5 shows the Eh±pH diagram for dissolved Pu species and the ranges of Eh and pH data measured in the experiments. Under these conditions, Pu exists only in tetravalent state and solid Pu(OH)4(am) controls solubility. The computed concentration of 610 10 mol (kg
H2O) 1 is obtained for the glass experiments. Fig. 4 shows a good agreement between the computed results and the experimental ®ndings for Pu. 5.5. Other actinides In the experiments dissolved Np and Am concentrations were determined, as well. Under the conditions relevant for these experiments, Am exists only in the trivalent state. Np, however, can occur in tetra- or pentavalent states. Fig. 6 shows the Eh±pH diagram for dissolved Np species and the measured Eh and pH data. The Eh±pH range observed in the experiments is just below the phase boundary between Np(V) and Np(IV) in the tetravalent region. In the case of Np(IV), solid Np(OH)4(am) would limit solubility to 110 9 mol (kg H2O) 1. This concentration is one order of magnitude below measured data which scatter between 310 8 and 610 9 mol (kg H2O) 1 (see Fig. 4). For the pentavalent state calculated Np concentrations are in the range of the measured values. For this condition, solid Np(OH)4(am) or Np2O5(cr) may control the dissolved concentration of 210 7 mol (kg H2O) 1 or of 210 6 mol (kg H2O) 1, respectively. Signi®cant sorption of Np(V) is reported in literature at the pH relevant for glass corrosion in NaCl solution
Fig. 2. Computed and measured concentrations of Sr and Mo released from HLW glass in NaCl brine. ^, Sr inactive glass sample (T=110 C); , Mo inactive glass sample (T=110 C); *, Sr inactive glass sample (T=190 C); & Mo inactive glass sample (T=190 C); &, Sr active R7T7 sample; ^, Mo active R7T7 sample.
B. Kienzler et al. / Waste Management 21 (2001) 741±752
Fig. 3. Eh±pH diagram of uranium (chemical activity of U: 10 of aqueous U species.
6
747
mol (kg H2O) 1). Dotted lines: stability ®eld of water; dashed lines: stability ®elds
Fig. 4. Computed pH and computed and measured uranium, neptunium, and plutonium concentrations released from the HLW glass in NaCl brine. *, uranium experimental ^, neptunium experimental &, plutonium experimental.
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Fig. 5. Eh±pH diagram of plutonium (chemical activity of Pu: 10 10 mol (kg H2O) 1). Dotted lines: stability ®eld of water, dashed lines: stability ®elds of aqueous Pu species, solid line: stability ®eld of Pu(OH)4(am).
Fig. 6. Eh±pH diagram of neptunium (chemical activity of U: 10 ®elds of aqueous Np species.
7
mol (kg H2O) 1). Dotted lines: stability ®eld of water, dashed lines: stability
B. Kienzler et al. / Waste Management 21 (2001) 741±752
[30]. A shift of Np to the pentavalent state may be caused by a-radiolysis. The total a-decay rate of the glass samples is about 2.5108 Bq/g. Taking a G-value of 2 into account, at the surface of the glass powder about 10 10 mol (kg H2O) 1 oxidising species are formed per second in a 20-mm surface layer of the glass powder. These oxidising species may change the redox conditions locally with respect to the bulk of the solution and transfer Np species into the pentavalent state. In the case of americium, the measured concentrations are in the range of 110 10 mol (kg H2O) 1, whereas the calculated concentrations are by three orders of magnitude higher (Fig. 7). Measured Am concentration is close to the detection limit and do not allow application of speciation techniques such as time resolved laser ¯uorescence (TRLF) spectroscopy. The in¯uence of solid solutions to limit the Am concentration has been investigated. Solid solutions are phases of variable compositions which can be considered as mixtures between two or more `end-member' minerals. EQ3/ 6 provides features to treat solid solutions. The thermodynamic activity of a solid solution component is de®ned by multiplication of the mole fraction of the component with its activity coecient in the solid phase. For the computations, an ideal solid solutions approach is chosen using activity coecients of 1. Two
749
types of end members are considered in this study: hydroxides and molybdates. One model considers the hydroxide end members of Am, lanthanides and aluminium, the other approach uses the two amorphous hydroxides Am(OH)3(am) and SiO2(am) (see Fig. 7). Calculations including possible precipitation of solid containing Am are not successful in resolving the discrepancy between the measured and calculated concentrations. Molybdate phases do not signi®cantly control the Am concentration. In addition, consideration of the presence of traces of CO2 gas in the leaching vessels resulting in the formation of Na±Am carbonates would not explain the low Am concentrations. The sorption experiments with Am onto pre-corroded simulated HLW glass powder show high retention factors (Rs) up to 105 ml/g independent on the initial concentration of the element. This ®nding indicates a sorption process. Details of the experiments and the results are reported elsewhere [31]. Measured distribution coecients for Am, Nd and Ce are given in Fig. 8 as a function of pH. These data can be used to interpret the discrepancy between measured and calculated Am concentrations. We assume that Am is sorbed onto precipitated, amorphous SiO2. The amount of precipitated SiO2(am) is computed by the model. The sorption data and the amount of precipitated SiO2(am)
Fig. 7. Americium concentrations released from the HLW glass in NaCl brine: measured [Am] and results of computations applying dierent model approaches. ~, americium experimental.
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Fig. 8. Measured distribution coecients for Ce, Nd and Am on the corrosion products of the corroded HLW glass matrix in NaCl brine. &, Am after 36 days experiment; *, Am after 10 days experiment.
are used to calculate an expected value for the Am concentration. As shown in Fig 7, better agreement with the measured Am concentrations is obtained by this model. Similar behaviour is also expected for Nd. Systematic investigations of the relevant processes, however, are extremely dicult, because the concentrations are close to the detection limit. As shown in Fig. 5, Pu concentration is controlled by solubility of Pu(OH)4(am). Sorption of Pu may be relevant, if Pu is shifted into the trivalent or pentavalent oxidation state. Corresponding experiments with HLW glass in NaCl brines have not been conducted up to now.
6. Summary and conclusions 6.1. Redox potential Dissolved concentrations computed for the actinides and for iron depend signi®cantly on the redox potential. The rather high calculated U concentrations are in agreement with the experiments. However, several U(VI) solids may account for solubility limiting phases. Experimental Eh and pH indicate clearly that Pu is well situated in the stability ®eld of Pu(IV). Based on measured pH and Eh of the bulk solution, one would expect Np to be in the tetravalent state, too. The measured Eh and pH are close to the stability ®eld of Np(V). One can assume that at the surface of the glass powder
radiolytically formed oxidizing species change locally the redox state of the solution and oxidize Np species. 6.2. Temperature eect Concentrations measured after cooling the solutions to ambient temperature can be explained by the computations for 25 C. The pH measured for 110 and 190 C is about 0.3 units lower than computed. The model is not able to describe the pH under radiation by the highly radioactive samples. The computed behaviour of Mo diers considerably from the experimental ®ndings. In the pH range between 3 and 4 the solubility constant Kosp of powellite is almost independent of the temperature [32]. However, at pH values above 6 preliminary experiments show a strong increase of solubility with temperature. We assume that the higher solubility at the experimental temperature may contribute to the dierences between measured and calculated Mo concentrations. High-temperature Pitzer data for important species and the identi®cation of temperature-dependent sorption reactions for the actinide-containing solids are not available. At present the model is not well de®ned for elevated temperatures. Ksp and stability constants of dissolved species can be extrapolated for higher temperatures by means of H and entropy data, such as performed in Ref. [33]. However, it has to be recognised, that for most of the equilibrium constants signi®cant errors persist in highly concentrated aqueous solutions.
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6.3. Thermodynamic data
References
The measured uranium concentrations indicate that Na2U2O7, schoepite or a silicate phase such as soddyite can be the limiting phase. Especially for silicate phases, the solubility products given in literature dier by several orders of magnitude. Under the prevailing conditions, these phases have been detected at the surface of the corroded glass. Applying the latest published log Kosp for soddyite [27], calculations result in a good agreement with measured U concentrations. However, the variation of the solubility products of silicate phases reported in literature is considerable. For the tetravalent state, neptunium concentrations are calculated to be one order of magnitude below the experimental ®ndings. Concentrations in the range of the experiments are computed only for Np(V). For solid Np phases, also rather large discrepancies of log Kosp for oxides and hydroxides are reported in literature. In the case of Am, the experiments show signi®cantly lower dissolved concentrations than predicted by assuming precipitation of an Am(OH)3 phase. Several mechanisms are investigated which may account for this, such as the formation of solid solutions. The discrepancy is resolved when sorption of Am onto SiO2(am) is considered. A simple solid solution approach, such as presented in this study, cannot yet describe the measured data. Additional work is required to qualify such an approach on the basis of the endmembers Am(OH)3(am) and SiO2(am). For plutonium a good agreement between the experimental ®ndings and the computed Pu(OH)4(am) solubility is obtained. It is important to note the in¯uence of the prevailing Eh and pH conditions: if Pu(III) would dominate, a signi®cantly higher solubility would be expected. In this case, sorption processes similar to those modelled for americium have to be considered in order to describe the experimental concentrations. The current study demonstrates the capability and the limits of geochemical modelling to predict dissolution of HLW glass, the development of the geochemical environment, as well as the concentrations and the speciation of mobilized radionuclides. Since thermodynamic data at higher temperatures are not experimentally veri®ed, further experimental work is needed to understand the mechanisms and reactions controlling dissolved actinide and ®ssion product concentrations.
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Acknowledgements The work is supported by the Bundesamt fuÈr Strahlenschutz (BfS) within the project: ``Erstellung eines integrierten Nahfeldmodells von Gebinden hochaktiver AbfaÈlle im Salzstock Gorleben: geochemisch fundierter Quellterm fuÈr HAW-Glas, abgebrannte Brennelemente und Zement''.
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