Uranium release from contaminated sludge materials and uptake by subsurface sediments: Experimental study and thermodynamic modeling

Uranium release from contaminated sludge materials and uptake by subsurface sediments: Experimental study and thermodynamic modeling

Accepted Manuscript Uranium release from contaminated sludge materials and uptake by subsurface sediments: Experimental study and thermodynamic modeli...

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Accepted Manuscript Uranium release from contaminated sludge materials and uptake by subsurface sediments: Experimental study and thermodynamic modeling Olga L. Gaskova, Anatoly E. Boguslavsky, Olga V. Shemelina PII: DOI: Reference:

S0883-2927(14)00331-X http://dx.doi.org/10.1016/j.apgeochem.2014.12.018 AG 3399

To appear in:

Applied Geochemistry

Please cite this article as: Gaskova, O.L., Boguslavsky, A.E., Shemelina, O.V., Uranium release from contaminated sludge materials and uptake by subsurface sediments: Experimental study and thermodynamic modeling, Applied Geochemistry (2015), doi: http://dx.doi.org/10.1016/j.apgeochem.2014.12.018

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Uranium release from contaminated sludge materials and uptake by subsurface sediments: Experimental study and thermodynamic modeling Olga L. Gaskovaa,b*, Anatoly E. Boguslavskya, Olga V. Shemelina a

Institute of Geology and Mineralogy, Siberian Branch of RAS, 3, Koptyug Ave., 630090 Novosibirsk, Russia Novosibirsk State University, 2, Pirogova Str., 630090 Novosibirsk, Russia [email protected]

1. Introduction

Mineral weathering and anthropogenic contamination release uranium into natural systems. Uranium is a common contaminant of concern in aquifers near nuclear waste management facilities around the world. The Electrochemical Plant (ECP) (Zelenogorsk, Krasnoyarsk Krai, Russia) and the Angarsk Electrolysis Chemical Complex (AEСC) (Angarsk, Irkutsk Region, Russia) have produced enriched uranium since the 1960’s. These plants produce solid and liquid radioactive wastes, all of which belong to the category of low-activity wastes. Liquid radioactive wastes (LRW) are formed during the neutralization of a nitric acid slurry after uranium extraction by hydrated lime Ca(OH)2, causing the pH to elevate above 9. Nitric acid partially bonds to calcium nitrate Ca(NO3)2; however, the residual NO3- content can reach several g/l in the resulting solution (Gaskova, Boguslavsky, 2013). Previous studies have shown that uranium in groundwater near the ECP and AECC is below the recommended MAC (maximum allowable concentration) standards, but alkaline nitrate solutions rich in Na, Ca and Mg penetrated and spread into aquifers and bogs (Gas’kova et al., 2011, Boguslavskiy et al., 2012). In this research, a number of experimental laboratory studies were conducted (a) to investigate the leaching of uranium-containing sludge by regional groundwater and (b) to measure the ability of host sediments to retain major and minor solution components as a function of the water/rock (W/R) ratio. A series of thermodynamic simulations were carried out _____________________________________________________________________ * Corresponding author. Tel.: +7-913-758-0557; fax: +7-383-333-2792. E-mail addresses: [email protected] (O.L. Gaskova)

to reveal the factors affecting the redistribution of the radioactive uranium from permanent repositories into the aquifers. The purpose of this work was to quantify the geochemical processes that control uranium leaching from the various sludges and to measure U(VI) sorption by two types of sediments (clay loam and decanted fine fraction of sand) from two solutions (natural neutral groundwater and technological alkaline waste water). The geochemical modeling code “HCh” (Shvarov, 2008) was applied to simulate these two chemical weathering processes. The experimental data serves as the basis for model parameterization. Thermodynamics alone is usually not sufficient to develop a realistic model of natural and technogeneous mineral dissolution/precipitation due to uncertainties in the thermodynamic data, the impossibility of accounting for heterogeneous transport behavior in clay rocks, and the absence of kinetic data for all of the phases considered. Therefore, the experimental results were compared with the thermodynamic

calculations

to

improve

the

experimental

procedures

and

model

parameterization. Uranium poses both radiological cancer risks and other, non-cancerous chemical hazards. The World Health Organization recently recommended a guideline level of 15 μg/l based on chemotoxic effects (World Health Organization, 2005). There is a higher guideline level of 100 μg/l for drinking water in Russia.

2. Experimental approach

2.1. Materials, methods and results of the dynamic experiments Under geological conditions, sludge materials change as they undergo various reactions, such as dissolution into groundwater and formation of secondary minerals. High-pH drainage solutions with alkaline components affect the physical and chemical properties of the surrounding rocks. As LRW (sludge) is formed during neutralization of nitric acid slurry by hydrated lime Са(ОН)2 after uranium extraction, rapid calcification may be an important process in the interaction of alkaline waste solutions with aquifer materials.

All experiments were performed with solid samples from AEСC and ECP. Preliminary treatment of the samples included drying to a constant weight and grinding in a porcelain mortar. As samples came from oxic environments, no special precautions were taken to isolate them from air during the experiments. The AEСC sample was characterized by a predominance of fluorite, with minor gypsum and calcite constituents, whereas the ECP sample was characterized by a predominance of gypsum over minor calcite and fluorite constituents (Table 1). All samples contained an appreciable fraction of structureless components (up to 20 %). The total uranium content of AECC reached ~0.2%; U isolations (own uranium phases) were found in sludge by scanning electron microscopy (SEM) with energy-dispersive X-ray (EDX) spectroscopy. No Uisolations were found in ECP sludge materials, which contained ~0.01% of total U. Small amounts

of

para-alumohydrocalcite,

CaAl2(CO3)2(OH)4⋅6H2O,

and

ettringite,

Ca6Al2(SO4)3(OH)12 26H2O, are characteristic of cementitious materials used for long-term disposal of radioactive wastes in repositories. It is interesting that rapidcreekite, Ca2(SO4)(CO3)·4H2O, was found in the ECP sludge materials. This mineral has only been found in natural conditions at a very few locations and usually in association with gypsum and carbonate minerals (Onac et al., 2013). Two sets of similar experiments were carried out in flow-through reactors to model the processes of infiltration zone leaching. Glass cylindrical reactors (50 cm3) were charged with ~20 g of solid and were subjected to a constant flow rate of ~2.5 ml per hour of the leach solution (natural pure groundwater). The experiments lasted for more than 80 days. During that time, the mass of the solid material in the tubes decreased to 6% (AEСC) and 39% (ECP) of the initial mass. A sequential chemical extraction procedure has been implemented to partition uranium into five fractions (solid species): soluble, exchangeable, carbonate-bound, Fe-Mn oxide-bound and residual (Tessier et al., 1985). A description of the sequential extraction procedure used for this experiment follows. Fifty milliliters of double-deionized water were added to 1 g (dry weight) of

sludge material and shaken for 1 h. To the remaining sediment, 50 mL of 1 M CH3COONH4 was added and shaken for 1 h. The residual sediment was then exposed to CH3COONH4 (1 M) + HNO3 (1 M) to remove the carbonate. After each extraction step, samples were centrifuged for 15 min, the supernatant was filtered through a membrane filter and measured using ICP MS. The sequential extraction data indicated that the main U(VI) solid species in the sludge was related to the “carbonate” fraction (Table 1), accounting for 82 and 93 % of the uranium in the AECC and ECP materials, respectively. Moreover, the results of the sequential analyses revealed Table 1 Mineral composition of wastes, conditions and results of the dynamic experiments, main U-solid species Angarsk ECC waste (sludge) Zelenogorsk ECP waste (sludge) CaF2, less CaCO3, CaSO4⋅2H2O, Mg(OH)2, trace CaSO4⋅2H2O, CaCO3, Ettringite, SodiumAlum of Quartz, Ettringite, Para-alumohydrocalcite Na2Al2(SO4)4⋅24H2O, minor Quartz (5-10%), Fe CaAl2(CO3)2(OH)4⋅6H2O, Fe- and Al-oxides, Illite- (hydro)oxides, CaF2, trace of CaSO4⋅0.5H2O, smectite. Micron-sized isolations of U-Ca (Na, K) Ca2(CO3)SO4⋅4H2O and Chlorite. U-minerals were phases. Residual U content is ~0.2%. not detected. Residual U content is ~0.01%. Weight of sample (g), volume (V) and fluid flow rate (R) 19 g, V 4500 ml, R ~2.3 ml per hour 15 g, V 5000 ml, R ~2.6 ml per hour Weight leaching of solid sample (%), the electrical conductivity (mSm/cm) and initial and final U concentration (μg/L) 6%, 0.35 mSm/cm, 68 - 230 μg/L 39%, 1.8 – 0.4 mSm/cm, 6.5 – 2.1 μg/L The percentage of U extracted by the CH3COONH4 + HNO3 solution (bound to carbonate fraction) 82% 93%

that the carbonate fraction dominated the samples by mass and was higher for AECC (65%) than for ECP (31%). In comparison, 35% of ECP material, and 10% of uranium, was bound to the water-soluble fraction. This suggests that there is a high probability of the polluted plume spreading into the aquifers after the sludge flood during high water events. Fig. 1a shows how uranium, Ca, Na and anion concentrations changed over time in conjunction with electrical conductivity (EC, right axis). The conductivity of the AECC solutions is rather low (~0.32±0.03 mSm/cm) and constant, except for the first point at 0.38 mSm/cm. The approximately constant value is due to a simultaneous decrease in the concentration of sulfate ions and an increase in the concentration of bicarbonate ions. The concentration of fluoride ions was 16 mg/L in the first sample and fell gradually until reaching a constant level of 10 mg/L (fluorite is the main phase). There is a clear indication that the W/R

ratio controls pH values: the pH increased from 7.4 to 8.4 over the course of the experiment. Fast dissolution of ettringite (1), sodium alum (2) and gypsum explains the initial pH of 7.4, with the increase over time due to solution equilibration with calcite in a carbon dioxide atmosphere (3). In another study, ettringite precipitated during the first days of the column experiments in highly alkaline solutions with high activities of dissolved Ca, SO42- and Al (Hampson, Bailey, 1982). The initial concentrations of SO42- are connected with the following phase’s dissolution during the first stage of leaching: Ca6(Al(OH)6)2(SO4)3·26H2O + 4H+ = 6Ca2+ + 2Al(OH)4- + 3SO42- + 30H2O

(1)

Na2Al2(SO4)4⋅24H2O + 8H+ = 2Na+ + 2Al(OH)4- + 4SO42- + 24H2O

(2)

CaCO3(s) + H+ = Ca2+ + HCO3-

(3)

Fig. 1. Variation in the dissolved concentrations of selected ions, uranium and electrical conductivity (EC) with the W/R ratio during interaction of (a) AECC sludge and (b) ECP sludge with natural groundwater having initial EC ~0.3 mSm/cm. The variation of pH in both cases is from 7.4 to 8.4. There is a strong correlation between the dissolved U and HCO3- concentrations. In this experiment, the uranium concentration gradually increased from 68 to 259 µg/l. In the first leaching step, Cu, Mo, and W were released in trace concentrations of 55, 2 and 5 µg/L, respectively, from the AECC waste materials. In general, the low EC is due to the solubility of the predominant CaF2 and CaCO3 minerals (the solubility constants of gypsum, calcite and fluorite are 10-4.58, 10-8.48 and 10-10.6

mol/L, respectively). The relatively low sludge solubility explains the low total percentage of solids that was leached (~6%). Fig. 1b shows the results of leaching the gypsum-dominated ECP sludge. The chemistry of the effluents was rather different due to the high initial EC value (1.85 mSm/cm) and SO42content (4 g/L). The concentration of SO42- was higher than expected based only on the CaSO4⋅2H2O solubility. As in the previous case, the other sulfate minerals (ettringite, sodium alum, etc.) are unstable under atmospheric conditions of 10-3.5 atm. pCO2 (Table 1). The Al content during the first stage of leaching (W/R = 8, 36, 58, 73) was relatively high (from 2012 to 120 µg/L), indicating a contribution from the dissolution of the named phases. The effect of the W/R ratio on the decrease in EC and sulfate ions was drastic: at W/R >100, these parameters reach natural groundwater levels. The uranium concentration was lower than 6.5 μg/L because the initial solid U content was ~0.01%. At the same time, the concentrations of W and Mo were highest for the samples from the first stage: 273 and 30 μg/L, respectively. Thus, the weakly leachable AECC sludge yields solutions with high uranium concentrations, whereas the highly soluble ECP sludge does not supply uranium in solution (<6.5 μg/L). The different uranium behaviors (Fig. 1a,b) are due to the solubilities of different U solid species. AECC sludge contains a fine mixture of uranium minerals — CaUO4, Na2U2O7, UO2CO3 (rutherfordine), K3NaUO2(CO3)3H2O (grimselite), Na2CaUO2(CO3)3(H2O)5 (andersonite) — as well as a small amount of uranium incorporated into the structure of other CO32- minerals (e.g., CaCO3) and fluorite. It is likely that the geometries of the uranyl species in the calcite cause a significant disruption of the local structure. This suggests that uranyl incorporation should increase the solubility of calcite (Reeder et al., 2001). ECP sludge primarily includes Ucontaining Ca carbonate phases and a small proportion of water-soluble sulfate minerals (< 5%); the U total content of the ECP slags was approximately 0.01%. The results of Tits et al. (2011) clearly show that U(VI), a radionuclide present in radioactive waste, is strongly retained by C-SH phases, which are major components of cementitious materials. The portlandite (Ca(OH)2)

fraction stabilizes both the mineral composition of hardened cement paste and the pH of cement pore water. In our case, the slow degradation/dissolution of sludge materials takes place in CO2containing solutions and at pH values of 7.4 – 8.4.

2.2. Materials, methods and results of the sorption experiments A second set of experiments was carried out in flow-through reactors to model the attenuation of uranium during infiltration of contaminated groundwater or waste solutions (the liquids remaining after the sludge settles in the tailings pond) through the host rocks. Both groundwater and waste solutions were contaminated with 35 ± 2 μg/L of U (1.5⋅10-7 mol/L). In this case, all of the U(VI) phases are undersaturated. The reagent used to prepare the test solutions was UO2(NO3)2 (Table 2). Table 2 Selected chemical data for the input solutions in column experiments EC Constituent рН Na Mg Ca Mn Ni Cu Unit of measurem. Groundwater +U Waste water +U

Zn

Zr

Mo

W

µSm/cm

mg/L

mg/L

mg/L

µg/L

µg/L

µg/L

µg/L

µg/L

µg/L

µg/L

U µg/L

7.4

308

12

11

73

14

<10

<10

<10

<0.1

2.0

0.15

36

9.4

7640

1400

6.9

130

110

1400

310

32

3.2

38

3.3

33

Cylindrical glass reactors (30 cm3) were charged with 2.5 g of a solid. Two types of host rocks were used: a C-2-9 clay loam of 0.1-0.25 mm size (sampling depth was 9 m) and a C-0 fine fraction of <0.005 mm in size decanted from a sand layer (Table 3). The major minerals in the sediments were smectite, kaolinite, micas and illite, quartz, and feldspars, with minor amounts of amphiboles, chlorite and pyrite, and little or no calcite (Table 4). The samples varied in particle size distribution, with a higher percentage of clay/silt in the C-0 sample; the higher abundance of fine-grained material may have caused a difference in U(VI) adsorption. The experiments lasted for 25 – 36 days. During that time, the solid material in the tubes (in the flow-through reactors) remained homogeneous without any noticeable change, but the infiltration coefficient was not constant.

Table 3 Weight concentrations of major constituents sorption, % Al2O3 Fe2O3 MnO Sample SiO2 С-0 47.16 19.02 14.25 0.11 С-2-9 60.50 17.53 5.22 0.05

in the samples used for column experiments on uranium MgO CaO Nа2O K2O SO3 TiO2 LOI U,ppm 3.87 1.78 1.06 2.50 0.04 0.95 8.80 <0.5 2.02 1.65 1.76 2.87 0.03 1.00 5.26 2

Table 4 Mineral composition of the wall-rock samples near the storage facilities of the Angarsk ECC (2.5 g by weight) and type of the input solutions C-0 - fine fraction decanted from a sand (<0.005 mm) C-2-9 - clay loam (0.1 - 0.25 mm) Kaolinite, Smectite, Illite are more then 30%, Quartz, Smectite, Kaolinite, Micas, Quatz, Kfsp, Ab-An, Kfsp, Ab-An, Chlorite, Vermiculite, traces of traces of Amphiboles, Apatite, Calcite or Siderite. Amphiboles and Pyrite. U content is lower than Clay-sized mineral fraction is 50%, U content is detection limit. ~0.2 g/t. Type of solutions, contaminated by adding of 0.035±0.002 mg/L U, Background groundwater, Diluted waste solution, Background groundwater, Diluted waste solution EC 300 μSm/cm, EC 7650 μSm/cm, EC 300 μSm/cm, EC 7650 μSm/cm, рН 7.4 (Fig. 2a) рН 9.4 (Fig. 2b) рН 7.4 (Fig. 2c) pН 9.4 (Fig. 2d)

Fig. 2 shows how the U (the curves are digitized), pH and Ca and Na concentrations changed with time as a function of W/R ratio. In the groundwater (a, c), calcium prevails over sodium, but in the waste solutions (b, d), Na prevails over Ca. Manganese was used as an additional W/R interaction marker that was independent of U behavior. The initial pH of experiments (a) and (c) was 7.4. The pH increased with time and with the W/R ratio, then stabilized at 8.5 ±0.2 after 5-7 days. The initial pH of experiments (b) and (d) was 9.4, and it both increased and decreased over time. In the case of the C-0 fine fraction (b), the pH values hovered at 9.1 ± 0.1 and remained alkaline after 27 days (pH = 9). The pH of the C-2-9 sample solution was lower at the end of the experiments, at 8.4 (the same as in groundwater experiments). This means that the buffering capacity of the C-2-9 sample was higher, probably due to a carbonate phase in the coarser fraction. The lower pH may also be due to a decrease in the rate of solution filtration and a longer W/R interaction. The infiltration coefficient (seepage velocity) was ~ 4.8 cm/day during the groundwater flow through C-2-9 clay compared to approximately 0.5 cm/day through the C0 sediment. The fluxes of Ca and Na did not fluctuate greatly: they decreased slightly in the beginning of the experiment due to the ion exchange reactions with clay materials and increased

slightly at the end of the experiment (Fig. 2a,c). The electrical conductivity increased from 308 to 350±25 μSm/cm for both sediment types.

Fig. 2. Total dissolved the Ca, Na, U and Mn concentrations versus pH in dynamic experiments: interaction of (a) C-0 with groundwater; (b) C-2-9 with groundwater; (c) C-0 with waste water; (d) C-2-9 with waste water. The Uranium concentrations are numbered. Concentrations of sodium and calcium were divided by 100 (b, d) for the best U representation. Every first point indicates the initial concentration.

In spite of rather low concentrations of dissolved uranium (1.5⋅10-7 mol/L), the only solutions that did not exceed concentrations of 1.4 μg/L were those resulting from groundwater interaction with C-0 sediments up to a W/R ratio of 71 (Fig. 2a). The subsequent increase of the U concentration up to the initial concentration indicates the saturation of the low sorption capacity of the C-0 sample and a negative influence of dissolved CO2 (CO2 10-3.5 atm). After the C-2-9 sample interaction with the ground water, the first solution had a uranium content of 1.5 μg/L, and there was a sharp increase to 45±1 µg/L after 120 hours, when the W/R ratio was equal to 168 due to the higher infiltration coefficient (Fig. 2c). This indicates that the previously adsorbed uranium was washed away by the subsequent portions of the inflow solution. Moreover, the 2 ppm of U in the C-2-9 clay (Table 3) explains the elevated U concentrations (45 μg/L compared

to the initial 33 μg/L). The observed trends for Mn indicate that it was mobilized from the sediments up to 250 μg/L (Figure 2a). During the C-0 and C-2-9 sample interactions with wastewater, only the first solutions had low uranium concentrations (3.5 and 5.6 µg/L at W/R ratios of 10 and 24, respectively; Figure 2b,d). Subsequent solutions increased in U concentrations up to 40-55 μg/L, with slight decreases in the U concentration at W/R ratios greater than 400. Unlike the (a) and (c) experiments, in experiments (b) and (d), electrical conductivity decreased with the increasing W/R ratio in accordance with the decrease in the Na and Ca concentrations. The explanation for the decrease in all of the parameters (pH, U, Na, Ca, and EC) at the end of experiments (b) and (d) is the reduction of the infiltration coefficient from 3.6 to 1.2 cm per day. The observed trends for Mn indicate that it was immobilized from the waste solution onto the sediments (from Mn concentrations of 110 μg/L to values below the detection limit of 10 μg/L). The results indicate that the sorption effect of clay samples in flow-through experiments is rather low and takes place only at low W/R ratios. The mobilization (desorption) of U appeared to be highly dependent on the water ingress. This indicates that long-term storage sites of LRW should install antifiltration screens; it is possible that the consolidation of sedimentary host rocks (reduction of the filtration coefficient) during drainage outflow could also improve the situation.

3. Geochemical modeling 3.1. Thermodynamic data Thermodynamic modeling was performed with the “HCh” (version 4.4) computer code at 25°C and 1 bar total pressure using a free energy minimization algorithm and the UNITHERM database (Shvarov, 2008). We modeled the heterophase 14-component system H-O-Ca-Na-SiAl-Fe-Mg-K-S-N-C-F-U. Dissolved and adsorbed U-species, as well as U solid phases, were incorporated into the model using data from the literature. Usually in solutions without carbonates, the soluble U(VI) species include UO2+, (UO2)3(OH)5+, and (UO2)3(OH)7-.

Carbonates in natural solutions cause the formation of species such as UO2CO30, UO2(CO3)22-, and UO2(CO3)34-. The anionic U-carbonate species dominate at neutral and alkaline pH, and these species tend to cause the desorption of U(VI) from mineral surfaces and solubilization of U(VI) solids. Soluble U(VI) aqueous complexes, such as nitrate, sulfate and fluoride, are included in the database but do not play an essential role. The high electrolyte concentrations in sludge waste salt solutions at the AECC and ECP sites can cause different speciations from those in dilute solutions, e.g., Ca2UO2(CO3)30 or CaUO2(CO3)32-. The whole system included 76 possible solid phases and 97 aqueous species in solution. The precipitation of specific minerals, such as UO2(NO3)2·5H2O or Na4UO2(CO3)3, could occur during neutralization of the technological sodium nitrate solutions. Only the Gibbs free energy data of those mineral phases and solution species related to the system under consideration were included in Table 5. This thermodynamic

equilibrium

modeling

simulated

the

coupled

dissolution/precipitation

phenomena that was possible in a LRW repository system after vadose/groundwater intrusion (inundation or submersion). The modeling was also used to consider U migration beyond nuclear waste storage management facilities and its adsorption onto clay host rocks underlying the facilities. The thermodynamic model is based on the technique suggested earlier for simulating oreforming hydrothermal systems. This technique exploits the principles of a flow-through multistep reactor and is valid for (a) column leaching simulations where pure groundwater interacts with the sludge materials and (b) simulation of the sorption experiments where contaminated groundwater or a waste solution is transported through sediment samples of increasing W/R ratio (Shvarov, 2008). Where the timescales of adsorption reactions are much less than the residence time of groundwater or the waste solution in sediments, equilibrium expressions can be used to describe the adsorption reactions in contaminant transport models. When describing uranium interactions at the mineral/water interface, the solid phase itself is generally considered to be invariant (total site concentration, exchange capacity): only surface

complexation with surface hydroxyl groups and Na-exchange with uranyl-ion and H+ are taken into account. This assumption is less well supported Table 5 Selected Gibbs free energy data of the mineral phases and solution species as well as sorption reactions with corresponding thermodynamic equilibrium constants* Uranium minerals ΔG0298, kJ mol-1 Uranium minerals and species (CaO)2(UO3)(SiO2)2.5(H2O)5 -5867.00 Na2U2O7 (CaO)2(UO3)1.5(SiO2)2(H2O)5 -6011.30 UO2CO3 -6806.42 UO3*2H2O (CaO)3(UO3)1.5(SiO2)2(H2O)5.5 -1888.10 U(OH)2SO4 CaUO4 -3653.00a MgUO4 [(UO2)2SiO4]*2H2O -1864.70 UO2(am) UO2(NO3)2*5H2O -6192.30 Ca[(UO2)(SiO3OH)]2*5H2O -10221.10 Na2Al2(SO4)4⋅24H2O Ca2UO2(CO3)30 -3737.80 Na4UO2(CO3)3 CaUO2(CO3)320 Exchange reactions U(VI) surface reactions logK -5.0 XH + Na+ = XNa+ H+ >SOH + UO22+ + H2O = >SOUO2OH + 2H+ 2XNa + UO22+ = X2UO2 + 2Na+ 2+ -1.0 Total sample weight >SOH + UO2 + H2CO3 = >SOHUO2CO3 + 2H+ Total site concentration 0.2 mg-equiv/g Exchange capacity >SOH represents a surface site for uranyl adsorption; X – ion exchange position * - Chen et al., 1999, Hummel, et al., 2002; Gaona et al., 2012; Davis et al., 2004 Christov, 2002 (SodiumAlum)

ΔG0298, kJ mol-1 -2975.46 -1563.18 -1629.95 -1766.24 -1749.30 -1003.60 -911.45 -773.00 logK0 -2.0 3.5 2.5 g 0.01 mg-equiv/g (U-Ca species);

for minerals such as calcite or gypsum that show fast surface dynamics in contact with electrolyte solutions, but they are a very small portion of the C-0 and C-2-9 samples. We applied the Generalized Composite (GC) approach and constants from (Davis et al., 2004). These constants were non-adjustable parameters, and for the optimization of the U(VI) surface complexation model, we used the total site concentration and exchange capacity of the sediment. In the surface complexation modeling, all surfaces (clay minerals, Fe hydroxides, etc.) were combined as a generic >SOH surface to calculate the fraction of total uranium content accounted for by >SOUO2OH and >SOHUO2CO3 using adsorption constants. The presence of strongly complexing inorganic ligands effectively competes with surface complexation. Carbonate, especially, can significantly suppress U(VI) uptake by solution complexation (UO2(CO3)34-) or competitive adsorption (>SHCO3), but can also favor formation of the ternary surface complex >SOHUO2CO3 (Geckeis et al., 2014). Therefore, the description of uranium sorption onto natural sediments in the presence of carbonate is only possible by postulating the

formation of various ternary uranium carbonate surface complexes. Different stoichiometries have been proposed in the literature for these species, but we consider only >SOHUO2CO3 (Table 5). While sorption by inner-sphere complexation is considered to occur on amphoteric surface hydroxyl groups and thus varies with pH, ion exchange to bound actinide species at exchange sites of clay minerals is pH-independent (Bradbury and Baeyens, 2005; Gaskova and Bukaty, 2008): >SOH + UO22+ + H2CO3 = >SOHUO2CO3 + 2H+

(4)

2XNa + UO22+ = X2UO2 + 2Na+

(5)

3.2. Modeling results 3.2.1. Thermodynamic simulation of the dynamic experiments During the “HCh” simulations, we assumed average sludge composition (Table 1) and assumed groundwater (Table 2) equilibrium with atmospheric CO2. The initial uranium solid species in this system were CaUO4 (0.02 g) and a small quantity of other soluble minerals from Table 5 (Na4UO2(CO3)3, UO2CO3, etc.). Calculations were based on 20 g of solid interacting with the rising quantity of groundwater up to a W/R ratio of 250 (as in the dynamic experiments). “Long-term” indicates extrapolation to W/R 5000, i.e., a situation predicted to occur after a long period of time. For a better graphical presentation, the scale in Fig. 3 was adjusted. For example, gypsum, fluorite and calcite are on the scale of one to ten, ettringite is on the scale of five to one, and CaUO4 is on the scale of fifty to one. Fig. 3a shows the results of simulated U-containing sludge dissolution by groundwater at different calculation steps. Our simulations indicate that if the effect of wastes is predominant (Initial and 5th steps), the ettringite, gypsum, calcite, fluorite and chlorite assemblage is stable. When the effect of water is predominant, the mineral assemblages include chlorite, calcite, fluorite and the same zeolite-like phase, NaCa2Al5Si5O20⋅6H2O. If surface- or groundwater interacts with the wastes for a long time (as 5000 solution portions in our model), a depleted chlorite-gibbsite-goethite material develops in place of the initial alkaline association. The

concentrations of the dissolved major constituents are summarized in Fig. 3b. The concentrations of SO42- and Ca2+ are high at the beginning of the calculation (approximately 5000 and 650

Fig. 3. Modeling data of U-containing sludge dissolution showing both (a) shear and composition of the solid phases and (b) solution composition versus W/R ratio. “Initial” means 20 g of solid + 10 g of solution (W/R = 0.5); “Long-term” means extrapolation to W/R = 5000, i.e., the predicted situation.

mg/L) and decrease to 123 and 74 mg/L, respectively, at a W/R ratio of 250. The concentrations of these elements are influenced by fast sulfate mineral dissolution at low W/R ratios; over a long period of time, the concentrations will be controlled by calcite and fluorite stabilities. We assume that the small difference in the decrease of the SO42- and Ca2+ concentrations, depending on the W/R ratio between the model calculations and experiments, is explained by the kinetics of gypsum dissolved. This implies that transport, rather than the rate of equilibration between solid and solution, is the rate-limiting process in the overall reaction. From the beginning of the flow-through reactor simulations (the first two W/R ratios, Fig. 3b), up to 10 μg/L of uranium was released into the solution due to dissolution of UO2CO3(s) and Na4UO2(CO3)3 phases, predominantly. Thereafter, only CaUO4 was stable in the system. Up to 300 μg/L of uranium may be released at pH 8.4 (W/R = 250) by complexation with HCO3- in groundwater that was in equilibrium with the atmospheric CO2(gas). In such a solution, the concentration of calcium is sufficient for Ca2UO2(CO3)03 to predominate and allow mobility of uranium. Kelly et al. (2003) used XAFS analysis to identify the stable position of uranyl in

uranyl-rich natural calcite. If the position of uranyl in natural calcite is generally as described above, calcite may provide a stable host for dispersed UO22+ over geological time scales. Our calculations did not confirm the possibility of the CaCO3-UO2CO3 ideal solid-solution precipitation, although it is possible to model solid solution formation with the “HCh” algorithm. Based on thermodynamics, we can explain the high uranium content in the AECC leachates as being due to the dissolution of uranium minerals in Ca-bicarbonate solutions. Low and nearconstant U concentrations in the ECP leachates (6.5 – 1.2 µg/L, Fig. 1b) were not modeled with sufficient reliability. Further studies of uranyl-bearing natural calcite models are needed to test the general applicability of this idea. Recall that ECP sludge material contained low concentrations of uranium (0.01 %). 3.2.2. Thermodynamic simulation of the sorption experiments Preliminary calculations showed that the waste water (Table 2) was supersaturated with respect to calcite, purolysite (MnO2) and goethite. However, as mentioned above, neither precipitation nor dissolution of minerals were taken into account in the uranium sorption model calculations. According to the adjustable parameters (Table 5), the effect of Na removal (up to 50 % removal) from the solution is most evident in the cases of C-0 and C-2-9 interactions with groundwater (initial concentration of 12 mg/L), but is negligible in the cases of their interactions with waste water (initial concentration of 1400 mg/L). Figure 4a shows the thermodynamic simulation of U(VI) sorption from contaminated low mineral content groundwater onto the C-0 sample versus the W/R ratio as well as the distribution of adsorbed species of uranyl (solid and empty squares). The main uranium species in solution are Ca2UO2(CO3)30 and UO2(CO3)34-, indicated as Usol. The results demonstrate that the presence of Ca2+ in solution (73 mg/L) can decrease U sorption under conditions in which the Ca-U-CO3 species dominates U(VI) aqueous speciation. Such conditions are prevalent in many uraniumcontaminated aquifers. As a whole, the results of these calculations predict that uranium

Fig. 4. (a) Total adsorbed (solid line) and total dissolved (dashed line) uranium versus the W/R ratio, which simulates the reaction time. The pH evolution corresponds to the calcite dissolution/precipitation; (b) Evolution of dissolved uranium and pH during the sorption experiments, where subindex “exp,sol” means experimental and “calc,sol” means calculated uranium.

is strongly adsorbed as a variety of U species (>SOUO2OH and >SOHUO2CO3) at low W/R ratios and neutral pH. The adsorption was insignificant at the end of the experiment at рН 8.4 (Fig. 2a). The results show that the model is capable of accurately estimating the main Utrapping trends observed in experiments by using thermodynamic data from Table 5. Figure 4b compares the experimental and calculated pH (pHexp and pHcalc) and U concentrations (Uexp,sol and Ucalc,sol) during sorption from the contaminated, highly mineralized waste solution onto the C-2-9 sample. Note that, unlike Fig. 4a, only uranium in solution is shown in this figure. The thermodynamic calculation could not accurately reproduce the Utrapping trends of the waste solution with high pH, high ionic strength and high mineralization: there was a difference between the calculated and experimental uranium concentrations (Fig. 4b). As indicated above, we observed uranium concentrations to decrease by as much as 5.6 µg/L in the beginning of the experiments (Fig. 2d, W/R = 24) and to decrease slightly at the end of the experiment (to 21 µg/L, W/R = 423). In the calculations, U was adsorbed insignificantly (~5%), with the pH decreasing to 8.4 at the end point. There are several possible explanations for the mismatch between model and experimental data. As the waste solution is supersaturated with respect to calcite, magnesite, purolysite and

goethite, 300 mg of these minerals could precipitate from 1 L of an alkaline solution. These minerals can capture uranium, and co-precipitation could be the major mechanism for uranium removal because sorption does not play a role at pH 9.4 where strong carbonate U complexes in solutions predominate. The prediction of trace element partitioning in growing minerals is a key issue in water-rock interaction processes related to environmental issues, such as the disposal of radioactive waste (Thien et al., 2014). Although the pH varies from 9.4 to 8.4, the availability of sorption sites decreases over time due to the sites filling and to mineral coatings that armor some clay phases from contact with infiltrating water. Much of the uncertainty in the model comes from a lack of relevant data to characterize the nature and concentration of sorption sites over time. Nevertheless, there was an unexpected increase in U(VI) sorption at higher W/R ratios in the experiment (Fig. 2d), and the calculations did not reproduce this phenomenon. This could be described only by the variation of the filtration coefficient. Over the course of the experiment, the filtration coefficient decreased from 3.6 to 2.1 cm per day due to a reduction of pore space. The slowing or interruption of filtration improved the sorption properties of the clay rocks. For comparison, in the case of C-0 sediment (Fig. 2a), the filtration coefficient was nearly constant. Thus, the electrolyte concentration influences the coefficient of permeability through its influence on dispersion (Mesri, Olson, 1971). During the experiment with C-0 samples (Fig 2a), the W/R ratio was equal to 6 for 24 hours, whereas in the case of C-2-9 samples, the W/R ratio was 24 (Fig. 2d). To better understand the processes governing the release and trapping of the elements in the column 2d, a comparison is presented in Fig. 5. Variation in electroconductivity is determined mainly by the decrease in sodium and calcium during the experiments, and it corresponds to the variation of the filtration coefficient. The U concentration and pH changed in a similar way. In general, the modeling simulation of uranium sorption in dynamic experiments was successful only in the case of a constant filtration coefficient in fine-grained clay rocks. Moreover, there is a high probability that under field conditions, drainage solutions will find

alternate flow paths around the geochemical clay barrier despite its impermeability and durability. 90 80 70 60 50 40 30 20 10 0

EC/10

Parameter

p

Kfiltr*10

Uso 0

24 144 168 216 288 360 504 648Hours

0

24

9.6 9.4 9.2 9 8.8 8.6 8.4 8.2 8

96 111 147 191 237 337 423 W / R

Fig. 5. Measured electroconductivity (μSm/cm), pH, U concentration (μg/L) and filtration coefficient (cm a day) as a function of the W/R ratio. The dotted lines correspond to the initial data of the 2d experiment.

4. Summary and conclusions

We examined subsurface drainage water plumes at and around two Siberian U-enrichment plant waste management facilities located in Angarsk, Irkutsk region (AEСC) and Zelenogorsk, Krasnoyarsk Krai (ECP). The results of the groundwater chemical analyses at both sites showed that uranium was below the MAC for drinking water in Russia and the U.S., and that only alkaline nitrate solutions rich in Na, Ca and Mg were penetrating and spreading into the aquifers and bogs. Nevertheless, anthropogenic and climatic changes can disrupt the local water balance. Undesirable consequences could arise in the case of an inflow of oxidizing bicarbonate waters into the sludge storage sites, which would promote Ca-U-CO3 species formation in alkaline solutions. Uranium mobility could be affected in the case of income of NO3- solutions or an increase in Eh. Experimental studies showed that the sludge materials of the two sites have characteristics in common (they formed during neutralization of nitric acid slurry by hydrated lime) but also different solid phase compositions. AEСC was found to contain predominantly fluorite with

minor gypsum and calcite components, a total uranium content of 0.2%, and a mineral similar to CaUO4 based on electron microscopy. The release of U from AECC sludge may result from the solubility of U minerals and amounts to 258 μg/L at solution equilibrium with atmospheric CO2. ECP sludge dominated by soluble gypsum contains only 0.01% of total uranium, mainly in the carbonate fraction, based on sequential extraction procedure data. The release of U from ECP sludge may result from the solubility of carbonate minerals with U impurities; the maximum U concentration was 6.5 μg/L. Thermodynamic simulations indicated that the model based on the flow-through reactor technique sufficiently described the experimental results for the solubility of high U-concentration slugs (0.2%): fast SO42- is released due to the gypsum and ettringite solubility and the subsequent solution equilibration with calcite and fluorite, accompanied by a gradual increase in the concentration of uranium in solution. The experimental and thermodynamic results indicated that the U(VI) release from the sludge containing U minerals (AECC) contrasted with the release from the fine-grained U(VI)-carbonate solid solutions (ECP). This indicates that U leaching from sludge materials is not dependent on the specific solubility of the sludge sediments, but rather on the solid uranium species in the sludge. We investigated the relationship between uranium (VI) sorption and the W/R ratio. Sorption plays a dominant role in determining the fate of uranium: below pH 7, sorption is generally on clay minerals, whereas at a higher pH, sorption occurs on iron and aluminum hydroxides. When U-doped regional groundwater (initial pH 7.4) reacted with fine-grained C-0 sediment as a function of the W/R ratio, uranium was retained at a level of 1 μg/L until W/R = 71, after which U increased up to its initial content. The extent of uranium sorption and Na+ exchange decreased with time, as the sediment surface sites equilibrated with the influent CO2-containing groundwater. The model simulations show sufficiently good approximation of experimental results. Sorption of U(VI) from the alkaline waste solution (pH 9.4) is potentially more complicated because of its high mineralization and supersaturation with respect to a number of minerals and

the undesirable change of the filtration coefficient. The model results demonstrate the inability of the “HCh” code to accurately simulate the water budget, affecting the capture of microelements by sediments. No significant retention of uranium was observed in any of the flow-through experiments. To more accurately describe the potential release of U from the residual waste sites, the experiments have to be conducted in an isothermal stirred batch reactor. We conclude that the release of contaminants to aquifers is limited by the rates of adsorption and desorption at the scale of individual sediment grains. If seasonal or catastrophic inundation or submersion were to occur, an environmental hazard will be posed by uranium mobilization. References Boguslavskiy, A.E., Gaskova, O.L., Shemelina O.V. 2012. Uranium Migration in the Ground Water of the Region of Sludge Dumps of the Angarsk Electrolysis Chemical Combine. Chemistry for Sustainable Development. 20, 465-478. Bradbury, M.H., Baeyens, B., 2005. Modelling the sorption of Mn(II), Co(II), Ni(II), Zn(II), Cd(II), Eu(III), Am(III), Sn(IV), Th(IV), Np(V) and U(VI) on montmorillonite: linear free energy relationships and estimates of surface binding constants for some selected heavy metals and actinides. Geochim. Cosmochim. Acta 69, 875–892. Chen, F., Ewing, R.C., Clark, S.B. 1999. The Gibbs free energies and enthalpies of formation of U6+ phases: An empirical method of prediction. Amer. Mineralogist. 84, 650-664. Christov, Ch. Thermodynamics of formation of ammonium, sodium and potassium alums and chromium alums. 2002. Calphad. 26 (1), 85-94. Davis, J.A., Meece, D.E., Kohler, M., Curtis, G.P., 2004. Approaches to surface complexation modeling of Uranium(VI) adsorption on aquifer sediments. Geochim. Cosmochim. Acta 68, 3621-3641. Fox, P.M., Davis, J.A., Zachara J.M., 2006. The effect of calcium on aqueous uranium(VI) speciation and adsorption to ferrihydrite and quartz. Geochim. Cosmochim. Acta 79, 13791387. Gaona, X., Kulik, D.A., Mace, N., Wieland, E., 2012. Aqueous–solid solution thermodynamic model of U(VI) uptake in C-S-H phases. Appl. Geochem. 27, 81-95. Gas'kova, O.L., Boguslavskii, A.E.; Sirotenko, T.G., 2011. Geochemical composition of natural waters near a storage site of low activity radioactive wastes. Water Resources 38 (5), 597607. Gaskova, O.L., Boguslavsky, A.E., 2013. Groundwater geochemistry near the storage sites of low-level radioactive waste: Implications for uranium migration. In: Hellmann, R; Pitsch, H. (Eds.), Proceedings of the fourteen International Symposium on Water-Rock Interaction, WRI 14 Book Series: Procedia Earth and Planetary Science. 7, 288-291.

Gaskova, O.L., Bukaty, M.B., 2008. Sorption of different cations onto clay minerals: Modelling approach with ion exchange and surface complexation. Physics and Chemistry of the Earth. 33 (14-16), 1050-1055. Geckeis, H., Lützenkirchen, J., Polly, R., Rabung, T., Schmidt T., 2013. Mineral-Water Interface Reactions of Actinides. Chem. Rev. 113, 1016-1062. Hampson, C.J., Bailey J.E. 1982. On the structure of some precipitated calcium alumino-sulphate hydrates. J. Mat. Sci. 17, 3341-3346. Hummel, W., Berner, U., Curti, E., Pearson, F.J., Thoenen, T. 2002. NAGRA/PSI Chemical Thermodynamic Data Base 01/01, Technical report 02-16, 589 p. Kelly, S.D., Newville, M.G, Cheng, L., Kemner, K.M., Sutton, S.R., Fenter, P., N . Sturchio N.C., Spotl C. 2003. Uranyl Incorporation in Natural Calcite Environ. Sci. Technol. 37, 1284-1287. Mesri, G., Olson, R.E. Mechanisms controlling the permeability of clays. 1971. Clays and Clay minerals. 19, 51-158. Onac, B.P., Effenberger, H.S., Wynn, J.G., Povara, I. 2013. Rapidcreekite in the sulfuric acid weathering environment of Diana Cave, Romania. Am. Mineral. 98, 1302–1309. Reeder, R.J., Nuget, M., Tait, C.D., Morris D.E., Heald, S.M., Beck, K.M., Hess, W.P., Lanzirotti A., 2001. Coprecipitation of Uranium(VI) with Calcite: XAFS, micro-XAS, and luminescence characterization. Geochim. Cosmochim. Acta. 65, 3491-3503. Shvarov, Yu.V. 2008. HCh: New potentialities for the thermodynamic simulation of geochemical systems offered by Windows. Geochem. International. 46, 834-839. Tessier, A., Rapin F., Carignan, R. 1985. Trace metals in oxic lake sediments: possible adsorption onto iron oxyhydroxides. Geochim. Cosmochim. Acta. 49:183-194. Thien, B.M.J., Kulik, D.A., Curti, E. 2014. A unified approach to model uptake kinetics of trace elements in complex aqueous - solid solution systems. Appl. Geochem. 41, 135-150. Tits, J., Geipel, G., Mace, N., Eilzer, M., Wieland, E. 2011. Determination of uranium(VI) sorbed species in calcium silicate hydrate phases: A laser-induced luminescence spectroscopy and batch sorption study J. Coll. Interface Sci. 359, 248–256. World Health Organisation (2005) Uranium in Drinking Water - Background document for development of WHO Guidelines for Drinking-water Quality, World Health Organisation, WHO/SDE/WSH/03.04/118, 26 p.

Highlights Leaching tests of the waste materials of two U-enrichment plants were conducted. The thermodynamic model using the HCh code was able to explain U release in both cases. No significant sorptive uptake of uranium by host clay sediments was observed. Thermodynamic modeling of U sorption was successful only in the case of groundwater.