Radionuclide migration at experimental polygon at Red Forest waste site in Chernobyl zone. Part 2: Hydrogeological characterization and groundwater transport modeling

Applied Geochemistry 27 (2012) 1359–1374

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Radionuclide migration at experimental polygon at Red Forest waste site in Chernobyl zone. Part 2: Hydrogeological characterization and groundwater transport modeling D. Bugai a,⇑, A. Skalskyy a, S. Dzhepo a, Yu. Kubko a, V. Kashparov b, N. Van Meir c, D. Stammose c, C. Simonucci c, A. Martin-Garin d a

Institute of Geological Sciences, 55-b, Gonchara Str., Kiev 01054, Ukraine Ukrainian Institute of Agricultural Radiology UIAR NUBiP of Ukraine, Mashinobudivnykiv Str. 7, Chabany, Kyiv-Svjatoshin Distr., Kiev Reg. 08162, Ukraine c Institute for Radioprotection and Nuclear Safety, IRSN/DEI/SARG/LR2S, BP 17, 92262 Fontenay aux Roses Cedex, France d Institute for Radioprotection and Nuclear Safety, IRSN/DEI/SECRE/LRE, BP 3, 13115 Saint-Paul-Lez-Durance Cedex, France b

a r t i c l e

i n f o

Article history: Available online 8 October 2011

a b s t r a c t This article represents the second of two articles, which review the main results of the international radioecological projects: Chernobyl Pilot Site Project (1999–2003) and Experimental Platform in Chernobyl (2004–2008). These projects studied radionuclide migration from the near-surface radioactive waste trench at the Red Forest waste dump in the Chernobyl zone, which contained nuclear fuel particles. This article presents results from the comprehensive hydrogeological site characterization program including the following issues: geological structure of the study site, hydraulic properties of the deposits, tracer tests in the aquifer, results of groundwater monitoring and unsaturated zone regime studies, as well as data on the 90Sr distribution in the unsaturated zone and aquifer, and analyses of 90Sr sorption behavior. The derived parameters were used to develop and calibrate 1D (flow tube) and 2D (cross-section) models describing the migration of 90Sr from the studied waste trench to the unsaturated zone and aquifer over a 16-a period (1986–2002). The models involved the following sub-models: (1) the geostatistical (structural) model for radioactivity distribution in the trench (using GSLIB); and (2) the radionuclide source term model (STERM1D) describing dissolution of fuel particles and a 1D of radionuclide redistribution in the trench body and unsaturated zone. The MODFLOW – MT3D codes were used to model the 2D 90Sr transport in the aquifer cross-section. Calibration of the 1D model with respect to Kds and dispersivities allowed quite accurate reproduction of 90Sr migration behavior for the early period (1995–1998). The less perfect fit between the 1D and 2D modeling results and monitoring data for the later period (1999–2002) suggests the need to improve the conceptual radionuclide migration model (i.e. to account for transient hydraulic and geochemical regimes of the waste site). Ó 2011 Elsevier Ltd. All rights reserved.

1. Introduction This article represents the second of two articles, which review the main results of the international radioecological research projects: Chernobyl Pilot Site Project (CPS project, 1999–2003) and Experimental Platform in Chernobyl (EPIC, 2004–2008). These projects were aimed at studying radionuclide migration from the near-surface radioactive waste trench, which contained Chernobyl nuclear fuel particles, as well as the development and experimental confirmation of the relevant radionuclide transport models (Dewière, 2000; Bugai and Dewière, 2004a, b). The studies were carried out at the experimental site (Chernobyl Pilot Site) established near waste trench no. 22T in the Red Forest waste dump ⇑ Corresponding author. Tel./fax: +380 44 486 30 23. E-mail address: [email protected] (D. Bugai). 0883-2927/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.apgeochem.2011.09.028

at the Chernobyl nuclear power plant site. The general description of the project objectives, of involved research teams and of the experimental site is given in the first part of this special issue (Kashparov et al., this issue). The research program included two interrelated directions of work, which have acquired within the project the names ‘‘Source Term’’ and ‘‘Aquifer’’. The questions addressed in the framework of the first block (‘‘Source Term’’) included the radionuclide inventory and the spatial distribution of radioactivity in the trench, as well as physical and chemical speciation and dissolution kinetics of nuclear fuel particles. Results of these studies are presented in the first part of this special issue (Kashparov et al., this issue). The second research direction ‘‘Aquifer’’ (or ‘‘hydrogeology block’’) included studies on the following main subjects (Dewière, 2000; Bugai et al., 2003; Ardois and Szenknect, 2004; Bugai and Dewière, 2004a,b): geological structure of the experimental site;

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Detailed data on the geological structure of the upper part of the geological cross-section of the experimental site were obtained by core material analysis of 8 m deep boreholes drilled at the experimental site. Grain-size distribution analysis of core material with depth allowed identification of different genetic types of deposits (Fig. 3a and b) (Dewière, 2000; Matoshko et al., 2004). In all, four different types of deposit were distinguished (see Fig. 3): (1) ‘‘technogenic’’ deposits (contemporary forest podzols or soil displaced during clean-up operations; from the ground surface down to a depth of 0.5–0.8 m); (2) eolian sandy deposits (well sorted finemedium grained sands with a content of clay <1%; down to a depth of 4–5 m); (3) sandy alluvial deposits of floodplain facies (content of clay up to 8%, depth of layer bottom 7–7.5 m); (4) loamy-clayey alluvial deposits of dead-channel facies appear in the bottom part of cross-section (down to the depth of 8 m). Based on drilling data from a deep observation well for the regional monitoring network (well no. 1–98, see Fig.2) the lower part of the geological cross-section (from 8 to 10 m down to the Kiev marls at 30 m) is composed of alluvial deposits of channel facies, which are represented by well-sorted medium-grained sands with a low clay content. Besides lithological studies of deposits from the experimental site, detailed studies of the geological analog of the Pilot Site, the ‘‘Pripyat Zaton’’ exposure were carried out (Matoshko et al., 2004). The Pripyat Zaton exposure is situated 2 km to NE from the study site (see Fig. 1). It represents the natural erosion ledge of the first terrace, which extends some 200 m along the Pripyat Zaton Inlet of the Pripyat River. Here the upper part of the crosssection of geological deposits is situated in unsaturated conditions.

regime and parameters of groundwater (soil moisture) flow in the system ‘‘atmosphere – unsaturated zone – aquifer’’; spatial distribution of 90Sr in the hydrogeological system of the waste site; estimation of parameters of 90Sr sorption by geological deposits; and, eventually, development and calibration of the models for groundwater flow and 90Sr subsurface migration from the waste trench to the surrounding hydrogeological environment. Below the main results of the hydrogeological characterization and monitoring studies listed above are reviewed, and the methodology and results of modeling 90Sr migration from the waste site to the unsaturated zone and aquifer are described. 2. Geological structure of the experimental site 2.1. Geological structure of the site and lithological properties of geological deposits With regard to the geomorphology, the experimental site is situated in the central part of the first terrace of the Pripyat River (Fig. 1). Its elevation is 112–115 m a.s.l. The upper part of the geological cross-section is composed of sandy Upper Pleistocene and Holocene deposits, which are subdivided into alluvial and eolian suites of deposits, which are overlain in many places by ‘‘technogenic’’ (or ‘‘man-made’’) deposits (such as soils and construction debris replaced during the clean-up operations). The total thickness of this upper part reaches 30 m, underneath, lies the regional aquitard layer composed of marls (carbonate clays) of the Kiev suite of the Eocene (Fig. 2).

N

105 106

106

10

8

Prip

yat

Pripyat Inlet

10

7

Riv

er

Infiltration Fields

109

Pripyat Zaton 110

Cooling pond

111

ChNPP 112 1/95-1

113

1/1

Pilot Site 31

114

groundwater flow direction

Groundwater table conturs (July 1999)

Ro

dv

ino

St

re

Monitoring wells

0

0.5 Kilometers

1

Bor

shy

am

Stre

am

Fig. 1. General hydrogeological scheme of the Chernobyl NPP site (groundwater table contours shown as in July 1999).

D. Bugai et al. / Applied Geochemistry 27 (2012) 1359–1374

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Fig. 2. Regional geological cross-section of the Chernobyl NPP site (from SW to NE). Legend: 1 – sands, 2 – silts, 3 – basal till, 4 – clay, 5 – marl, 6 – inter-bedded sands and silts, 7 – peat and peaty sand, 8 – boreholes (numbered), 9 – inferred fault, 10 – boundaries between suites: supposed (upper) and established (lower), 11 – boundaries between depositional facies, 12 – facial replacement, 13 – groundwater level (generalized) (Matoshko et al., 2004).

This simplifies sample collection and allows visually investigating the sedimentary structures of deposits. In particular, detailed sample collection on a two dimensional grid with a spatial resolution of 10 cm was carried out at this exposure, which allowed developing the geostatistical structural model of the geology (Dewière et al., 2004). The geostatistical characteristics obtained were used to develop a priori estimates for hydrodispersion parameters of dissolved species in the geological environment (for the purpose of interpreting the aquifer tracer test data; see Section 4.2). The porosity of the sandy deposits of the unsaturated zone and the aquifer at the experimental site constitutes 33–42%, while bulk density is 1.6–1.72 g/cm3. The minimal values of porosity are inherent for eolian sands, while the maximum ones – for alluvial deposits of dead-channel facies. The mineralogy of deposits is mainly quartz with admixtures of feldspars (5–9%) and accessory minerals (<0.5%). Cation-exchange capacity (CEC) of eolian sands is about 1 meq/100 g. The CEC values of alluvial deposits reach 5–10 me 9/100 g. A good correlation is observed between the CEC value and clay content. Different lithological properties of eolian and alluvial deposits (in particular, the higher content of clay fractions in the alluvium) cause different hydraulic properties (see next section) and sorption properties of these deposits (see Section 5.2).

2.2. Hydraulic properties of geological deposits The methods used to characterize hydraulic properties of the geological deposits of the study site included laboratory column tests, field hydraulic tests (slug-tests and pump tests) and mathematical modeling (calibration of groundwater flow model) (Dewière, 2000; Bugai et al., 2001). Results of the hydraulic tests are summarized in Table 1. It should be noted, that alluvial deposits of floodplain facies are essentially anisotropic with respect to hydraulic conductivity: jX  jZ (where index X denotes horizontal, Z – vertical component, see Table 1). This is caused by the layered structure of these deposits: the presence of higher permeability horizontal sub layers in the generally relatively low

permeability deposit strata. The estimates of hydraulic parameters of eolian sands, i.e. the upper portion of the unconfined aquifer, were cross-validated in the course of the natural gradient tracer tests in the aquifer (see Section 4.2). 3. Field setup for monitoring of hydrogeological process The hydrogeological monitoring system established at the Chernobyl Pilot Site in 2000–2001 consisted of the following main components: the monitoring well network, soil water samplers (SWS), automated station for unsaturated zone monitoring (pit), and the automated weather station (Dewière, 2000; Bugai and Dewière, 2004a). These components are described below in more detail. 3.1. Monitoring well network The monitoring well network at the CPS is composed of several sub-systems. The groundwater level regime was monitored by means of an observation well network encompassing the study area (see Fig. 2 in Kashparov et al., this issue). The 5–6 m deep observation wells were constructed from 5 cm diameter PVC tubes, and were equipped with a 1 m screen. Observation wells were equipped with TD-Diver automated water level loggers (Van Essen Instruments). Observations on these wells allow determining groundwater flow direction and the horizontal gradient of hydraulic head. Spatial distribution of radionuclides and chemical constituents in the aquifer cross-section was studied using several multilevel well profiles, oriented along the groundwater flow direction. In this publication monitoring data collected using the so-called ‘‘Laboratory’’ (LAB) well profile established in the vicinity of the laboratory module at the experimental site are presented. The profile is composed of 6 clusters of four 2.5 cm diameter PVC wells of different depths equipped with 20 cm long screens in the depth range from 2 to 8 m below ground surface (Fig. 4a). To avoid cross-contamination, the individual screens within the borehole were isolated from each other by 20 cm thick bentonite sealing layers. The well

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3.3. Pit for monitoring of moisture flow in the unsaturated zone Water flow in the unsaturated zone was studied using a specially designed automated unsaturated zone monitoring station (pit) put into operation in June 2000. The design of the pit is shown in Fig. 4b. One wall of the pit partially cuts the body of the trench with radioactive waste, while the opposite wall provides access to the soil profile outside of the trench. Such an arrangement of the pit allowed monitoring of processes of moisture flow both in the body of the trench and outside the trench in the ‘‘background’’ undisturbed soil profile. The pit was equipped with arrays of electronic sensors for measuring soil water content, suction pressure and temperature, and with a data logger to store sensor readings (Fig. 4b) The sensors were inserted in the soil profile from inside the pit to a distance of 1 m (horizontally) from the pit walls. The following sensors manufactured by Delta-T Devices Ltd. (UK) were used: Theta Probe ML2x TDR (Time Domain Reflectometry) type sensors for measurement of volumetric moisture content in soil; Equitensiometer EQ2 electric impedance sensors for measurement of suction pressure (matric potential) in the range from 0 to 1000 kPa (‘‘dry’’ soil conditions); SWT6 analog pressure transducer sensors for measurement of suction pressure in the range from 0 to 100 kPa (‘‘wet’’ soil conditions); ST1 thermistor sensors for measurement of soil temperature. The Delta-T DL2e logger was used for automated recording of sensor readings. The monitoring data collected using the pit station are discussed in the Section 4.3. 3.4. Automated weather station

Fig. 3. Example profile grain-size distribution diagram (a), and geologic sections along the borehole profile (b), based on analysis of core material from bore holes dilled at 1999 tracer test site (notation for deposits: mc–ms – man made/modern soil; eol – eoloian, a – alluvial; ob – overbank) (Dewière, 2000).

clusters are spaced horizontally 4 m from each other. Monitoring data from this profile were used to calibrate the radionuclide transport model, which is described in Section 6. Special-purpose observation well networks were installed at the experimental site to carry out tracer tests in the aquifer (see Section 4.2 for more detail). 3.2. Soil water samplers Soil pore solutions were sampled using clusters of ceramic vacuum soil water samplers (SWS) (1900L model series of Soilmoisture Equipment Corp.) installed at various depths in the soil profile (in the range of 20 cm to 2.5 m). Samplers were installed in several locations of the site both within the trench, and outside the trench area. Sampling was carried out at about 1–2 month intervals (excluding periods with air temperature below freezing point).

The meteorology observations were carried out using automated Delta-T weather station (Delta-T Devices Ltd.). The weather station is situated at the 6 m  6 m plot in the central part of the experimental site. The list of controlled parameters included: air temperature and relative humidity, wind direction and speed, amount of rainfall, solar radiation, and soil temperature. The meteorological parameters collected allowed estimation of the amount of rainfall and potential evapo-transpiration at the ground surface, which are important parameters for calculating the water balance of the site and for specifying boundary conditions for the unsaturated zone flow and transport models. 4. Hydrogeological conditions of the experimental site 4.1. General features of groundwater regime The hydrogeology studies were focused on the unconfined aquifer in Quaternary eolian and alluvial sandy deposits (see Section 2.1) with a total thickness of 25–30 m. The Quaternary deposits are underlain by low permeability (hydraulic conductivity 0.001 m/day) Eocene marls, which represent the regional aquitard layer (see Fig. 2). The experimental site is situated in the area of transit of the regional groundwater flow system from the region of the elevated Chistogalovka moraine hills (water divide of the Uzh River and Pripyat River) towards the main discharge contour, the Pripyat

Table 1 Representative values of physical, lithological, hydraulic and geochemical properties for the main types of geological deposits of the experimental site.

a

Type of deposits

Bulk density, kg/dm3

Porosity, %

Clay fraction content (<0.01 mm), %

Hydraulic conductivity, m/day

CEC, meq/100 g

Eolian Alluvial, flood plain facies Alluvial, channel facies

1.72 1.73 1.67

34–36 35 37

1–2 10–20 2–8

3–5 1(jX), 0.01(jZ)a 5–15

0.5–1.2 5–10 0.5–4

Anisotropic parameter; subscript X denotes horizontal direction, Z – vertical direction.

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(a) Cluster of multilevel monitoring wells.

(b) Pit for unsaturated zone monitoring (vertical cross-section) Fig. 4. Field facilities for hydrogeological monitoring.

River (Fig. 1). More details on the general hydrogeological settings of the Chernobyl NPP site can be found in (Dzhepo and Skalskii, 2002). Below data are given on the groundwater regime of the experimental site, obtained from the local observation well network (see Section 2.1). 4.1.1. Groundwater level regime The example groundwater level (GWL) hydrograph for piezometer no. 2-99 of the experimental site for the period of 2000–2004 along with atmospheric precipitation data is shown at Fig. 5. The

GWL regime is characterized by well-defined seasonal cycles of rising and lowering (recession). The main increase in GWL is usually observed at the end of winter – beginning of spring (February– April) due to snowmelt and spring rains. The recession of GWL is often observed during the period from the end of spring to the beginning of winter, which is related to decreased infiltration recharge due to the increase of moisture evapotranspiration in summer (as a consequence of higher air temperatures, seasonal development of vegetation, etc.). A good correlation is observed between the GWL and rainfall events. The increase of GWL starts almost immediately after large rainstorms, while the maximum

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Fig. 5. Groundwater level regime at Chernobyl Pilot Site in 2000–2004 (TD-Diver level logger data from the well 2–99).

levels of the groundwater table are reached in 5–10 days. Seasonal fluctuations of GWL are usually about 1 m. Based on observations in 2000–2004 the average distance between the bottom of the trench and the groundwater table was 50–70 cm, while during some episodes the groundwater table rose up to 30 cm above the trench bottom (Bugai and Dewière, 2004a). 4.1.2. Infiltration recharge to groundwater The available detailed GWL hydrograph data allowed estimating recharge to the groundwater due to infiltration of atmospheric precipitation using the Water Table Fluctuation (WTF) method (Helay and Cook, 2002). The important parameter in the WTF method is the specific yield coefficient of the deposits (l). This coefficient was estimated for the eolian sand from the laboratory column drainage experiments and pump tests at l = 0.14 ± 0.01. Results of the calculation of integral yearly recharge for the period of 2000–2003 are summarized in Table 2 (Bugai and Dewière, 2004a). The mean yearly infiltration recharge is estimated at about 200 ± 50 mm/a, or 40 ± 5% of the total amount of atmospheric precipitation. The main inflow of infiltration water to the aquifer usually occurs in spring following the snowmelt. Intensive rains in spring,

summer or autumn can also provide significant inflow of infiltration water to the aquifer. Therefore, infiltration recharge is related to specific meteorology events and is ‘‘focused’’ (rather than ‘‘distributed’’) in time. The values obtained for infiltration recharge (160–260 mm/a, see Table 2) generally agree with results of estimating this parameter using the Cl balance method, and with results of calibrating the regional groundwater flow model of the study area (Dewière, 2000; Bugai et al., 2001). 4.1.3. Areal and vertical distribution of groundwater head at the experimental site Based on groundwater monitoring, during the periods of stable recession (lowering) of the water table the groundwater flow in the unconfined aquifer is directed to the NNE, while the lateral gradient of hydraulic head constitutes 0.001–0.0015 m/m. During periods of intensive infiltration recharge the lateral gradient increases to 0.002–0.003 m/m, and groundwater flow direction deviates to the NNW. Fluctuations of groundwater flow direction constitute ±10–15°. Changes in groundwater flow direction are likely caused by the topography of the experimental site and by related changes in timing and values of infiltration recharge to the groundwater (depending on the thickness of the unsaturated zone).

Table 2 Estimates of infiltration recharge rate to groundwater at the Red Forest site for the period 2000–2003 obtained using the WTF method (Bugai and Dewière, 2004a).

a b

Year

Meteorological event

Months

Dhb, mm

Infiltration, mm

Precipitation, mm

Integral infiltrationa, mm

Infiltration as a part of total precipitation, %

2000 2000 2000 2001 2001 2001 2001 2001 2002 2002 2002 2003 2003

Snowmelt Summer rains Autumn rains Snowmelt Snowmelt Spring rains Summer rains Autumn rains Snowmelt Autumn rains Autumn rains Snowmelt Autumn rains Autumn rains

February July September January March May July November February June November March October November

670 675 380 430 620 370 120 235 535 100 587 705 650 240

87–101 88–101 49–57 56–65 81–93 48–56 16–18 31–35 70–80 13–15 76–88 91–106 84–98 31–36

564

224–259 (2000 year)

40–46

554

232–266 (2001 year)

42–48

487

159–183 (2002 year)

33–38

451

206–240 (2003 year)

46–53

Specific yield coefficient l is assumed to vary in the range 0.13–0.15. Dh – increase of GWT due to meteorological event.

D. Bugai et al. / Applied Geochemistry 27 (2012) 1359–1374 Table 3 Results of natural gradient tracer tests in the eolian sand aquifer at Chernobyl Pilot Site.

a

Parametera

Experiment of 1999

Experiment of 2001

Real groundwater flow velocity Vx, cm/day Vz:Vx Longitudal dispersivity aX, cm Transverse dispersivity aZ, cm

3

3.1

Not determined 9 9

1:3 2–3 2–7

X – horizontal direction, Z – vertical direction.

The piezometric observations have revealed a specific distribution of hydraulic head in the aquifer cross-section, which is related to the vertical heterogeneity of the lithological and hydraulic properties of aquifer deposits. In the higher permeability eolian sands located in the upper part of the geological section the vertical head gradient is fairly small (104 m/m). In the alluvial deposits of the floodplain – dead channel facies in the middle part of the crosssection (110–106 m a.s.l.) the vertical gradient of the hydraulic head constitutes 0.02–0.04, which exceeds the lateral (horizontal) head gradient by a factor 10–20. As a result, groundwater flow (and dissolved species transport) in the eolian sand layer is close to horizontal. In the sub-layer of low permeability alluvial deposits (in the depth interval 110–106 m a.s.l.) the groundwater flow direction is nearly vertical (Bugai et al., 2001; Bugai and Dewière, 2004a). 4.2. Tracer test in the aquifer

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groundwater flow velocity (see Table 3) is in good agreement with the results of 1999 experiment. In addition, it was estimated that the vertical component (Vz) of the real groundwater flow velocity is Vz  1/3  Vx. Estimates for dispersivities based on geostatistical analyses of the spatial distribution of hydraulic properties of the eolian sands ranging between 1 and 6 cm (Dewière et al., 2004), were in reasonable agreement with estimates of these parameters based on the observed dispersion of the tracer plume in the aquifer (see Table 3). It should be noted, that the presented values of dispersivity coefficients correspond to a conservative tracer (Cl). Larger dispersivity values are expected for 90Sr due to sorption–desorption processes of the radionuclide on geological deposits (Brusseau, 1994). 4.3. Moisture flow in the unsaturated zone In order to model radionuclide migration from the trench to the aquifer it is necessary to characterize the moisture flow in the unsaturated zone (from the ground surface to groundwater table). The program of unsaturated zone studies included monitoring of the unsaturated zone in the field using the pit (see Section 3.3), laboratory studies of soil characteristic functions, and interpretation of data using water balance methods and inverse modeling techniques (Bugai and Dewière, 2004a; Bugai et al., 2008; Van Meir et al., 2009). 4.3.1. Laboratory determinations of unsaturated soil characteristic functions The following parameters of the soils of the experimental site were studied in laboratory tests (Bugai et al., 2008):

Two natural gradient tracer tests in the aquifer (in the eolian sand layer) were carried out at the experimental site in 1999– 2001 in order to estimate real groundwater flow velocity, groundwater flow direction and hydrodispersion parameters of dissolved species in groundwater (Bugai et al., 2001, 2002; Dewière et al., 2004). The tracer tests used 36Cl, which is a conservative tracer (that is non-interacting with aquifer deposits). Special observation well networks were installed in order to carry out the tracer tests, which included injection wells and downstream rows of observation wells (perpendicular to the expected flow direction). Automated water samplers (ISCO, USA) were used to monitor the tracer content in groundwater. 4.2.1. Tracer test in 1999 The first tracer test was started in August 1999 (Dewière, 2000; Bugai et al., 2001). Distance between the injection well and the row of three downstream observation wells was 1 m. The duration of the experiment was 3 months. Results of the experiment are listed in Table 3. The established real groundwater flow velocity from the tracer test is 3 cm/day (10 m/a). This value is in agreement with the independent estimate of real flow velocity in the eolian sand layer of 2.4 cm/day calculated using the hydraulic conductivity value of 5 m/day (see Section 2.2), the mean value of the horizontal hydraulic head gradient of 0.0014 m/m (Bugai and Dewière, 2004a) and effective porosity of 0.3. 4.2.2. Tracer test in 2001 A more detailed observation well network was installed for this test, which included two rows of observation wells at a distance of 1 m and 2.5 m downstream from the injection well (Dewière et al., 2004). The experiment lasted 5 months. The aim was the estimation of hydrodispersion parameter values of Cl, and comparison of the derived dispersivity values with the a priori estimates based on the geostatistical theory of solute hydrodispersion in porous media. The estimate for the horizontal component (Vx) of the real

Fig. 6. Results of estimation of unsaturated hydraulic properties of eolian sand: (a) soil water retention function; (b) unsaturated hydraulic conductivity (Bugai et al., 2008).

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inter-dependence of soil suction pressure (W) and soil moisture content (h), and unsaturated hydraulic conductivity of soils as a function of soil moisture content K(h) (or soil suction pressure K(W)). In particular, the parameters mentioned (which are termed soil characteristic functions) are needed to parameterize numerical models of moisture flow in the unsaturated zone. The laboratory studies utilized pressure cells, column drainage tests and steady state flow column tests. Example characteristic soil functions for the eolian sand are shown in Fig. 6. 4.3.2. Field monitoring of the unsaturated zone regime and estimation of infiltration recharge to groundwater The unsaturated zone monitoring data collected using the PIT monitoring station (see Section 3.3) allowed characterization of general features of the moisture flow in the unsaturated zone of the experimental site and estimation of infiltration recharge to the aquifer (Dewière, 2000; Bugai and Dewière, 2004a). According to the moisture flow regime, the soil profile can be subdivided into three zones: from the soil surface to a depth of 1 m the moisture flow changes direction throughout the year (during moist periods the flow is downward, during dry periods – upward); the depth interval from 1 m to 2 m is characterized by stable downward moisture flow towards the water table during the whole year; lastly the depth range below 2 m is located within the range of fluctuations of the water table capillary fringe. To study the infiltration recharge regime at the Pilot Site the approach described in Freeze and Banner (1970) and Sitnikov (1978) was used, which is referred to as the ‘‘hydro-physical method’’ (after Sitnikov (1978)). The infiltration recharge rate (soil moisture flux) at a given depth was calculated using the Darcy law in unsaturated conditions (i.e., the Richards formula, Bear, 1972) using data on the soil matrix head gradient (provided by tensiometers) and the unsaturated hydraulic conductivity of soil (corresponding to the suction pressure value measured in soil profile), derived in laboratory tests, which were outlined in the beginning of this section (Bugai and Dewière, 2004a). Results of the calculation of the daily infiltration recharge rate through the eolian sand profile using the hydro-physical method for the hydrological year (11.2000–10.2001) is shown in Fig. 7. Calculations use data provided by the PIT unsaturated zone monitoring station for the depth interval 1.25–1.75 m. The maximum daily infiltration recharge rate during the considered period reached 5 mm/d. During the dry periods of the year the infiltration flux dropped to less that 0.1 mm/d. The cumulative annual infiltration

recharge to the aquifer constituted about 240 mm. The infiltration recharge from 1 January to 10 May is about 175 mm, or 75% of the integral annual value. The estimated real evapotranspiration for winter–spring months (1 January–10 May) is about 140 mm, and for the rest of the year it is 220 mm. The estimate obtained for cumulative infiltration recharge of 240 mm/a is in good agreement with previously established estimates using the WTF method (see Section 4.1). While the WTF method appeared to be effective in determining the cumulative infiltration values, the hydro-physical method presented above is unique in a sense that it provides detailed information on the daily distribution of the recharge through the year. 4.4. Geochemical monitoring of the waste site The hydrogeological characterization program of the Chernobyl Pilot Site included monitoring the chemical composition of groundwater (Bugai et al., this issue). Below the main features of the geochemical regime of the waste site are briefly reviewed, with the main focus on the data from the LAB multilevel well profile (as the data set from this profile is used to calibrate radionuclide transport models given in Section 6). With respect to geochemical conditions the aquifer can be subdivided into two zones: (1) an aquifer zone influenced by the trench, and (2) a zone representative of background geochemical conditions. In the first zone groundwater composition is influenced by trench pore solutions, which infiltrate the aquifer from the trench body. Data on the typical composition of groundwater in the zone of influence of the trench and in background conditions are given in Table 4. The zone of influence of the trench in the aquifer was characterized by elevated concentrations of Ca, Mg, SO4, NO3 and lower pH values. Analysis of long-term monitoring time series has shown that the geochemical impact of the trench was more pronounced in 1998–1999 compared to the conditions observed in 2004–2005 (see below). The observed geochemical patterns can be explained by a process of organic matter decomposition (vegetation, litter, organic soil layer, etc.), that was disposed of in trench no. 22T. Oxidation of organic matter inside the trench caused acidification of the trench pore solutions (in particular, due to generation and dissolution of CO2 in pore waters and due to the nitrification process). This resulted in leaching of cations adsorbed on the exchange complex of soils buried in the trench. At the same time, balance calculations using rainwater chemistry suggest that the geochemical

Fig. 7. Daily infiltration recharge rate through the unsaturated zone at Chernobyl Pilot Site (eolian sand profile) for the period 11.2000–10.2001 estimated using hydrophysical method (Bugai and Dewière, 2004a).

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D. Bugai et al. / Applied Geochemistry 27 (2012) 1359–1374 Table 4 Mean concentrations (and range of values) of chemical constituents in groundwater of Chernobyl Pilot Site (data from the LAB monitoring well profile; September 2004). Constituent mg/L Na+ K+ Ca2+ Mg2+ Cl SO2 4 HCO 3 NO 3 pH

Background, mg/L

Influenced by the trench

2.3(0.7–4.7) 1.5(1–2) 6.9(3.6–9.4) 1.8(0.5–3.6) 3.5(0.8–6.4) 22.5(15.2–41.9)

1.6(0.5–5.3) 1.7(1–3) 11.5(4–23.2) 2.0(0.6–3.9) 3.8(2.8–5.4) 27.4(8.2–56.8)

10.6(6.1–19.5) 1.1(0.1–2.6) 6.2(5.6–6.95)

24.2(6.1–62)) 2.3(0.5–6.8) 6.0(5.35–6.9)

composition of groundwater in ‘‘background’’ conditions is mainly due to evapotranspiration of atmospheric precipitation. The higher ionic strength of the groundwater in the zone of influence of the trench caused a decrease in 90Sr sorption by geological deposits (due to competition effects from major ions, especially Ca, for exchange sites on the soil matrix) and promoted radionuclide migration in the aquifer. Observation data for 1998– 2008 indicate a progressive decrease of concentrations of the leached chemical constituents in the aquifer in the zone of influence of the trench accompanied by an increase in pH. An example time series of concentrations of K, Ca, Mg, SO4 and pH in well 6-95-1 situated immediately downstream from the trench is shown at Fig. 8. The observed long-term trends can be explained by a gradual decrease of inventory, humification of the original OM inside the trench, by depletion in the inventory of cations in the exchange

complex of buried soil, as well as by an increased uptake of nutrient chemicals by vegetation on the trench (Bugai et al., this issue). 5. Strontium-90 migration in the unsaturated zone and aquifer 5.1. Two-dimensional distribution of aquifer

90

Sr in the unsaturated zone and

The typical distribution of 90Sr in the aquifer based on observations in the LAB multilevel well profile (in June 2002) is shown in Fig. 9. The Figure also presents data on radionuclide activity in trench pore solutions (obtained using ceramic soil water samplers, see Section 3.2). The 90Sr plume in the aquifer with concentrations of 1000–2000 Bq/L extends from the trench approximately 10 m downstream (see Fig. 9). Maximum 90Sr activity concentrations within the plume are similar to pore solution concentrations inside the trench. The background activity in the top part of the aquifer, in the order of n  100 Bq/L, is due to vertical migration into the unsaturated zone from the topsoil layer containing residual contamination. In accordance with groundwater flow patterns, the radionuclide plume is sinking slightly with distance from the trench, being displaced from the top by infiltration recharge water. The downstream edge of the plume reaches the interface zone between the eolian and alluvial layers. Here, dissolved 90Sr encounters alluvial deposits with a higher adsorption capacity compared to eolian deposits (see Section 2.2). Taking into account the results of the tracer experiments (see Section 4.2) the 90Sr migration velocity in the eolian layer was estimated in 2002 to be 9% of groundwater velocity (the respective retardation factor for 90Sr transport in the aquifer is R  12). 5.2. Parameters of strontium-90 sorption by geological deposits The key geochemical process, which controls 90Sr migration in groundwater, is radionuclide sorption by geological deposits. The parameter, which is most commonly used to characterize radionuclide sorption, is the distribution coefficient (Kd). The distribution coefficient represents the ratio of equilibrium concentration of radionuclide in the solid phase (sediment matrix) and in the liquid phase (groundwater). It is known (Pavlotskaya, 1974; Lefevre et al., 1996; Szenknect et al., 2005) that 90Sr sorption by geological deposits is mostly governed by an ion exchange process. It should be noted, that the Kd value is not a constant, but is dependent on the cation content of the solution, competing with 90Sr for exchange centers on the soil matrix, in particular Ca and stable Sr (see references listed above). The 90Sr Kds for the main types of geological deposits of the study site were characterized using laboratory batch tests and dynamic column tests, as well as special field experiments on determination of in situ partitioning of 90Sr between the matrix and porous solutions of the geological materials (Bugai et al., 2001, 2002). In addition to the purely empirical studies, a dedicated program on sorption studies (Ardois and Szenknect, 2004; Szenknect et al., 2005) was carried out with the aim of parameterizing the geochemical model of 90Sr sorption using the PHREEQC code (Parkhurst and Appelo, 1999), which accounted for the chemical composition of groundwater. These studies are reviewed briefly below.

Fig. 8. Long-term changes in mean annual concentrations of trench ‘fingerprint’ chemicals and pH in well 6-95-1, located immediately downstream from the trench no. 22T: (a) K and Mg; (b) Ca, SO4 and pH.

5.2.1. Field experiments on the in situ determination of the 90Sr Kd This method was used to estimate 90Sr Kd for eolian sands. Sediment cores were extracted from the water saturated top part of eolian layer from the contaminated part of the aquifer downstream from the trench (6 core samples in total). The core sampling was

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Fig. 9.

90

Sr distribution in cross-section of the aquifer at the CPS in June 2002 (analytical error 10–20%) (Dewière et al., 2004).

carried out using a specially designed split auger sampler, and the porous solution was immediately separated from the matrix by applying a vacuum. Next, the sample solutions were stabilized, and 90Sr activity was determined in the laboratory separately for the liquid and solid phase, and eventually Kd values were estimated (Bugai et al., 2002). The resulting Kd values ranged from 0.2 to 5 mL/g, with a mean value of 2 mL/g. 5.2.2. Laboratory batch determinations of 90Sr Kd Laboratory batch tests on determination of Kd were carried out using eolian and alluvial sediment samples obtained from borehole drilling at the experimental site, as well as with the sediment samples collected at the Pripyat Zaton exposure, which represents the geological analog of the Chernobyl Pilot Site (see Section 2.1). The contaminated groundwater from the aquifer at the experimental site was used as a test solution (Bugai et al., 2001). Results on 90 Sr Kds for different types of deposits are summarized in Table 5. The distribution coefficient for eolian sand obtained using different methods are in good agreement (on average 1–3 mL/g). Estimated 90 Sr Kds for alluvial deposits are approximately 10 times higher. This is apparently caused by the fact that the cation exchange capacity (CEC) of alluvial deposits (on average 3–10 meq/100 g) is significantly higher compared to the CEC of eolian deposits (0.5–1.5 meq/100 g). It should be pointed out that the layer of alluvial deposits has a regional spread at the Chernobyl NPP site (Matoshko et al., 2004), and, therefore, may represent an important natural sorption barrier for radionuclide migration from the Red Forest waste dumps. 5.2.2.1. Development and parameterization of the geochemical model for 90Sr sorption (Ardois and Szenknect, 2004; Szenknect et al., 2005). The experimental methods included: batch tests, dynamic stirred flow-through reactor experiments, and dynamic column tests. Sorption of the radioactive Sr label (85Sr) on eolian deposits was studied as a function of the concentration of the stable Sr ([Sr]) in the solution. The chemical composition of the test solution

corresponded to the ‘‘background’’ conditions of the experimental site ([Ca] = 7.7  105 mol/L, pH = 6.4). In the range of [Sr] from 1013 to 106 mol/L radiostrontium sorption was adequately described by the linear isotherm (Kd = 20 mL/g), however upon further increase of [Sr] from 106 to 105 mol/L radiostrontium Kd decreased by a factor of 3. Thus it was shown that Sr sorption over the wide range of concentrations is a non-linear process. The experiments in the stirred flow-through reactors have shown also that radiostrontium sorption is a ‘‘fast’’ and reversible process (characteristic sorption time <1 min; desorption 100%). The PHREEQC – based radiostrontium sorption model was parameterized with results of the stirred flow-through reactor experiments, which allowed estimation of Sr distribution in the soil–solution system depending on the chemical composition of the solution (Ardois and Szenknect, 2004; Szenknect et al., 2005). The developed model was experimentally confirmed (‘‘validated’’) by column tests using varying concentrations of stable Sr in test solutions. Application of the model to predict 90Sr Kd in geochemical conditions, which are typical for the aquifer in the zone of influence of the trench ([Ca] = 103,42 mol/L, [Sr] = 106 mol/L), has given Sr Kd values in the range of 1.7–3.5 mL/g, which are in good agreement with the results of field studies described in the beginning of this section (Ardois and Szenknect, 2004). 5.2.2.2. Radionuclide sorption in conditions of partial saturation of soils by water. The set of dynamic column tests was aimed at studying 90Sr and 137Cs sorption in conditions of partial saturation of eolian sands by water (Ardois and Szenknect, 2005). The important conclusion from these studies is that the 90Sr sorption distribution coefficient determined in saturated conditions (Kdsat), can be applied to describe migration behavior of radionuclides in conditions of partial saturation. The expression for calculating the sorption retardation coefficient for 90Sr transport in the unsaturated zone (Runs) is as follows (Ardois and Szenknect, 2005)

Runs ¼ 1 þ

q huns

Kdsat ;

D. Bugai et al. / Applied Geochemistry 27 (2012) 1359–1374 Table 5 Sorption parameters of

90

Sr for eolian and alluvial deposits.

Type of deposits

Site

Method

Mean Kd (range), mL/g

Eolian Eolian Alluvial Eolian Alluvial

Pilot Site Pilot Site Pilot Site Pripyat Zaton Pripyat Zaton

in situ partition test Batch tests Batch tests Batch tests Batch tests

2 (0.2–5.5) 3 (0.8–4.7) 5 (3.4–10) 2.8 (2.7–2.9) 20 (6–50)

where huns is the moisture content in the unsaturated zone, and q is the bulk density of soil. Results of the sorption studies were used to parameterize the mathematical model of 90Sr migration from trench no. 22T to the unsaturated zone and aquifer (see Section 6). 5.3. Strontium-90 balance in the system ‘‘waste burial – aquifer’’ The data on 90Sr distribution in groundwater and derived Kd estimates allowed direct evaluation of the radionuclide inventory in the aquifer (Bugai et al., 2001). The amount of 90Sr in the aquifer (both dissolved in groundwater and adsorbed on deposits) in 2001 was calculated via spatial integration:

AAQ ¼

ZZ

1369

Cðx; zÞðm þ qK d Þdxdz;

where AAQ is 90Sr inventory in the aquifer (Bq/m of trench cross-section), C(x, z) is 90Sr concentration in groundwater, m is sediment porosity, q is deposit bulk density, and Kd is the spatially dependent distribution coefficient. Assuming that the average 90Sr Kd value for the eolian layer ranges from 2 to 3 mL/g, and for the alluvial layer from 10 to 20 mL/g, the resulting estimate for the aquifer cross-section is:

AA Q ¼ 200—350 MBq ðper 1 m of cross-sectionÞ: The initial inventory of 90Sr within the trench for the aquifer cross-section corresponding to the profile of multilevel observation wells was estimated using data on 137Cs activity in the trench soil and fission product activity correlations in Chernobyl NPP nuclear fuel (Kuriny et al., 1993). The resulting estimate (corrected for decay for 2000) constituted:

5:5  2:2 GBq ð=m of cross-sectionÞ: Therefore, the amount of 90Sr in the aquifer in the considered cross-section in 2001 was 4–6.5% of the initial 90Sr inventory in the trench. The last value is in good agreement with that presented in Kashparov et al. (this issue) for an independent estimate of 90Sr release from trench of 7 ± 5%, which was based on comparing 90Sr to the 154Eu ratio in Chernobyl reactor fuel and in trench material (the last one being depleted in 90Sr due to leaching to groundwater by infiltration flow). It should be pointed out that the estimate of Kashparov et al. (this issue) includes both 90Sr in the unsaturated zone and in the aquifer. Thus a good cross-check of 90Sr release to the aquifer using independent experimental methods was obtained. Lastly, taking into account data from Kashparov et al. (this issue) on 90Sr distribution inside the trench between different families of fuel particles, the amount of 90Sr in the aquifer in 2001 constituted 12–23% of the initial 90Sr activity associated in the trench with relatively higher solubility U–O matrix fuel particles.

to the unsaturated zone and the aquifer (Bugai and Dewière, 2004b; Dewière et al., 2005; Van Meir et al., 2009). The numerical modeling summarized all knowledge and data collected in the course of the project studies; it is an integration of separate subsystems or compartments (model for dissolution of fuel particles, filtration model, geochemical sorption model, etc.), and represents a quantitative validation of hypotheses on the main hydrodynamic and geochemical processes which control radionuclide migration in the ‘‘waste site – unsaturated zone – aquifer’’ system. The developed model is of applied interest, as it can be used as a tool for predicting long-term radionuclide migration from the Red Forest waste dumps for risk assessment purposes. 6.1. Overview of modeling approach The 90Sr transport model for the waste site was based on a modular approach. The spatial distribution of radionuclides in the trench was described by means of the 3-dimensional (3D) structural (geostatistical) model (Bugai et al., 2005). Dissolution of fuel particles and 90Sr redistribution in the trench body and in the unsaturated zone was modeled by a specially developed one-dimensional (1D) transport code STERM1D (see Section 6.3). This code uses the model for release of 90Sr from fuel particles developed in Kashparov et al. (this issue). The 90Sr transport in the aquifer was simulated using the Visual Modflow (Waterloo Hydrogeologic, Inc.) pre-/post-processor groundwater modeling software that uses MODFLOW (McDonald and Harbaugh, 1984) for flow and MT3D (Zheng, 1990) for transport calculations. The calculation assumed steady-state hydrodynamic conditions corresponding to the long term averaged values of the relevant hydrogeological parameters – GWL position, infiltration recharge rate, and groundwater flow velocity and direction. The model also assumed steady-state geochemical conditions (i.e., constant in time 90Sr Kd values for different geological units). Fig. 10 shows data exchanges between the sub-models schematically. The 1D transport calculations using the STERM1D code

6. Modeling of 90Sr transport from the trench no. 22T to the unsaturated zone and aquifer The ultimate objective of the Chernobyl Pilot Site project consisted of developing an integrated (so called ‘‘global’’) mathematical model of 90Sr migration from the studied waste trench no. 22T

Fig. 10. Schematic representation of the numerical realization and data exchanges of the radionuclide transport model for the Chernobyl Pilot Site (Dewière et al., 2005).

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were carried out for separate columns of the 3D numerical grid, describing the initial distribution of radioactivity within the trench. The output of STERM1D (90Sr concentration in water outflowing from the unsaturated zone to the groundwater table) served as a boundary condition for the MODFLOW – MT3D model, describing radionuclide transport in the aquifer. The simulation period covered 16 a: from 1987 (date of the ChNPP site clean-up operations following the Chernobyl accident) to 2002. Though extensive characterization studies have been carried out at the Chernobyl Pilot Site, essential uncertainties remained regarding a number of radionuclide migration parameters (in particular, due to imprecise knowledge of initial conditions and past stresses on the studied waste site system). To avoid unnecessary complexity the most sensitive parameters: the trench sorption distribution coefficient (Kd,tr), the aquifer sorption distribution coefficient (Kd,aq), and dispersivites for radionuclide transport in the aquifer have been treated as uncertain. For other parameters the ‘‘best guess’’ values were used (see Tables 6 and 7). The infiltration recharge rate value of 300 mm/a used in calculations (which was higher than the present day infiltration values of 200 ± 50 mm/a, see Table 2) accounted for the lower evapotranspiration at the waste site in the early period after trench no. 22T construction due to the lack of vegetation. Calibration calculations for the radionuclide transport model used the following data: (1) time series of 90Sr activity in groundwater in monitoring wells at the CPS, which has been monitored since 1995, and (2) data on the two-dimensional 90Sr distribution in the aquifer cross-section, which were obtained using multilevel observation wells installed at the site in 2000–2002 at the ‘‘Laboratory’’ profile (see Section 5.1). Calibration calculations proceeded from the 1D to the 2D case. The 1D stream-tube model (based on STERM1D) was used for model calibrations and sensitivity analyses using a single well time series of 90Sr measurements (with respect to Kds). Next, calibration calculations were carried out using the 2D aquifer cross-section model using MODFLOW-MT3D (with respect to both Kds and dispersivities). Both 1D and 2D model calibrations were carried out manually using the ‘trial and error’ approach. The most sensitive model parameter was the distribution coefficient (Kd). The range of suitable Kd values was fitted first. In the next step the dispersivities Table 6 Hydrogeological parameters of the groundwater flow model for the Chernobyl Pilot Site. Parameter

Value

Groundwater level elevation (multi-year mean value), m a.s.l. Flow direction in eolian layer Horizontal hydraulic head gradient in eolian layer Vertical hydraulic head gradient in alluvial layer Infiltration recharge rate, mm/y Hydraulic conductivity of eolian layer (isotropic), m/d Hydraulic conductivity of alluvial layer (anisotropic), Kx, m/d Kz, m/d

111.5 North (±15°) 0.0015 0.03 300 3.6 0.5 0.0275

Table 7 Fuel particle dissolution parameters in trench no. 22T. Parameter

Value 90

Initial content of Sr in exchangeable form (E0), % Initial content of 90Sr in UO2+X fuel particles (E1), % Initial content of 90Sr in UO2 fuel particles (E2), % Initial content of 90Sr in U–Zr–O fuel particles (E3), % Dissolution rate constant for UO2+X fuel particles (a1), a1 Dissolution rate constant for UO2 fuel particles (a2), a1 Dissolution rate constant for U–Zr–O fuel particles (a3), a1

2 19 56 23 0.28 0.018 0

were adjusted (for the 2D model). The separate sub-models and results of calibration studies are described in more detail below. The described numerical approach also allowed the 3D simulation of the long-term 90Sr transport in the ‘trench–aquifer’ system (Bugai and Dewière, 2004b). 6.2. Geostatistical model for radioactivity distribution in the waste site The initial activity of 90Sr in waste was inferred from data on distribution of 137Cs activity in trench no. 22T measured in 2000–2002 using the gamma-logging technique (Kashparov et al., this issue). The calculation procedure was based on the assumption of low mobility of 137Cs in the trench soil, and used the radionuclide correlation ratio between 90Sr and 137Cs in the nuclear fuel of the Chernobyl NPP Unit 4 at the time of the accident. The 3dimensional 137Cs distribution in the trench was evaluated by means of geostatistics, which included the development of a structural model for the radioactivity distribution in the trench (using the semi-variogram function), and interpolation of data (using the kriging method). Details can be found in Bugai et al. (2005). Analyses used the geostatistical software library GSLIB (Deutch and Journel, 1998). 6.3. A one-dimensional stream-tube model for radionuclide transport in the ‘‘trench–aquifer’’ system (STERM1D) The 1D radionuclide transport in a flow tube in the ‘‘trench– aquifer’’ system was described by the following advection–dispersion equation in the curvilinear coordinate system (Fig. 11a):

H ðxÞ @C VðxÞ @t

¼ aL

@ 2 C @C H ðxÞ H ðxÞ  kC þ Fðt; xÞ;  2 @x @x VðxÞ VðxÞ

H ðxÞ ¼ HðxÞ þ qðxÞK d ðxÞ where x is the curvilinear coordinate along the flow tube (m); t is time (day); C(x, t) is the 90Sr concentration in groundwater (Bq/L); V(x) is the Darcy velocity (V(x) is variable along the tube; however, it holds at V(x) S(x) = Q = const, where S(x) is the cross-section area of the tube, and Q is the discharge rate of the flow tube) (m/day); aL is the longitudinal dispersivity coefficient (m); k is the 90Sr radioactive decay constant (day1); F(t, x) is the source term describing radionuclide leaching from fuel particles inside the trench (Bq/(dm3 day)), (see formula (7) in Kashparov et al., this issue) H(x) is the soil water content (unitless); q(x) is the soil bulk density (kg/dm3); and Kd(x) is the sorption distribution coefficient (L/kg). The transport equation presented above was solved numerically using the explicit finite-difference scheme and implemented in the STERM1D computer code (Bugai et al., 2003). The described model also allowed simulating radionuclide transport in the unsaturated soil column only (in order to specify boundary condition for the 2D/3D aquifer transport problem; see Section 6.1). When simulating the 90Sr transport in the unsaturated zone the following hydraulic parameters were used: V = 300 mm/a (i.e. the infiltration recharge rate value; see Table 6), and h = 0.12. Parameters for the trench source-term model (F(t, x)) were established from the purposeful experimental program (Kashparov et al., this issue), and are listed in Table 7. 6.4. Radionuclide transport model calibration results 6.4.1. Preliminary 1D modeling results (STERM1D) The calibration process for the 1D model is illustrated in Fig. 11b. Calibration calculations used monitoring data on 90Sr activity in well 6-95-1. This well is located in the immediate vicinity of trench no. 22T, with a screen approximately 1 m below the

D. Bugai et al. / Applied Geochemistry 27 (2012) 1359–1374

1371

best fit was achieved with Kd,tr = 8 mL/g, and Kd,aq = 1 mL/g. The ’model – observations’ fit is not perfect. The monitoring data show a decrease in the 90Sr concentration in groundwater during the period from 1999 to 2001, which is not reproduced by the model. The likely explanation is that the above-described behavior is caused by changes in geochemical conditions within the aquifer during the modeled period resulting in an increase of 90Sr Kds and a decrease in the radionuclide leaching rate from the trench. These changes were supposedly caused by bio-geochemical factors outlined in Section 4.4 (see next paragraph for more discussion).

Fig. 11. One-dimensional radionuclide transport model STERM1D in the ‘trench– aquifer’ system: (a) schematic representation of 1D flow-tube model domain; (b) 1D model calibration results using observation data for well no. 6-95-1 (Van Meir et al., 2009).

trench bottom. The dispersivity coefficient (aL) was set at 2 cm. The value of the distribution coefficient for the trench material (Kd,tr) regulated the timing of the 90Sr breakthrough to the aquifer, while the distribution coefficient for the aquifer (Kd,aq) influenced the rate of migration and maximum values of radionuclide concentrations in the aquifer during the simulation period (see Fig. 11b). The

Fig. 12. Two-dimensional

90

6.4.2. Two-dimensional transport using MODFLOW/MT3D For the 2D simulation the best fit with field data on 90Sr distribution in the aquifer cross-section was achieved for the scenario assuming Kd,tr = 5 mL/g, Kd,aq = 0.5 mL/g (for the eolian sand layer), aL = 10 cm, and aT = 1 cm (Fig. 12). The fitted longitudinal dispersivity value (aL = 10 cm) is large compared to experimental data from the natural-gradient tracer tests using 36Cl (see Section 4.2). This may be because the 2D model scale (20 m) is an order of magnitude larger than the tracer tests scale (1–2 m). One other argument is that 90Sr, which is a reactive tracer, is expected to be subject to a higher dispersion than a non-reactive tracer (36Cl), as a weak negative correlation is observed between the sorption capacity and the hydraulic conductivity of the aquifer deposits (Dewière et al., 2004). Similarly to the 1D model calibrations, in the 2D case some reversal in the 90Sr concentration distribution in the plume along the flow path can seen: the 2002 field observations show a higher activity ‘‘core’’ in the downstream part of the plume, while the modeling suggests a monotonous decrease from the source zone in the downstream direction. As already discussed, this pattern can be caused by long-term temporal changes in the geochemical conditions in the ‘trench– aquifer’ system leading to an increase with time of the 90Sr sorption by the trench matrix and the aquifer deposits. The same hypothesis can explain a relatively low fitted Kd value for the aquifer of 0.5 mL/g. In the context of the used model, the fitted value represents the time-averaged ‘‘effective’’ parameter. If it is assumed that in reality a temporal evolution of aquifer Kds may have taken place from lower to higher values, the ‘‘effective’’ parameter should be lower than Kd estimates based on present-day conditions. The 90Sr sorption on soils is governed by ion-exchange processes, and is sensitive to the presence in solution of other cations,

Sr transport model calibration results for t = 16 a (Van Meir et al., 2009).

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D. Bugai et al. / Applied Geochemistry 27 (2012) 1359–1374

especially Ca and stable Sr, which compete with 90Sr for exchange sites on the soil matrix (Pavlotskaya, 1974; Lefevre et al., 1996; Szenknect et al., 2005). As outlined in Section 4.4, the long-term monitoring data series show a decrease of concentrations of base cations (Ca, Mg), leached from the trench to groundwater. Ardois

and Szenknect (2004) evaluated the impact of changes in Ca concentration in solution on 90Sr sorption by eolian sand by means of geochemical modeling using the PHREEQC-based ion-exchange model. The simulated decrease of Ca in solution from 2  103 mol/L (80 mg/L; i.e., conditions observed downstream

Fig. 13. Modeling predictions of flow lines in the unconfined aquifer (green arrows) and of the long-term evolution of the 90Sr plume from trench no. 22T in the aquifer crosssection (horizontal and vertical scales in meters) (IGS, 2005). (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

D. Bugai et al. / Applied Geochemistry 27 (2012) 1359–1374

from the trench in 1998, see Fig.8) to 2.7  104 mol/L (10 mg/L; conditions observed in 2003–2004) caused an increase of 90Sr Kd from 0.9 mL/g to 14 mL/g. Such an increase in Kd corresponds to the increase of 90Sr retardation factor from R = 6 to R = 80, which clearly shows the importance of the discussed changes in the geochemical regime of groundwater for the 90Sr transport in the unsaturated zone and aquifer. For more discussion of factors influencing the geochemical regime of the aquifer downstream from trench no. 22T and implications for radionuclide migration the reader is referred to Bugai et al. (this issue). 6.5. Long-term predictions of

90

Sr transport

The radionuclide transport model described was used to carry out modeling predictions of the long-term transport of 90Sr in groundwater from waste trench no. 22T (IGS, 2005). Such calculations are of interest, as they allow estimating the spatial and temporal scale of the hazards caused by radioactive contamination of groundwater from the Red Forest waste dumps. Calculations used ‘‘best guess’’ model parameter values (Tables 6 and 7). The following 90Sr Kd values were used: eolian sand layer (112–109 m a.s.l.) – Kd = 1 mL/g, alluvial sand layer of dead-channel facies (109–106 m a.s.l.) – Kd = 4 mL/g; alluvial sand layer of channel facies (106–82 m a.s.l.) – Kd = 1 mL/g (Fig. 13). According to the modeling 90Sr activity in porous solutions below the trench bottom reaches its maximum (20–25 kBq/L) around 20–30 a after the accident, and decreases to 1 Bq/L in 240 a. The maximum amount of 90Sr in the aquifer is reached 40 a after the accident, and it constitutes 18% of the initial inventory within trench no. 22T. After this time radioactive decay prevails over the infiltration of radionuclide to the hydrogeological environment from the waste trench. The results of the long-term (300 a) modeling predictions of 90 Sr migration in the aquifer has shown that: the maximum distance of 90Sr migration from the waste site in concentrations >2 Bq/L (which is the 90Sr drinking standard in Ukraine) constitutes 200 m; the aquifer is contaminated to the depth of about 15 m from the groundwater table; in about 250 a 90Sr concentrations throughout the aquifer will decrease below 2 Bq/L (Fig. 13). These results generally confirm the conclusions from previous screening-level studies (Bugai et al., 1996a,b) that 90Sr migration from the Red Forest waste dumps essentially does not represent a risk of contamination of Pripyat River and the underlying the confined aquifer in Eocene deposits, which is used as a potable water source at the Chernobyl NPP site. It should be pointed out that modeling predictions were determined using the assumption of steady-state hydrodynamic and geochemical conditions (constant Kds) within the hydrogeology system of the waste site. In reality, the increase of pH and decrease of ionic strength of groundwater takes place in the geochemical plume emerging from the trench due to the bio-geochemical transformations of organic matter inside the trench, uptake of nutrient species by roots of newly planted forest at the top of the trench and other factors (see Section 4.4 and Bugai et al. (this issue), for more detail). This causes an increase in 90Sr Kds, attenuation of radionuclide concentrations in groundwater, and increased retardation of migration. Therefore, the modeling predictions presented (which neglect the geochemical evolution of the waste site) can be considered as upper boundary estimates of groundwater contamination hazards. 7. Conclusions Carrying out the Chernobyl Pilot Site (1999–2003) and Experimental Platform in Chernobyl (2004–2008) projects, described in

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this and the accompanying article (Kashparov et al., this issue) has broadly improved understanding of the hydrogeological, geochemical and biotic processes of radionuclide migration in the Red Forest waste dumps in the Chernobyl exclusion zone, as well as raised a number of new issues and questions. The modeling case presented in this article has shown that despite the high degree of effort involved in characterization and the numerical model complexity, the developed models were only partially successful in accurately reproducing migration behavior of 90 Sr from the waste trench (i.e., the 1D model for the period of 1995–1998). An imperfect fit was observed between the modeling results and monitoring data for both 1D and 2D model for the period 1999–2002. Also the 90Sr Kd values fitted during the model calibration for the eolian sand aquifer (0.5–1 mL/g) were lower than the a priori predicted range (2–3 mL/g). The numerical model limitations are presumably related to the shortcomings of the conceptual model, in particular to the assumption of steady-state geochemical conditions in the aquifer. One other difficulty is related to the uncertainty in the initial conditions for the transport modeling (distribution of 90Sr in the trench in 1987) (Bugai et al., 2003). Modeling has lead however to a new insight in factors controlling long-term transport of 90Sr in the studied waste site, and to the identification of how to improve the conceptual model. Development and experimental validation of a more detailed radionuclide transport model accounting for a transient hydrological regime and evolving geochemical conditions is among the research subjects which are currently being pursued by the French and Ukrainian institutes involved in the framework of the continued collaborative projects at the EPIC site. The developed model has proved to be a useful tool for assessing risks from radionuclide migration to groundwater from Red Forest waste dumps, and getting an upper boundary estimate of groundwater contamination hazards. The important ‘‘lesson learned’’ from the decade of radioecology studies reported is that the complex interplay between the hydrological, geochemical and biotic processes observed in a ‘‘real world’’ radioactively contaminated site such as the Red Forest dictates the need for multidisciplinary research and integrated radionuclide fate and transport models accounting for the whole spectrum of cross-influence between the radionuclides, atmosphere and groundwater, the soil geochemical environment, microorganisms and higher plant species. It should be noted, that the reported project studies resulted in a detailed experimental data set on the dissolution behavior of fuel particles, radiological and hydrogeological conditions, as well as on biotic migration processes (see also Kashparov et al., this issue) in the Red Forest waste dump site. The established database could be used for the international radioecological and radionuclide migration model inter-comparison and validation exercises. The experimental polygon created in the framework of the project equipped with the up-to-date facilities and equipment for hydrogeological, meteorological and radiological monitoring of the radioactively contaminated site may serve as a basis for further multidisciplinary radioecological research projects, which could benefit from the previously created research infrastructure and the established parameter database. Acknowledgements Studies presented in this paper were supported by French– Ukrainian collaborative radioecology projects ‘‘Chernobyl Pilot Site Project’’ (1999–2004) and ‘‘Experimental Platform in Chernobyl’’ (2005–2008) funded by the French Institute for Radiation protection and Nuclear Safety (IRSN, Fontenay-aux-Roses), and by Research Theme No. III-11-06 of the Ukrainian National Academy

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of Sciences. We also commemorate our colleague and personal friend Dr. Lionel Dewière, the founder of the Chernobyl Pilot Site and one of the initiators of the described radioecology research, who died in 2009. We thank Jean-Michel Matray and two anonymous reviewers for their remarks, which helped to improve the manuscript.

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