Applied Geochemistry 27 (2012) 1320–1338

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Applied Geochemistry

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Geochemical influence of waste trench no. 22T at Chernobyl Pilot Site at the aquifer:Long-term trends, governing processes, and implications for radionuclide migration

D. Bugai a,⇑, E. Tkachenko a, N. Van Meir b, C. Simonucci b, A. Martin-Garin c, C. Roux b,d, C. Le Gal La Salle d,Yu. Kubko a

a Institute of Geological Sciences, 55-b, Gonchara Str., Kiev 01054, Ukraineb Institute for Radioprotection and Nuclear Safety, IRSN/DEI/SARG/LR2S, BP 17, 92262 Fontenay aux Roses Cedex, Francec Institute for Radioprotection and Nuclear Safety, IRSN/DEI/SECRE/LRE, BP 3, 13115 Saint-Paul-Lez-Durance Cedex, Franced University of Nimes, Laboratory of Environmental Geochemistry, 150, rue Georges Besse, 30035 Nîmes Cedex, France

a r t i c l e i n f o

Article history:Available online 28 September 2011

0883-2927/$ - see front matter � 2011 Elsevier Ltd. Adoi:10.1016/j.apgeochem.2011.09.021

⇑ Corresponding author. Tel./fax: +380 44 486 30 2E-mail address: dmitri.bugay@gmail.com (D. Buga

a b s t r a c t

This article discusses and analyzes data for geochemical monitoring of groundwater, collected in1998–2008 at the international experimental radioecological study site (Chernobyl Pilot Site) locatedin the Red Forest radioactive waste dump site in the Chernobyl exclusion zone. Groundwater in the zoneof influence of the studied waste trench no. 22T was characterized by a specific geochemistry. Along witha high content of 90Sr (of an order of n � 1000–n � 10,000 Bq/L) groundwater showed elevated concen-trations of Ca, K, NO�3 , SO2�

4 and of some trace elements (in particular stable Sr), and had more acidicpH values compared to ‘‘background’’ aquifer conditions. The observed water quality changes are appar-ently related to degradation of organic matter (pine forest remnants, litter, humus containing topsoillayer) buried inside trench no. 22T, which have lead to acidification of groundwater and leaching of cat-ions adsorbed on the exchange complex of soils buried inside the trench. Regime monitoring data for theproject period of 1998–2008 shows a progressive decrease of concentrations of the leached ions accom-panied by an increase of pH in the chemical plume emerging from the trench. This can be explained by acombined effect of a gradual decrease of the inventory, humification of the original organic matter insidethe trench, and of nutrient element uptake by roots of the newly planted pine forest on top of the trench.The identified trends of evolution of geochemical regime favor attenuation of subsurface migration of 90Srfrom Red Forest waste dumps due to an increase of the 90Sr distribution coefficients, and stabilization ofplumes of contaminated groundwater.

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1. Introduction

The Chernobyl accident on 26 April 1986 has resulted in severeradioactive contamination of the territory immediately adjacent tothe destroyed Chernobyl Reactor Unit 4. The remedial measureimplemented shortly after the accident (1987) in order to reducethe external exposure to the ‘‘liquidators’’ was the in situ burialof contaminated soil, vegetation, and other radioactive debris,including the ‘‘Red Forest’’ pine trees (killed by high levels of radi-ation), within about an 8 km2 area around the Chernobyl nuclearpower plant (ChNPP). The hastily constructed unlined trencheswith clean-up wastes became with time serious sources of radioac-tive contamination of groundwater (Dzhepo and Skalskyy, 2002).

Starting from 1998 the migration behavior of radionuclides inChernobyl waste dumps has been studied jointly by an interna-tional team of French and Ukrainian radioecologists (Bugai et al.,2002; Dewiere et al., 2004, 2005; Szenknect et al., 2005; Van Meir

ll rights reserved.

3.i).

et al., 2009). For this purpose the experimental site (Chernobyl Pi-lot Site) was established in the vicinity of trench no. 22T at the RedForest waste dump site (Fig. 1), and a multidisciplinary site charac-terization and monitoring program was initiated. These studiesallowed characterization of the geological structure and hydrogeo-logical conditions of the study site, water flow regime in the unsat-urated zone and aquifer, radionuclide distribution and speciationin the trench, dissolution rates of dispersed nuclear fuel particlesinside the trench, etc., and eventually development and calibrationof a groundwater flow and radionuclide transport model for thestudied waste site (Dewiere et al., 2005; Van Meir et al., 2009).

This article is devoted to the analysis and the interpretation ofdata for geochemical monitoring of groundwater at the experimen-tal site collected during the period of 1998–2008. This is an impor-tant and interesting subject as geochemical conditions in theaquifer influence migration behavior and mobility of radionuclidesfrom the waste dumps in the subsurface environment. Below,changes to groundwater chemistry caused by the trench, long-term trends in geochemical data are analyzed, hydrological andbio-geochemical process controlling the impact of the waste site

Fig. 1. Schematic map of experimental site near trench no. 22T with groundwater monitoring well network.

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1321

on the aquifer discussed, along with consequences of the evolutionof the geochemical environment for 90Sr migration in groundwater.

2. Materials and methods

2.1. Experimental site

The experimental site is located at the Red Forest radioactivewaste dump site in the so-called ‘‘Near Zone’’ (approx. 10 km ra-dius) of the Chernobyl Nuclear Power Plant (ChNPP). The researchis focused on waste trench no. 22T created during the clean-upoperations at the site in 1987. The trench contains pine forest(and possibly some other vegetation remnants), topsoil layer, litteretc. contaminated by radioactive fallout fuel particles originatingfrom the destroyed Unit 4 of ChNPP (Fig. 1). Upon completion ofthe clean-up operations in 1987 the territory of the waste sitewas covered by a 30–50 cm thick ‘‘clean’’ soil cover layer (thesandy soil excavated from the trenches during their constructionwas used for this purpose). In addition, forest planting (pines, birchtrees) was carried out in the territory of the waste site in1988–1989 in order to prevent wind resuspension and watererosion of the soil cover layer (Arkhipov et al., 1994).

The studied trench no. 22T is approximately 70 m long, 5–6 mwide and 2–3 m deep. The specific activity of buried waste with re-spect to 90Sr and 137Cs constitutes on average n � 105–106 Bq/kg.The total radioactivity inventory within the trench is estimatedto be 600 ± 240 GBq for 137Cs and 290 ± 140 GBq for 90Sr (Bugaiet al., 2002). An important feature of trench no. 22T is the hetero-geneous character of the radioactivity distribution inside thetrench with variation of 137Cs activity of more than an order of

magnitude, which likely resulted from mixing of contaminatedtopsoil when bulldozing it to the trench during clean-up (Bugaiet al., 2005).

Based on analyses of soil sample weight loss on ignition carriedout in 2006 the organic matter (OM) content in the trench rangedfrom 2.6% to 8.6% (see in more detail Section 3.3). Pine trunkswhich were not fully decomposed were encountered until recentlyduring drilling of characterization boreholes through the trenchbody. For comparison OM content of host sediments (eolian sands)was estimated using the same method at 0.17%.

Determination of extractable (using soda solution) organic C(Corg,ext) in the trench soil carried out in 2008–2009 on several tensof samples (taken at constant depth) in the framework of microbi-ology studies have given Corg,ext values in the range of 0.5–4 g/kg.An important finding is the high correlation between the contentof Corg,ext and the 137Cs activity of trench soil (r2 = 0.78), whereasthis correlation is less pronounced for the soils sampled outsidethe trench (r2 = 0.38) (Chapon et al., 2011).

The vegetation cover at the trench surface of about 400 m2 wasrepresented in 1999–2000 by 47 pines (Pinus sylvestris), 14 birchtrees (Betula pendula) and 49 bushes (Swida alba, Physocarpusopulifolius, etc.); about 30 m2 was covered by grass (Calamagrostisepigeios).

With respect to the geomorphology, the study site is located inthe central part of the 1st terrace of the Pripyat River. Ground sur-face elevations are 112–115 m a.s.l. The depth to the groundwatertable varies from 1 to 4 m (depending on site topography and sea-sonal groundwater level fluctuations). The water saturated geolog-ical sediments are represented by eolian and alluvial sands. Theupper part of the geological cross-section (to the elevations of

Fig. 2. Groundwater level regime at experimental site in 1999–2008 (well 2-99). Dotted line shows elevation of trench no. 22T bottom at the LAB profile.

Fig. 3. Laboratory (LAB) multilevel well profile and soil water samplers (in front).View from trench no. 22T to the North (October 2005).

1322 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

109–110 m a.s.l.) is composed of well permeable eolian sands(hydraulic conductivity �3–5 m/day). The relatively low-perme-ability alluvial sediments (horizontal hydraulic conductivity�0.5–1 m/day, vertical �0.03 m/day) are located below. The aqui-fer sediment porosity is 33–42%, the bulk density is 1.6–1.7 g/cm3.The sand mineralogical composition is quartz with admixtures offeldspars (up to 5–9%) and accessory minerals (<0.5%). The cationexchange capacity (CEC) of the eolian sands is �1 meq/100 g, whilethe CEC of alluvial deposits reaches 5–10 meq/100 g (Bugai et al.,2002; Matoshko et al., 2004).

The infiltration recharge rate to the aquifer is estimated, basedon data of studies in 2000–2004, at 200–300 mm/a, or about �30–50% of the yearly amount of meteoric water precipitation (Bugaiet al., 2001). Most infiltration recharge usually occurs in springfollowing snowmelt, however, large rainstorms in the summer–au-tumn period can also significantly contribute to yearly recharge.

The groundwater level regime at the experimental site in1999–2008 registered by automated level meter is shown atFig. 2. It is characterized by well-defined seasonal cycles of risesand recessions, linked to the yearly recharge patters discussedabove.

The groundwater flow in the eolian layer occurs approximatelyto the North with seasonal fluctuations of 10–15� (in the directionof the Prypiat River), with an average head gradient of 0.001–0.0015 m/m. The flow in the alluvial layer is nearly vertical witha head gradient of about 0.02–0.04 m/m. The horizontal flowvelocity in the layer of eolian sands was estimated from tracer testsat �3 cm/day (�10 m/a) (Bugai et al., 2002; Dewiere et al., 2004).

Hydrogeological conditions and waste dump design features(lack of engineering barriers) favor intensive migration of radio-nuclides and other chemical constituents from the waste trenchto groundwater. The most mobile radionuclide in groundwater is90Sr (half life 28.8 a). The plume of 90Sr contaminated groundwaterwith concentrations in the range of n � 102–104 Bq/L has formeddownstream from the trench (see below Section 3.2). It was esti-mated that about 4–7% of 90Sr initial trench inventory had mi-grated by 2000 from the waste site to the aquifer, and that 90Srmigration velocity in the eolian sand aquifer was retarded by sorp-tion to �9% of real groundwater flow velocity (Bugai et al., 2002;Dewiere et al., 2004).

2.2. Groundwater monitoring system and geochemistry dataset

Three profiles of multilevel monitoring wells oriented along thegroundwater flow direction were constructed during the course ofthe project to characterize radionuclides and chemical species dis-tribution in the aquifer cross-sections (Fig. 1). These well profileswere designated ‘‘1995’’ (1995 profile; constructed in 1995–1998), ‘‘Laboratory’’ (LAB profile; constructed in 2000–2002;

Fig. 3) and ‘‘Weather station’’ (WS profile; constructed in2002–2006). The orientation of the 1995 profile differs from otherwell profiles, as the mean groundwater flow direction had not beendefined accurately enough at the early stage of research, when thisprofile was constructed.

Each multilevel well profile is composed of 4–6 clusters of wellsof different depth, which are spaced approximately 4 m from eachanother. In its turn each well cluster is composed of 3–4 PVC wellsof 2.54 cm (1 in.) diameter with 20 cm long stainless steel meshscreens, in the depth range from �2–3 m to �8 m with a verticalinterval between well screens of �1–1.5 m (see Fig. 1). In LABand WS well profiles to avoid cross-contamination, the individualscreens within the boreholes were isolated from each other by�20 cm thick bentonite sealing layers. In the 1995 well profile suchisolation measures have not been carried out.

In addition to monitoring wells, two sets of ceramic vacuummoisture samplers (1900L series model of Soilmoisture EquipmentCorp.) were installed in 2005 to various depths (from 0.5 m to2.25 m) in trench no. 22T at LAB (see Fig. 3) and WS locations tosample the soil pore solutions from the trench.

Below geochemical data acquired in 1998–2008 are discussed.The geochemical data for the period of 1998–2002 are less compre-hensive and systematic (at that time groundwater sampling wasdone on wells of the 1995 profile). Since 2004 geochemical moni-toring became more systematic, and geochemical surveys werecarried out with a frequency of twice a year – usually in spring(May–June) and autumn (September–October), with the main

Table 1Upstream aquifer concentrations of major ions in eolian sand layer (LAB profile,September 2004) and meteoric water concentrations (mean values on 30 rainfall andsnow samples collected in 2005–2006).

Ion Upstream groundwater(Cbg,i), mg/L mean (min–

Meteoric water(Crain,i), mg/L mean

‘‘Concentrationfactor’’ Cbg,i/

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1323

focus on sampling LAB and WS well profiles. In addition to ground-water, about 30 meteoric water (rainfall, snow) samples were col-lected at the study site and analyzed for chemical composition in2005–2006.

The monitoring program for radionuclides and chemical com-position of groundwater included determination of 90Sr, major cat-ions (Ca, K, Na, Mg) and anions (Cl, NO3, HCO3, SO4), as well as traceelements (Si, Al, Mn, Fe, stable Sr, etc.), DOC and some low-molec-ular weight organic acids (ex. lactic, acetic, oxalic acids).

2.3. Field sampling methods and analytical procedures

Groundwater sampling in the field was done by a vacuum pumpor peristaltic pump. Observation wells were purged prior to samplecollection by pumping 3–5 L of groundwater. Groundwater sam-ples were filtered through a 0.45 lm filter immediately upon col-lection. Samples for cations were stabilized by adding HNO3. Thegroundwater conductivity, pH, Eh and dissolved O2 were measuredin situ in the field using a portable meter.

Chemical and radionuclide analyses were carried out in thehydrochemical laboratories of the Institute of Geological Sciences(IGS) and the Institute of Radioprotection and Nuclear Safety(IRSN). The inter-calibration exercises carried out during the pro-ject have shown satisfactory agreement in analytical data providedby participating laboratories.

2.3.1. Procedures used by IGSChloride concentration was determined by titration using

AgNO3 in the presence of potassium chromate in neutral or alka-line solution. (For meteoric water samples pre-concentration byevaporation was used in view of low concentrations of Cl� in suchsamples). Sulfate was determined by weight analysis (weight ofBaSO4 precipitated by reaction with BaCl). Ammonium was deter-mined by colorimetry of the yellow–brown compound formed bythe reaction between NHþ4 and Nessler’s reagent. Nitrate wasdetermined by colorimetric analysis of the solution resulting fromthe reaction between NO�3 and phenoldisulfonic reagent. Potas-sium and Na+ were determined by flame atomic emission spec-trometry. For Ca2+ and Mg2+, chelatometry titration was used.Bicarbonate was determined by titration using HCl and methylor-ange indicator. The 90Sr activity in water samples was measuredusing scintillation beta-spectrometer SEB-01 (Atomprylad, Kiev).

2.3.2. Procedures used by IRSNIn 1998 hydro-chemical samples were analyzed by means of

plasma emission spectrometry (for cations and anions) and byatomic absorption spectrometry (AAS) (for cations). All other sam-ple series were measured by AAS. From 2006 major ions and somelow-molecular weight organic acids were determined by ion chro-matography (Dionex and Metrohm), trace elements by ICP-MS, andDOC with a Total Organic Carbon analyser (Shimatzu). The 90Sractivity in groundwater samples was determined using a liquidscintillation spectrometer Quantulus 1220.

max) (min–max) Crain,i

Na+ 0.85 (0.7–1.0) 0.4 (0.1–1.0) �2.1K+ 1.5 (1.1–2.0) 0.4 (0.1–1.1) �3.8Ca2+ 6.0 (5.9–6.2) 2.1 (0.4–36) �2.9Mg2+ 1.35 (1.0–1.7) 0.5 (0.1–1.0) �2.7Cl� 1.65 (0.8–2.5) 0.74 (0.2–2.5)* �2.2

SO2�4

19.1 (18.9–19.3) 6.6 (1.6–11.5) �2.9

NO�3 0.85 (0.6–1.1) 1.1 (0.1–6.2) �0.8HCO�3 7.95 (6.1–9.8) 7.3 (2.4–15.3) �1.1pH 5.9 (5.6–6.2) 5.6 (4.6–6.9)**

* Mean values based on measuring seven water samples pre-concentrated byevaporation.** Off-site measurement in laboratory.

3. Results

The main source of replenishment of groundwater within thestudy area is atmospheric precipitation. In the course of transportof meteoric water from the ground surface through the unsatu-rated soil zone to the aquifer, its chemical composition changesas a result of evapotranspiration, leaching of chemicals from thetrench body, and other bio-geochemical factors.

When presenting the geochemical data below, more detailedconsideration is given to data collected in 2004–2005. This is be-cause monitoring studies in 2004–2005 used the newly con-

structed LAB and WS multilevel profiles (which provided betterspatial resolution of geochemical plumes), and addressed a largerlist of chemical constituents and parameters in groundwater. Onthe other hand impact of the waste site on groundwater was morepronounced in 2004–2005 compared to the later period (see Sec-tion 3.4 for more detail). The data collected in 1998–2002 (1995profile) and in 2006–2008 (LAB and WS profiles) are still usefulfor statistical treatment and for analyzing long-term trends in geo-chemical influence.

3.1. Chemical composition of groundwater upstream from the trench

Data on typical chemical composition of groundwater from thetop part of the aquifer (eolian sand layer) upstream from trench no.22T are shown in the Table 1 (first column). The upstream ground-water is characterized by rather low concentrations of all dissolvedions and close to neutral (or slightly acidic) pH values.

The eolian sands (of predominantly quartz mineralogy) com-posing the unsaturated zone and upper part of the aquifer at theexperimental site represent a low-reactivity environment. There-fore, it appears that the chemistry of groundwater in the ‘‘back-ground’’ aquifer conditions is determined mainly by evapo-transpiration (concentration) of meteoric waters. This hypothesisis supported by simple balance calculations presented in Table 1.

The data on composition of meteoric waters (rainfall, snow) areshown in the second column of Table 1, while the third columngives values of ‘‘concentration factors’’, representing ratios of back-ground aquifer concentrations of major ions to respective meteoricwater concentrations. The resulting ‘‘concentration factors’’ (seeTable 1) for most major ions are close to the independently esti-mated ‘‘evapo-transpiration coefficient’’ (or precipitation concen-tration factor), representing the ratio of net yearly precipitation(R, mm) to net yearly infiltration recharge rate (e, mm).

Kevap ¼ R=e ¼ 2:5ð�0:5Þ:

The above given Kevap value was estimated from groundwaterrecharge analysis using the water table fluctuation method (Helayand Cook, 2002) and unsaturated zone regime studies carried outin the course of the Chernobyl Pilot Site project (Bugai et al., 2001).

The low values of ‘‘concentration factors’’ for NO�3 and HCO�3 canbe explained by non-conservative behavior of these ions in theaquifer.

The preliminary analysis has shown that chemical compositionof groundwater downstream from the trench substantially differsfrom samples taken from upstream monitoring wells. In order toelucidate the geochemical impact of the waste site on the aquifer,

Fig. 4. Frequency histograms for Ca concentrations in groundwater for the initial (a) and log transformed (b) data sets.

1324 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

cross-correlation analyses of concentrations of chemical constitu-ents in groundwater have carried out, as described in the nextsection.

3.2. Cross-correlation analysis of concentrations of chemical elementsin groundwater

In order to quantify the geochemical impact of the waste trenchon the aquifer it was decided to sub-divide the groundwater sam-ples (or observation wells) into two categories (‘‘hydro-chemicalfacies’’): (1) influenced by trench no. 22T, and (2) representativeof ‘background’ aquifer conditions.

Table 2Cross-correlation coefficients for log transformed* concentrations of chemicalelements in groundwater of experimental site for the whole groundwater monitoringsample set (for the period 1998–2005). Highly significant (P < 0.001) values ofcorrelation coefficients are marked in bold.

90Sr Na+ K+ Ca2+ Mg2+ Cl� NO�3 SO2�4

pH

1.00 �0.30 0.71 0.45 0.0001 0.06 0.69 �0.04 �0.43 90Sr1.00 0.11 0.4 0.65 0.59 �0.2 0.48 0.26 Na+

1.00 0.68 0.17 0.17 0.65 0. 39 0.25 K+

1.00 0.52 0.44 0.55 0.23 �0.15 Ca2+

1.00 0.49 0.26 0.59 0.15 Mg2+

1.00 0.29 0.47 0.01 Cl�

1.00 �0.03 �0.33 NO�31.00 �0.05 SO2�

4

1.00 pH

* Logarithmic transformation was applied to all parameters except pH.

Observation wells influenced by the trench can be establishedusing ‘‘fingerprint’’ (or ‘‘indicator’’) chemical elements, which orig-inated from the trench in elevated (or contrasting) concentrationscompared to background aquifer concentrations. The influence oftrench no. 22T on the aquifer is first of all manifested in elevated90Sr concentrations in groundwater. Therefore, it is thought thatchemical constituents leached from the trench in elevated concen-trations should be correlated with 90Sr.

A correlation matrix of all variables was produced and thenexamined for significant relationships to explore this question,which is a common ‘‘first step’’ in many data analyses. To meetassumptions of normality, which are important for assessingstatistical significance of correlations (Walpole et al., 2007),

Table 3Cross-correlation coefficients of log transformed* concentrations of chemical ele-ments in groundwater samples from the ‘1995’ multilevel well profile (for the period1998–2000). Highly significant (P < 0.001) values of correlation coefficients aremarked in bold.

90Sr Na+ K+ Ca2+ Mg2+ Cl� SO2�4

pH

1.00 0.15 0.91 0.85 0.04 �0.028 0.59 �0.22 90Sr1.00 0.31 0.49 0.77 0.79 0.89 0.14 Na+

1.00 0.72 0.15 0.39 0. 43 �0.001 K+

1.00 0.57 0.55 0.72 �0.19 Ca2+

1.00 0.64 0.86 0.082 Mg2+

1.00 0.695 0.19 Cl�

1.00 �0.11 SO2�4

1.00 pH

* Logarithmic transformation was applied to all parameters except pH.

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1325

geochemical data (except pH) were log transformed prior toanalysis (Fig. 4). Next, Pearson product-moment correlationcoefficients (correlation coefficients) were calculated for allpossible pairs of log transformed hydro-chemical parameters(concentrations of 90Sr, major anions, cations) and pH (Table 2).The accepted level of significance was P < 0.001.

Examination of the resulting cross-correlation matrix for thewhole geochemical sample set corresponding to the period be-tween 1998 and 2005 (see Table 2) shows significant positive cor-relation of 90Sr with three components: NO3, K and Ca. The highestcorrelation coefficient is with NO3 (r = 0.69). Cross-correlationcoefficients between NO3, K and Ca are also important (r = 0.55and higher). A significant negative correlation of 90Sr with pH(r = �0.43) was also noted.

One other interesting observation from Table 2 is the significantcorrelation coefficients of Cl with Na, Mg and SO4 (r = 0.47 and

Fig. 5. Distribution of 90Sr (a), Ca (b) and K (c) in aquifer c

higher), as well as between each of these components. Chlorineis known to be an element which originates mainly from rainfall,and which is conservative in the subsurface environment. Changesin Cl concentrations are usually governed by evapotranspirationprocesses. Therefore, high correlation coefficients of Na, Mg andSO4 with Cl indicate that concentrations of these elements ingroundwater are also likely to be mostly determined by the evapo-transpiration process.

The Ca concentration represents a special case. It has significantpositive correlations (r = 0.45 and higher) both with the elementsof the ‘‘90Sr pool’’ (90Sr, K, NO3), as well as with some elements ofthe ‘‘Cl pool’’ (Cl, Na, Mg and SO4; r = 0.4 and higher).

To summarize, based on cross-correlation analysis of the geo-chemical data set for the period between 1998 and 2005, it canbe concluded that NO3, Ca and K can be used as trench no. 22T ‘‘fin-gerprint’’ species. Also, pH in the zone of the trench influence is

ross-section in wells of 1995 profile in October 1999.

1326 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

lower, than in the background aquifer zone. In contrast, the princi-pal source for Na, Mg and SO4 ions is likely meteoric water, andvariations in their concentrations are mainly caused by the eva-po-transpiration process. The Ca concentration in groundwaterseems to be equally influenced by both factors (by the trenchsource-term and by atmospheric inputs).

Fig. 6. Distribution of 90Sr (a), Ca (b) and NO3 (c) in aqui

Consideration of a similar correlation table for the monitoringwell profile 1995 (where the geochemical data set encompassesthe early project period of 1998–2000; Table 3) shows that corre-lation coefficients of 90Sr with Ca and K are noticeably higher(0.85–0.91) than for the whole sample set (Table 2), which sug-gests a more pronounced geochemical influence of the waste

fer cross-section in wells of WS profile in May 2005.

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1327

trench on the groundwater composition in the ‘‘1995’’ profile in1998–2000. In addition Table 3 shows a positive correlation of90Sr with SO4 (0.59; this correlation however is significant at a low-er level P < 0.01), which suggests that in the early period the trenchmay have also served as a source of SO4 to the aquifer (along withatmospheric inputs, which are evidenced by a high correlationcoefficient of this anion with Cl). Unfortunately, NO3 concentra-tions in groundwater were not systematically measured in1998–2000, and, therefore, no conclusions can be reached regard-ing the behavior of this element in wells of the 1995 profile.

In the later period (2006–2008) concentrations of ‘‘indicator’’elements downstream from the trench gradually decreased,approaching background values (see Section 3.4 for more discus-sion), and correlations between chemical parameters became morevague and less statistically significant.

Data for cross-correlation analysis were further checked by con-struction of two-dimensional distributions of trench ‘‘fingerprint’’elements in the aquifer cross-sections. For example, plots of 90Sr,Ca, K and NO3 distributions in groundwater for 1995 and WS wellprofiles (where geochemical plumes are most pronounced) areshown at Figs. 5 and 6.

An example of groundwater pH distribution in wells of the WSprofile in October 2005 showing a well-defined acidic plume, ispresented in Fig. 7.

The distribution patterns of chemicals are consistent with thehydrodynamics of groundwater flow below the trench and visuallycorrelated with the 90Sr plume. It should be noted, that the back-ground 90Sr activity in the top part of the aquifer, of the order ofn � 100 Bq/L, is due to vertical migration into the unsaturated zonefrom the topsoil layers containing residual contamination. Inaccordance with groundwater flow patterns, the radionuclide andchemicals plumes show slightly sinking with downstream distancefrom the trench, being displaced from the top by ‘‘clean’’ infiltra-tion recharge water. The downstream edge of the geochemicalplume reaches the interface zone between the eolian and alluviallayers. Here, dissolved 90Sr encounters alluvial sediments with ahigher adsorption (and retardation) capacity compared to eoliansediments.

Using information on the two-dimensional distribution of ‘‘indi-cator’’ – parameters (90Sr, NO3, Ca, K and pH) in aquifer cross-sec-tions, and taking into account the hydrodynamics of groundwaterflow, monitoring wells have been subdivided into two categories:(1) ‘‘influenced’’ by the trench, and (2) representing ‘‘background’’aquifer conditions.

Fig. 7. Distribution of groundwater pH in aquifer cr

The resulting classification of multi-level wells for the LAB pro-file is shown at Fig. 8. The figure also shows the steady-state flowlines constructed using the groundwater flow model for this cross-section described in Bugai et al. (2001). The well classificationbased on ‘‘indicator’’ species is generally consistent with thegroundwater flow net presented at Fig. 8. Similar schemes for theWS and 1995 profiles are shown on Figs. 5a and 6a. While forthe 1995 and LAB profiles deepening of the plume of chemical ele-ments originating from the trench to the alluvial layer is observed(see Fig. 8), in the case of the WS profile sub-horizontal transport ofdissolved species in the eolian layer is seen (see Figs. 5a and 6a).Such discrepancies in plume migration patterns may be causedby lower hydraulic conductivity of alluvial sediments at the WSlocation compared to the 1995 and LAB locations.

3.3. Geochemical impact of the waste trench at the aquifer

Data on chemical impact of trench no. 22T on groundwater aresummarized in Table 4 and Fig. 9. Plotted concentrations representmedian values on the respective sets of observation wells from the‘‘influenced’’ and ‘‘background’’ aquifer zones.

Data from Table 4 and Fig. 9 generally confirm conclusions fromcross-correlation analysis. In all multilevel well profiles the influ-ence of the trench leads to an increase of mean Ca, K and NO3 con-centrations in the geochemical plumes downstream from thetrench. In wells of the 1995 profile, an increase of SO4 downstreamfrom the trench is also observed, and in the WS wells the trenchcauses increased Mg concentrations in groundwater.

Somewhat elevated Cl� concentrations in the aquifer in thezone of influence of the trench in the 1995 and WS profiles canbe explained by Cl concentration (‘‘entrapment’’) in the root zoneof trees inside the trench, and may be indicative of higher roottranspiration rates in the trench soil compared to background con-ditions. It is known that Cl is excluded to a large extent by rootsduring adsorption of water and nutrients, and its concentrationin soils rises as a function of evapotranspiration (Grimaldi et al.,2009; Moore et al., 2010). Preferential penetration of plant rootsto the trench is not surprising as the organic soil in the trenchserves as a repository for plant nutrient species.

Median plume concentrations of Ca (20.7 mg/L) and K (4.2 mg/L)in the 1995 profile are noticeably higher than in LAB (Ca – 10.5 mg/L; K – 1.5 mg/L) and WS profiles (Ca – 14.8 mg/L; K – 2.35 mg/L).This reflects the fact that plotted data for the 1995 profile

oss-section in wells of WS profile (May 2005).

Fig. 8. Aquifer cross-section at the LAB monitoring site showing positions of screens of multilevel monitoring wells and groundwater flow lines obtained by means of the flowmodel described in Bugai et al. (2001).

Table 4Median concentration values (mg/L) of major ions in groundwater samples from different zones of the aquifer. (Values in parentheses represent 0.1–0.9 percentile range.)

Ion 1995 profile (1998–2000) LAB profile (2004–2005) WS profile (2004–2005)

Background Influenced by the trench Background Influenced by the trench Background Influenced by the trench

Na+ 0.9 1.4 1.65 1.0 1.0 1.15(0.4–1.9) (0.6–2.4) (0.7–3.6) (0.5–3.2) (1–3.5) (0.7–1.6)

K+ 1.7 4.2 1.45 1.5 1.0 2.35(0.9–2.5) (3–8.9) (1.0–2.0) (1.0–2.1) (0.9–1.1) (1.7–4.2)

Ca2+ 5.4 20.7 7.2 10.5 5.3 14.8(3.2–15.8) (13–47.7) (5.1–9.6) (4.9–20) (4–7.7) (5.3–22.4)

Mg2+ 0.9 1.5 1.7 1.5 1.2 2.7(0.4–4.8) (0.8–3.3) (0.7–2.7) (0.8–3.3) (0.9–2.3) (0.9–5.5)

Cl� 2.7 3.5 3.5 3.5 3.7 4.25(2.5–2.9)* (3.2–4.8)* (2–5.7) (2.7–4.3) (2.5–4.2) (3–5.4)

SO2�4

16 38.9 20.2 21.4 16.3 15.7

(13.6–22.6)* (22.2–46.6)* (10.9–33.3) (9.1–41.3) (9.3–26.2) (5.7–27.8)HCO3 No data No data 9.8 11 10.9 8.2

(6.1–24.5) (6.7–53.7) (5.3–15.5) (6.3–12.2)NO�3 No data No data 1.3 2.7 0.7 38.4

(0.5–2.4) (0.7–5.3) (0–1.6) (10.9–65.3)pH 5.5 5.5 5.9 5.85 5.9 5.6

(5–5.6) (5–5.6) (5.6–6.6) (5.5–6.6) (5.3–6.2) (5.1–6.2)

* Median values are based on a small number of observations.

1328 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

correspond to the earlier period (1998–2000) than for the LAB andWS profile (2002–2005) (see Section 3.4 for more discussion).

One other important observation is that concentrations of NO3

in the LAB profile (median plume concentration 2.7 mg/L) aremuch lower than in WS profile (median plume concentration38.4 mg/L). The LAB profile was characterized also by lower 90Srconcentrations and somewhat higher groundwater pH values thanthe WS profile. (Maximum 90Sr concentration in the LAB profile in2004–2005 was 6190 Bq/L, while in WS profile 90Sr activityreached 24,920 Bq/L.)

The data on the chemistry of the trench pore solutions at theLAB and WS sites provided by soil water samplers (SWS) (Table5) are generally consistent with the groundwater monitoring dataon aquifer chemistry downstream from the trench (see Table 4,Fig. 9), and also show distinctive geochemical patterns at the LABand WS sites.

The geochemical discrepancies between the LAB and WS siteslikely resulted from differences in the quality of the trench soilsubstrate and from the related differences in the leaching processat different locations of trench no. 22T.

This explanation is supported by the fact that 137Cs activity intrench no. 22T at the WS site is noticeably higher than at theLAB site (see data in Bugai et al., 2005). The radioactivity disposedin trench no. 22T was mainly associated with contaminated vege-tation and topsoil (organic materials). The higher 137Cs activity ofsoil implies that the organic matter content of waste at the WS siteis higher than at the LAB site (see also data on correlation between137Cs and Corg,extr presented in Section 2.1). The higher radioactivityof waste also resulted in higher activity of 90Sr in the trench poresolutions and groundwater in the WS profile.

The above-discussed explanation was confirmed by compara-tive characterization of trench soil from LAB and WS locations

(a) 1995 well profile (1998-2000)

(b) LAB well profile (2002-2005)

(c) WS well profile (2004-2005)

Fig. 9. Median values of concentrations of major ions for different ‘geochemicalfacies’ of groundwater in the vicinity of the trench no. 22-T. (Vertical bars represent0.1 and 0.9 data percentiles).

Table 5Concentrations of major ions (mg/L), pH and 90Sr (Bq/L) in soil water samplers (SWS)installed in trench no. 22T at LAB and WS sites (May 2007).

Constituent LAB site* mean (min–max) WS site** mean (min–max)

Na+ 2.5 (1.4–3.6) 0.8 (0.5–1.0)K+ 1.1 (0.7–1.5) 1.3 (0.7–1.7)Ca2+ 19.0 (18–20) 8.1 (3–15.4)Mg2+ 1.2 (1.1–1.3) 1.4 (0.6–2.4)NHþ4 0.35 (0.3–0.4) 0.32 (0.2–0.5)Cl� 3.4 (2.8–3.9) 2.9 (2.8–3.2)

SO2�4

31.8 (25–38.7) 14.4 (13.2–15.2)

HCO3 10.1 (9.1–11) 2.0 (1.2–4.8)NO�3 2.7 (2.4–2.9) 12.9 (±14.4)pH 5.25 (5.2–5.3) 4.63 (4.6–4.7)90Sr 6580 (6570–6600) 13,160 (8670–15,600)

* Mean value of the three samples from depths of 1.25 m, 1.75 m and 2.25 m.** Mean value of the two samples from depths of 1.25 m and 2.25 m (other SWS didnot yield solution).

Table 6Comparative characteristics (mean values and min–max range) of trench soil from theLAB and WS sites (June 2006).

Site* Weight loss on ignition (%) pH of ‘‘soil–water’’ suspension**

LAB 3.2 (2.6–3.9) 5.1 (4.9–5.4)WS 6.3 (3.1–8.3) 4.7 (4.6–4.75)

* Sample collection depths 1 m, 1.5 m and 2 m.** Suspension of the 40 g soil sample at field water content with 50 mL distilledwater.

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1329

carried out in June 2006 (Table 6). In fact, the trench soil from theWS site has an organic matter content which is twice as high andhas a more acidic pH of ‘‘soil–water’’ suspension. The data on pHfor the trench soil water extracts (Table 6) are in good agreementwith the SWS geochemistry sampling data (Table 5).

In addition to major ions, concentrations of some trace ele-ments, in particular Al, Mn and stable Sr (88Sr) in the zone of influ-ence of the trench, are an order of magnitude higher compared tobackground aquifer concentrations. In particular, based on moni-toring studies in 2006–2007 mean 88Sr concentrations in the geo-chemical plume in the LAB and WS well profiles were 57–104 lg/L,while mean background concentration of this element was 35–40 lg/L.

The bio-degradation of the buried organic materials in thetrench is manifested also in elevated concentrations of dissolvedorganic (DOC) downstream from the trench (Fig. 10).

3.4. Long-term trends of evolution of chemical composition ofgroundwater

Regime monitoring data for the project period of 1998–2008shows a progressive decrease of concentrations of the leached‘indicator’ – ions accompanied by an increase of pH in the chemicalplume emerging from the trench.

The long-term trends in groundwater geochemistry are illus-trated below by the dynamics of the plume-averaged chemical ele-ment concentrations in groundwater:

CmeanðtÞ ¼1N

XN

i¼1

CiðtÞ;

where N is the number of downstream wells influenced by thetrench in the aquifer cross-section, and Ci(t) is element concentra-tion in ith well. The behavior of Cmean(t) is representative of thedynamics of mass of a chemical element in groundwater down-stream from the trench.

The longest observation series (starting from 1998) are avail-able for the 1995 well profile. These data show mean Ca concentra-tion in the trench plume decreased from 41 mg/L in 1998 to 9 mg/L(close to the background level) in 2002 (Fig. 11a). In parallel an in-crease was observed of the groundwater pH in observation well 6-95-1 (immediately downstream from the trench) from 4.7 in 1998to �5.8 in 2008 (Fig. 11b).

Same trends are observed for Ca concentrations downstreamfrom the trench in the LAB and WS profiles, where Ca concentra-tions in 2008 approached background levels (Fig. 12a).

The mean NO3 concentration behaves the same way in thetrench geochemical plume in the WS profile, decreasing from44.5 mg/L in 2004 to <10 mg/L in 2008 (however mean NO3

Fig. 10. Distribution of dissolved organic C (DOC) in aquifer cross-section in wells of WS profile (May 2005).

(a) Mean plume Ca concentration in 1995 profile in 1998-2002

(b) Groundwater pH in well 6-95-1 (yearly averaged values)

Fig. 11. Long-term trends in chemical composition of groundwater in 1995 profile downstream from trench no. 22T.

1330 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

concentrations downstream from the trench are still much higherthan the background level of 1.3 mg/L) (Fig. 12b).

In addition to a long-term trend, the graph of mean plume Caconcentration in the wells of the LAB profile in 2004–2008 (seeFig. 12a) shows the major influence of short-term fluctuations ofgroundwater table on aquifer chemistry (Van Meir et al., 2009).

The increase of mean Ca concentration in the geochemicalplume emerging from the trench in 2005 was caused by flooding

of the trench bottom at the LAB cross-section by an extremely highgroundwater table in spring 2005 (when the groundwater levelmaximum was observed for the whole observation period; seeFigs. 2 and 13). This extreme event caused a drastic increase of ele-mental and 90Sr concentrations in wells sampling the upper part ofthe aquifer immediately downstream from the trench (exampledata for Ca in well 19-00-1 are shown at Fig. 13; location of screensof wells in LAB profile is shown in Fig. 7). The deeper wells have not

(a) Mean plume Ca concentration in LAB profile in 2004-2008

(b) Mean plume NO3 concentration in WS profile in 2004-2008

Fig. 12. Trends in chemical composition of groundwater in LAB and WS profilesdownstream from trench no. 22T in 2004–2008.

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1331

been influenced by this event (e.g., well 19-00-2 at Fig. 13). As thegroundwater level regime stabilized in the following years, thegeochemical signal returned to the decreasing trends.

4. Discussion

4.1. Overview of factors controlling chemical leaching losses from thetrench

The groundwater geochemistry downstream from trench no.22T was apparently controlled in 1998–2008 by bio-geochemicalprocesses inside the trench (such as decomposition of organic

Fig. 13. Calcium concentrations in wells 19-00-1 and 19-00-2 in LAB profile as influencescreens).

materials, microbial and plant root respiration and nutrient up-take) and resultant leaching of chemical species and radionuclidesfrom the trench soil to groundwater.

The ‘indicator elements’ (Ca, K and NO3) leached in elevatedconcentrations from trench no. 22-T are also known as basic plantnutrient species. In undisturbed forest conditions these elementsare stored in the exchange complex of forest soils (cations) or areincorporated to soil humus (N), and are involved in bio-geochem-ical cycling in the ‘soil–plant’ system.

Unfortunately, data on initial characteristics of the wastes dis-posed in trench no. 22T are not available, as the radioactiveclean-up work in 1987–1988 was not properly documented. How-ever some estimation can be made based on general characteristicsof the site vegetation cover prior to clean-up, and based on thegross characteristics of waste in the Red Forest waste dumps.

The territory adjacent to the ChNPP was covered at the time ofthe accident by 30–40 a-old pine forests (P. silvestris) (Kozubovet al., 1989). The area of complete mortality of pine forest due toradiation at the ChNPP site at the beginning of 1988 was approxi-mately 500 ha, with the following amounts of dead biomass indifferent compartments: tree stems – 99,600 m3 (200 m3/ha),branches – 9050 m3 (18 m3/ha), and needles – 3960 m3 (8 m3/ha)(Vasilchenko et al., 1989).

Ilyichev and Akhunov (1992), who carried out characterizationof Red Forest waste dumps in 1991, estimated the total amountof disposed wastes at 1350 m3/ha. By comparing the ‘‘waste sitearea–waste volume’’ ratio they concluded that the trenches con-tained an approximately 15–20 cm thick layer of initially contam-inated topsoil. The above waste volume estimates shows thatvegetation remnants (stems, branches, needles) initially consti-tuted <17% of the buried waste volume, while the rest of the wastein the Red Forest dumps was composed of local soil.

Analogy can be clearly seen between the leaching of chemicalspecies from trench no. 22T and case studies describing chemicalnutrient leaching losses from soils after forest clear-cutting (Vito-usek and Melillo, 1979; Jordan, 1986). Indeed, conditions at theRed Forest site after the Chernobyl accident and following clean-up works, which included removal and burial of dead pine trees,resembled conditions occurring after clear-cutting in a pine standin many aspects. Trench no. 22-T itself, as shown above, repre-sented the repository of forest topsoil mixed with organic debris.

Commonly accepted ecological and environmental factors caus-ing increased leaching losses from forest soils after clear-cuttinginclude (Jordan, 1986, and references therein):

(i) reduced transpiration and increased infiltration rate throughsoil (as a consequence of removal of the vegetation cover);

d by the trench bottom flooding event in May 2004. (See Fig. 7 for position of well

Fig. 14. Correlation between Ca2+ and HCO�3 ions in LAB well profile.

1332 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

(ii) elimination of biogeochemical cycling of nutrient elementsby vegetation;

(iii) addition to topsoil organic substrate available for minerali-zation (dead roots, branches, etc.).

The usual picture in a longer-term period (e.g., after 2–3 a ormore) is that following the period of soil depletion, nutrient uptakeby the newly developing vegetation on the top of the clear-cut areareestablishes control over the leaching losses (Vitousek and Mel-illo, 1979).

The factors and mechanisms listed above are pertinent to thecase of trench no. 22T. However the bio-geochemical system oftrench no. 22 along with similarities has a number of distinctiveand unique peculiarities, which are discussed in more detail below.

4.2. Groundwater acidification mechanisms

The rates of leaching of chemical species from trench no. 22T togroundwater were likely controlled by the process of mineraliza-tion of organic materials inside the trench leading to acidificationof the soil pore solutions.

The common leaching agents responsible for cation (i.e., Ca2+,Mg2+, K+) losses from forest soils are H2CO3, HNO3 and organicacids (Jordan, 1986; Johnson and Cole, 2005). The H2CO3 is formeddue to release and dissolution in soil water of CO2 resulting fromdecomposition of organic matter and other soil respiration pro-cesses (i.e., plant root respiration, etc.). The nitrification and asso-ciated acidification process occurs, when NH4 from themineralization of organic matter is oxidized to NO3 (Vitousek andMelillo, 1979). This leads to release from soil of NO3 and cations,which are exchanged for H+ on soil surfaces.

The geochemical data for trench no. 22T suggest that both pro-cesses may play important roles in the studied case.

In particular, the H2CO3 leaching process appear to influenceaquifer geochemistry at the LAB profile. The ion-exchange processof H2CO3 with Soil Matrix (SM) can be represented by the followinggeneric equation:

nH2CO3 þ IðSMÞ $ Inþ þ nHðSMÞ þ nHCO�3 ; ðIÞ

where In+ is a general term for K and/or Ca (Mg, Sr etc.) ions, andn = 1, 2. (The same mechanism applies to radionuclides in cationicform, e.g. 90Sr2+).Eq. (I) predicts release of additional HCO�3 ions tosoil solution. The corresponding stoichiometric ratio for leachingof Ca2+ ions with H2CO3 is:

D½HCO�3 � ¼ 61=20 � D½Ca2þ�; ðIIÞ

where D[Ca2+] and D½HCO�3 � are increments of ionic species concen-trations above the initial levels in solution, and 61/20 is the ratio ofthe equivalent weight of HCO�3 to the equivalent weight of Ca2+.

The observed values of Ca (dominant cation) and HCO�3 (domi-nant anion) in the LAB well profile are generally consistent with Eq.(II), as shown by the correlation plot in Fig. 14. The [Ca2+]–[HCO�3 ]correlation slope factor (i.e., 2.7) is close to the coefficient pre-dicted by the stoichiometric Eq. (II).

On the other hand, in the WS profile, where the NO�3 is the dom-inant anion in solution, soil leaching appears to be controlled bythe nitrification process. As already discussed in Section 3.3 suchdifferences between the LAB and WS sites can be related to the dif-ferences in the quality of the organic substrate buried in the differ-ent locations of trench no. 22T.

4.3. Peculiarities of organic matter decomposition in trench no. 22

Though a large amount of organic matter (OM) decompositionstudies has been carried out over the last several decades, these

studies have usually emphasized near-surface litter horizons innatural ecosystems and decay rates in the first few years, whenthe litter mass is lost most rapidly (Gregorich and Janzen, 2000;Currie et al., 2002; Berg and McClaugherty, 2008). Therefore,results of these studies are not fully instructive with regard tothe conditions in trench no. 22T, where decomposition evolvedin specific physical and chemical conditions (burial of initial OMat 0.2–2.5 m depth in soil; high radiation levels, etc.), and wherethe monitoring dataset encompasses the period from 10 to 20 aafter disposal of waste.

The principal factors regulating organic matter decompositionrates in soils include: substrate quality, physico-chemical environ-ment and decomposer community (Swift et al., 1979). Below thesefactors are briefly discussed in the context of trench no. 22T.

4.3.1. Organic substrate and physical environmentThe initial OM inside the trench likely represented a complex

mixture of pine stem wood, branches, needles, litter and forest soil(see Section 4.1), with maybe a relatively small amount of othervegetations remnants (herbs, bushes, etc.). It is known that pine-wood and needles represent rather stable OM, while herbaceousmaterial is composed of faster degradable OM. In its turn the woo-dy debris generally decays more slowly than pine needles and fineroot litter, and smaller diameter pieces decompose faster than thelarge diameter ones (Currie et al., 2002). Mineralization rates of soilOM are known to increase with increasing moisture content up tofield capacity but decline with greater moisture. Decompositionrates generally increase with increasing temperatures (Currieet al., 2002, and references therein).

The yearly-averaged values of soil temperature in the trench isclose to 11 �C; the upper and intermediate depth soil layers arecharacterized by soil water content values of 0.1–0.15 m3/m3

(Dewiere, 2002), and thus provide generally suitable conditionsfor the decomposition process. The deeper layers of the trenchmay have, however, impeded O2 access (or even an anaerobic envi-ronment) compared to surface conditions. Also these soil layersperiodically experience water content values in excess of fieldcapacity (up to 0.3–0.32) due to the closeness of the groundwatertable.

Thus, with respect to composition of organic substrate as wellas physical parameters in the subsurface, trench no. 22T likely rep-resented a spectrum of conditions from those favoring relativelyhigh decomposition rates (e.g., fine litter material with larger N:Cratios, good O2 access and moisture content conditions in near-sur-face and intermediate-depth soil layers) to less favorable ones (e.g.,larger woody debris composed of lignified pine tissues; deeper,possibly anaerobic and higher water content environments).

Table 7Estimates of cation leaching losses by infiltration flux from trench no. 22T.

Element Initialconcentrationin trench soil*

Concentrationin infiltrationwater, mg/L**

Yearly leachingloss, g/(m2 a) (%of init. trenchinventory)

Losses over20 a,% ofinit. trenchinventory

Ca 1–2 meq/100 g

20.7 5.2 9–19

(0.5–0.9%)K 30–76 mg/kg 4.2 1.1 10–25

(0.5–1.3%)

* Data from Arkhipov et al. (1989) for a typical range of concentrations of Ca and Kin undisturbed pine forest soils at the ChNPP site.** Data from the 1995 well profile corresponding to the period of 1998–2000.

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1333

The relatively slow dynamics of OM degradation (at least for thesignificant part of OM in the trench) is indirectly confirmed by thecontinued impact of the trench on groundwater quality some 20 aafter waste disposal (see Section 3.3), which is supposedly con-trolled by OM decomposition.

4.3.2. Impact of radioactivityRelatively little is currently known about the possible impact of

radioactivity on the decomposition process. Soil radioactivity cancause slowing of decomposition presumably due to toxicity to-wards the decomposer microbe community. Berg and McClaugher-ty (2008) who reviewed the few studies carried out so far on thissubject did not identify unambiguous evidence of radioactivity im-pact on decomposition. In the reviewed cases, gamma-irradiationstrong enough to kill vegetation had much less of an effect ondecomposition than did moisture level and litter chemical quality.

A microbiological study by Chapon et al. (2011) of the soil fromtrench no. 22T did not identify significant differences in diversity ofbacterial communities between the highly-contaminated samplesfrom the trench and ‘control’ low-activity soil samples from thenearby area. Thus no evidence was found of the impact of radioac-tivity on the soil bacterial community. This finding is coherent withthe above-cited conclusions of Berg and McClaugherty (2008).However the study of Chapon et al. (2011) was merely a firstmicrobiological assessment of trench no. 22T, and did not focuson the specific decomposer microbe communities. Therefore, fur-ther microbiological and geochemical studies are needed to under-stand the possible impact of radioactivity on the decompositionprocess in trench no. 22T.

4.3.3. Time dynamics of decompositionDecomposition of most plant residues is well described by a

two-phase kinetic model: a rapid initial mass loss (easily decom-posable ‘liable’ material), followed by a slower mass loss rate (re-calcitrant chemical compounds, e.g. lignified plant tissues)(Gregorich and Janzen, 2000). In the later stages once the litterhas been transformed to humus or mineral soil OM, the rate of de-cay slows, being constrained by the recalcitrant chemical proper-ties of the residual material (Berg and McClaugherty, 2008).

Some representative values of first order decomposition rateconstants for the pine OM are from k = 0.18–0.23 a�1 (half life 3–3.8 a) for the fine litter material (Pausas, 1997), to k = 0.016–0.096 a�1 (half life 7.2–42 a) for the large woody debris (Currieet al., 2002 and references therein).

It should be pointed out that in the case of trench no. 22T wasteburial occurred quite long ago (1987). The ‘fast’ phase of buried OMdecomposition likely lasted for first several years after the creationof the trench. During the period of the reported project studies(1998–2008) apparently a geochemical environment resultingfrom less intensive and gradually attenuating process of transfor-mation and humification of the secondary OM degradation prod-ucts inside the trench was observed.

The groundwater monitoring data collected in the course of theproject (see Section 3.4) indeed indirectly indicate attenuation ofOM transformation processes inside the trench (and a correspond-ing decrease in groundwater acidification and cation leaching),which is consistent with the above-discussed general OM decom-position patterns.

4.4. Infiltration leaching losses and nutrient element uptake byvegetation

4.4.1. Hydrological leaching lossesApart from geochemical evolution of OM in the trench, changes

in the geochemical regime of the aquifer can be caused by gradualdepletion of the inventory of cations in the exchange complex of

trench soils due to constant leaching by infiltration waters. Esti-mates of leaching losses for Ca and K are presented in Table 7.The initial content of cations in trench soil was calculated usingdata from Arkhipov et al. (1989) on a typical range of concentra-tions of Ca and K in undisturbed pine forest soils at the ChNPP site.The volume of buried soil per 1 m2 of trench area is assumed toequal �2 m3, and the infiltration rate through the soil 250 mm/a.Table 7 uses data on cation concentrations in pore water belowthe trench from the 1995 well profile corresponding to the periodof 1998–2000. The results suggest that the leaching losses from thetrench during the 20-a period (i.e., 1988–2008) may constitute 10–20% for Ca, and 10–25% for K, of the initial trench inventory. Theseare order of magnitude estimates, as the detailed time dynamics ofcation concentrations in groundwater below the trench and otherparameters are not known precisely.

4.4.2. Pine root uptake of nutrient elementsThe forest plantations at the Red Forest site in 1988–1989 con-

sisted primarily of pine trees located in large strips (of about 12 mwidth, comprising 6 rows of pines). The average pine density isestimated at 3300 trees/ha (Thiry et al., 2009). Initially birch treesand bushes were planted between these strips; however their sur-vival was poor (presumably as a result of poor nutrient status ofresidual soils), and inter-strip zones are now colonized by only afew young birches, bushes and grasses. The birch trees growing di-rectly on the trench are characterized by noticeably larger above-ground biomass, which reflects favorable nutrient conditionsprovided by the trench soil. In particular, such a row of birch treesis growing on the top of the trench at the LAB monitoring well pro-file (see Fig. 3).

A useful insight into nutrient element uptake fluxes by pinetrees at the Red Forest site can be found using results of studiesof Goor and Thiry (2004) and Thiry et al. (2009) on 137Cs, K and90Sr cycling in forests in the Chernobyl contaminated areas. A 17-a-old pine forest at the Vetka Site in Gomel district (Belarus) stud-ied by Goor and Thiry (2004) is similar to the Red Forest site withregard to the pine trees age, geographical region, soil type (sandypodzol) and climatic conditions. The estimated uptake rate of Kfor this forest was 0.8 g/m2 a (including immobilization of 0.38 g/m2 a in pine tissues), which is comparable to the hydrologic leach-ing losses from trench no. 22T (see Table 7). The Vetka Site, how-ever, may possibly have differed from the Red Forest site inplantation density, so the given values should be considered asan approximate order of magnitude estimate, if extrapolated tothe Chernobyl Pilot Site.

Study of 90Sr accumulation in the vegetation on the top oftrench no. 22T by Thiry et al. (2009) concluded that the upwardflux of 90Sr through pine root uptake from the trench (0.82%/a ofthe radionuclide pool in the trench in 2005) was at least of thesame magnitude as 90Sr downward migration losses from thetrench (i.e., 0.14–0.5% of inventory/a, as estimated by Dewiere

Fig. 15. Distribution patterns of 90Sr (Bq/L) in the LAB multilevel well profile in 2004–2008.

1334 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1335

et al. (2004) for the period of 1987–2000). Because of geochemicalsimilarities between Sr and Ca (Åberg, 1995), 90Sr can be consid-ered as a tracer of Ca fluxes in the ‘vegetation–trench–aquifer’ sys-tem at the Red Forest site.

Thus, the above-presented data suggest that root uptake of Kand Ca by pine trees was comparable in 1998–2008 to the hydro-logic fluxes from the trench to the aquifer. As biological fluxes tendto increase with the growth of forest and a rise in its biomass pro-ductivity, the element ‘pumping’ and cycling by vegetation is

(a) Mean 90Sr plume concentrations in

(b) Mean 90Sr plume concentrations in w

Fig. 16. Long-term trends in 90Sr concentratio

Fig. 17. Evolution of the conditions of 90Sr migration from trench no. 22T in 1988–200physical values).

imposing limits on nutrient element leaching losses from thetrench.

In addition to nutrient uptake by roots, the forest growth has po-tential to influence the hydrological balance of the site. The canopyinterception and the transpiration by pine forest stands is known topeak at the age of 20–40 a (Murakami et al., 2000, and referencestherein), which suggests that during the period of 1998–2008 thedeveloping forest cover was increasing evapotranspiration (and,respectively, decreasing infiltration) at the Red Forest site.

wells of 1995 profile in 1997-2001.

ells of LAB profile in 2002-2008.

ns downstream from the trench no. 22T.

8. (The arrows depicting hydrologic and radionuclide fluxes are not to scale with

1336 D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338

4.5. Influence of geochemical conditions on strontium-90 migrationfrom the trench

Geochemical conditions in the aquifer caused by the trenchsource term and their evolution in time has important conse-quences for 90Sr migration in the subsurface.

In the geochemical conditions of the site, Sr in solution is as-sumed to be present mainly as an uncomplexed Sr2+ ion (Szenknectet al., 2005). In such conditions, 90Sr sorption on soils and sedi-ments is controlled by the ion-exchange mechanism, and the 90Srsorption distribution coefficient (Kd) value is strongly influencedby presence in pore solutions of other cations, especially Ca andstable Sr, which compete with 90Sr for exchange sites on the soilmatrix (Kipp et al., 1986; Konoplev et al., 1992; Lefevre et al.,1996; Marinin and Brown, 2000; Szenknect et al., 2005).

Laboratory tests and geochemical modeling using eolian sandfrom the ChNPP site by (Szenknect et al., 2005) have shown thatan increase of stable Sr concentration in water from 10�6 to 10�5

mol/L (which corresponds to conditions observed in the aquiferat the Chernobyl Pilot Site) causes a decrease of Sr sorption retar-dation coefficient by a factor of 3. (Other cation concentrations inthe test solution corresponded to ‘background’ aquifer conditionsat Chernobyl Pilot Site). Ardois and Szenknect (2004) studied thesame ‘soil – solution’ system by means of geochemical modelingand established, that the decrease of Ca in solution from2 � 10�3 mol/L (80 mg/L) to 2.7 � 10�4 mol/L (10.8 mg/L) causedan increase of 90Sr Kd from 0.9 mL/g to 14 mL/g (the other cationconcentrations in the modeled case corresponded to the ‘‘mean’’groundwater composition downstream from trench no. 22T:[K+] = 9.8 � 10�5 mol/L, [Na+] = 4.6 � 10�5 mol/L, [Mg2+] = 4.4 � 10�5 mol/L, [Sr2+] = 8.4 � 10�6 mol/L).

Thus, the elevated concentrations of Ca and stable Sr in the geo-chemical plume emerging from the trench likely represented themajor factors promoting mobility of 90Sr in the subsurface environ-ment in the early period following the disposal of radioactivelycontaminated organic debris in 1987.

The monitoring data for the LAB profile show that 90Sr in theaquifer had migrated a distance of �14–16 m downstream fromthe trench by 2002 (Fig. 15a), which is evidence of relatively high90Sr mobility in groundwater during the first 15 a after wastedisposal.

Calibration of the numerical radionuclide transport model forthe waste site accounting for dissolution of fuel particles insidethe trench, radionuclide transport in the unsaturated zone, and uti-lizing Modflow – MT3D for the radionuclide transport in the aqui-fer has given an ‘‘effective’’ (i.e., fitted time-averaged for the periodof 1987–2002) Kd value of 0.5 mL/g (Dewiere et al., 2005; Van Meiret al., 2009). The corresponding 90Sr retardation factor is R = 4, andthe radionuclide migration velocity in the eolian sand aquifer isV90Sr = 2.5 m/a (the real groundwater velocity in the eolian aquiferis estimated at 10 m/a, see Section 2.1).

The in situ determinations of 90Sr Kd by means of sampling sed-iment cores from the contaminated zone of the aquifer carried outin 2000 (Bugai et al., 2002), gave Kd values of 2–3 mL/g (R = 12–17,V90Sr � 0.7 m/a), which indicated increased radionuclide sorptionand retardation compared to the earlier period.

Groundwater observation data for the period of 2004–2008(Fig. 15) show limited tendencies for dispersion and displacementof the radionuclide plume in the subsurface compared to the aboveestimated higher migration rates in the period 1987–2002, whichis consistent with the above discussed data on the evolution ofthe 90Sr Kds.

Eventually, the laboratory batch tests using synthetic ground-water representative of background aquifer conditions gave 90SrKd values in the range 25–43 mL/g (corresponding retardation fac-tor is R = 134–230, V90Sr < 0.1 m/a) (Van Meir et al., 2009). This

suggests that 90Sr sorption and retardation will be further in-creased, as the chemical leaching rates from the trench decreaseand downstream groundwater quality approaches backgroundconditions.

Gradual increase of 90Sr Kds in the aquifer caused by decreasedcompetition from Ca and stable Sr for the exchange sites on the soilmatrix has two-side impacts on radionuclide migration. First, as al-ready discussed, radionuclide migration by advective transport isretarded. Secondly, radionuclide concentration in groundwater de-creases due to redistribution of 90Sr between the liquid and solidphase (increased adsorption on soil matrix) in accordance withthe evolving ion-exchange sorption equilibrium conditions ingroundwater.

In accordance with the above discussion the groundwater mon-itoring data demonstrate clear long-term trends of attenuation ofmean 90Sr concentrations downstream from the trench (Fig. 16).As already mentioned in Section 3.4 flooding of the trench bottomby an extremely high groundwater table in May–June 2005 causeda short-term increase of 90Sr (as well as Ca; see Fig. 12a) concentra-tions in groundwater, followed by a return of concentrations to thelong-term decreasing trend (Fig. 16b).

The scheme summarizing the long-term evolution of the condi-tions of 90Sr migration from trench no. 22T discussed above isshown in Fig. 17.

5. Conclusions

Groundwater downstream of the studied waste trench at theRed Forest site in 1998–2008 showed a specific geochemical re-gime. Along with the elevated content of 90Sr, groundwatershowed high Ca, K, NO3, SO4 and some trace element (stable Sr,Al) concentrations and more acidic pH values compared to ‘‘back-ground’’ aquifer conditions. These changes in groundwater geo-chemistry apparently resulted from decomposition of organicmatter (pine forest remnants, litter, humus containing topsoillayer) buried inside trench no. 22T.

Decomposition of organic matter inside the trench followingthe pine forest debris burial in 1987 caused acidification of ground-water (presumably due to H2CO3 production and nitrificationmechanisms). This resulted in leaching of cations adsorbed onthe exchange complex of soils buried inside the trench.

Monitoring data for the period of 1998–2008 show a progres-sive decrease of concentrations of ‘‘fingerprint’’ ions (Ca, K, NO3)in the chemical plume in the aquifer emerging from the trenchaccompanied by an increase in pH. This long-term trend can be ex-plained by gradual decrease of the inventory, humification of theoriginal OM inside the trench and a corresponding decrease inthe acidification process. In parallel, an increased root uptake ofnutrient chemicals (as well as 90Sr) has been taking place becauseof development of the newly planted pine forest on the top of thetrench.

The identified features of the hydrochemical regime and itstime dynamics have important implications for 90Sr migrationfrom Red Forest waste dumps to groundwater. The progressive de-crease during last decade of concentrations of major cations (inparticular Ca and stable Sr) in groundwater leads to an increaseof 90Sr Kd values in the waste trench and in the aquifer, to an atten-uation of radionuclide concentrations in groundwater and retarda-tion of subsurface migration processes.

Accounting for evolving geochemical conditions in the aquiferappears to be an essential prerequisite for adequately modelingradionuclide transport in groundwater from Chernobyl wastedumps. The groundwater modeling predictions of radionuclidemigration from the Red Forest waste dumps carried out previouslyfor risk assessment purposes (Bugai et al., 1996), assumed

D. Bugai et al. / Applied Geochemistry 27 (2012) 1320–1338 1337

stationary hydrochemical conditions in the aquifer. This has lead toan overestimation of radionuclide release from Chernobyl radioac-tive waste dumps to groundwater and the river network. Theresults presented in this paper may serve also as an argument infavor of Monitored Natural Attenuation (MNA) strategy for radioac-tively contaminated groundwater (IAEA, 2006), which wasdiscussed as a candidate for management strategy for the RedForest waste site in Shestopalov et al. (2006).

Further quantitative geochemical analysis (e.g., using geochem-ical modeling) and integrated ecosystem modeling studies areneeded to gain better founded insights into mechanisms of OMdegradation impact on the aquifer geochemistry, to explain thegeochemical differences between different locations of trench22T, and to predict long-term bio-geochemical interactions andradionuclide fluxes in the Chernobyl Red Forest waste dumps.

Acknowledgements

Studies presented in this paper were supported by French–Ukrainian collaborative radioecology projects ‘‘Chernobyl Pilot SiteProject’’ (1999-2004) and ‘‘Experimental Platform in Chernobyl’’(2005–2008) funded by the French Institute for Radiation protec-tion and Nuclear Safety (IRSN, Fontenay-aux-Roses), by the CNRSand IRSN through the GNR TRASSE (Research action 2008-1B;GDR TRASSE contribution 2011/06), and by Research Theme no.III-11-06 of the Ukrainian National Academy of Sciences. We thankSergei Zvarich for helpful discussions of the mechanisms of thegeochemical impact of trench no. 22T on the aquifer. We also thankLuc Aquilina and two anonymous reviewers for their constructiveremarks and suggestions, which helped to improve the manuscript.

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