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Journal of Environmental Radioactivity 100 (2009) 468–476

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Journal of Environmental Radioactivity

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137Cs, 239,240Pu and 241Am in bottom sediments and surface water of LakePaijanne, Finland

M. Lusa*, J. Lehto, A. Leskinen, T. JaakkolaLaboratory of Radiochemistry, A.I. Virtasen aukio 1, P.O. Box 55, 00014 University of Helsinki, Finland

a r t i c l e i n f o

Article history:Received 1 May 2008Received in revised form2 March 2009Accepted 4 March 2009Available online 11 April 2009

Keywords:CesiumAmericiumPlutoniumBottom sedimentsLake Paijanne

* Corresponding author. Tel.: þ358 9 191 50518; faE-mail address: merja.lusa@helsinki.fi (M. Lusa).

0265-931X/$ – see front matter � 2009 Elsevier Ltd.doi:10.1016/j.jenvrad.2009.03.006

a b s t r a c t

The concentrations and vertical distribution of 239,240Pu, 241Am and 137Cs in the bottom sediments andwater samples of Lake Paijanne were investigated. This lake is important, since the Paijanne area receiveda significant deposition from the Chernobyl fallout. Furthermore Lake Paijanne is the raw water sourcefor the Helsinki metropolitan area. In addition no previous data on the distribution of plutonium andamericium in the sediment profiles of Lake Paijanne exist. Only data covering the surface layer (0–1 cm)of the sediments are previously available. In the sediments the average total activities were 45� 15 Bq/m2 and 20� 7 Bq/m2 for 239,240Pu and 241Am, respectively. The average 241Am/239,240Pu ratio was0.45� 0.14. The 241Am/239,240Pu ratio is lowest in the surface layer of the sediments and increases asa function of depth. The 238Pu/239,240Pu ratio of the sediment samples varied between 0.012� 0.025 and0.162� 0.079, decreasing as a function of depth. The average activity in water was 4.9� 0.9 mBq/m3 and4.1� 0.2 mBq/m3 for 239,240Pu and 241Am, respectively. The 241Am/239,240Pu ratio of water samples was0.82� 0.17. 239,240Pu originating from the Chernobyl fallout calculated from the average total activitiescovers approximately 1.95� 0.01% of the total 239,240Pu activity in the bottom sediments. The averagetotal 137Cs activity of sediment profiles was 100� 15 kBq/m2 and 19.3� 1.4 Bq/m3 in water samples.

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1. Introduction

Deposition of 137Cs, 239,240Pu and 241Am from the Chernobylaccident was very unevenly distributed in Finnish lakes andcatchment areas (Kansanen et al.,1991; Paatero et al., 2002; Ilus andSaxen, 2005; Salminen et al., 2005). This was caused by the differ-ences in the areal rainfall conditions. The highest deposition valuesof 239,240Pu and 241Am in Finland were located in the southwesternand central parts of the country (Paatero et al., 2002; Salminen et al.,2005). In August 1986 239,240Pu and 241Am concentrations in thesurface layer of the bottom sediments (0–1 cm) of Lake Paijannewere 1.5� 0.1–2.5� 0.2 Bq/m2 and 0.58� 0.05–1.6� 0.2 Bq/m2,respectively (Suutarinen et al., 1993). The maximum 137Cs deposi-tion values, 45–78 kBq/m2, were found in the same region (Arvelaet al., 1989). Chernobyl-derived 137Cs deposition is rather highcompared to that from the atmospheric nuclear weapons tests inthe 1950s and 1960s: 1700 Bq/m2 (decay corrected to 1986) (AMAP,1998). Instead the major source of transuranium nuclides in Finlandis from nuclear weapons test fallout (Paatero et al., 2002; Salminenet al., 2005). The total amount of the Chernobyl-derived 239,240Pu

x: þ358 9 191 50121.

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deposited in Finland was approximately half of a percent of theplutonium fallout of the nuclear tests. Similarly the 241Am activityoriginating from the Chernobyl accident was approximately 1.7% ofthe total activity of 241Am in Finland (Salminen et al., 2005).

The radioactivity status of Asikkalanselka, the southernmostbasin of Lake Paijanne, is an important topic since the Helsinkimetropolitan area takes its raw water from this basin. This paperdescribes the present radioactivity situation on the Asikkalanselka,twenty years after the Chernobyl accident. Also the radionuclidedistribution and the radionuclide sources are discussed.

In the hydrosphere the prevailing aqueous species of cesium isthe uncomplexed Csþ ion and changes in the pH and Eh do notaffect the speciation of cesium (Lieser and Steinkopff, 1989).Nonetheless, cesium may be adsorbed on surfaces of colloids andsuspended particles, which deposit onto lake bottoms. The abilityof bottom sediments to bind cesium varies with particle size, and anincrease in 137Cs activity with a decrease in particle size has beenobserved (Lujanien _e et al., 2004). According to Lujanien _e et al. thehighest calculated Kd values, i.e. the equilibrium ratio of 137Cs inparticles compared to 137Cs in water phase, were obtained forparticles smaller than 4 mm. This is because the high adsorption ofCsþ ions is mainly determined by the clay minerals present insediments (Lieser and Steinkopff, 1989). The main mechanism ofthe adsorption is ion exchange and in natural sediments the

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476 469

sorption behavior is dominated by the highly selective exchangesites in clay minerals (Lieser and Steinkopff, 1989). The adsorptionhas been reported to be virtually irreversible (Gutierrez andFuentes, 1996). Although the cations with similar charge and ionicradii are expected to compete with cesium for adsorption sites, thebonding strength found for clay minerals decreases in the sequenceCsþ> Rbþ>Kþ>Naþ (Lieser and Steinkopff, 1989), which is inagreement with the decrease in ionic radius and the increase in thehydration enthalpies.

The single most important property affecting the characteristicof transuranium elements is their oxidation state. Precipitation,complexation, sorption and colloid formation processes depend onthe prevailing oxidation state (Silva and Nitsche, 1995). All fourprocesses contribute to the chemical behavior and environmentaltransport properties of actinides, including plutonium and ameri-cium, in the environment. In aqueous solutions with pH and Ehranges of natural waters plutonium can be present in four oxidationstates: III, IV, V, VI (Choppin, 2006). The main part of the dissolvedplutonium in natural waters exists in oxidation state Pu(V), whilePu(IV) is present in colloidal form (Choppin, 2006). Low pH valuespromote lower oxidation states while higher oxidation statesbecome more general as the pH increases. Pu(III) is possible inprevailing anoxic conditions, e.g. in lower sediments layers.Americium exists in nature in oxidation state III and Am(OH3) sorbsreadily to nearly all surfaces.

2. Study area

Asikkalanselka is situated in the southern part of Lake Paijanne(Fig. 1). The total area of Lake Paijanne is 1100 km2 and that ofAsikkalanselka 52.8 km2. The maximum depth in the Asikkalanselkaarea is 53 m and its mean depth is 10 m. Paijanne is ice-covered an

Fig. 1. Location of Lake Paijanne in Finland and the 137Cs fallout from

average of five months a year from mid December to the beginningof May. The lake is dimictic and oligotrophic.

3. Materials and methods

3.1. Sampling

Sediment samples were taken in 2007 from three sampling sites, one at thedeepest place of the basin at 51 m (samples A1 and A2) and at two other 32 mdeep sites 400 m away from this site, the other to the southeast (A3) and theother to the northwest (A4) (Fig. 2). The locations in Finnish National CoordinateSystem and in EUREF-FIN geographical coordinate system are represented inTable 1. Sediment samples were taken with a Limnos gravity corer. The corerconsisted of a series of 1 cm-high rings and an inner tube diameter of 9.3 cm.Sediment samples were divided in situ to 1 cm-thick slices and brought to thelaboratory in plastic containers. In the laboratory the sediment samples werefrozen and freeze-dried. In addition to sediment samples 400 L of surface waterwas collected in 20–50 L plastic canisters from the western shore of Asikka-lanselka for 239,240Pu, 241Am and 137Cs analyses. 8 M HNO3 was added to thewater samples and their pH was adjusted to 1–2. Samples were kept in a cold-storage room (þ4 �C) until analyzed.

3.2. Analysis of americium and plutonium from sediment and water samples

The method for the separation of plutonium and americium from the sedimentsamples included wet ashing, coprecipitation with calcium oxalate and threeextraction chromatography steps using UTEVA�, TRU� and TEVA� resins (EichromIndustries) (Fig. 3). 242Pu and 243Am tracers were used for yield determinations. The243Am tracer used in these analyses contained 2.5% 241Am as an impurity. This wassubtracted from the 241Am peak in the spectra of the samples. Furthermore, blankanalyses were prepared to investigate possible contaminations during the pluto-nium and americium purification steps.

242Pu and 243Am tracers were added to the dried four-gram aliquot samples andthe samples were wet-ashed with concentrated HNO3 and HCl to bring the analytesinto the solution before further chemical separations. Wet-ashed samples werefiltered and plutonium as well as americium were co-precipitated as oxalates asdescribed earlier by Paatero, Outola and Salminen (Paatero, 2000; Outola, 2002;Salminen et al., 2005). 100–200 mg of calcium carrier and 2–4 g of oxalic acid wereused for the precipitation. In the case of the upper parts of the sediment profile

Chernobyl in Finland (fallout map reworked from www.stuk.fi).

Fig. 2. Location of sampling sites A1, A2, A3 and A4 in Asikkalanselka (permission to publish from National Land Survey of Finland).

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476470

calcium oxalate precipitate was formed when lower amounts of calcium carrier(100 mg) and oxalic acid (2 g) were used. In the lower parts, however, higheramounts, 200 mg and 2 g, respectively, were needed to precipitate calcium oxalate,which was probably due to the differences in the matrix concentrations in differentdepths. Calcium oxalate precipitation was used to remove disturbing ions such asKþ, Fe3þ, Al3þ, Ti4þand PO4

3� from the samples (Sidhu, 2006; Paatero, 2000). Theseions remained in solution as the actinides were co-precipitated. In the first placeFe3þ was important to remove since it substantially decreases the retention ofamericium to TRU� extraction chromatography resin.

The precipitate was calcinated overnight and the ashed sample was dissolvedin a small amount of concentrated HNO3 and the solution was evaporated todryness. The residue was dissolved into 3 M HNO3þ1 M Al(NO3)3 solution and0.6 M Fe(SO3)2 and ascorbic acid were added to separate americium and pluto-nium from other interfering nuclides with extraction chromatography resinsUTEVA�, TRU� and TEVA� as described earlier by Salminen et al. (2004, 2005)(Fig. 3). Al3þ was added, because it complexed with phosphate and thereby pre-vented the interference of phosphate to neptunium or thorium uptake to UTEVA�.Fe(SO3)2 reduced Pu(IV) to Pu(III). Ascorbic acid was added to ensure that iron was

Table 1The locations of sampling sites A1, A2, A3 and A4 in the Finnish National CoordinateSystem (FNCS) and in EUREF-FIN geographical coordinate system.

FNCS FNCS EUREF-FIN EUREF-FIN

X (p) Y(i) N/lat E/lon

A1 and A2 6,792,301 3,421,983 61� 130 53.66500 25� 320 39.42200

A3 6,792,059 3,422,278 61� 130 46.06100 25� 320 59.54700

A4 6,792,507 3,421,706 61� 140 0.11800 25� 320 20.55500

in oxidation state II. The solution was poured into a UTEVA� column to separateAm(III) and Pu(III) from tetra- and hexavalent impurities. Am(III) and Pu(III)remained in the eluent while U(VI), Th(IV) and Np(IV) were retained by the resin.Eluent containing Am(III) and Pu(III) was poured to TRU� resin to separateamericium and plutonium from each other. Americium and plutonium was elutedfrom TRU� resin in stages. First Pu(III) was oxidized with NaNO2 to Pu(IV) whichremained in the column while, Am(III) was eluted with 4 M and 9 M HCl.Thereafter, plutonium was reduced back to Pu(III) with TiCl3 and eluted from thecolumn with 4 M HCl. Americium fraction was evaporated to dryness and thendissolved into 2 M NH4SCNþ 0.1 M HCOOH. This fraction was further purified fromlanthanides with TEVA� resin (Salminen et al., 2005). Lanthanides form tiocyanatecomplexes and were not retained in TEVA� resin. Am(III) retained into the resinand was eluted from the resin with 2 M HCl. Americium fraction was treated withaqua regia and evaporated to dryness. Furthermore americium fraction wastreated twice with conc. HNO3 and evaporated to dryness. The residue was dis-solved into 1 M HNO3. Plutonium and americium fractions were co-precipitatedwith neodymium as trifluorides and the precipitate was collected on a membranefilter for alpha counting.

The analyses of plutonium and americium from the unfiltered water werestarted by coprecipitating these nuclides with Fe(OH)3 from the water phase (Paa-tero, 2000; Pilvio, 1998; Suutarinen et al., 1993). The precipitate was filtered anddissolved in 8 M HNO3. Dissolved samples were co-precipitated with calciumoxalate and their analysis was continued in the same manner as the sedimentsamples.

The americium and plutonium activities were measured from the filterscontaining the NdF3 precipitates with surface barrier semiconductor detectors.The nominal resolution for 241Am and 239,240Pu of the three detectors variedbetween 20–30 keV. A typical counting time was 5000 min. For a typicalcounting time and counting efficiency of 27% the minimum detectable activity(MDA) was 0.16 Bq/m2 and 0.6 mBq kg�1 for the sediment samples and watersamples, respectively.

Fig. 3. The extraction chromatography procedure used for separating plutonium and americium from sediment samples.

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476 471

3.3. Analysis of cesium from the sediment and water samples

137Cs activity of the sediment samples were determined with a NaI(Tl)detector, which included an automatic sample changer, from the freeze-dried andhomogenized samples in standard 20-ml polyethene counting vials with a samplevolume of 15 ml. For the counting time of 120 min and the counting efficiency of20% with the typical background counts of 39.9 cpm the lowest limit of detection(LLD) for the sediment samples was 9.6 Bq L�1. The self-adsorption of gamma rayswas calculated to induce an error of 1–2% but this was not taken into consider-ation in experimental arrangements or calculations. Although the resolution of theNaI(Tl) detector is lower than with semiconductor detectors, the advantage ofthe automatic sampler changer used is that all samples can be loaded into theapparatus at the same time and since the counting efficiency of the detector ishigh, the counting time is reduced compared with the counting time needed withsemiconductors. All samples could be measured in less than a week. A disad-vantage of the NaI(Tl) detector is the higher inaccuracy of the results. Forcomparison 137Cs from one sediment core (A2) was determined with a semi-conductor detector (Table 6). For profile A2, the average error for the NaI(Tl)detector was 15%. For the same samples the error was on average 1%, whensemiconductor detector was used.

For the analysis of 137Cs from the unfiltered water 30 L of water was evaporatedto dryness in a 2 L beaker. The residue was dissolved into 8 M HNO3 and 137Csactivity was measured with a semiconductor detector.

4. Results and discussion

4.1. Plutonium and americium in sediments and unfiltered watersamples

The activity concentrations of 239,240Pu and 241Am in sedimentprofiles from the depths of 0–9 cm are presented in Table 2 and inFigs. 4 and 5. The total activity of 239,240Pu ranged between29.6� 5.7 and 65.1�7.1 Bq/m2 and the average total activitywas 45�15 Bq/m2. The average total activity of 241Am was19.8� 7.1 Bq/m2 and the range was from 11.3�1.9 to 28.7�1.7 Bq/m2. The peak activities were in depths 5–7 cm for both nuclides and

Table 2239,240Pu and 241Am activity concentrations (Bq/m2), 241Am/239,240Pu activity ratios in the bottom sediments of Asikkalanselka at depth 0–9 cm in 2007 and the total activity(S, Bq/m2) of sampling sites A1, A2, A3 and A4. The error indicated for the individual depths is 1s counting error from alpha spectrometry and the error for average values is thestandard deviation of the values. The error for the total activities is the combined error of the individual depths in quadrature. The results under detection limit are indicatedwith a dash (–).

Depth A1 A2

239,240Pu 241Am 241Am/239,240Pu 239,240Pu 241Am 241Am/239,240Pu

0–1 1.0� 0.2 0.5� 0.2 0.48� 0.27 0.8� 0.1 0.2� 0.1 0.18� 0.091–2 1.5� 0.2 0.8� 0.1 0.56� 0.07 1.4� 0.2 0.7� 0.1 0.50� 0.062–3 2.0� 0.2 1.0� 0.1 0.52� 0.06 1.9� 0.2 1.1� 0.1 0.56� 0.073–4 3.8� 0.5 1.8� 0.1 0.46� 0.04 2.0� 0.2 0.9� 0.1 0.46� 0.064–5 4.9� 0.6 3.0� 0.2 0.60� 0.04 4.0� 0.5 2.8� 0.2 0.70� 0.065–6 11.4� 1.3 4.8� 0.6 0.42� 0.06 12.3� 1.3 5.6� 0.3 0.45� 0.036–7 17.3� 1.8 7.1� 0.9 0.41� 0.05 13.3� 1.4 6.2� 0.3 0.47� 0.037–8 2.0� 0.3 0.7� 0.2 0.34� 0.09 5.8� 0.7 1.8� 0.2 0.31� 0.048–9 – – – 0.5� 0.1 0.4� 0.1 0.80� 0.31S 43.9� 2.4 19.6� 1.1 42.0� 2.1 19.6� 0.6

Average 0.47� 0.09 Average 0.49� 0.19A3 A4

0–1 0.5� 0.1 0.1� 0.2 0.18� 0.31 0.4� 0.1 0.1� 0.1 0.19� 0.251–2 1.5� 0.2 0.5� 0.1 0.36� 0.09 2.6� 0.3 1.2� 0.1 0.44� 0.042–3 2.9� 0.4 1.4� 0.2 0.47� 0.08 3.2� 0.3 1.9� 0.1 0.59� 0.043–4 4.0� 1.0 2.3� 0.3 0.58� 0.10 5.1� 0.6 2.8� 0.2 0.55� 0.054–5 7.1� 1.5 1.7� 0.2 0.25� 0.04 9.6� 1.0 4.7� 0.2 0.49� 0.025–6 10.3� 1.5 4.1� 0.5 0.39� 0.05 17.2� 1.9 7.4� 0.3 0.43� 0.026–7 3.4� 0.6 1.0� 0.2 0.31� 0.06 19.2� 2.0 7.0� 0.3 0.36� 0.027–8 0.0� 0.8 0.1� 0.2 – 6.9� 0.8 3.2� 0.3 0.47� 0.048–9 – – – 0.9� 0.2 0.4� 0.2 0.45� 0.19S 29.6� 2.6 11.3� 0.7 65.1� 3.1 28.7� 0.6

Average 0.36� 0.13 Average 0.44� 0.11

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476472

in these layers 239,240Pu and 241Am originate from the nuclear tests,as will later be described. The activity concentrations increasedexponentially as a function of depth up to the peak value whereafter the activities decreased rapidly. The mineralogy of the sedi-ment layers was not further characterized in this study, but it isknown that the upper layers of the sediment profiles are composedof fine organic matter. In lower stratums the structure graduallychanges to more clayey.

The activity concentrations of 239,240Pu and 241Am in theunfiltered water were low: the average activities being 4.9� 0.9mBq/m3 and 4.1�0.2 mBq/m3, respectively (Table 3).

Fig. 4. Vertical distribution of 239,240Pu (Bq/m2) in sampling sites A1, A2, A3 and A4 inAsikkalanselka in 2007.

4.2. 241Am/239,240Pu activity ratio in sediments and unfiltered water

No previous data on the distribution of plutonium and ameri-cium in the sediment profiles of Lake Paijanne exist. Only datacovering the surface layer (0–1 cm) of the sediments are previouslyavailable. Calculated from the Suutarinen et al. (1993) results fromthe year 1986 the average 239,240Pu activity concentrations ofsurface sediments (0–1 cm) from five sampling sites with waterdepths of 19–51 m was 2.0� 0.4 Bq/m2 and the average 241Amactivity concentration 1.0� 0.1 Bq/m2. In the sampling site corre-sponding to the depths of sites A1 and A2 in our study the

Fig. 5. Vertical distribution of 241Am (Bq/m2) in sampling sites A1, A2, A3 and A4 inAsikkalanselka in 2007.

Table 3The activity concentrations (mBq/m3) of 239,240Pu and 241Am and the 241Am/239,240Puactivity ratio in the unfiltered water samples in 2007.

Sample 239,240Pu (mBq/m3) 241Am (mBq/m3) 241Am/239,240Pu

W1 4.3� 1.0 3.0� 1.7 0.71� 0.47W2 5.5� 1.4 5.2� 1.4 0.94� 0.36Average 4.9� 0.9 4.1� 0.2 0.82� 0.17

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476 473

241Am/239,240Pu – activity ratio was 0.60� 0.12 in 1986 (Suutarinenet al., 1993). In our study the average 241Am/239,240Pu – activity ratioin the surface layer (0–1 cm) of the sediments in these sites was0.33� 0.21 in 2007 (Table 2). As 241Pu decays to 241Am with a half-life of 14.35 years, the 241Am/239,240Pu activity ratio is presumed toincrease from 1986 to 2007. Based on our results and those ofSuutarinen et al. (1993) the 241Am/239,240Pu activity ratio in theuppermost sediment layer has on the contrary decreased betweenyears 1986 and 2007. Furthermore the 241Am/239,240Pu activity ratiois lowest in the surface layer and increases as a function of depth(Fig. 6). This is probably a consequence of the partial solubility ofamericium from surface sediment layer to the water body. Sanadaet al. have reported a corresponding increase of 241Am/239,240Puactivity ratio in the lower layers of bottom sediment samples fromPripyat River (Sanada et al., 2002). According to Sanada et al. (2002)this ‘‘implies that its (241Am) behavior is not necessarily similar tothat of 239,240Pu’’.

In our study the 241Am/239,240Pu activity ratio of the watersamples ranged between 0.71�0.47 and 0.94� 0.36 and theaverage ratio was 0.82� 0.17. Based on the 241Am/239,240Pu activityratio both in the sediment and water samples, the relativeconcentration of 241Am was almost two times higher in the waterthan in the sediments. This, in addition to the low 241Am/239,240Puratio in the surface sediment layer, strongly indicates to the highersolubility of americium compared to plutonium. These results arein good agreement with the results reported by Lujanien _e et al.(2002) where higher mobility of Am in comparison with Puisotopes in the Chernobyl soil was estimated via speciation analyses(Lujanien _e et al., 2002). Lujanien _e et al. (2002) have suggested thatthe organic matter of soil can reduce the mobility of plutonium notonly due to the sorption of plutonium on it, but also by affecting thevalence state of plutonium. Humic substances can reduce Pu(V) andPu(VI) to less soluble Pu(IV) and in most natural waters plutoniumexists as Pu(IV) (Choppin, 2006). After reduction plutoniumforms hydroxides with very low solubility at oxidation state IV.

Fig. 6. 241Am/239,240Pu activity ratio in the bottom sediments of Lake Paijanne in 2007.

Concurrently americium can form water soluble lower molecularweight complexes with organic matter (Lujanien _e et al., 2002). Asmentioned above the upper layers of the sediment samples in ourstudy are composed of fine organic matter. Assuming theenhancement of the solubility of americium due to complexationand the reduction of plutonium to the less soluble state the abovementioned reactions could explain the differences found in241Am/239,240Pu activity ratio of our samples.

Actinides in the different oxidation states may have differentmigration rates as solubility, sorption and interactions with organicor inorganic ligands depend on the oxidation state. According toSokolik et al. (2004) an increase in the amount of soluble organicmatter favors the mobility of americium and plutonium in soil.According to another study of Sokolik et al. (2002) plutonium andamericium species in soil solutions have different electrical chargescompared to each other, which results in different migration abilityof these radionuclides. Furthermore in the Sokolik et al. (2002)study the relative content of anionic species was higher for amer-icium than for plutonium. This may result in higher migration rateof americium. When comparing the chemistry of radionuclides insoil and sediments it should however be taken into account that theredox and flow conditions in these two media may differ.

According to Salminen et al. the 241Am/239,240Pu ratio in theChernobyl deposition was 0.37 (Salminen et al., 2005). In 2007 thiswould correspond to a 241Am/239,240Pu ratio of 2.3. In the weaponstest fallout in 1963–1965 the 241Am/239,240Pu ratio in lichensamples was 0.11 (Jaakkola et al., 1981). Deriving from the physicaldecay of 241Pu this ratio has increased from 0.11 to a ratio of 0.44 in2007. Salminen et al. (2005) calculated that the total deposition of241Am from the Chernobyl accident was only 1.7% of the totalamericium deposition in Finland and the rest originated from thenuclear weapons tests (Salminen et al., 2005). In our study theaverage 241Am/239,240Pu ratio in the 0–9 cm depth of bottom sedi-ments was 0.45� 0.14. This corresponds well to the ratio origi-nating from the nuclear tests and indicates that they are the mainsource of plutonium and americium in the sediments, assumingthat the ratio in lichens is the same as in sediments.

4.3. 238Pu/239,240Pu activity ratio in sediments and unfiltered water

According to Holm et al. the 238Pu/239,240Pu activity ratio in theChernobyl fallout was 0.47 and in nuclear test fallout 0.026 (Holmet al., 1992). In our study the 238Pu/239,240Pu ratio of the sedimentsamples varied between 0.012� 0.025 and 0.162� 0.079. Theaverage 238Pu/239,240Pu ratio of the four sampling sites as a functionof depth is presented in Fig. 7. In the lowest layers the activity ratioof plutonium isotopes corresponds to the plutonium originatingfrom nuclear weapon tests. In the upper layers the correspondingactivity ratio is higher and decreases as a function of depth. On thebasis of the 238Pu/239,240Pu activity ratio we deduce that the mainportion of plutonium originates from the nuclear weapons falloutand plutonium from the Chernobyl fallout is located in the upper-most parts of the sediment profile.

Paatero et al. (2002) have estimated that the Chernobyl-derived239,240Pu is approximately 0.5% of the activity of 239,240Pu fromnuclear test fallout in Finland. In this study the percentage ofChernobyl-derived plutonium activity was calculated by takinga logarithm from the measured 238Pu/239,240Pu ratios and fittingthese values to a linear function (Fig. 8) with an equation (1)

R ¼ 10 expð�0:08239z� 1:0468Þ (1)

where R is 238Pu/239,240Pu and z depth. This was done to describebetter the dominating trend in different depths of the sedimentprofile.

Fig. 7. The average 238Pu/239,240Pu activity ratio of four sampling sites A1, A2, A3 andA4 as a function of depth in the bottom sediments of Asikkalanselka in 2007.

Table 4The average fraction of Chernobyl-derived plutonium isotopes 238 and 239,240 inAsikkalanselka bottom sediments (%) and the average 239,240Pu activity originatingfrom the Chernobyl fallout (Bq/m2).

Depth(cm)

The Chernobyl fallout fraction ofthe total activity of plutonium(%)

239,240Pu activity originating fromthe Chernobyl fallout (Bq/m2)

0.5 11.8� 2.6 0.08� 0.031.5 8.8� 1.9 0.15� 0.052.5 6.4� 1.4 0.16� 0.043.5 4.3� 0.9 0.16� 0.064.5 2.6� 0.6 0.17� 0.075.5 1.2� 0.3 0.15� 0.046.5 0.00� 0.01 0.000� 0.0027.5 0.0� 0.2 0.00� 0.038.5 0.0� 1.2 0.00� 0.02

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476474

The fraction of Chernobyl-derived plutonium in different depthswas then calculated from the 238Pu/239,240Pu ratios determinedwith equation (1) using equation (2)

Chernobyl derived Puð%Þ ¼ ½0:47Rþ 0:026ð1� RÞ� � 100 (2)

where R is 238Pu/239,240Pu ratio. The percentages of total plutoniumactivity originating from the Chernobyl fallout in various depths ofbottom sediments of Asikkalanselka are presented in Table 4. Thetotal plutonium activities (Bq/m2) from the Chernobyl and nucleartest fallout are illustrated in Fig. 9. The 239,240Pu originating fromthe Chernobyl fallout (Table 4) calculated from the average totalactivities from the depths of 0–9 cm covers 1.95� 0.01% of the total239,240Pu activity in the bottom sediments of Asikkalanselka. Thiscorresponds to a total activity of 0.9� 0.3 Bq/m2 which corre-sponds well to the levels found in this area by Paatero et al. (2002).

The majority of the 239,240Pu and 241Am activities have beentransported to the bottom sediments. The water volume of

Fig. 8. The logarithmic 238Pu/239,240Pu activity ratios in the bottom sediments of LakePaijanne in 2007. The 238Pu/239,240Pu activity ratio decreases as a function of depthfollowing a linear function y¼�0.08239x–1.0468.

Asikkalanselka is 8.05�108 m3 and the total area 5.28� 107 m2.The average activity in unfiltered water was 4.9� 0.9 mBq/m3 and4.1�0.2 mBq/m3 for 239,240Pu and 241Am, respectively. Thepercentage of the average total activity of 239,240Pu and 241Am inwater compared to the total activity in Asikkalanselka was calcu-lated using equation (3):

A% ¼�

abðcdþ abÞ

�� 100 (3)

where a is the nuclides’ average activity in water (Bq/m3), b is thetotal volume of Asikkalanselka (m3), c is the nuclides’ average totalactivity in bottom sediments (Bq/m2) and d is the total area ofAsikkalanselka. It was calculated that in Asikkalanselka only0.17� 0.07% and 0.32� 0.05% of the total activity of 239,240Pu and241Am, respectively, is in the water column. These values are roughestimates since there are two simplifications in the calculations.First, the lake mirror area was used in the calculations instead of theactual bottom surface area, which is not known. Secondly, thedistribution of radionuclides in the bottom was assumed to be even,which is certainly not the case. Fortunately these simplificationsmodify the results into opposite directions, since when the lakemirror area is used the calculated total activity of the bottomsediments is smaller than if the actual bottom surface area wasused. Furthermore the activities in the central parts of the lake are

Fig. 9. Plutonium activity (Bq/m2) originating from the Chernobyl and nuclear testfallout in the bottom sediments of Asikkalanselka in 2007. The broken line is nuclearweapons test plutonium and the solid line Chernobyl-derived plutonium.

Table 5137Cs activity concentrations in the bottom sediments of Asikkalanselka in the depthof 0–9 cm in 2007 and the total activity (S, Bq/m2) of sampling sites A1, A2, A3 andA4. The errors indicated for various depths are 1s counting error from gammaspectrometry and for the total activities the error is the combined error of theindividual depths in quadratur.

Depth (cm) A1 A2 A3 A4

A kBq/m2 A kBq/m2 A kBq/m2 A Bq/m2

0–1 2.6� 0.4 0.8� 0.1 1.5� 0.2 1.5� 0.21–2 5.2� 0.8 5.1� 0.7 5.9� 0.8 10.7� 1.52–3 8.0� 1.2 8.8� 1.3 11.6� 1.6 17.0� 2.43–4 15.5� 2.2 11.8� 1.7 20.4� 2.9 33.5� 4.74–5 28.6� 4.0 43.3� 6.1 22.1� 3.1 41.3� 5.95–6 15.4� 2.2 45.7� 6.5 8.1� 1.1 11.7� 1.76–7 3.8� 0.5 5.2� 0.7 3.0� 0.4 3.4� 0.57–8 2.4� 0.3 5.0� 0.7 1.0� 0.1 1.5� 0.28–9 1.1� 0.2 1.9� 0.3 0.6� 0.1 1.3� 0.29–10 0.7� 0.1 1.4� 0.2 0.4� 0.1 0.6� 0.110–11 0.5� 0.1 0.6� 0.1 0.3� 0.1 0.4� 0.1S 83.8� 5.3 129.4� 9.3 74.9� 4.8 122.9� 8.3

Table 6A comparison between the results of the sediment profile A2 measured withboth semiconductor detector and NaI(Tl) detector. The error indicated for the depthsof 0–9 cm is 1s counting error from gamma spectrometry. The error indicated for thetotal activities is calculated from the sum of the individual counting errors of depthsof 0–9 cm.

Depth Semiconductor Counting error NaI(Tl) detector Counting error

Bq/g % Bq/g %

0–1 2.3 1.1 4.8 181–2 6.5 1.0 6.3 152–3 6.7 1.0 7.7 143–4 10.7 0.9 12.0 144–5 36.3 0.9 32.4 145–6 21.6 0.8 24.8 146–7 3.2 0.9 3.7 147–8 1.6 1.0 1.7 148–9 0.7 1.1 0.8 159–10 0.3 1.1 0.4 1610–11 0.1 3.1 0.2 14

Total 90.1 0.9 94.9 14

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476 475

presumably larger than the activities on the beach front. As thesevalues are used the calculated total activity of the bottom sedi-ments is larger than the actual activity.

4.4. Cesium in sediments and unfiltered water

The activity concentrations of 137Cs in sediment samplesfrom the depths of 0–9 cm are presented in Table 5 and inFig. 10. The total activity of sediment profiles ranged between75�11 and 129�18 kBq/m2 and the average total activity was100�15 kBq/m2.

The peak activities for 137Cs were in depths of 4–5 cm. Theactivities decreased rapidly in the sediment layers below the peakvalue. In 1988 (Saxen et al., 1998) the peak value of 137Cs in a pointcorresponding the sampling sites A1 and A2, was on the surface ofthe sediment profile. In years 1994 and 1997 the peak value was ina depth of 1.5–2.5 cm and 2–3 cm, respectively. In 2000 the peakwas found in a depth of 3–4 cm (Ilus and Saxen, 2005). 29% of thetotal 137Cs activity was settled to the layers below the maximumconcentrations and in proportion 37% above the peak values. Anaverage of 8% more 137Cs activity was in the sediment layers abovethe stratums with maximum concentrations than in the lowerlayers.

Fig. 10. The vertical distribution of 137Cs (Bq/m2) in sampling sites A1, A2, A3 and A4 inAsikkalanselka on 9.3.2007.

The depth profiles of 137Cs in the sampling sites studied showedconsiderable variety. The total activities varied from 75 to 129 kBq/m2 and the average value was 100�15 kBq/m2. In the deepestbasin the total activity in sediment from A1 was 65% of the totalactivity of sediment from A2 even though they were taken onlya few meters from each other. On the other hand, on the basinfringes the activity of site A3 was only 61% of the activity on theother site of the basin at point A4. In former studies other authorshave reported the uneven distribution of radionuclides in thesediments. Saxen et al. (1998) and Kansanen et al. (1991) reportedthat the 137Cs activity concentrations showed considerable varia-tions in the bottom sediments of southern Lake Paijanne and otherFinnish lakes. According to Kansanen et al. (1991) the total 137Csactivity concentration in Asikkalanselka was 102 400 Bq/m2 in year1989. Ilus and Saxen (2005) reported that in 2003 the total activitywas approximately 90 000 Bq/m2. According to the results of Saxenet al. (1998) the total activity of 137Cs in a point corresponding toour sampling sites A1 and A2 was 135 500 Bq/m2 in 1997. Fromphysical decay this activity would correspond to an activity of108 000 Bq/m2 in year 2007. In our study the average activity of thesampling sites A1 and A2 was 106 600� 5400 Bq/m2. The activityconcentrations of 137Cs in the analyzed unfiltered surface watersamples were 18.3� 0.6–20.3� 0.5 Bq/m3. In 1998 137Cs activity inthe water of Lake Paijanne was 69 Bq/m3 (Saxen et al., 1998). Takinginto account the physical decay this would correspond to theactivity of 56 Bq/m3 in 2007. This indicates considerable transfer of137Cs from the water body to the bottom sediments.

Kansanen et al. (1991) reported observations which support theview that redistribution and resuspension of 137Cs had significantimpact in Asikkalanselka. These would cause the focusing ofradioactivity in the depressions. In our study clear evidence of theredistribution or focusing of the radioactivity could not be proven.This might be a consequence of low sample number.

One sediment profile was also measured with semiconductordetectors to assure the results measured with NaI(Tl) detector(Table 6). The activities measured with the semiconductordetector correspond to the activities determined with the NaI(Tl)detector within the limits of accuracy. The error in NaI(Tl)measurements was, however, more than ten times higher.

5. Conclusions

Based on the 238Pu/239,240Pu and 241Am/239,240Pu activity ratios,the majority of the 239,240Pu and 241Am activities in the bottom

M. Lusa et al. / Journal of Environmental Radioactivity 100 (2009) 468–476476

sediments of Asikkalanselka originates from the 1950s and 1960snuclear test fallout. The impact of Chernobyl fallout to the 239,240Puand 241Am activities in Asikkalanselka is minor compared to thenuclear test fallout. Only 1.95� 0.01% of the total activity of239,240Pu in the bottom sediments of Asikkalanselka was calculatedto have originated from the Chernobyl fallout. This corresponds tothe total activity of 0.9� 0.3 Bq/m2. 239,240Pu and 241Am from theChernobyl fallout are located in the uppermost parts of the bottomsediments. The peak values are found in depths of 5–7 cm for bothelements and in these layers they originate from the nuclear tests.

The average 241Am/239,240Pu activity ratio in the water body was0.82� 0.17 and correspondingly the average value in the bottomsediments was 0.45� 0.14. In the sediments the 241Am/239,240Puactivity ratio was lowest in the surface layer and increased asa function of depth. Furthermore the 241Am/239,240Pu ratio of thesurface layer of the sediments has decreased during the last 11years. This probably reflects the solubility of 241Am from the sedi-ments into the water body.

The depth profiles of 137Cs in the sampling sites studied showedconsiderable variety. The total activities ranged between 75�11 and129�18 kBq/m2 and the average total activity was 100�15 kBq/m2.The peak activities for 137Cs were in a depth of 4–5 cm. This peakcorresponds to the Chernobyl fallout in 1986. The activitiesdecreased rapidly in the sediment layers below the peak value. Noclear evidence of the redistribution or focusing of 137Cs activity couldbe proven. This might be a consequence of a low sample number.

In bottom sediment samples the peak values of 137Cs were onaverage 2 cm farther up the sediment profile than those of 239,240Puand 241Am. This is because 137Cs originated mainly from the Cher-nobyl fallout while plutonium and americium came from nuclearweapons tests.

Acknowledgements

The authors wish to thank Stewart Makkonen-Craig for thelanguage revision of the manuscript.

References

AMAP, 1998. AMAP Assessment Report, Arctic Pollution Issues, Arctic Monitoringand Assessment Programme. Norway, Oslo.

Arvela, H., Markkanen, M., Lemmela, H., Blomqvist, L., 1989. Environmental GammaRadiation and Fallout Measurements in Finland, 1986–87, Report STUK-A76.Radiation and Nuclear Safety Authority, Finland.

Choppin, G.R., 2006. Environmental behaviour of actinides. Czechoslovak Journal ofPhysics 56 (Suppl. D), D13–D21.

Gutierrez, M., Fuentes, H.R., 1996. A mechanistic modeling of montmorillonitecontamination by cesium sorption. Applied Clay Science 11, 11–24.

Holm, E., Rioseco, J., Pettersson, H., 1992. Fallout of transuranium elementsfollowing the Chernobyl accident. Journal of Radioanalytical and NuclearChemistry 156 (1), 183–200.

Ilus, E., Saxen, R., 2005. Accumulation of Chernobyl-derived 137Cs in bottom sedimentsof some Finnish Lakes. Journal of Environmental Radioactivity 82, 199–221.

Jaakkola, T., Keinonen, M., Hakanen, M., Miettinen, J.K., 1981. Investigation on thetransfer of plutonium and americium from plants to Reindeer and Man in FinnishLapland. In: Wrenn, M.E. (Ed.), Actinides in Man and Animals. RDPress, pp. 509–523.

Kansanen, P.H., Jaakkola, T., Kulmala, S., Suutarinen, R., 1991. Sedimentation anddistribution of gamma-emitting radionuclides in bottom sediments ofsouthern Lake Paijanne, Finland, after the Chernobyl accident. Hydrobiology222, 121–140.

Lieser, K.H., Steinkopff, T.H., 1989. Chemistry of radioactive Cesium in the hydro-sphere and in the geosphere. Radiochimica Acta 46, 39–47.

Lujanien _e, G., Plukis, A., Kimtys, E., Remeikis, V., Jankunait _e, D., Ogorodnikov, B.I.,2002. Study of 137Cs, 90Sr, 239,240Pu, 238Pu and 241Am behavior in the Chernobylsoil. Journal of Radioanalytical and Nuclear Chemistry 251 (1), 59–68.

Lujanien _e, G., Joksas, K., Silobritiene, B., Morkuniene, R., 2004. Physical andchemical characteristics of 137Cs in Baltic Sea. In: Radionuclides in the Envi-ronment, International Conference on Isotopes in Environmental Studies:Aquatic Forum 2004. Elsevier, Monaco, pp. 25–29.

Outola Iisa, 2002. Effect of industrial pollution on the behaviour of 239,240Pu, 241Amand 137Cs in forest ecosystems. Report Series in Radiochemistry 21/2002,Helsinki.

Paatero, J., 2000. Deposition of Chernobyl Derived Transuranium Nuclides andShort-Lived Radon-222 Progeny in Finland. Finnish Meteorological InstituteContributions No. 28. Finnish Meteorological Institute, Helsinki, p.55þAppendixes.

Paatero, J., Jaakkola, T., Ikaheimonen, T.K., 2002. Regional distribution of Chernobyl-derived plutonium deposition in Finland. Journal of Radioanalytical and NuclearChemistry 252 (2), 407–412.

Pilvio R., 1998. Methods for the determination of low-level actinide concentrationsand their behaviour in the aquatic environment. Report Series in Radiochem-istry 10/1998, Helsinki, p. 43þAppendixes.

Salminen, S., Outola, I., Jaakkola, T., Pulli, S., Zilliacus, R., Lehto, J., 2004. Method fordetermining plutonium in air filters in detection of nuclear activities. Radio-chimica Acta 92, 467–473.

Salminen, S., Paatero, J., Jaakkola, T., Lehto, J., 2005. Americium and curiumdeposition in Finland from the Chernobyl accident. Radiochimica Acta 93,771–779.

Sanada, Y., Matsunaga, T., Yanase, N., Nagao, S., Amano, H., Takada, H.,Tkachenko, Y., 2002. Accumulation and potential dissolution of Chernobyl-derived radionuclides in river bottom sediment. Applied Radiation andIsotopes 56, 751–760.

Saxen, R., Jaakkola, T., Rantavaara, A., 1998. 137Cs and 90Sr in the Southern Part ofLake Paijanne and its catchments. Radiochemistry 40 (6), 522–528.

Sidhu, R., 2006. Radiochemical Procedures for the Determination of Sr, U, Pu, Amand Cm. Nordic nuclear safety research, Radchem, ISBN 87-7893-185-1.

Silva, R.J., Nitsche, H., 1995. Actinide environmental chemistry. Radiochimica Acta70/71, 377–396.

Sokolik, G.A., Ovsyannikova, S.V., Kimlenko, I.M., 2002. Effect of humic substanceson Plutonium and Americium speciation in soils and soil solutions. Radio-chemistry 45 (2), 176–181.

Sokolik, G.A., Ovsiannikova, S.V., Ivanova, T.G., Leinova, S.L., 2004. Soil-plant transferof plutonium and americium in contaminated regions of Belarus after theChernobyl catastrophe. Environment International 30, 939–947.

Suutarinen, R., Jaakkola, T., Paatero, J., 1993. Determination of Pu, Am and Cmconcentrations and the oxidation states of Pu from aquatic samples. The Scienceof the Total Environment 130/131, 65–72.