Lake sediment and water radioisotope analysis

Journal of Environmental Radioactivity 100 (2009) 468–476

Contents lists available at ScienceDirect

Journal of Environmental Radioactivity
journal homepage: www.elsevier.com/locate/jenvrad

137
Cs, 239,240Pu and 241
Am in bottom sediments and surface water of Lake
¨ ¨
Paijanne, Finland
M. Lusa*, J. Lehto, A. Leskinen, T. Jaakkola
Laboratory 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 a b s t r a c t

Article history: The concentrations and vertical distribution of 239,240Pu, 241Am and 137Cs in the bottom sediments and
Received 1 May 2008 ¨ ¨ ¨ ¨
water samples of Lake Paijanne were investigated. This lake is important, since the Paijanne area received
Received in revised form ¨ ¨
a significant deposition from the Chernobyl fallout. Furthermore Lake Paijanne is the raw water source
2 March 2009
for the Helsinki metropolitan area. In addition no previous data on the distribution of plutonium and
Accepted 4 March 2009
¨ ¨
americium in the sediment profiles of Lake Paijanne exist. Only data covering the surface layer (0–1 cm)
Available online 11 April 2009
of the sediments are previously available. In the sediments the average total activities were 45 Æ 15 Bq/
m2 and 20 Æ 7 Bq/m2 for 239,240Pu and 241Am, respectively. The average 241Am/239,240Pu ratio was
Keywords:
Cesium 0.45 Æ 0.14. The 241Am/239,240Pu ratio is lowest in the surface layer of the sediments and increases as
Americium a function of depth. The 238Pu/239,240Pu ratio of the sediment samples varied between 0.012 Æ 0.025 and
Plutonium 0.162 Æ 0.079, decreasing as a function of depth. The average activity in water was 4.9 Æ 0.9 mBq/m3 and
Bottom sediments 4.1 Æ 0.2 mBq/m3 for 239,240Pu and 241Am, respectively. The 241Am/239,240Pu ratio of water samples was
¨ ¨
Lake Paijanne 0.82 Æ 0.17. 239,240Pu originating from the Chernobyl fallout calculated from the average total activities
covers approximately 1.95 Æ 0.01% of the total 239,240Pu activity in the bottom sediments. The average
total 137Cs activity of sediment profiles was 100 Æ 15 kBq/m2 and 19.3 Æ 1.4 Bq/m3 in water samples.
Ó 2009 Elsevier Ltd. All rights reserved.

1. Introduction deposited in Finland was approximately half of a percent of the
plutonium fallout of the nuclear tests. Similarly the 241Am activity
Deposition of 137Cs, 239,240Pu and 241Am from the Chernobyl originating from the Chernobyl accident was approximately 1.7% of
accident was very unevenly distributed in Finnish lakes and the total activity of 241Am in Finland (Salminen et al., 2005).
catchment areas (Kansanen et al., 1991; Paatero et al., 2002; Ilus and ¨
The radioactivity status of Asikkalanselka, the southernmost
´
Saxen, 2005; Salminen et al., 2005). This was caused by the differ- ¨ ¨
basin of Lake Paijanne, is an important topic since the Helsinki
ences in the areal rainfall conditions. The highest deposition values metropolitan area takes its raw water from this basin. This paper
of 239,240Pu and 241Am in Finland were located in the southwestern ¨
describes the present radioactivity situation on the Asikkalanselka,
and central parts of the country (Paatero et al., 2002; Salminen et al., twenty years after the Chernobyl accident. Also the radionuclide
2005). In August 1986 239,240Pu and 241Am concentrations in the distribution and the radionuclide sources are discussed.
surface layer of the bottom sediments (0–1 cm) of Lake Paijanne¨ ¨ In the hydrosphere the prevailing aqueous species of cesium is
were 1.5 Æ 0.1–2.5 Æ 0.2 Bq/m2 and 0.58 Æ 0.05–1.6 Æ 0.2 Bq/m2, the uncomplexed Csþ ion and changes in the pH and Eh do not
respectively (Suutarinen et al., 1993). The maximum 137Cs deposi- affect the speciation of cesium (Lieser and Steinkopff, 1989).
tion values, 45–78 kBq/m2, were found in the same region (Arvela Nonetheless, cesium may be adsorbed on surfaces of colloids and
et al., 1989). Chernobyl-derived 137Cs deposition is rather high suspended particles, which deposit onto lake bottoms. The ability
compared to that from the atmospheric nuclear weapons tests in of bottom sediments to bind cesium varies with particle size, and an
the 1950s and 1960s: 1700 Bq/m2 (decay corrected to 1986) (AMAP, increase in 137Cs activity with a decrease in particle size has been
1998). Instead the major source of transuranium nuclides in Finland _ _
observed (Lujaniene et al., 2004). According to Lujaniene et al. the
is from nuclear weapons test fallout (Paatero et al., 2002; Salminen highest calculated Kd values, i.e. the equilibrium ratio of 137Cs in
et al., 2005). The total amount of the Chernobyl-derived 239,240Pu particles compared to 137Cs in water phase, were obtained for
particles smaller than 4 mm. This is because the high adsorption of
Csþ ions is mainly determined by the clay minerals present in
* Corresponding author. Tel.: þ358 9 191 50518; fax: þ358 9 191 50121. sediments (Lieser and Steinkopff, 1989). The main mechanism of
E-mail address: merja.lusa@helsinki.fi (M. Lusa). the adsorption is ion exchange and in natural sediments the

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

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

sorption behavior is dominated by the highly selective exchange average of five months a year from mid December to the beginning
sites in clay minerals (Lieser and Steinkopff, 1989). The adsorption of May. The lake is dimictic and oligotrophic.
has been reported to be virtually irreversible (Gutierrez and
Fuentes, 1996). Although the cations with similar charge and ionic 3. Materials and methods

radii are expected to compete with cesium for adsorption sites, the
3.1. Sampling
bonding strength found for clay minerals decreases in the sequence
Csþ > Rbþ > Kþ > Naþ (Lieser and Steinkopff, 1989), which is in Sediment samples were taken in 2007 from three sampling sites, one at the
agreement with the decrease in ionic radius and the increase in the deepest place of the basin at 51 m (samples A1 and A2) and at two other 32 m
hydration enthalpies. deep sites 400 m away from this site, the other to the southeast (A3) and the
other to the northwest (A4) (Fig. 2). The locations in Finnish National Coordinate
The single most important property affecting the characteristic System and in EUREF-FIN geographical coordinate system are represented in
of transuranium elements is their oxidation state. Precipitation, Table 1. Sediment samples were taken with a Limnos gravity corer. The corer
complexation, sorption and colloid formation processes depend on consisted of a series of 1 cm-high rings and an inner tube diameter of 9.3 cm.
the prevailing oxidation state (Silva and Nitsche, 1995). All four Sediment samples were divided in situ to 1 cm-thick slices and brought to the
laboratory in plastic containers. In the laboratory the sediment samples were
processes contribute to the chemical behavior and environmental
frozen and freeze-dried. In addition to sediment samples 400 L of surface water
transport properties of actinides, including plutonium and ameri- was collected in 20–50 L plastic canisters from the western shore of Asikka-
cium, in the environment. In aqueous solutions with pH and Eh lanselka for 239,240Pu, 241Am and 137Cs analyses. 8 M HNO3 was added to the
¨
ranges of natural waters plutonium can be present in four oxidation water samples and their pH was adjusted to 1–2. Samples were kept in a cold-
states: III, IV, V, VI (Choppin, 2006). The main part of the dissolved storage room (þ4 C) until analyzed.

plutonium in natural waters exists in oxidation state Pu(V), while
3.2. Analysis of americium and plutonium from sediment and water samples
Pu(IV) is present in colloidal form (Choppin, 2006). Low pH values
promote lower oxidation states while higher oxidation states The method for the separation of plutonium and americium from the sediment
become more general as the pH increases. Pu(III) is possible in samples included wet ashing, coprecipitation with calcium oxalate and three
prevailing anoxic conditions, e.g. in lower sediments layers. extraction chromatography steps using UTEVAÒ, TRUÒ and TEVAÒ resins (Eichrom
Industries) (Fig. 3). 242Pu and 243Am tracers were used for yield determinations. The
Americium exists in nature in oxidation state III and Am(OH3) sorbs 243
Am tracer used in these analyses contained 2.5% 241Am as an impurity. This was
readily to nearly all surfaces. subtracted from the 241Am peak in the spectra of the samples. Furthermore, blank
analyses were prepared to investigate possible contaminations during the pluto-
nium and americium purification steps.
2. Study area 242
Pu and 243Am tracers were added to the dried four-gram aliquot samples and
the samples were wet-ashed with concentrated HNO3 and HCl to bring the analytes
¨ ¨ ¨nne
Asikkalanselka is situated in the southern part of Lake Paija into the solution before further chemical separations. Wet-ashed samples were
filtered and plutonium as well as americium were co-precipitated as oxalates as
(Fig. 1). The total area of Lake Pa ¨ nne is 1100 km2 and that of
ïja described earlier by Paatero, Outola and Salminen (Paatero, 2000; Outola, 2002;
Asikkalanselka 52.8 km2. The maximum depth in the Asikkalanselka
¨ ¨ Salminen et al., 2005). 100–200 mg of calcium carrier and 2–4 g of oxalic acid were
area is 53 m and its mean depth is 10 m. Pa ¨ nne is ice-covered an
ïja used for the precipitation. In the case of the upper parts of the sediment profile

137
¨ ¨nne in Finland and the
Fig. 1. Location of Lake Paija Cs fallout from Chernobyl in Finland (fallout map reworked from www.stuk.fi).

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

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

calcium oxalate precipitate was formed when lower amounts of calcium carrier in oxidation state II. The solution was poured into a UTEVAÒ column to separate
(100 mg) and oxalic acid (2 g) were used. In the lower parts, however, higher Am(III) and Pu(III) from tetra- and hexavalent impurities. Am(III) and Pu(III)
amounts, 200 mg and 2 g, respectively, were needed to precipitate calcium oxalate, remained in the eluent while U(VI), Th(IV) and Np(IV) were retained by the resin.
which was probably due to the differences in the matrix concentrations in different Eluent containing Am(III) and Pu(III) was poured to TRUÒ resin to separate
depths. Calcium oxalate precipitation was used to remove disturbing ions such as americium and plutonium from each other. Americium and plutonium was eluted
Kþ, Fe3þ, Al3þ, Ti4þand PO3À from the samples (Sidhu, 2006; Paatero, 2000). These
4 from TRUÒ resin in stages. First Pu(III) was oxidized with NaNO2 to Pu(IV) which
ions remained in solution as the actinides were co-precipitated. In the first place remained in the column while, Am(III) was eluted with 4 M and 9 M HCl.
3þ
Fe was important to remove since it substantially decreases the retention of Thereafter, plutonium was reduced back to Pu(III) with TiCl3 and eluted from the
americium to TRUÒ extraction chromatography resin. column with 4 M HCl. Americium fraction was evaporated to dryness and then
The precipitate was calcinated overnight and the ashed sample was dissolved dissolved into 2 M NH4SCN þ 0.1 M HCOOH. This fraction was further purified from
in a small amount of concentrated HNO3 and the solution was evaporated to lanthanides with TEVAÒ resin (Salminen et al., 2005). Lanthanides form tiocyanate
dryness. The residue was dissolved into 3 M HNO3 þ 1 M Al(NO3)3 solution and complexes and were not retained in TEVAÒ resin. Am(III) retained into the resin
0.6 M Fe(SO3)2 and ascorbic acid were added to separate americium and pluto- and was eluted from the resin with 2 M HCl. Americium fraction was treated with
nium from other interfering nuclides with extraction chromatography resins aqua regia and evaporated to dryness. Furthermore americium fraction was
UTEVAÒ, TRUÒ and TEVAÒ as described earlier by Salminen et al. (2004, 2005) treated twice with conc. HNO3 and evaporated to dryness. The residue was dis-
(Fig. 3). Al3þ was added, because it complexed with phosphate and thereby pre- solved into 1 M HNO3. Plutonium and americium fractions were co-precipitated
vented the interference of phosphate to neptunium or thorium uptake to UTEVAÒ. with neodymium as trifluorides and the precipitate was collected on a membrane
Fe(SO3)2 reduced Pu(IV) to Pu(III). Ascorbic acid was added to ensure that iron was filter for alpha counting.
The analyses of plutonium and americium from the unfiltered water were
started 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 and
Table 1 dissolved in 8 M HNO3. Dissolved samples were co-precipitated with calcium
The locations of sampling sites A1, A2, A3 and A4 in the Finnish National Coordinate oxalate and their analysis was continued in the same manner as the sediment
System (FNCS) and in EUREF-FIN geographical coordinate system. samples.
The americium and plutonium activities were measured from the filters
FNCS FNCS EUREF-FIN EUREF-FIN
containing the NdF3 precipitates with surface barrier semiconductor detectors.
X (p) Y(i) N/lat E/lon The nominal resolution for 241Am and 239,240Pu of the three detectors varied
between 20–30 keV. A typical counting time was 5000 min. For a typical
A1 and A2 6,792,301 3,421,983 61 130 53.66500 25 320 39.42200
counting time and counting efficiency of 27% the minimum detectable activity
A3 6,792,059 3,422,278 61 130 46.06100 25 320 59.54700
(MDA) was 0.16 Bq/m2 and 0.6 mBq kgÀ1 for the sediment samples and water
A4 6,792,507 3,421,706 61 140 0.11800 25 320 20.55500
samples, respectively.


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

3.3. Analysis of cesium from the sediment and water samples For the analysis of 137Cs from the unfiltered water 30 L of water was evaporated
to dryness in a 2 L beaker. The residue was dissolved into 8 M HNO3 and 137Cs
137
Cs activity of the sediment samples were determined with a NaI(Tl) activity was measured with a semiconductor detector.
detector, which included an automatic sample changer, from the freeze-dried and
homogenized samples in standard 20-ml polyethene counting vials with a sample
volume of 15 ml. For the counting time of 120 min and the counting efficiency of
20% with the typical background counts of 39.9 cpm the lowest limit of detection
4. Results and discussion
(LLD) for the sediment samples was 9.6 Bq LÀ1. The self-adsorption of gamma rays
was calculated to induce an error of 1–2% but this was not taken into consider- 4.1. Plutonium and americium in sediments and unfiltered water
ation in experimental arrangements or calculations. Although the resolution of the samples
NaI(Tl) detector is lower than with semiconductor detectors, the advantage of
the automatic sampler changer used is that all samples can be loaded into the
apparatus at the same time and since the counting efficiency of the detector is The activity concentrations of 239,240Pu and 241Am in sediment
high, the counting time is reduced compared with the counting time needed with profiles from the depths of 0–9 cm are presented in Table 2 and in
semiconductors. All samples could be measured in less than a week. A disad- Figs. 4 and 5. The total activity of 239,240Pu ranged between
vantage of the NaI(Tl) detector is the higher inaccuracy of the results. For 29.6 Æ 5.7 and 65.1 Æ 7.1 Bq/m2 and the average total activity
comparison 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)
was 45 Æ 15 Bq/m2. The average total activity of 241Am was
detector was 15%. For the same samples the error was on average 1%, when 19.8 Æ 7.1 Bq/m2 and the range was from 11.3 Æ 1.9 to 28.7 Æ 1.7 Bq/
semiconductor detector was used. m2. The peak activities were in depths 5–7 cm for both nuclides and


Table 2
239,240
Pu 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 the
standard 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 indicated
with a dash (–).

Depth A1 A2
239,240 241 241 239,240 239,240 241 241
Pu Am Am/ Pu Pu Am Am/239,240Pu
0–1 1.0 Æ 0.2 0.5 Æ 0.2 0.48 Æ 0.27 0.8 Æ 0.1 0.2 Æ 0.1 0.18 Æ 0.09
1–2 1.5 Æ 0.2 0.8 Æ 0.1 0.56 Æ 0.07 1.4 Æ 0.2 0.7 Æ 0.1 0.50 Æ 0.06
2–3 2.0 Æ 0.2 1.0 Æ 0.1 0.52 Æ 0.06 1.9 Æ 0.2 1.1 Æ 0.1 0.56 Æ 0.07
3–4 3.8 Æ 0.5 1.8 Æ 0.1 0.46 Æ 0.04 2.0 Æ 0.2 0.9 Æ 0.1 0.46 Æ 0.06
4–5 4.9 Æ 0.6 3.0 Æ 0.2 0.60 Æ 0.04 4.0 Æ 0.5 2.8 Æ 0.2 0.70 Æ 0.06
5–6 11.4 Æ 1.3 4.8 Æ 0.6 0.42 Æ 0.06 12.3 Æ 1.3 5.6 Æ 0.3 0.45 Æ 0.03
6–7 17.3 Æ 1.8 7.1 Æ 0.9 0.41 Æ 0.05 13.3 Æ 1.4 6.2 Æ 0.3 0.47 Æ 0.03
7–8 2.0 Æ 0.3 0.7 Æ 0.2 0.34 Æ 0.09 5.8 Æ 0.7 1.8 Æ 0.2 0.31 Æ 0.04
8–9 – – – 0.5 Æ 0.1 0.4 Æ 0.1 0.80 Æ 0.31
S 43.9 Æ 2.4 19.6 Æ 1.1 42.0 Æ 2.1 19.6 Æ 0.6
Average 0.47 Æ 0.09 Average 0.49 Æ 0.19
A3 A4
0–1 0.5 Æ 0.1 0.1 Æ 0.2 0.18 Æ 0.31 0.4 Æ 0.1 0.1 Æ 0.1 0.19 Æ 0.25
1–2 1.5 Æ 0.2 0.5 Æ 0.1 0.36 Æ 0.09 2.6 Æ 0.3 1.2 Æ 0.1 0.44 Æ 0.04
2–3 2.9 Æ 0.4 1.4 Æ 0.2 0.47 Æ 0.08 3.2 Æ 0.3 1.9 Æ 0.1 0.59 Æ 0.04
3–4 4.0 Æ 1.0 2.3 Æ 0.3 0.58 Æ 0.10 5.1 Æ 0.6 2.8 Æ 0.2 0.55 Æ 0.05
4–5 7.1 Æ 1.5 1.7 Æ 0.2 0.25 Æ 0.04 9.6 Æ 1.0 4.7 Æ 0.2 0.49 Æ 0.02
5–6 10.3 Æ 1.5 4.1 Æ 0.5 0.39 Æ 0.05 17.2 Æ 1.9 7.4 Æ 0.3 0.43 Æ 0.02
6–7 3.4 Æ 0.6 1.0 Æ 0.2 0.31 Æ 0.06 19.2 Æ 2.0 7.0 Æ 0.3 0.36 Æ 0.02
7–8 0.0 Æ 0.8 0.1 Æ 0.2 – 6.9 Æ 0.8 3.2 Æ 0.3 0.47 Æ 0.04
8–9 – – – 0.9 Æ 0.2 0.4 Æ 0.2 0.45 Æ 0.19
S 29.6 Æ 2.6 11.3 Æ 0.7 65.1 Æ 3.1 28.7 Æ 0.6
Average 0.36 Æ 0.13 Average 0.44 Æ 0.11

in these layers 239,240Pu and 241Am originate from the nuclear tests, 4.2. 241
Am/239,240Pu activity ratio in sediments and unfiltered water
as will later be described. The activity concentrations increased
exponentially as a function of depth up to the peak value where No previous data on the distribution of plutonium and ameri-
after the activities decreased rapidly. The mineralogy of the sedi- ¨ ¨
cium in the sediment profiles of Lake Paijanne exist. Only data
ment layers was not further characterized in this study, but it is covering the surface layer (0–1 cm) of the sediments are previously
known that the upper layers of the sediment profiles are composed available. Calculated from the Suutarinen et al. (1993) results from
of fine organic matter. In lower stratums the structure gradually the year 1986 the average 239,240Pu activity concentrations of
changes to more clayey. surface sediments (0–1 cm) from five sampling sites with water
The activity concentrations of 239,240Pu and 241Am in the depths of 19–51 m was 2.0 Æ 0.4 Bq/m2 and the average 241Am
unfiltered water were low: the average activities being 4.9 Æ 0.9 activity concentration 1.0 Æ 0.1 Bq/m2. In the sampling site corre-
mBq/m3 and 4.1 Æ 0.2 mBq/m3, respectively (Table 3). sponding to the depths of sites A1 and A2 in our study the

Fig. 4. Vertical distribution of 239,240Pu (Bq/m2) in sampling sites A1, A2, A3 and A4 in Fig. 5. Vertical distribution of 241
Am (Bq/m2) in sampling sites A1, A2, A3 and A4 in
¨
Asikkalanselka in 2007. ¨
Asikkalanselka in 2007.


Table 3 Concurrently americium can form water soluble lower molecular
The activity concentrations (mBq/m3) of 239,240Pu and 241Am and the 241Am/239,240Pu _
weight complexes with organic matter (Lujaniene et al., 2002). As
activity ratio in the unfiltered water samples in 2007.
mentioned above the upper layers of the sediment samples in our
239,240
Sample Pu (mBq/m3) 241
Am (mBq/m3) 241
Am/239,240Pu study are composed of fine organic matter. Assuming the
W1 4.3 Æ 1.0 3.0 Æ 1.7 0.71 Æ 0.47 enhancement of the solubility of americium due to complexation
W2 5.5 Æ 1.4 5.2 Æ 1.4 0.94 Æ 0.36 and the reduction of plutonium to the less soluble state the above
Average 4.9 Æ 0.9 4.1 Æ 0.2 0.82 Æ 0.17
mentioned reactions could explain the differences found in
241
Am/239,240Pu activity ratio of our samples.
241
Am/239,240Pu – activity ratio was 0.60 Æ 0.12 in 1986 (Suutarinen Actinides in the different oxidation states may have different
et al., 1993). In our study the average 241Am/239,240Pu – activity ratio migration rates as solubility, sorption and interactions with organic
in the surface layer (0–1 cm) of the sediments in these sites was or inorganic ligands depend on the oxidation state. According to
0.33 Æ 0.21 in 2007 (Table 2). As 241Pu decays to 241Am with a half- Sokolik et al. (2004) an increase in the amount of soluble organic
life of 14.35 years, the 241Am/239,240Pu activity ratio is presumed to matter favors the mobility of americium and plutonium in soil.
increase from 1986 to 2007. Based on our results and those of According to another study of Sokolik et al. (2002) plutonium and
Suutarinen et al. (1993) the 241Am/239,240Pu activity ratio in the americium species in soil solutions have different electrical charges
uppermost sediment layer has on the contrary decreased between compared to each other, which results in different migration ability
years 1986 and 2007. Furthermore the 241Am/239,240Pu activity ratio of these radionuclides. Furthermore in the Sokolik et al. (2002)
is lowest in the surface layer and increases as a function of depth study the relative content of anionic species was higher for amer-
(Fig. 6). This is probably a consequence of the partial solubility of icium than for plutonium. This may result in higher migration rate
americium from surface sediment layer to the water body. Sanada of americium. When comparing the chemistry of radionuclides in
et al. have reported a corresponding increase of 241Am/239,240Pu soil and sediments it should however be taken into account that the
activity ratio in the lower layers of bottom sediment samples from redox and flow conditions in these two media may differ.
Pripyat River (Sanada et al., 2002). According to Sanada et al. (2002) According to Salminen et al. the 241Am/239,240Pu ratio in the
this ‘‘implies that its (241Am) behavior is not necessarily similar to Chernobyl deposition was 0.37 (Salminen et al., 2005). In 2007 this
that of 239,240Pu’’. would correspond to a 241Am/239,240Pu ratio of 2.3. In the weapons
In our study the 241Am/239,240Pu activity ratio of the water test fallout in 1963–1965 the 241Am/239,240Pu ratio in lichen
samples ranged between 0.71 Æ 0.47 and 0.94 Æ 0.36 and the samples was 0.11 (Jaakkola et al., 1981). Deriving from the physical
average ratio was 0.82 Æ 0.17. Based on the 241Am/239,240Pu activity decay of 241Pu this ratio has increased from 0.11 to a ratio of 0.44 in
ratio both in the sediment and water samples, the relative 2007. Salminen et al. (2005) calculated that the total deposition of
241
concentration of 241Am was almost two times higher in the water Am from the Chernobyl accident was only 1.7% of the total
than in the sediments. This, in addition to the low 241Am/239,240Pu americium deposition in Finland and the rest originated from the
ratio in the surface sediment layer, strongly indicates to the higher nuclear weapons tests (Salminen et al., 2005). In our study the
solubility of americium compared to plutonium. These results are average 241Am/239,240Pu ratio in the 0–9 cm depth of bottom sedi-
in good agreement with the results reported by Lujaniene et al. _ ments was 0.45 Æ 0.14. This corresponds well to the ratio origi-
(2002) where higher mobility of Am in comparison with Pu nating from the nuclear tests and indicates that they are the main
isotopes in the Chernobyl soil was estimated via speciation analyses source of plutonium and americium in the sediments, assuming
_ _
(Lujaniene et al., 2002). Lujaniene et al. (2002) have suggested that that the ratio in lichens is the same as in sediments.
the organic matter of soil can reduce the mobility of plutonium not
only due to the sorption of plutonium on it, but also by affecting the 238
valence state of plutonium. Humic substances can reduce Pu(V) and 4.3. Pu/239,240Pu activity ratio in sediments and unfiltered water
Pu(VI) to less soluble Pu(IV) and in most natural waters plutonium
exists as Pu(IV) (Choppin, 2006). After reduction plutonium According to Holm et al. the 238Pu/239,240Pu activity ratio in the
forms hydroxides with very low solubility at oxidation state IV. Chernobyl fallout was 0.47 and in nuclear test fallout 0.026 (Holm
et al., 1992). In our study the 238Pu/239,240Pu ratio of the sediment
samples varied between 0.012 Æ 0.025 and 0.162 Æ 0.079. The
average 238Pu/239,240Pu ratio of the four sampling sites as a function
of depth is presented in Fig. 7. In the lowest layers the activity ratio
of plutonium isotopes corresponds to the plutonium originating
from nuclear weapon tests. In the upper layers the corresponding
activity ratio is higher and decreases as a function of depth. On the
basis of the 238Pu/239,240Pu activity ratio we deduce that the main
portion of plutonium originates from the nuclear weapons fallout
and 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-derived
239,240
Pu is approximately 0.5% of the activity of 239,240Pu from
nuclear test fallout in Finland. In this study the percentage of
Chernobyl-derived plutonium activity was calculated by taking
a logarithm from the measured 238Pu/239,240Pu ratios and fitting
these values to a linear function (Fig. 8) with an equation (1)

R ¼ 10 expðÀ0:08239z À 1:0468Þ (1)
238 239,240
where R is Pu/ Pu and z depth. This was done to describe
better the dominating trend in different depths of the sediment
241
Fig. 6. Am/239,240Pu activity ratio in the bottom sediments of Lake Paijanne in 2007.
¨ ¨ profile.


Table 4
The average fraction of Chernobyl-derived plutonium isotopes 238 and 239,240 in
Asikkalanselka bottom sediments (%) and the average 239,240Pu activity originating
¨
from the Chernobyl fallout (Bq/m2).
239,240
Depth The Chernobyl fallout fraction of Pu activity originating from
(cm) the total activity of plutonium the Chernobyl fallout (Bq/m2)
(%)
0.5 11.8 Æ 2.6 0.08 Æ 0.03
1.5 8.8 Æ 1.9 0.15 Æ 0.05
2.5 6.4 Æ 1.4 0.16 Æ 0.04
3.5 4.3 Æ 0.9 0.16 Æ 0.06
4.5 2.6 Æ 0.6 0.17 Æ 0.07
5.5 1.2 Æ 0.3 0.15 Æ 0.04
6.5 0.00 Æ 0.01 0.000 Æ 0.002
7.5 0.0 Æ 0.2 0.00 Æ 0.03
8.5 0.0 Æ 1.2 0.00 Æ 0.02

Asikkalanselka is 8.05 Â 108 m3 and the total area 5.28 Â 107 m2.
¨
The average activity in unfiltered water was 4.9 Æ 0.9 mBq/m3 and
4.1 Æ 0.2 mBq/m3 for 239,240Pu and 241Am, respectively. The
Fig. 7. The average 238Pu/239,240Pu activity ratio of four sampling sites A1, A2, A3 and percentage of the average total activity of 239,240Pu and 241Am in
¨
A4 as a function of depth in the bottom sediments of Asikkalanselka in 2007. ¨
water compared to the total activity in Asikkalanselka was calcu-
lated using equation (3):

The fraction of Chernobyl-derived plutonium in different depths ab
A% ¼ Â 100 (3)
was then calculated from the 238Pu/239,240Pu ratios determined ðcd þ abÞ
with equation (1) using equation (2)
where a is the nuclides’ average activity in water (Bq/m3), b is the
total volume of Asikkalanselka (m3), c is the nuclides’ average total
¨
Chernobyl derived Puð%Þ ¼ ½0:47R þ 0:026ð1 À RÞŠ Â 100 (2)
activity in bottom sediments (Bq/m2) and d is the total area of
where R is 238Pu/239,240Pu ratio. The percentages of total plutonium ¨
Asikkalanselka. It was calculated that in Asikkalanselka only ¨
activity originating from the Chernobyl fallout in various depths of 0.17 Æ 0.07% and 0.32 Æ 0.05% of the total activity of 239,240Pu and
¨ 241
bottom sediments of Asikkalanselka are presented in Table 4. The Am, respectively, is in the water column. These values are rough
total plutonium activities (Bq/m2) from the Chernobyl and nuclear estimates since there are two simplifications in the calculations.
test fallout are illustrated in Fig. 9. The 239,240Pu originating from First, the lake mirror area was used in the calculations instead of the
the Chernobyl fallout (Table 4) calculated from the average total actual bottom surface area, which is not known. Secondly, the
activities from the depths of 0–9 cm covers 1.95 Æ 0.01% of the total distribution of radionuclides in the bottom was assumed to be even,
239,240
¨
Pu activity in the bottom sediments of Asikkalanselka. This which is certainly not the case. Fortunately these simplifications
corresponds to a total activity of 0.9 Æ 0.3 Bq/m2 which corre- modify the results into opposite directions, since when the lake
sponds well to the levels found in this area by Paatero et al. (2002). mirror area is used the calculated total activity of the bottom
The majority of the 239,240Pu and 241Am activities have been sediments is smaller than if the actual bottom surface area was
transported to the bottom sediments. The water volume of used. Furthermore the activities in the central parts of the lake are

Fig. 8. The logarithmic 238Pu/239,240Pu activity ratios in the bottom sediments of Lake Fig. 9. Plutonium activity (Bq/m2) originating from the Chernobyl and nuclear test
¨ ¨nne in 2007. The 238Pu/239,240Pu activity ratio decreases as a function of depth
Paija ¨
fallout in the bottom sediments of Asikkalanselka in 2007. The broken line is nuclear
following a linear function y ¼ À0.08239x–1.0468. weapons test plutonium and the solid line Chernobyl-derived plutonium.


Table 5 Table 6
137
¨
Cs activity concentrations in the bottom sediments of Asikkalanselka in the depth A comparison between the results of the sediment profile A2 measured with
of 0–9 cm in 2007 and the total activity (S, Bq/m2) of sampling sites A1, A2, A3 and both semiconductor detector and NaI(Tl) detector. The error indicated for the depths
A4. The errors indicated for various depths are 1s counting error from gamma of 0–9 cm is 1s counting error from gamma spectrometry. The error indicated for the
spectrometry and for the total activities the error is the combined error of the total activities is calculated from the sum of the individual counting errors of depths
individual depths in quadratur. of 0–9 cm.

Depth (cm) A1 A2 A3 A4 Depth Semiconductor Counting error NaI(Tl) detector Counting error

A kBq/m2 A kBq/m2 A kBq/m2 A Bq/m2 Bq/g % Bq/g %
0–1 2.6 Æ 0.4 0.8 Æ 0.1 1.5 Æ 0.2 1.5 Æ 0.2 0–1 2.3 1.1 4.8 18
1–2 5.2 Æ 0.8 5.1 Æ 0.7 5.9 Æ 0.8 10.7 Æ 1.5 1–2 6.5 1.0 6.3 15
2–3 8.0 Æ 1.2 8.8 Æ 1.3 11.6 Æ 1.6 17.0 Æ 2.4 2–3 6.7 1.0 7.7 14
3–4 15.5 Æ 2.2 11.8 Æ 1.7 20.4 Æ 2.9 33.5 Æ 4.7 3–4 10.7 0.9 12.0 14
4–5 28.6 Æ 4.0 43.3 Æ 6.1 22.1 Æ 3.1 41.3 Æ 5.9 4–5 36.3 0.9 32.4 14
5–6 15.4 Æ 2.2 45.7 Æ 6.5 8.1 Æ 1.1 11.7 Æ 1.7 5–6 21.6 0.8 24.8 14
6–7 3.8 Æ 0.5 5.2 Æ 0.7 3.0 Æ 0.4 3.4 Æ 0.5 6–7 3.2 0.9 3.7 14
7–8 2.4 Æ 0.3 5.0 Æ 0.7 1.0 Æ 0.1 1.5 Æ 0.2 7–8 1.6 1.0 1.7 14
8–9 1.1 Æ 0.2 1.9 Æ 0.3 0.6 Æ 0.1 1.3 Æ 0.2 8–9 0.7 1.1 0.8 15
9–10 0.7 Æ 0.1 1.4 Æ 0.2 0.4 Æ 0.1 0.6 Æ 0.1 9–10 0.3 1.1 0.4 16
10–11 0.5 Æ 0.1 0.6 Æ 0.1 0.3 Æ 0.1 0.4 Æ 0.1 10–11 0.1 3.1 0.2 14
S 83.8 Æ 5.3 129.4 Æ 9.3 74.9 Æ 4.8 122.9 Æ 8.3
Total 90.1 0.9 94.9 14

presumably larger than the activities on the beach front. As these
values are used the calculated total activity of the bottom sedi- The depth profiles of 137Cs in the sampling sites studied showed
ments is larger than the actual activity. considerable variety. The total activities varied from 75 to 129 kBq/
m2 and the average value was 100 Æ 15 kBq/m2. In the deepest
4.4. Cesium in sediments and unfiltered water basin the total activity in sediment from A1 was 65% of the total
activity of sediment from A2 even though they were taken only
The activity concentrations of 137Cs in sediment samples a few meters from each other. On the other hand, on the basin
from the depths of 0–9 cm are presented in Table 5 and in fringes the activity of site A3 was only 61% of the activity on the
Fig. 10. The total activity of sediment profiles ranged between other site of the basin at point A4. In former studies other authors
75 Æ 11 and 129 Æ 18 kBq/m2 and the average total activity was have reported the uneven distribution of radionuclides in the
100 Æ 15 kBq/m2. ´
sediments. Saxen et al. (1998) and Kansanen et al. (1991) reported
The peak activities for 137Cs were in depths of 4–5 cm. The that the 137Cs activity concentrations showed considerable varia-
activities decreased rapidly in the sediment layers below the peak ¨ ¨
tions in the bottom sediments of southern Lake Paijanne and other
value. In 1988 (Saxen et al., 1998) the peak value of 137Cs in a point
´ Finnish lakes. According to Kansanen et al. (1991) the total 137Cs
corresponding the sampling sites A1 and A2, was on the surface of activity concentration in Asikkalanselka was 102 400 Bq/m2 in year
¨
the sediment profile. In years 1994 and 1997 the peak value was in ´
1989. Ilus and Saxen (2005) reported that in 2003 the total activity
a depth of 1.5–2.5 cm and 2–3 cm, respectively. In 2000 the peak was approximately 90 000 Bq/m2. According to the results of Saxen ´
´
was found in a depth of 3–4 cm (Ilus and Saxen, 2005). 29% of the et al. (1998) the total activity of 137Cs in a point corresponding to
total 137Cs activity was settled to the layers below the maximum our sampling sites A1 and A2 was 135 500 Bq/m2 in 1997. From
concentrations and in proportion 37% above the peak values. An physical decay this activity would correspond to an activity of
average of 8% more 137Cs activity was in the sediment layers above 108 000 Bq/m2 in year 2007. In our study the average activity of the
the stratums with maximum concentrations than in the lower sampling sites A1 and A2 was 106 600 Æ 5400 Bq/m2. The activity
layers. concentrations of 137Cs in the analyzed unfiltered surface water
samples were 18.3 Æ 0.6–20.3 Æ 0.5 Bq/m3. In 1998 137Cs activity in
the water of Lake Paijanne was 69 Bq/m3 (Saxen et al., 1998). Taking
¨ ¨ ´
into account the physical decay this would correspond to the
activity of 56 Bq/m3 in 2007. This indicates considerable transfer of
137
Cs from the water body to the bottom sediments.
Kansanen et al. (1991) reported observations which support the
view that redistribution and resuspension of 137Cs had significant
¨
impact in Asikkalanselka. These would cause the focusing of
radioactivity in the depressions. In our study clear evidence of the
redistribution 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 semiconductor
detectors to assure the results measured with NaI(Tl) detector
(Table 6). The activities measured with the semiconductor
detector 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

Fig. 10. The vertical distribution of 137Cs (Bq/m2) in sampling sites A1, A2, A3 and A4 in Based on the 238Pu/239,240Pu and 241Am/239,240Pu activity ratios,
¨
Asikkalanselka on 9.3.2007. the majority of the 239,240Pu and 241Am activities in the bottom


¨
sediments of Asikkalanselka originates from the 1950s and 1960s Holm, E., Rioseco, J., Pettersson, H., 1992. Fallout of transuranium elements
following the Chernobyl accident. Journal of Radioanalytical and Nuclear
nuclear test fallout. The impact of Chernobyl fallout to the 239,240Pu
Chemistry 156 (1), 183–200.
and 241Am activities in Asikkalanselka is minor compared to the
¨ Ilus, E., Saxen, R., 2005. Accumulation of Chernobyl-derived 137Cs in bottom sediments
´
nuclear test fallout. Only 1.95 Æ 0.01% of the total activity of of some Finnish Lakes. Journal of Environmental Radioactivity 82, 199–221.
239,240 Jaakkola, T., Keinonen, M., Hakanen, M., Miettinen, J.K., 1981. Investigation on the
¨
Pu in the bottom sediments of Asikkalanselka was calculated
transfer of plutonium and americium from plants to Reindeer and Man in Finnish
to have originated from the Chernobyl fallout. This corresponds to Lapland. In: Wrenn, M.E. (Ed.), Actinides in Man and Animals. RD Press, pp. 509–523.
the total activity of 0.9 Æ 0.3 Bq/m2. 239,240Pu and 241Am from the Kansanen, P.H., Jaakkola, T., Kulmala, S., Suutarinen, R., 1991. Sedimentation and
Chernobyl fallout are located in the uppermost parts of the bottom distribution of gamma-emitting radionuclides in bottom sediments of
¨ ¨
southern Lake Paijanne, Finland, after the Chernobyl accident. Hydrobiology
sediments. The peak values are found in depths of 5–7 cm for both 222, 121–140.
elements and in these layers they originate from the nuclear tests. Lieser, K.H., Steinkopff, T.H., 1989. Chemistry of radioactive Cesium in the hydro-
The average 241Am/239,240Pu activity ratio in the water body was sphere and in the geosphere. Radiochimica Acta 46, 39–47.
_ ¨ _
Lujaniene, G., Plukis, A., Kimtys, E., Remeikis, V., Jankunaite, D., Ogorodnikov, B.I.,
0.82 Æ 0.17 and correspondingly the average value in the bottom 2002. Study of 137Cs, 90Sr, 239,240Pu, 238Pu and 241Am behavior in the Chernobyl
sediments was 0.45 Æ 0.14. In the sediments the 241Am/239,240Pu soil. Journal of Radioanalytical and Nuclear Chemistry 251 (1), 59–68.
activity ratio was lowest in the surface layer and increased as _ ´
Lujaniene, G., Joksas, K., Silobritiene, B., Morkuniene, R., 2004. Physical and
chemical characteristics of 137Cs in Baltic Sea. In: Radionuclides in the Envi-
a function of depth. Furthermore the 241Am/239,240Pu ratio of the ronment, International Conference on Isotopes in Environmental Studies:
surface layer of the sediments has decreased during the last 11 Aquatic Forum 2004. Elsevier, Monaco, pp. 25–29.
years. This probably reflects the solubility of 241Am from the sedi- Outola Iisa, 2002. Effect of industrial pollution on the behaviour of 239,240Pu, 241Am
and 137Cs in forest ecosystems. Report Series in Radiochemistry 21/2002,
ments into the water body.
Helsinki.
The depth profiles of 137Cs in the sampling sites studied showed Paatero, J., 2000. Deposition of Chernobyl Derived Transuranium Nuclides and
considerable variety. The total activities ranged between 75 Æ 11 and Short-Lived Radon-222 Progeny in Finland. Finnish Meteorological Institute
129 Æ 18 kBq/m2 and the average total activity was 100 Æ 15 kBq/m2. Contributions No. 28. Finnish Meteorological Institute, Helsinki, p.
55 þ Appendixes.
The peak activities for 137Cs were in a depth of 4–5 cm. This peak ¨
Paatero, J., Jaakkola, T., Ikaheimonen, T.K., 2002. Regional distribution of Chernobyl-
corresponds to the Chernobyl fallout in 1986. The activities derived plutonium deposition in Finland. Journal of Radioanalytical and Nuclear
decreased rapidly in the sediment layers below the peak value. No Chemistry 252 (2), 407–412.
¨
Pilvio R., 1998. Methods for the determination of low-level actinide concentrations
clear evidence of the redistribution or focusing of 137Cs activity could and their behaviour in the aquatic environment. Report Series in Radiochem-
be proven. This might be a consequence of a low sample number. istry 10/1998, Helsinki, p. 43 þ Appendixes.
In bottom sediment samples the peak values of 137Cs were on Salminen, S., Outola, I., Jaakkola, T., Pulli, S., Zilliacus, R., Lehto, J., 2004. Method for
determining plutonium in air filters in detection of nuclear activities. Radio-
average 2 cm farther up the sediment profile than those of 239,240Pu chimica Acta 92, 467–473.
and 241Am. This is because 137Cs originated mainly from the Cher- Salminen, S., Paatero, J., Jaakkola, T., Lehto, J., 2005. Americium and curium
nobyl fallout while plutonium and americium came from nuclear deposition in Finland from the Chernobyl accident. Radiochimica Acta 93,
771–779.
weapons tests. Sanada, Y., Matsunaga, T., Yanase, N., Nagao, S., Amano, H., Takada, H.,
Tkachenko, Y., 2002. Accumulation and potential dissolution of Chernobyl-
Acknowledgements derived radionuclides in river bottom sediment. Applied Radiation and
Isotopes 56, 751–760.
The authors wish to thank Stewart Makkonen-Craig for the Saxen, R., Jaakkola, T., Rantavaara, A., 1998. 137Cs and 90Sr in the Southern Part of
´
¨ ¨
Lake Paijanne and its catchments. Radiochemistry 40 (6), 522–528.
language revision of the manuscript. Sidhu, R., 2006. Radiochemical Procedures for the Determination of Sr, U, Pu, Am
and Cm. Nordic nuclear safety research, Radchem, ISBN 87-7893-185-1.
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