Migration and bioavailability of137Cs in forest soil of southern Germany
I. Konoplevaa, E. Klemta, A. Konoplevb, G. Zibolda,*
aHochschule Ravensburg-Weingarten, University of Applied Sciences, 88250 Weingarten, Germany
bScientific Production Association ‘‘TYPHOON’’, Obninsk, Russia
a r t i c l e i n f o
Available online 23 January 2009
a b s t r a c t
To give a quantitative description of the radiocaesium soil–plant transfer for fern (Dryopteris carthusiana)
and blackberry (Rubus fruticosus), physical and chemical properties of soils in spruce and mixed forest
stands were investigated. Of special interest was the selective sorption of radiocaesium, which was
determined by measuring the Radiocaesium Interception Potential (RIP). Forest soil and plants were
taken at 10 locations of the Altdorfer Wald (5 sites in spruce forest and 5 sites in mixed forest). It was
found that the bioavailability of radiocaesium in spruce forest was on average seven times higher than in
mixed forest. It was shown that important factors determining the bioavailability of radiocaesium in
forest soil were its exchangeability and the radiocaesium interception potential (RIP) of the soil. Low
potassium concentration in soil solution of forest soils favors radiocaesium soil–plant transfer. Ammo-
nium in forest soils plays an even more important role than potassium as a mobilizer of radiocaesium.
The availability factor – a function of RIP, exchangeability and cationic composition of soil solution –
characterized reliably the soil–plant transfer in both spruce and mixed forest. For highly organic soils in
coniferous forest, radiocaesium sorption at regular exchange sites should be taken into account when its
bioavailability is considered.
? 2008 Elsevier Ltd. All rights reserved.
The majority of soils in semi-natural ecosystems (forests and
bogs) of the northern hemisphere are acidic, rich in organic
material and poor in nutrients such as potassium (Delvaux et al.,
2000). Soils in southern Germany contain present radiocaesium
inventories in the range 10–60 kBq m?2(a legacy from the Cher-
nobyl accident, 1986), which appears in solution only at trace
contamination level. For surface soils rich in clays (as is the case in
soils used for agriculture), most of this radiocaesium is adsorbed at
the selective sites of clay minerals and is, therefore, notavailable for
plant uptake. In forest soils, however, competition for the radio-
caesium between plant roots and the relatively small number of
selective sorption sites available leads to bio-recycling of the radi-
ocaesium by root uptake. To date, it has not been possible to
considerably minimize the137Cs soil–plant transfer in forest soil
and thus reduce the radiation dose to humans due to ingestion of
forest products such as mushrooms and game meat. One reason is
that the process of137Cs transfer from forest soil to plants is rather
complex and still a matter of considerable research.
As an example, enhanced activity concentrations of137Cs inwild
boar from Upper Swabia (south-west Germany) have been reported
since 2003, with measured values reaching above 8000 Bq/kg of
fresh mass (Klemt et al., 2005). The increase with respect to
previous years has been attributed tothe consumption by the boars
of highly contaminated deer truffles (Elaphomyces granulatus fr.)
(Fielitz, 2005). Deer truffles grow in spruce forest soil at depths of
about 7 cm under the surface, in the humus horizon. This organic
soil horizon contains the maximum fraction of the radiocaesium in
the soil profile which is taken up and accumulated by deer truffles.
In forest soils, humus is formed by litter decomposition. Litter is of
keyimportance in radionuclide redistribution in forestsoil. The rate
of litter decomposition is one of the most important factors in the
137Cs downward migration. This rate is governed by the type of
vegetation, climatic conditions, and the composition of the
underlying mineral layer.
Litter decomposes and transforms into humus with different
rates depending on its composition and, in particular, on its
nitrogen content (Dushofur, 1998). Deciduous litter decomposes
faster than coniferous litter and forms a thin humus layer near the
soil surface. Also, in deciduous forest the137Cs downward migra-
tion to the mineral layer has been shown to be faster (Scheglov,
1997). In contrast, coniferous litter, which is generally depleted in
nitrogen content and releases antibacterial substances, changes
very slowlyand forms a thick humus horizon. This horizon, situated
* Corresponding author. University of Applied Science, P.O. Box 1261, D-88241
Weingarten, Doggenriedstr., Germany. Tel.: þ49 751 94011.
E-mail address: email@example.com (G. Zibold).
Contents lists available at ScienceDirect
Journal of Environmental Radioactivity
journal homepage: www.elsevier.com/locate/jenvrad
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Journal of Environmental Radioactivity 100 (2009) 315–321
near the soil surface, accumulates radionuclides for a long time and
is the main source for the
The retention of the137Cs in the humus layer of forest soils has
been attributed mainly tothe presence of clay minerals (Maes et al.,
1998; Kruyts et al., 2004) and microbiological immobilization
(Guillitte et al., 1994; Brukmann and Wolters, 1994). The137Cs has
been shown to be selectively sorbed and fixed by clay minerals (2:1
structure) if present in organic rich soils even in small quantities
(Valcke, 1993; Rigol et al., 1998). The highly selective sites are
located at the expanded edges of the clay particle interlayers and
are called ‘‘frayed edge sites’’ (FES) (Cremers et al., 1988). The
bioavailability of the137Cs in soils depends on the number of these
selective sorption sites (FES), the cationic composition of soil
solution and the portion of exchangeable radiocaesium in soil
(Konoplev et al., 1999).
The137Cs concentration in the soil solution is the key charac-
teristic determining its uptake by plants (Konoplev et al.,1993). The
common measure of the radionuclide exchange between soil and
soil solution is the solid–liquid distribution coefficient (Kd). Since
part of the caesium in the soil solid phase is not available for an
exchange with the solution because of its irreversible sorption by
minerals, it is preferable to use the exchangeable solid–liquid
distribution coefficient (Kd
between soil and soil solution and to assess the137Cs concentration
in the soil solution. The Kd
tration of radionuclide reversibly sorbed by the solid phase to its
concentration in the liquid phase (Konoplev et al., 1992). The Kd
may be quantitatively estimated in terms of the radiocaesium
interceptionpotential, RIP, and the concentration of competing ions
in the soil solution(Sweecket al.,1990). Note thatRIP is aproductof
the capacity of the highly selective frayed edge sites (FES) and
137Cs uptake by grazing plants and
ex) to describe the partition of
exis defined as the ratio of the concen-
FES(Cs/M), the selectivity coefficient of Cs in relation to competing
RIPðMÞ ¼ KFES
where M¼Kþor NH4
Values of RIP(M) characterize the potential ability of soils to
selectively and reversibly adsorb Csþand the values allow the
calculation of Kd
ðCs=MÞ ? ½FES?
þdepending on cationic scenario.
exfor other ion scenarios on the basis of the
Therefore, the solid–liquid distribution of137Cs is essentially
governed bythe FES capacityand the soil solution concentrations of
concentrations of stable Cs and137Cs are generally very small in
comparison to the number of competing ions in the soil solution
(e.g. 1 kBq137Cs corresponds to about 10?12M). Therefore,137Cs in
most soils and sediments is a trace contaminant and is more or less
completely sorbed on FES. In this case the strong influence of the
capacity of FES on137Cs soil–plant transfer becomes plausible. In
soils with a large content of organic matter the FES are not the only
controlling factor and the regular exchange complex plays an
important role. In the high organic matter soils, the Kd
calculated on the basis of the equation:
þ. The FES are highly selective for
137Cs and the
dðCsÞ ¼ KFES
where RES are the Regular Exchange Sites on the external surfaces
of clay particles and humus substances. Therefore, Kd
ðCsÞ þ KRES
Forest soils have important characteristic features: (1) the
presence of litter and organic horizons with different thicknesses,
(2) a depth distribution of roots characterized by plant species, and
(3) a heterogeneous distribution of radiocaesium in the soil profile
depending on sorption characteristics of the soil horizons. Never-
theless it is still common to use the aggregated transfer factor Tagto
quantify the transfer of137Cs from soil to plants, mushrooms or
The aggregated transfer factor Tag(in m2/kg) is defined by the
caesium activity concentration (in Bq/kg) of the dry mass of the
plants, divided by the total inventory (in Bq/m2) of the soil.
Thus, the value of Tagis related to the total inventory, which
means that the geometry of plant roots, the depth distribution of
137Cs, and the availabilityof the137Cs toplants in the rootzone layer
are not taken into consideration. This causes high variability of Tag.
Measured values of Tagfor a plant can vary by a factor 100–1000,
depending on conditions in soil and sampling (IAEA, 1994).
An improved description of the soil–plant transfer can be ach-
ieved if the activity concentration of the plant is related to the
activity concentration in the soil horizon where the plant roots are
located (Konoplev et al., 1996,1998). According to Fig. 1 this would
be the Of-horizon for wood sorrel and the Ohhorizon for fern. This
transfer factor is named TF which is also known as the concentra-
tion ratio, CR. Several efforts have been undertaken to predict TF
values using soil chemical properties, e.g. the exchangeability of
radionuclides in the soil (Oughton et al., 1992; Konoplev et al.,
1993), the concentrations of radionuclides and of competing
cations in the soil solution (Shaw et al.,1992; Smolders et al.,1997;
Roca et al.,1997). Concerning137Cs, a fruitful hypothesis has been to
assume that the TF is proportional tothe fraction of137Cs in the root
exchange complex, which depends on the composition of the soil
solution (Smolders et al., 1997). Using this hypothesis, Konoplev
et al. (1996, 1998) and Konoplev and Konopleva (1999) developed
a model which describes the bioavailability of
expressed in the form:
137Cs in soil as
TF ¼ A ? B ¼aex
ðNH4=KÞ ? ½NH4?SS
In this formula, A is the ‘‘availability’’ factor, B characterizes the
ability of a specific plant to sorb radiocaesium on the root exchange
complex and transfer it through the cell wall, aexis the exchange-
ability of radiocaesium in soil, RIP is a measure for the selective
Fig. 1. The root distribution of fern (Dryopteris carthusiana) and wood sorrel (Oxalis
acetosella) in the different soil horizons as compared to the depth distribution of137Cs
in the soil of a spruce forest (Klemt et al., 1996).
I. Konopleva et al. / Journal of Environmental Radioactivity 100 (2009) 315–321316
sorption of137Csþ, Kc
ions with respect to the competing ion Kþ, in rectangular brackets [
] are the cation concentrations in the soil solution. This model was
successfully tested for forest soils from Upper Swabia, Russia,
Sweden and Switzerland (Konoplev et al., 1999, 2000). Delvaux
et al. (2000) discovered low concentrations of soluble K in the
rhizosphere to be the cause of increased uptake of137Cs and they
verified the above relation betweenTFandRIP for ryegrass (Lolium).
Kruyts et al. (2004) showed that the accumulation of organic
material in the top soil can cause a decrease of RIP in the thick
humus horizons of forest soil and thus lead to an increase of the
The objective of this paper was to study the migration and
bioavailability of137Cs in soils of ‘‘spruce forest’’ and ‘‘mixed spruce
and beech forest’’ in the pre-alpine region in south-western
Germany in connection with soil characteristics and soil solution
composition. Special attention was devoted to the effect of poten-
tially elevated ammonium concentrations in humified layers of
FES(NH4/K) is the selectivity coefficient of NH4
2. Materials and methods
2.1. Sampling sites, altitude, precipitation, soil type, geology
The area is located 30 km north of Lake Constance in the south of Germany
(Center of the forest: Gauss Krueger coordinates (PD) 3552980; 5300190, see Fig. 2).
It is characterized by a mixture of forest areas (about 25%), agricultural land, bogs
and small lakes at an altitude of about 650 m. Mean annual temperature is about
8?C. At Bad Schussenried the following average values for the years 1980–2005
were recorded: maximum temperature: 17.8?C in July, minimum temperature:
?1.1?C in January, maximum precipitation: 116 mm in July, minimum precipitation:
49 mm in February, and a total precipitation of 916 mm on average per year. The
annual precipitation varies locally between 700 and 1400 mm. Altdorfer Wald
comprises about 60 km2of forest mainlyspruce, Picea abies and mixed forest (beech,
Fagus, and spruce, P. abies). The main type of soil is Luvisol with a tendency to
podsolic Luvisol belonging to the soil family mottled loam. The geology of the
bedrock is mainly moraine. A schematic map of the area under study and the
sampling sites are presented in Fig. 2.
Soil and plant sampling: Soil material was taken as a monolith. A volume of
dimension of about 30 cm ?20 cm area, and depth of about 25 cm was dug and
transported to the laboratory. In the laboratory this sample was divided according to
the different horizons. Area, thicknesses, and weightof the horizons were measured.
After removal of stones and tree roots, the soil material was dried in air and sieved
using a mesh size of 2 mm. At each sampling site all plants growing on the monolith
were collected, separated according to their species (fern, blackberry), dried at
105?C and crushed in a mixer. Plant samples located within a circle of about 20 m
around the monolith were also harvested and prepared in the same way. The137Cs
activity concentration was determined by gamma spectrometry using HPGe detec-
tors. Measuring times were chosen in order to achieve a statistical uncertainty
smaller than 5%. On September 15th 2005, the137Cs inventory in the forest soil
varied at the sites studied between 8000 and 26,000 Bq/m2.
Soil pH was measured in 0.01 M CaCl2using a solid–liquid (S/L) ratio of 1:2.5 for
mineral layers and in S/L ratio of 1:5 for organic layers after 2 h of equilibration.
Soil texture was determined using a hydrometer method after destruction of
organic matter with 30% hydrogen peroxide (Gee and Bauder,1986). Classification of
the soil by grain size was done according to USDA (US Department of Agriculture)
Theorganic mattercontent in samples was determinedbyloss on ignition (10 h at
RIP determination (Wauters et al., 1996): The soil sample (about 1 g) was
equilibrated with a mixed potassium–calcium solution (0.5 mM KClþ100 mM
CaCl2) to mask the RES by Ca2þand to saturate FES by Kþ. After pre-saturation
(3?20 h), a phase separation was conducted by centrifugation and the soil sample
was equilibrated with the same K–Ca solution spiked with137Cs. After 24 h the
distribution coefficient Kd (137Cs) was obtained by measuring the Cs activity
remaining in the solution. The productof Kd(137Cs) and the potassium concentration
represents the value of RIP. The RIP in a NH4scenario, RIP(NH4), was measured in the
same way, only NH4Cl was used instead of KCl.
137Cs exchangeability and exchangeable cations: After determining its
activity, the soil sample (50 g) was equilibrated with 1 M NH4OAc during 24 h using
a solid–liquid ratio of 1:10 for mineral layers and of 1:20 for organic layers. The soil
suspension was centrifuged and the solution was filtered through a 0.45-mm
membrane filter. The ratio of
exchangeability. Contents of exchangeable Ca, Mg, and K were measured in a 1 M
NH4OAc extract. Exchangeable ammonium was determined in the same way using
a 2 M KCl extract.
Soil solution isolation: To collect soil solution, a syringe without plunger was
supplied with a paper filter and with glass fiber as a plug. Then it was filled with
a soil sample. The syringe with the soil was centrifuged for 1 h at 1700 g. The soil
solution was filtered through a 0.45-mm membrane filter.
Cation content (K, Ca, and Mg) in soil solution was determined using atomic
absorption spectrophotometry (AAS). The uncertainties are 0.1 mg/L for K and Mg
and 0.2 mg/L for Ca. The NH4in soil solution was measured with a colorimetric
method by indophenol reaction (Krom, 1980). The uncertainty was about 5% as
tested by replicates.
137Cs in the solution to that in the soil is its
3. Results and discussion
3.1. Depth distribution of137Cs in forest soil
Typical depth distributions of137Cs in soil of mixed forest(a) and
spruce forest (b) are shown in Fig. 3. It can be noted that 19 years
afterdeposition still more than 50% of the137Cs activity is located in
the upper 10-cm soil layer with a peak of the activity concentration
in the Ahhorizon. The organic horizons in spruce forest soils are
thickerand richer in137Cs than those in mixed forest soils. In spruce
forest soils, plant roots are located mainly in the Oh horizon,
whereas in mixed forest soils roots are mainly in the Ahhorizon.
Our data show that in spruce forest about 50% of the total137Cs
inventory is located in the root zone; in mixed forest it is about 30%.
Fig. 3 shows that the organic Ohhorizon in spruce forest soil
accumulates137Cs and thus prevents its downward migration. As
a result, the penetration of137Cs to deeper soil layers of spruce
forest is more limited in comparison to that in a mixed forest.
3.2. Main soil characteristics
Main soil characteristics of the root zone layers are given in
Table 1. All soils were acidic with pH values between 3.6 and 5.8 in
mixed forest, and pH values between 2.8 and 3.4 in spruce forest.
Soil textures ranged from sandy clay loam to sandy loam. Sand is
Fig. 2. Sites under study in forest Altdorfer Wald. The location of the cities Weingarten
and Ravensburg are indicated.
I. Konopleva et al. / Journal of Environmental Radioactivity 100 (2009) 315–321 317
the dominant grain size fraction with 46–65% in the mineral
Organic humified layers of forest soils could be characterized by
elevatedlevels of ammoniumconcentrations. Ammonium
effectively mobilizes137Cs from soil sorption sites causing its higher
mobility and plant availability (Sanchez et al., 1999).
At all 10 sampling sites the vertical distribution of ammonium
concentration in the soil solution was measured. It is known that
137Cs, Bq/(m2 cm)
0 500 100015002000
137Cs, Bq/(m2 cm)
Fig. 3. Depth distribution of the137Cs activity concentration in soil of a mixed forest (a) and of a spruce forest (b). Organic horizons are above the maximum of the distribution,
mineral horizons are below.
Texture, grain size and physico-chemical soil properties of root zone layers.
OMa, % Grain sizeb
TextureExchangeable cations TRBc,
Clay, %Silt, % Sand , %K, cmol kg?1
Ca, cmol kg?1
Mg, cmol kg?1
Sandy clay loam
Sandy clay loam
Sandy clay loam
Sandy clay loam
Sandy clay loam
Sandy clay loam
Sandy clay loam
Sandy clay loam
Sandy clay loam
aOrganic matter content determined as losses on ignition at 450?C.
bGrain size distribution was determined for B soil layers.
cTRB is total reserve in bases.
I. Konopleva et al. / Journal of Environmental Radioactivity 100 (2009) 315–321 318
ammonification is a bacterial mineralization of organic substances
and it is accelerated with increasing concentrations of dissolved
organic matter and oxygen. In agreement with this idea, highest
ammonium concentrations were measured in well aerated Of
horizons. Root zone layers in spruce and mixed forest differ
essentially in organic matter content (Table 1); however, differ-
ences in soil solution ammonium concentrations in those layers are
not pronounced in both types of forest (Fig. 4). On average, the soil
solution potassium concentration in the root zone exceeded the
Potassium concentrations ranged from 0.07 to 0.53 mM and
ammonium concentrations from 0.05 to 0.25 mM. Low soil solution
Kþconcentrations in the root zone have an essential influence on
the plant availability of137Cs (Zhu and Smolders, 2000).
The dependence of the137Cs uptake by plants on potassium in
the Kþconcentration range of less than 1 mM has beenwell studied
(Smolders et al.,1996; Waegeneers et al., 2001). The CF (the plant to
solution concentration ratio) for137Cs is reduced with increasing K
in solution and the largest effect of K on Cs uptake was found in the
Kþconcentration range less than 0.25 mM (Zhu and Smolders,
2000). So, observed low concentrations of potassium in soil solu-
tion may cause higher137Cs soil–plant transfer.
þconcentrations by about a factor of 1.7 as shown in Table 2.
3.3. Radiocaesium interception potential of the soils under study
The RIP value characterizes the ability of soils to selectively
adsorb Csþ. The RIP values were determined for Ohand Ahhorizons
of the soils under study. Fig. 5 illustrates the dependence of RIP on
the fraction of clay in soil. Although the clay mineralogy can differ
for different layers and sites, a linear dependence between RIP and
the clay fraction was found. This allows the assumption that the
clay mineralogy is more or less similar at all investigated sites.
The exchangeability of
NH4OAc varies in a range from 1.8 to 5.7% in spruce forests and is
generally higher in comparison with mixed forests with 0.4 to 3.4%.
Our data show that the higher values of the137Cs exchangeability
correspond to lower values of RIP; however, this correlation is
In the series of mixed forest soils, we have found a negative
linear correlation (r¼?0.9) between exchangeability and RIP(K)
values. This fact indicates that the same kind of clay minerals could
be responsible for137Cs selective sorption and interlayer fixation,
and that137Cs is adsorbed in the mineral part of mixed forest soil.
For coniferous forests, RIP(K) values appeared to be inversely
proportional to the exchangeability (not shown in the figures).
A set of RIP(NH4) was measured in order to estimate the selec-
tivitycoefficient Kc(NH4/K) on FES (Table 2). The values of Kc
K) are in a range from 1.4 to 2.8. Taking into account the measured
potassium and ammonium concentrations this means that in most
sites ammonium plays a more important role than potassium as
a competing cation of radiocaesium for the selective sorption sites
FES (Table 2).
The RIP(K) values ranged from 23 to 149 mequiv./kg in the Oh
horizon in coniferous forest (sites 6–10) and decreased with
increasing organic matter content. The RIP values are proportional
to the portion of clay minerals in Ohand Ahlayers (see Fig. 5). The
small portion of clay minerals in the humus layers of spruce forest
137Cs obtained by extraction with
Soil solution NH4 concentration, mM
Fig. 4. Mean NH4concentration in the soil solution of the root zone of forest soil.
Selected characteristics of soils,137Cs exchangeability aex, soil solution Kþ, NH4
137Cs soil–plant transfer factors attributed to root zone TF.
þ, Ca2þand Mg2þconcentrations, percentage of137Cs on RES, calculated availability factors (A),
FES(NH4/K)K, mM NH4, mM Ca, mM Mg, mM% Kd
A, 10?4mM1/2kg/mequiv.TF, Bqkg?1/Bqkg?1
Clay content, %
Fig. 5. Dependence of RIP on the fraction of clays in the soil.
I. Konopleva et al. / Journal of Environmental Radioactivity 100 (2009) 315–321319
is caused by the low rate of litter decomposition. Coniferous litter
decomposes very slowly and forms a thick humus horizon like
mor or moder. In the soil samples taken in the Ohlayers with the
mor humus, the RIP values were almost negligible, and notable
values were observed in the samples from Oh horizons with
For soils in mixed forests essentially higher RIP(K) values as
compared with spruce forests have been observed in the range
from 224 to 632 mequiv./kg. Just on the basis of this comparisonwe
may expect that radiocaesium soil–plant transfer for the mixed
forests should be lower than that for spruce forests.
3.4. Predicted availability factor A and measured values of RIP
In Fig. 6 the predicted ‘‘availability’’ factor A is plotted versus the
measured values of RIP, in logarithmic scale. A dependence
(squares) is shown taking into consideration the selective sorption
of Csþon FES only. It can be fitted by a straight line with a negative
slope of (?1.7?0.2). We also considered sorption both on FES and
RES and this dependence is shown as triangles. In this case factor A
was calculated according to the following equation:
For spruce forest with small FES values, the A values are now
smaller and the slope of the straight line decreased slightly
(?1.5?0.2). In soils of mixed forests the selective sorption of Csþ
on FES dominates as indicated by the negligible decrease of the A
values for higher RIP values.
Fig. 7 presents the dependence of TF attributed to the root zone
for fern and blackberry on the RIP of the correspondent soil layer
for both types of forest. The graph clearly shows that RIP plays
a predominant role in characterizing forest soils in terms of radi-
In Fig. 8 predicted values of the availability factor A versus
measured values of the transferfactorTFare plotted in linear scales.
The linear dependency between A and TF is better fulfilled for
blackberry than for fern. It can be seen from Figs. 7 and 8 that as
expected (Drissner et al., 1998) fern has substantially higher TF
values in comparison to blackberry. It is important to note that the
dependencies presented in Figs. 7 and 8 are comprising both spruce
and mixed forests. These dependencies can be described bya single
function for both forest types.
RES) was calculated according to Eq. (4).
1. Soil–plant transfer factors of137Cs attributed tothe root zone in
soils of spruce forest are on average an order of magnitude
higher than that in soils of mixed forest.
2. Relatively low potassium concentrations in the soil solution of
the root zone (less than 0.25 mM in most cases) for all inves-
tigated sites are in favor of an elevated soil–plant transfer.
3. Concentrations of ammonium and potassium in root zones
have been found to be close to each other. However, taking into
account that ammonium is a more efficient competitor to
radiocaesium than potassium (Kc(NH4/K) varies from 1.4 to 2.8)
we conclude that in most investigated sites ammonium is
a more important mobilizer of radiocaesium than potassium.
4. Besides the ratio of the ammonium to the potassium concen-
tration in soil solution, the most important factors determining
the radiocaesium bioavailability in forest soils and finally the
aggregated transfer factor Tagare the137Cs exchangeability and
the RIP value in the root zone.
RIP (K) meq/kg
Availability factor A 10-4mM1/2kg/meq
FES + RES
Fig. 6. Availability factor A versus RIP(K) in logarithmic scale. Mixed forest stands are
found to the lower right.
10 100 1000
TF, Bq kg-1/Bq kg-1
Fig. 7. Dependence of the137Cs transfer factor TF attributed to the root zone for fern
(upper) and blackberry (lower) on radiocaesium interception potential RIP of the
corresponding soil layer.
R2 = 0.56
R2 = 0.90
Availability factor A, 10-4mM1/2kg/meq
TF, Bq kg-1/Bq kg-1
Fig. 8. Transfer factor TF for the root zone versus availability factor A for fern and
I. Konopleva et al. / Journal of Environmental Radioactivity 100 (2009) 315–321320
5. In soils of spruce forest with low RIP, the role of the regular Download full-text
exchange sites RES is comparable to that of FES concerning the
sorption of radiocaesium and finally its bioavailability. Kd
accounted for 11–49% of the total Kd
6. A linear dependence of the radiocaesium transfer factor TF
attributed to the root zone for fern and blackberry on the
bioavailability factor was found for the whole set of forest sites,
spruce sites as well as mixed sites. This means that the
bioavailability of radiocaesium is determined mostly by phys-
ico-chemical characteristics of the root layer. Some physico-
chemical properties of the soil profile (thickness of humus
layer, pH, selective sorption capacity of Cs) are characterized by
the type of forest which can be used to predict the radio-
caesium bioavailability and its transfer factor.
exfor spruce forest sites
Funding by Baden-Wu ¨rttemberg Projektra ¨gerschaft Lebens-
grundlage Umwelt und ihre Sicherung (BWPLUS) project No. BWR
24018 ‘‘Migration und Bioverfu ¨gbarkeit von Radioca ¨sium in Bo ¨den
Su ¨ddeutschlands’’ and continuous support by FD Dr. Bosch and FD
Maluck and their co-workers during sampling are gratefully
Brukmann, A., Wolters, V., 1994. Microbial immobilization and recycling of137Cs in
the organic layers of forest ecosystems: relationship to environmental condi-
tions, humification and invertebrate activity. Sci. Total Environ. 157, 249–256.
Cremers, A., Elsen, A., De Preter, P., Maes, A., 1988. Quantitative analysis of radio-
caesium retention in soils. Nature 335, 247–249.
Delvaux, B., Kruyts, N., Cremers, A., 2000. Rhizospheric mobilization of radiocesium
in soils. Environ. Sci. Technol. 34, 1489–1493.
Drissner, J., Bu ¨rmann, W., Enslin, F., Heider, R., Klemt, E., Miller, R., Schick, G.,
Zibold, G., 1998. Availability of caesium radionuclides for plants – classification
of soils and role of mycorrhiza. J. Environ. Radioact. 41, 19–32.
Dushofur, F., 1998. New data on humification in forest soils of temperate climate.
Pochvovedenie (Soil Sci.) 7, 883–889 (in Russian).
Fielitz, U., 2005. Untersuchungen zum Verhalten von Radioca ¨sium in Wild-
schweinen und anderen Biomedien des Waldes, Schriftenreihe Reaktorsicher-
heit und Strahlenschutz, BMU-2005-675, ISSN 1612–6386.
Gee, G.W., Bauder, J.W., 1986. Particle-size analysis. In: Klute, A. (Ed.), Methods of
Soil Analysis. Part 1-Physical and Mineralogical Methods. Agronomy Mono-
graph, No. 9. American Society of Agronomy, Inc., Soil Science of America, Inc.
Madison, Wisconsin USA, pp. 383–411.
Guillitte, O., Melin, J., Wallberg, L., 1994. Biological pathways of radionuclides
originating from the Chernobyl fallout in a boreal forest ecosystem. Sci. Total
Environ. 157, 207–215.
in Temperate Environments. Technical Reports Series No. 364. IAEA, 74 pp.
Klemt, E., Gu ¨ner, I. Putyrskaya, V., Semizhon, T., Zibold, G., 2005. Datenerhebung zur
Radioca ¨sium Kontamination im Jahr 2005. Abschlussbericht 2005, LfU Werk-
lived Chernobyl radionuclides in a soil–water system. Analyst 117,1041–1047.
Konoplev, A.V., Viktorova, N.V., Virchenko, E.P., Popov, V.E., Bulgakov, A.A.,
Desmet, G.M., 1993. Influence of agricultural countermeasures on the ratio of
different chemical forms of radionuclides in soil and soil solution. Sci. Total
Environ. 137, 147–162.
Konoplev, A.V., Drissner, J., Klemt, E., Konopleva, I.V., Zibold, G., 1996. Parameter-
isation of radiocaesium soil–plant transfer using soil characteristics. In:
Proceedings of XXVIth Annual Meeting of ESNA. Working Group 3: Soil–Plant
Relationships, Busteni (Romania), 12–16 September 1996, 147–153.
Konoplev, A.V., Avila, R., Bulgakov, A.A., Drissner, J., Johanson, K.-J., Klemt, E.,
Konopleva, I.V., Miller, R., Nikolova, I., Popov, V.E., Zibold, G., 1998. Modelling
radiocaesium bioavailability in soil. In: Proceedings of International Union of
Radioecologists Topical Meeting, SCK-CEN, Mol, Belgium, 1–5 June 1998,
Konoplev, A.V., Konopleva, I.V., 1999. Characteristics of steady-state selective
sorption of radiocesium on soils and bottom sediments. Geochemistry Inter-
national 37, 177–183.
Konoplev, A.V., Avila, R., Bulgakov, A.A., Drissner, J., Johanson, K.-J., Klemt, E.,
Konopleva, I.V., Miller, R., Nikolova, V.E., Popov, V.E., Zibold, G., 1999. Modelling
radiocaesium bioavailability in forest soils. In: Linkov, I., Schell, W.R. (Eds.),
Contaminated Forests. Proceedings of the NATO Advanced Research Workshop,
Kiev (Ukraine), 24–28 June 1998, pp. 217–229.
Konoplev, A.V., Avila, R., Bulgakov, A.A., Johanson, K.-J., Konopleva, I.V., Popov, V.E.,
2000. Quantitative assessment of radiocaesium bioavailability in forest soils.
Radiochim. Acta 88, 789–792.
Krom, M.D., 1980. Spectrophotometric determination of ammonia: a study of
a modified Berthelot reaction using salicylate and dichloroisocyanurate. Analyst
Kruyts, N., Titeux, H., Delvaux, B., 2004. Mobility of radiocaesium in three distinct
forest floors. Sci. Total Environ. 319, 241–252.
Maes, E., Delvaux, B., Thiry, Y., 1998. Fixation of radiocaesium in acid brown forest
soil. Eur. J. Soil Sci. 49, 133–140.
Oughton, D.H., Salbu, B., Riise, G., Lien, H., Oestby, G., Noeren, A.,1992. Radionuclide
mobility and bioavailability in Norwegian and Soviet soils. Analyst 117,
Rigol, A., Vidal, M., Rauret, G.J., Shand, C.A., Cheshire, M.V., 1998. Competition of
organic and mineral phases in radiocaesium partitioning in organic soils of
Scotland and the area near Chernobyl. Environ. Sci. Technol. 32, 663–669.
Roca, M.C., Vallejo, V.R., Roig, M., Tent, J., Vidal, M., Rauret, G., 1997. Prediction of
cesium-134 and strontium-85 crop uptake based on soil properties. J. Environ.
Qual. 26, 1354–1362.
Sanchez, A.L., Wright, S.M., Smolders, E., Naylor, C., Stevens, P.A., Kennedy, V.H.,
Dodd, B.A., Singleton, D.L., Barnett, C.L.,1999. High plant uptake of radiocaesium
from organic soils due to Cs mobility and low soil K content. Environ. Sci.
Technol. 33, 2752–2757.
Scheglov, A.I., 1997. Biogeochemistry of technogenic radionuclides in forest
ecosystems in central regions of Eastern-European plain. Thesis for doctor
degree, Moscow, 45 pp.
Shaw, G., Hewamanna, R., Lillywhite, J., Bell, J.N.B., 1992. Radiocaesium uptake and
translocation in wheat with reference to the transfer factor concept and ion
competition effects. J. Environ. Radioact. 16, 167–180.
Smolders, E., Kiebooms, L., Buysse, J., Merckx, R., 1996.137Cs uptake in spring wheat
(Triticum aestivum L. cv Tonic) at varying K supply. I. The effect in solution
culture. Plant Soil 181, 205–210.
Smolders, E., Sweeck, L., Merckx, R., Cremers, A., 1997. Cationic interactions in
radiocaesium uptake from solution by spinach. J. Environ. Radioact. 34,
Sweeck, L., Wauters, J., Valcke, E., Cremers, A., 1990. The specific interception
potential of soils for radiocaesium. In: Desmet, G., Nassimbeni, P., Belli, M.
(Eds.), Transfer of Radionuclides in Natural and Semi-Natural Environments.
Elsevier Applied Science, pp. 249–258.
Valcke, E.,1993. The behaviour dynamics of radiocesium and radiostrontium of soils
rich in organic matter. PhD thesis, Faculty of Agricultural and Applied Biological
Sciences. Katholieke Universiteit Leuven, Belgium.
Waegeneers, N., Camps, M., Smolders, E., Merckx, R., 2001. Genotypic effects in
phytoavailability of radiocaesium are pronounced at low K intensities in soil.
Plant Soil 235, 11–20.
Wauters, J., Elsen, A., Cremers, A., Konoplev, A.V., Bulgakov, A.A., Comans, R.N.J.,
1996. Prediction of solid/liquid distribution coefficients of radiocaesium in soils
and sediments. Part one: a simplified procedure for the solid phase character-
isation. Appl. Geochem. 11, 589–594.
Zhu, Y.G., Smolders, E., 2000. Plant uptake of radiocaesium: a review of mecha-
nisms, regulation and application. J. Exp. Bot. 51, 1635–1645.
I. Konopleva et al. / Journal of Environmental Radioactivity 100 (2009) 315–321 321