Introduction

The Chernobyl reactor breakdown disturbed global gamma-ray radiation for a long time and by varying intensities. The breakdown also changed the radiation environment in Poland. Radioactive air pollution in Poland at the end of April through May of 1986 exceeded the values registered before the breakdown many times over. The main components of the pollution were inert gases, iodine and caesium isotopes. Pollutants were transferred to the soil by atmospheric deposition, and local variation in rainfall resulted in uneven soil pollution.

After the Chernobyl breakdown, most countries introduced monitoring of post-Chernobyl 137Cs concentrations in soil. Studies were carried out in countries including Belarus (Knatko et al. 1996), Chile (Schuller et al. 2000), Israel (Lavi et al. 2006), Japan (Takenaka et al. 1998), Jordan (Hamarneh et al. 2003), Canada (Blagoeva and Zikovsky 1995), Mexico (Gaso et al. 1998), Germany (Schimmack et al. 1997), Poland (Jagielak et al. 1997), Russia (Strand et al. 1999), Serbia and Montenegro (Bikit et al. 2005), Switzerland (Riesen et al. 1999), Sweden (Melin et al. 1994) and Italy (Giovani et al. 1994).

From 1980 to 1990 in Poland, the geographical distribution of radiocaesium pollution (137Cs and 134Cs) was studied by three independent teams from the Central Laboratory for Radiological Protection (Biernacka et al. 1991), the Institute of Nuclear Physics of the Polish Academy of Science PAN (Mietelski et al. 1996) and the National Geological Institute (Strzelecki et al. 1993). Maps that were produced from these studies agreed with the geographical pollution distribution, with most 137Cs deposition (more than 50 kBq m−2) in the Opole–Nysa–Kłodzko region in southwest Poland, known as the Opole Anomaly.

The objective of this study was to assess historical changes in the surface 137Cs concentrations in woodland and wasteland soils in the Opole Anomaly and to assess the potential influence of 137Cs translocation mechanisms on these changes. To allow comparison, 137Cs activity in epigeic mosses (Pleurozium schreberi) and epiphytic lichens (Hypogymnia physodes) was tested.

Materials and methods

The locations of the woodland and wasteland study sites were in the area within the co-ordinates 50°–51° N and 15.5°–18.5° E and are shown in Fig. 1.

Fig. 1
figure 1

Location of the study area

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The surface 137Cs concentrations were determined for randomly selected locations, distributed evenly throughout the study area. We sampled 87 measurement points in 2007 and 35 in 2010. We collected samples of P. schreberi mosses and H. physodes lichens in 2010 at 31 randomly selected points in woodlands in the study area. We sampled from locations only when both mosses and lichens were present. Approximately 20 g of mosses and 20 g of lichens were collected from each sampling site. Lichens and mosses were cleaned and dried at a temperature of up to 303 K.

Equipment and reagents

The surface 137Cs activities were detected using a Canberra portable scintillation spectrometer InSpector1000. This is a portable, digital, multichannel analyser and is suitable for monitoring in situ. It can be used to determine radiation dose and power, identify the source location, for nuclide identification with activity measurements in the field, and for spectrum acquisition and analysis. Genie 2000 software (Gamma Analysis Option model S501C) was used to analyse the spectrum.

The 137Cs activity measurements were determined using a gamma spectrometer with a high-resolution germanium detector HPGe (Canberra) as follows: 1.29 keV (FWHM) at 662 keV and 1.70 keV (FWHM) at 1332 keV. The relative efficiency was 21.7 %. Energy and efficiency calibration of the gamma spectrometer energy and efficiency was done using a MBSS 2 standard solution (Czech Metrological Institute, Prague, Czech Republic), which covers an energy range from 59.54 to 1836.06 keV. The calibration source geometry was Marinelli (447.7 ± 4.5 cm3), had a density of 0.99 ± 0.01 g/cm3, and contained 241Am, 109Cd, 139Ce, 57Co, 60Co, 137Cs, 113Sn, 85Sr, 88Y and 203 Hg. The Marinelli sample container had a volume of 450 cm3. GENIE 2000 software was used to measure and process the spectra 24 h a day. The measurement uncertainty for 137Cs, estimated by the GENIE 2000 software, was approximately 1–2 %.

An ordinary kriging technique was used to produce isoline maps of 137Cs surface activity distribution and the specific activity of 137Cs in mosses and lichens. An exponential model was used to describe the experimental semi-variance. The geostatistical software Surfer (Golden Software Ltd.) was used to prepare the maps based on the WGS 1984 reference co-ordinate system.

Results

Table 1 presents information on the surface activity of 137Cs in the study area for 1994, 2007 and 2010, and the specific activity of 137Cs accumulated in the sampled mosses and lichens in 2010. The values of the surface activity of 137Cs registered in 1994 were read from the map (Jagielak et al. 1997), which displays surface activity 137Cs isolines in steps of 5 kBq m−2. The data from this period do not distinguish values of surface activities greater than 65 kBq m−2.

Table 1 The activity factors of 137Cs accumulated in soil during the years 1994, 2007 and 2010 and in mosses and lichens in 2010
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There was a decrease in the surface 137Cs activity in soil over time. To assess the influence of translocation on 137Cs activity, we compared our results for 2007–2010 with those reported by Jagielak et al. (1997). The measured 137Cs activities at 2 April 2013 were calculated using formula 1:

$$ A(t)={A}_0\cdot \exp \frac{- \ln 2\cdot t}{t_{1/2}} $$
(1)

where A 0 = initial activity of 137Cs, A(t) = activity after time t and t 1/2 = half-life (t 1/2Cs =30.08 years (National Nuclear Data Center)).

The distribution of the surface activity of 137Cs in 1994 and for 2007–2010 is presented in Figs. 2 and 3. The letters A–D mark the parts of the study area that had the highest surface 137Cs activity.

Fig. 2
figure 2

Distribution of surface activity of 137Cs in the area of the Opole Anomaly in 1994 (Jagielak et al. 1997), calculated for 2 April 2013

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Fig. 3
figure 3

Distribution of surface 137Cs activity in the Opole Anomaly for 2007–2010, calculated for 2 April 2013

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It should be noted that the 137Cs activity distribution shown in the map for 1994 (Fig. 2) does not cover the whole area that was studied in 2007 and 2010, but that it mainly covers the area bounded by the co-ordinates: 50°–51° N and 15.5°–18.5° E.

Figure 4 presents a semi-variogram of how the surface 137Cs activity changed in the study area in the years 2007–2010.

Fig. 4
figure 4

Variability in surface 137Cs activity in the study area for 2007–2010

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The theoretical model of the semi-variogram accurately describes the dependence defined by the experimental semi-variogram and suggests that the assumed model of the spatial distribution of surface 137Cs activities that was used to prepare the map shown in Fig. 3 was accurate.

The distribution of the specific 137Cs activity accumulated in mosses and lichens is shown in Figs. 5 and 6. As mentioned before, mosses and lichens are good biosorbents of heavy metals and 137Cs.

Fig. 5
figure 5

Distribution of the specific 137Cs activity in mosses collected in 2010

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Fig. 6
figure 6

Distribution of the specific 137Cs activity in lichens collected in 2010

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The maps (Figs. 5 and 6) were based on theoretical models of semi-variograms, which showed good agreement with experimental semi-variograms as shown in Fig. 4.

Compared with the surface 137Cs activity of soil, the highest specific activities in mosses and lichens were recorded at locations A and B.

Discussion

To meet the objectives of the study, we analysed soils to help us understand historical changes in the surface 137Cs activity accumulated in soil. The statistical parameters of surface 137Cs activities in the study area, calculated using formula 1 as at 2 April 2013, based on data from 1994, 2007 and 2010, are shown in Table 2. For comparison, data from 2007 and 2010 were back-calculated for 1 June 1986.

Table 2 Statistical parameters regarding the surface activity concentration of 137Cs in the studied area (kBq m−2), calculated as at 2 April 2013 and 1 June 1986, on the basis of the data from 1994 and the years 2007 and 2010
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There was little variation in the values of statistical parameters for surface 137Cs activity calculated for 2 April 2013, based on data from 1994, 2007 and 2010. It has been shown that changes in the surface activity of 137Cs in soil are mainly due to radioactive disintegration of the radionuclide. This can be confirmed by comparing the surface activity concentrations of 137Cs, measured in 2007 and 2010. If we assume that the reduction in the amount of the radionuclide is due only to radioactive decay, then, based on formula 1, 137Cs activity concentrations have decreased such that its activity ratio a2007/a2010 in those two measurement periods lies within the range 1.09–1.12. This activity ratio was calculated for five measurement points located within 2 km of each other. The mean of the calculated ratios was 1.13, which is slightly higher than that estimated by the model that assumes that change in radionuclide concentrations at a given location is due only to radioactive disintegration.

The results show limited 137Cs migration into the soil profiles. It has been demonstrated that approximately 90 % of 137Cs is stored in the top 10 cm of soil (e.g. Takenaka et al. 1998; Beli et al. 1994; Ziembik et al. 2010).

The values of surface 137Cs activity in soil, back-calculated for 1 June 1986 (Table 2), show that the amount of 137Cs deposited after the Chernobyl nuclear power plant breakdown in the Opole Anomaly area (mean 26.4, max. 125 kBq m−2) is comparable to that deposited in areas of Southern Belarus (Knatko 1996).

We compared the specific 137Cs activity accumulated in mosses and lichens using values of their respective specific 137Cs activities from the maps (Figs. 5 and 6) for locations where the geographical network coincided with the map. This comparison showed a linear correlation as follows: amoss = 0.952·alichen–0.03. This linear correlation method has been used in other studies to estimate recent deposition of various analytes (Kłos et al. 2010, 2011). Numerous examples show that recent deposition of analytes can be defined by the comparison factor (CF) >0.62 (CF = 2 × (clichen − cmoss)/(clichen + cmoss), where clichen and cmoss = concentration of an analyte in lichens and mosses, respectively). The mean CF value (0.15) calculated from measurements of the specific 137Cs activity in mosses and lichens shows that it did not drift in 2010, which agrees with the results published by Message (2010).

Surface 137Cs activity in soil (Fig. 3) and radionuclide activity in mosses and lichens were not significantly correlated (Figs. 5 and 6), nearly 25 years after the 137Cs deposition. As stated before, the maximum activities overlap only at points A and B marked on the map. Therefore, the specific 137Cs activity in mosses and lichens cannot be used as an indicator of the surface activity of soil. Studies have suggested that metal cations may translocate from soil to epigeic mosses and epiphytic lichens via dust from soil and, in the case of mosses, via water (Kłos et al. 2012). However, the bioavailability of 137Cs differs, depending on the physical and chemical characteristics of soil and of dust that originates from the soil. Statistically significant correlations between the specific 137Cs activity in mosses and lichens and the specific 137Cs activity of mobile forms in the surface layer of soil released in cation form to a solution with a pH of 3.9 were reported by Kłos et al. (2009). Various studies of the specific 137Cs activity in mosses, carried out in the Yamal Peninsula (Nifontova 1995), large areas of the Ural and Siberia (Nifontova 2006) and near Ekaterinburg, where the radionuclide was deposited in large quantities after the Chernobyl disaster (Nifontova 2006), have shown that the level of 137Cs accumulation is influenced by biological half-life time (t B1/2). The completed studies clearly suggest that t B1/2 < t 1/2. In summary, accumulation of radiocaesium and other pollutants in mosses and lichens depends on the biological half-life and climate conditions, and pollutants in soil reflect polluting atmospheric aerosols. This study, similar to other studies that have documented translocation of 137Cs from soil to higher plants (Tagami et al. 2012; Lukšienė et al. 2013), demonstrates that there may be local dispersion of 137Cs accumulated in soil. 137Cs in the environment is tied up mainly in topsoil, and its circulation is limited to genetic levels O–A, from which it is integrated into the plant structure via the root system, creating a new sub-level Ol over time (Ziembik et al. 2010; Dołhańczuk-Śródka et al. 2006; Dołhańczuk-Śródka and Wacławek 2007). Such local dispersion can be illustrated by wider zones, limited by isolines, created by comparing the results from studies carried out in 2007 and 2010 (Fig. 3) with those from 1994 (Fig. 2). Other studies have shown that radioactive disintegration and migration are the main causes of decreased 137Cs activity in woodlands (Isajenko et al. 2010, 2012).

The lack of correlation between the 137Cs activity in soil, mosses or lichens does not diminish the importance of using these organisms as biomonitoring indicators of radionuclide environmental pollution. Indeed, biomonitoring studies, initiated as early as the 1960s, are still being continued (Yoshida et al. 2004; Adelinė et al. 2006; Celik et al. 2009).

Conclusions

Results from this study show that the decreasing activity of 137Cs deposited in woodlands and wastelands after the breakdown of the Chernobyl nuclear power plant is mainly due to radioactive disintegration of 137Cs. These changes are also, to a lesser degree, caused by the migration of 137Cs with dust from soil and translocation to plants.

Relative to other parts of Poland, high 137Cs activities are still detected in the Opole Anomaly area. The average surface activity of 137Cs calculated for 2 April 2013 was 14.3 kBq m−2, and the maximum surface activity concentration was 125 kBq m−2. 137Cs concentrations detected in the mosses sampled in 2010 were also high, with mean specific concentrations in lichens and mosses of 190 and 155 Bq kg−1, respectively.