Radionuclide transfer to wildlife at a ‘ Reference site ’ in the Chernobyl Exclusion Zone and resultant radiation exposures

This study addresses a signi ﬁ cant data de ﬁ ciency in the developing environmental protection framework of the International Commission on Radiological Protection, namely a lack of radionuclide transfer data for some of the Reference Animals and Plants (RAPs). It is also the ﬁ rst study that has sampled such a wide range of species (invertebrates, plants, amphibians and small mammals) from a single terrestrial site in the Chernobyl Exclusion Zone (CEZ). Samples were collected in 2014 from the 0.4km 2 sampling site, located 5km west of the Chernobyl Nuclear Power complex. We report radionuclide ( 137 Cs, 90 Sr, 241 Am and Pu-isotopes) and stable element concentrations in wildlife and soil samples and use these to determine whole organism-soil concentration ratios and absorbed dose rates. Increasingly, stable element analyses are used to provide transfer parameters for radiological models. The study described here found that for both Cs and Sr the transfer of the stable element tended to be lower than that of the radionuclide; this is the ﬁ rst time that this has been demonstrated for Sr, though it is in agreement with limited evidence previously reported for Cs. e ﬀ ects on wildlife in the CEZ generally relate observations to ambient dose rates determined using handheld dose meters. For the ﬁ rst time, we demonstrate that ambient dose rates may underestimate the actual dose rate for some organisms by more than an order of magnitude. When reporting e ﬀ ects studies from the CEZ, it has previously been suggested that the area has comparatively low natural background dose rates. However, on the basis of data reported here, dose rates to wildlife from natural background radio- nuclides within the CEZ are similar to those in many areas of Europe.


Introduction
In environmental radiation protection, the estimation of activity concentrations in organisms is one of the largest uncertainties in the prediction of dose rates received by wildlife (e.g. Vives i Batlle et al., 2007;Beresford et al., 2008a;Johansen et al., 2012). Furthermore, transfer parameters are not available for many radionuclide-organism combinations (ICRP, 2009;IAEA, 2014;Brown et al., 2016). To address this lack of data, the International Commission on Radiological Protection (ICRP) (2009) suggested identifying a series of sites where each of the 'Reference Animals and Plants' (RAPs) considered in the ICRP assessment framework (ICRP, 2008) could be collected and analysed.
To date such sites, in the terrestrial environment, have been sampled in Norway (Thørring et al., 2016), Spain (Guillén et al., 2018) and England ; the Norwegian study also sampled marine and freshwater RAPs. The RAPs are defined at the taxonomic level of family and for terrestrial ecosystems they are: Reference Wild grass (Poaceae); Reference Pine tree (Pinaceace); Reference Earthworm (Lumbricidae); Reference Bee (Apidae); Reference Rat (Muridae); Reference Deer (Cervidae); Reference Duck (Anatidae); and Reference Frog (Ranidae).
The approximately 4760 km 2 area abandoned after then 1986 Chernobyl accident is heterogeneously contaminated by a range of radionuclides, including 90 Sr, 137 Cs, 241 Am and Pu-isotopes (Kashparov https et al., 2017). The area gives the opportunity to study the transfer of these radionuclides to a range of wildlife (e.g. Ryabokon et al., 2005;Barnett et al., 2009;Beresford et al., 2016;Gaschak et al., 2018). It also allows studies of the effects of radiation on different wildlife taxa (e.g. Chesser and Baker, 2006;Møller et al., 2013). However, there is considerable contention over the interpretation of effects studies conducted in the vicinity of Chernobyl and dose rates are often poorly estimated (see Beresford et al. this issue; Beaugelin-Seiller et al. this issue).
In this paper we determine transfer parameters and radiation dose rates at a site in the Ukrainian Chernobyl Exclusion Zone (CEZ) from which a range of species were sampled, including those falling within the definition of the terrestrial RAPs.
All the data from the study (including individual measurements of radionuclides and stable elements) are freely available from Beresford et al. (2018).

Sample site
The sampling site (0.4 km 2 ) was located towards the western edge of the 'Red Forest', approximately 5 km west-southwest of the Chernobyl Unit Number 4 (Fig. 1). The site was not within the areas where pine trees were killed by high exposure levels in 1986. Most of the site was formerly used as kitchen gardens ('dacha') by the residents of Pripyat and it still has fruit trees. With the exception of Pinus sylvestris, all samples were collected from an area of the former kitchen gardens of about 0.06 km 2 in area (Fig. 2); this is subsequently referred to as the 'inner sampling area'. The predominant soil type of the sampling site was soddy-podzolic sandy loam and the surrounding habitats were largely deciduous woodland (some of which was previously agricultural land) and marsh.

Sampling
All samples were collected over a period of about 1 month in May/ June 2014. Although sampling was focussed on species falling into the ICRP RAP definitions (after Barnett et al., 2014), additional species caught were analysed for 90 Sr and 137 Cs activity concentrations.

Wild grass
The perennial Poaceae species Agrostis gigantea (black bent grass) was sampled from the inner sampling area. The area was walked on a grid pattern with A. gigantea being sampled to approximately 1 cm above the ground surface at regular intervals on the grid. The sample was placed into one of three collection bags (the first sample being placed into bag #1, the second sample into bag #2, the third into bag #3, the fourth into bag #1, etc.). The samples were air dried (20-25°C) and then homogenised prior to analyses.

Pine tree
Trunk wood from Pinus sylvestris (Scots pine), a species in the Pinaceace family, was sampled from three felled trees which were estimated to be > 28 year old (i.e. to predate the 1986 accident); trunk is   Gaschak, Chornobyl Center). With the exception of the pine trees all samples were collected from the inner sampling area (the points marked within the inner sampling area denote location of earthworm collection). the ICRP RAP geometry. Additionally, samples of needles, cones and branches were collected from each felled tree. The samples were dried at 75°C before being homogenised for subsequent analyses.

Earthworm
Earthworms (Lumbricidae family, most likely Eisenia hortensis), were collected from six sites within the inner sampling area predominantly from under old fruit (pear, apple, plum and cherry) trees. In the laboratory the earthworms were rinsed in water to remove external adhering soil and then kept in aerated containers with damp tissue paper to allow evacuation of gut contents for approximately three days. One set of sub-samples were used for radioanalyses and the other subsample (comprising 15 individuals from each site) were freeze-dried prior to stable element analyses.

Bees and other insects
Bees were collected using12 pan collectors ¾ filled with water (Westphal et al., 2008); the collectors had been sprayed with fluorescent yellow paint before being deployed within the inner sampling area. The traps were checked and emptied at least every three days between mid-May to Mid-June. Species other than bees were collected in the pan traps and sampled animals were separated by taxa. Bees sampled falling within the ICRP definition of the RAP (i.e. in the Apidae family) were Xylocopa spp. (carpenter bee) and Bombus spp. (bumblebee). Other insect species collected and retained as samples were: Tropinota spp. (scarab beetle); species in Elateridae family (click beetle); Cetonia spp. (chafer beetle); Vespa spp. (hornet). The number of individuals collected per species ranged from 13 (bumblebee) to 96 (chafer beetle). Separated samples were stored frozen, prior to drying at 20-25°C for subsequent analyses; only samples of bees were analysed to determine stable element concentrations.

Small mammals
Small mammals were trapped from the inner sampling area using Sherman human traps over five nights in June 2014; 200 traps were deployed each night. A total of 166 animals of seven species were caught with Apodemus agrarius (striped field mouse, n = 94) and Myodes glareolous (bank vole; n = 46) being the most abundant. Other species caught were Apodemus flavicollis (yellow-necked mouse; n = 12), Microtus agrestis (field vole, n = 1) Microtus spp., (vole; n = 3), Muscardinus avellanarius (common dormouse; n = 1), Sorex araneus (common shrew; n = 8) and Sorex minutus (Eurasian pygmy shrew; n = 2). Because of the relatively high numbers caught, some A. agrarius and M. glareolous were released without further processing. All other animals were live-monitored (see below) and mass, sex and approximate age recorded. With the exception of the A. flavicollis, all animals were subsequently released at the study site. Nine individuals of A. flavicollis (a species falling within the ICRP Rat RAP definition) were euthanised and then ashed for subsequent radiochemical analyses. The remaining three A. flavicollis were dissected and the following samples removed: hind-leg muscle, hind-leg bone, liver, testes or embryo depending on sex, and a bulked sample comprising the spleen, kidneys and lungs. These samples were stored frozen, prior to freeze-drying and subsequent stable element analyses.

Amphibians
A plastic amphibian fence was erected in the inner sampling area with a number of pit traps being created at gaps in the fence. However, this was not a very efficient way to collect samples and catching by hand was used to collect most of the animals caught. Species caught were: Rana arvalis (moor frog; n = 12); Bombina bombina (European fire-bellied toad; n = 6); Bufo bufo (European toad; n = 4) and Pelobates fuscus (common spadefoot toad; n = 7). All animals were live-monitored (see below) and their mass recorded. Apart from individuals of R. arvalis all animals were subsequently released at the study site. Nine R. arvalis were ashed for subsequent radiochemical analyses and the remaining three, all males, were dissected to obtain the samples of the same tissue type as collected from A. flavicollis; samples were stored frozen, prior to freeze-drying for subsequent analyses to determine stable element concentrations.
Frogspawn (egg mass) was collected in early April 2015 from an area of flooded bog and freeze-dried prior to stable element analyses.

Soils
Fifteen, 10 cm deep soil cores (2.5 cm diameter) were collected from an area of 3-4 m radius around each of the three sampled pine trees; these were then bulked into one sample per tree. Ambient dose rate was determined, using a MKS-01R meter, at a height of 1 m above the ground at each soil sampling site. Soil samples were collected in a similar manner from each of the six earthworm sampling sites and from 19 further sites in the vicinity of the various animal traps. In total 28 soils samples were collected from the site. Samples were dried at 80°C before being homogenised.
All samples were analysed to determine 137 Cs and 90 Sr activity concentrations. For actinide ( 241 Am and Pu-isotopes) analysis soil samples collected from the inner sampling area were bulked to give five samples (each bulk sample comprising five individual samples with consecutive sampling numbers); all three pine tree bulk soil samples were analysed for actinides. Individual pine tree soil samples and the five bulks from the inner sampling area were also analysed to determine stable element concentrations (note these bulks each comprised five different samples selected at random and hence were not the same bulk samples as used for actinide analyses).

Live-monitoring
In total 118 animals were live-monitored including: 37 A. agrarius, 24 M. glareolous and 12 R. arvalis (i.e. species falling within the definitions of the ICRP Rat and Frog respectively). The wholebody 137 Cs and 90 Sr concentrations were determined using the method described by Bondarkov et al. (2011) as previously summarised in Beresford et al. (2016). Prior to counting, the animals were placed in a small, disposable, cardboard box (70 × 40 × 40 mm), the upper side of which was made from < 0.1 mm thick polyethylene. The box was then placed inside a lead shielded counting container. The detectors comprised a hyper-pure germanium detector and thin-film (1 mm) NaI scintillation detector to measure 137 Cs and 90 Sr, respectively. The 137 Cs spectra were analysed using the Canberra Genie-2000 software package. The activity concentration of 90 Sr was determined from that of its daughter nuclide, 90 Y. The method has previously been calibrated against phantoms containing 137 Cs and 90 Sr and validated against traditional radiochemical extraction and analysis methodologies. Counting times varied from 150 to 1200 s depending upon the radioactivity in the animal. Counting errors were typically < 3% for 90 Sr and < 7% for 137 Cs.

Radioanalyses
To determine 90 Sr and 137 Cs activity concentrations in samples, other than those in small mammals and amphibians that were livemonitored, the methods described in Penrose et al. (2016) were used. Samples were first homogenised using a domestic coffee grinder and then 10 g dry mass (DM) aliquots accurately weighed into petri-dishes.
Caesium-137 activity concentrations were measured using a Canberra-Packard gamma-spectrometer with a high-purity germanium (HPGe) detector (GC 3019). A standard 152 Eu source (OISN-16; Applied Ecology Laboratory of Environmental Safety Center, Odessa, Ukraine) comprising epoxy granules (< 3.0 mm) with the density of 1 g cm −3 with used for calibration. The minimally detectable activity was 0.18 Bq per sample with uncertainties of around 10-15% depending on the sample type.
The 90 Sr concentrations in soil, plant and invertebrate samples were measured spectrometrically without any radiochemical pretreatment.
The procedure used a β-spectrometer EXPRESS-01 with a thin-filmed (0.1 mm) plastic scintillator detector. Spectra were processed by a correlation with the measured spectra from standard sources, such as: 90 Sr+ 90 Y, 137 Cs and the 90 Sr + 90 Y, and 137 Cs combinations as well as from background. Daily calibrations were conducted. A more detailed description of method principle can be found in Bondarkov et al. (2002Bondarkov et al. ( , 2011 and Gaschak et al. (2011); uncertainties were around 20%.
For actinide analyses, before processing of the samples, 242 Pu and 243 Am were added as yield tracers. To determine Pu and Am isotopes all samples, except for soils, were initially dissolved in 65% HNO 3 . Soil samples were dissolved in HF followed by treatment with HNO 3 , HCl, H 3 BO 3 +HNO 3 and then 8M HNO 3 . Plutonium and Am were separated using anion exchange resin (Bio Rad AG 1 × 8, 100-200 mesh). The Pu fraction was evaporated and thin alpha sources prepared for measurement on an alpha spectrometer. Americium was precipitated with calcium oxalate and then separated using TRU resin columns (IAEA, 1999); lanthanides were then removed using an anion exchange resin column. Subsequently, thin alpha sources were prepared for measurement of 241 Am on an alpha spectrometer. Thin alpha sources of each separated actinide element were prepared by micro-coprecipitation with neodymium fluoride and measured using a Canberra Alpha Analyst alpha spectrometer. Counting errors were typically < 20% for the Pu and Am isotopes.

Stable element analyses
Acid digestions were undertaken to determine concentrations of 29 elements by ICP-MS. Though not discussed in this paper, I concentrations were also determined if the sample size was sufficient; the methodology and results for I can be found in Beresford et al. (2018).

Extractions
Approximately 0.2 g of dry soil was accurately weighed into a Savillex™ vial, adding concentrated Primar grade HNO 3 (4 mL) and heating at 80°C overnight using a teflon-coated graphite hot block (to pre-digest the organic matter contained in soils). The next step consisted of adding concentrated Primar grade HF (2.5 mL), HNO 3 (2 mL) and HClO 4 (1 mL). A stepped heating program up to 160°C overnight was applied to fully digest silicate and oxide phases. The dry residue was re-constituted after warming with 2.5 mL ultrapure MilliQ water and 2.5 mL HNO 3 and the final volume made up to 50 mL. The National Institute of Standards and Technology) NIST Standard Reference Material (SRM) 2711a 2 (Montana soil) in duplicate and four blanks were all digested in a similar manner to check the accuracy and precision of the digestion and analysis methods. All the digests were diluted 1-in-5 before analysis.
Plant material (approximately 0.2 g dry matter (DM)) was accurately weighed into digestion vessels and 6 mL concentrated Primar grade HNO 3 added. The samples were digested using a Multiwave PRO Anton Paar microwave reaction system, with heating at 140°C for 20 min and further cooling to 55°C for 15 min. Once the digestion was complete, the samples were made to a final volume of 20 mL. Digestion of NIST SRM 1573a (Tomato Leaves) and four blanks were all undertaken for quality control. Prior to analysis, the acid digests were diluted 1-in-15 to give a final matrix of 2% HNO 3 .
A portion of animal tissue (up to circa 0.2 g DM where available) was accurately weighed into digestion vessels and a mixture of 3 mL Primar grade HNO 3 + 3 mL MilliQ ultrapure water + 2 mL 30% v/v H 2 O 2 was added. The samples were allowed to froth for 20 min in uncovered vessels and they were then microwave digested at 140°C for 20 min. Once the digestion was complete, the extracts were made to a final volume of 20 mL. Two replicates of NIST Controlled Reference Material (CRM) 1577c (Bovine Liver) and five blanks were all prepared in a similar manner. Prior to analysis, the acid digests were diluted 8fold to give a final HNO 3 concentration of approximately 2%. Full dissolution was achieved for all samples with the exception of earthworms, which appeared to contain traces of soil.
In general, satisfactory elemental recoveries for the soil, plant and animal certified reference materials were obtained (see Fig. S1). Specifically for NIST 1573a and NIST 1577c, recoveries of 100 ± 15% were reported for the majority (18 out of 24 and 22 out of 23, respectively) of certified and non-certified elements. A slightly broader range was reported for NIST 2711a (the soil) with recoveries of typically 100 ± 25% for the majority (18 out of 24). None of the elements discussed in this work showed recoveries outside of the quoted ranges for the three different reference materials analysed.

Analyses
Multi-element analysis of diluted solutions in acid matrix was undertaken by ICP-MS (Thermo-Fisher Scientific iCAP-Q). Further technical detail of the ICP-MS runs can be found in Beresford et al. (2018).
Detection limits reported were calculated as three times the standard deviation of the reagent blanks for each extraction form and sample type. Results were reported for the following elements: Al, As, B, Ba, Be, Ca, Cd, Co, Cr, Cs, Cu, Fe, K, Li, Mg, Mn, Mo, Na, Ni, P, Pb, Rb, S, Se, Sr, Tl, U, V and Zn.

Dose assessment
To determine the total exposure of organisms at the study site Tier 3 (probabilistic assessment) of the ERICA Tool  version 1.2.1 was used. To determine external exposure rates to all organisms the arithmetic mean and standard deviation of the bulked soil dry matter activity concentrations (see Table 1 below) were used assuming a soil dry matter content of 100% and a lognormal distribution (see Brown et al., 2008). For animals and plants the arithmetic mean and standard deviation of the measured data (fresh mass (FM)) (see Table 2 below) were used for each species generally assuming a lognormal distribution; where the number of samples was less than three an exponential distribution was assumed. For P. sylvestris the activity concentration in trunk wood was used (trunk being the default ERICA Tool and ICRP (ICRP, 2008) geometry). For both bee (Apidae) and Microtus species data for the sampled species were averaged and then used as the input activity concentrations for the assessment. As a combined result for 239,240 Pu was reported, for the dose assessment it was assumed that each isotope contributed 50% of the total activity concentration. If for a given species there were no data for either 241 Am or Pu-isotopes then the same value was assumed as for the species of  For each small mammal and amphibian species, specific geometries were created in the ERICA Tool using species masses determined in the study and dimensions obtained from literature and on-line sources (see Supplementary Information). The relevant ERICA Tool default geometries were used for the bee species, earthworm and plants. Occupancy factors (fraction of time spent in soil, on soil or in air) were derived for each species based upon relevant information (see Supplementary  Information); amphibian species were assumed to spend 100% of their time in the terrestrial environment. The Tools default radiation weighting factors of 10 for alpha, 3 for low energy beta and 1 for other beta/gamma were used.

Results
In the text below, we present summarised data; all of the underlying data for this study are presented in the accompanying dataset . This includes individual sample radionuclide and stable element concentrations, together with information such as animal live mass, sex and approximated age, and dose estimates (external, internal and total). Results for wildlife are presented on a fresh matter (FM) basis and those for soil on a dry matter (DM) basis. Statistical comparisons discussed below were performed using Mintab 17.

Radionuclide activity concentrations in soil
Summarised radionuclide activity concentrations in bulked soil samples are presented in Table 1. Individual soil sample results, for 90 Sr and 137 Cs, can be found in the accompanying dataset ; actinide results are only available for the bulked samples. Radionuclide activity concentrations in soils collected in the vicinity of the sampled pine trees were in the range of those collected from the inner sampling area, with the exception of higher Pu-isotope concentrations (by a factor of two to three). Activity concentrations of both 90 Sr and 137 Cs ranged over two orders of magnitude across the site (E +3 to E+5 Bq kg −1 DM).

Radionuclide activity concentrations in wildlife
Summarised radionuclide activity concentrations in the different species of wildlife are presented in Table 2. Activity concentrations for all animals are presented as whole-body values.
A general linear model was used to test for significant differences (Tukey pairwise comparisons; 95% confidence) in radionuclide activity concentrations between species with sufficient sample numbers: 137 Cs -B. bombina and M. glareolus had significantly higher activity  Replication was insufficient to statistically compare Pu-isotope activity concentrations. However, it is worth noting that, as for 241 Am, Pu-isotope activity concentrations were comparatively high in Lumbricidae spp.
For the majority of samples 137 Cs and 90 Sr activity concentrations were comparable, with a tendency for 137 Cs to be higher (Table 2). However, there were some exceptions, notable was that 90 Sr concentrations in the trunk wood of P. sylvestris were more than an order of magnitude higher than 137 Cs activity concentrations. Whilst 90 Sr concentrations were also comparatively high in branches and needles of this species, for cones the 137 Cs activity concentrations were about an order of magnitude higher than 90 Sr values (Table 3). Both A. gigantea and Lumbricidae spp. had consistently higher 90 Sr than 137 Cs activity concentrations as did the majority of samples of both shrews analysed and also two of the amphibian species (B. bufo and P. fuscus).

Concentration ratios
The majority of available wildlife assessment models use concentration ratios (CR wo-soil ) (Beresford et al., 2008a) Table 4 presents summarised CR wo-soil values estimated for organisms at the study site. For all species with the exception of P. sylvestris the mean soil activity concentrations calculated from the individual samples from the inner sampling site was used for calculating CR wo-soil values; for P. sylvestris, activity concentrations in the soils taken from the tree sampling locations were used. For Pu, the CR wo-soil values are based on 239,240 Pu data.
Statistical differences between CR wo-soil values for different species were the same as those reported above for activity concentrations.

Stable element data
The focus of this paper is to present the measured radionuclide activity concentrations and use these to estimate concentration ratios and dose rates. The stable element data are only used here in the discussion of these results, i.e. comparisons of stable element and radionuclide transfer to organisms and the status of important analogues such as Ca and K. Consideration is also given to: (i) elements which inform on natural background exposure rates (namely K and U) as it has been suggested that background dose rates in the area of the CEZ are comparatively low (Møller and Mousseau, 2011) though there is little available data to support this, and also (ii) pollutant elements given the increasing interest in multi-stressor exposure (e.g. Hinton et al., 2013), the potential for elevated concentrations of elements such as Pb or B as a consequence of these having been dropped on the burning reactor in 1986 (Jagoe et al., 1998) and the lack of data for such elements within the CEZ. All the stable element data, including tissue specific values for the vertebrate species, are presented within Beresford et al. (2018).
Summarised concentrations of K, Ca, Sr, Cs, Pb and U in the bulked soil samples analysed are presented in Table 5; the Ca concentration in soil from the P. sylvestris collection points was circa 25% of that from soil in the inner sampling area. Data for the same elements in wildlife samples are presented in Table 6. To estimate the activity concentrations of 40 K and 238 U in soil and wildlife (Tables 5 and 6) we have assumed 31.6 Bq 40 K g −1 K and 12.21 Bq 238 U mg −1 U (Beresford et al., 2008b).
Concentration ratios for stable Cs and Sr were estimated as above for radionuclides and are presented in Table 7. To determine fresh mass concentrations in wildlife dry:fresh weight ratios from Barnett et al. (2013Barnett et al. ( , 2014 were used; the exception was frogspawn for which a dry:fresh weight ratio of 0.37 was used (Barnett, unpublished). Wholebody concentrations were estimated from the sum of the total content of each element in sampled tissue, and dividing this by the mass of the sampled tissues assuming this was representative of the whole-body; an approach previously used in similar studies Guillén et al., 2018). To estimate total muscle and bone masses of A. flavicollis and R. arvalis data on the proportions these tissues contribute to the live-weight of Apodemus  and Anura (Barnett, unpublished) species were used respectively.
There is greater variability across the species in CR wo-soil values for 90 Sr and 137 Cs compared to their stable elements. Stable element CR wosoil values also tended to be lower than values for the radioisotopes. This was most noticeable for 90 Sr for which the CR wo-soil value for P. sylvestris trunk wood was more than 70-times higher than the stable element CR wo-soil value; whilst the P. sylvestris CR wo-soil value for 90 Sr was an order of magnitude higher than other species considered in Table 7, that for stable Sr was less than those for some of the other species.
Figs. 3 and 4 show the contributions of internal and external exposure to total dose rate, and the different radionuclides to internal dose rates respectively. For P. sylvestris, internal dose dominated the total dose because of the comparatively high 90 Sr activity concentrations in trunk wood. Internal dose was comparatively more important for the amphibian species (41-74% of total dose rate), again this reflected the comparatively high 90 Sr activity concentrations in these species. Internal dose was estimated to be comparatively unimportant for Apidae and Lumbricidae spp., contributing less than 10% of the total dose rate. This is largely the consequence of the comparatively low 90 Sr and 137 Cs activity concentrations in these species, though their Table 5 Mean ± SD dry matter (DM) concentrations of K, Ca, Sr, Cs, Pb and U in the bulked soil samples form the inner sampling area (n = 5) and the Pine tree site (n = 3); estimated activity concentrations of 40 K and 238 U are also shown. With the exception of M. glareolus, Apidae and Lumbricidae spp., 90 Sr contributed more than 50% of the internal dose rate (Fig. 4). In the case of P. sylvestris and two of the amphibian species > 90% of the internal dose rate was due to 137 Cs. The contribution of actinide radionuclides was < 10% of the internal dose rate for most of the species. The only organism for which actinide radionuclides were estimated to contribute significantly to internal dose was Lumbricidae spp. for which they comprised > 40% of the internal dose rate and about 40% of the total dose rate (the external dose rate being relatively low for this organism).

Discussion
To our knowledge, this is the first reported study to compare the radionuclide activity concentrations in a wide range of wildlife sampled from a given location within the CEZ.

Activity concentrations and concentration ratios
For some of the organisms sampled there are comparatively few published CR wo-soil values. For example, ICRP (2009) presents no data for bee species and data for amphibians are limited (ICRP, 2009;IAEA, 2014).
There were significant differences in the radionuclide activity concentrations (and hence CR wo-soil values) between some species. It is possible that, in part, the comparatively high 241 Am and Pu concentrations in Lumbricidae spp. were due to residual soil in the gastrointestinal tract. For the other species there is no obvious explanation for the differences (e.g. diet, see Supplementary Information) and the same applies to variation in the 90 Sr: 137 Cs ratio between vertebrate species. There was a tendency for most amphibian species to have comparatively high 90 Sr concentrations compared to mammal species (Table 2) supporting the observations of an earlier study in the CEZ . Published collations of CR wo-soil values suggest a similar transfer for the two wildlife groups (IAEA, 2014).
Comparing the CR wo-soil values to the updated version (see Brown et al., 2016) international wildlife transfer databases (WTD) , in the cases of 137 Cs and 90 Sr all of the CR wosoil values measured at the site are within the ranges for the appropriate wildlife group (see IAEA, 2014). However, there is a tendency for 241 Am CR wo-soil values from the study site to be low in comparison with the WTD for all sampled species (there are no data for tree species in the WTD). In the case of Pu the CR value presented here for A. gigantea is comparatively low. However, data for the other species for which comparisons are possible (Apidae spp., Lumbricidae spp. and A. flavicollis) are within the WTD ranges.
With respect to 241 Am, the data in the WTD tend to be for specific source terms. Most of the data originate around the Sellafield reprocessing plant in north-west England including ecosystems contaminated by sea-spray (e.g. Wood et al., 2009). For grass CR wo-soil values in the WTD, values from ecosystems impacted by sea-spray close to Sellafield are more than an order of magnitude higher than data collected elsewhere. In the case of mammals, data also originate from waste disposal sites in the USA and the Maralinga nuclear bomb test site in Australia. There are mammalian data in the WTD for 241 Am from the sites in the CEZ for M. glareolus with mean CR wo-soil values in the range 2.7 × 10 −3 to 4.5 × 10 −2 compared to the values reported here for A. flavicollis which had a mean of 2.5 × 10 −3 .
Previous studies have suggested that the transfer of 90 Sr to organisms decreases with increasing level of contamination within the CEZ (see discussion in Beresford et al. (2016)). This is likely due to Sr at the most contaminated sites being in particulate form. The study site used here was in one of the more contaminated areas of the CEZ. Comparison with CR wo-soil values in the WTD for less contaminated areas of the CEZ (calculated from data presented by Beresford et al. (2008aBeresford et al. ( , 2016 and Table 6 Mean ± SD fresh matter (FM) concentrations of K, Ca, Sr, Cs, Pb and U in wildlife samples (n = 3, with the exception of Lumbricidae spp. for which n = 6); estimated activity concentrations of 40  Ryabokon et al. (2005)) show values at the study site are generally lower for small mammal species (typically by approximately an order of magnitude). The ICRP (2008) have included different live-stages for their Reference Frog including frogspawn. However, no data were available for this life-stage in ICRP (2009). The data presented here enable comparative concentrations between the adult and frogspawn life-stages to be determined, which may be useful in assessments of dose rates to amphibians throughout their lifespan. For the majority of the 27 elements for which comparisons could be made (see Beresford et al., 2018 for individual data), frogspawn had lower (stable element) concentrations than the adult life-stage (including for Cs and Sr). Exceptions with any radiological significance were Ni, Fe and U, for which, concentrations were similar for the two life-stages or highest for frogspawn.

A comparison of stable-and radio-element concentrations ratios for Sr and Cs
There is an increasing use of stable element data to provide transfer parameter data for both human (e.g. Tagami and Uchida, 2010;Sheppard et al., 2010) and wildlife assessment models (e.g. Takata et al., 2010;Barnett et al., 2014;Thørring et al., 2016;Guillén et al., 2018). There is an assumption that stable element values will represent steady-state conditions (Sheppard et al., 2010). However, Barnett et al. (2014) and Thørring et al. (2016) report differences in 137 Cs and stable 133 Cs CR wo-soil values for wild grass and pine tree species; Barnett et al. also observed this difference for roe deer (Capreolus capreolus). Furthermore, both Beresford et al. (2013) and Wood et al. (2013) observed significant differences between radio-and stable-caesium CR wo-soil values extracted from the WTD, although biases in the data may have been the reason for this (e.g. stable element data being biased to one geographical region and radiocaesium to another). The data presented for the study site here tend to show lower CR wo-soil values for both stable Cs and Sr compared to 137 Cs and 90 Sr, the difference being most noticeable for the Sr transfer to P. sylvestris. It would appear we need to more fully investigate the validity of using stable element data to provide parameters for radiological models and to identify factors which determine when this commonly used assumption is valid or not.

Dose rates
With the exceptions of A. gigantea, Apidae spp. and Lumbricidae spp. all estimated dose rates were either within or above the relevant ICRP Derived Consideration Reference Levels (DCRLs). The DCRLs are an order of magnitude band of dose rate (defined for each RAP) within which there is likely to be some chance of deleterious effects of ionising radiation occurring to individuals of that type of RAP (ICRP, 2008). The sampling site was at the edge of the Red Forest and it is likely that organisms in more contaminated areas will be receiving considerably higher dose rates.
In many studies of the potential effects of radiation on wildlife in the CEZ the only measure of dose reported is ambient dose rate determined using a handheld dose rate meter (e.g. Møller et al., 2012Møller et al., , 2013. There are few other estimates of dose to organisms within the CEZ derived Table 7 Stable Cs and Sr concentration ratios (CR wo-soil ) for individual wildlife samples (n = 6 for Lumbricidae spp., for all other organisms n = 3; with the exception of plants, CR wo-soil values are for the whole-organism).

Species
Arithmetic  from measurements of radionuclides in soils and organisms. Whilst it appears likely that ambient dose rate measurements will give a reasonable approximation of external dose to at least some organisms (Chesser et al., 2000;Beresford et al., 2008c), results present here demonstrate that they will give no indication of total dose rates (i.e. external plus internal exposure). With the exception of Apidae and Lumbricidae spp. ambient dose rate measurements taken from across the sampling site (mean ≈ 12 μSv h −1 ) are 3-13 times lower than the total absorbed dose rate estimate. For vertebrate species, estimated external dose rates were approximately three times higher than the ambient dose rate. Differences in the relative contributions of isotopes to the internal dose rate of different species, as demonstrated in Fig. 4, further highlight why ambient dose rate measurements do not provide any meaningful estimate of actual absorbed dose rates received by different organisms.
The dose to P. sylvestris wood was dominated by exposure to 90 Sr which contributed approximately 97% of the internal and 90% of the total dose rate. The relative radionuclide activity concentrations in P. sylvestris cone samples were considerably different to those in wood, with 137 Cs activity concentrations being about an order of magnitude higher than 90 Sr values (Table 3). In wood 90 Sr activity concentrations were more than an order of magnitude higher than those of 137 Cs. This infers that the dose to cones will be somewhat different to that for wood. Using a geometry for cone (see Supplementary Information) a total internal dose rate of approximately 43 μGy h −1 is estimated which is approximate one-third of the internal dose estimated for wood (Table 8). Caesium-137 comprised 74% of the total internal dose rate for cones.

Exposure to natural background radionuclides
Potassium-40 and 238 U activity concentrations in soils are lower than average values presented for much of Europe though within the ranges reported for most countries (UNSCEAR, 2000;Beresford et al., 2008b). A similar observation has been made for soils from Ivankov district, which is adjacent to the CEZ and has similar soil types.
As would be expected for an element that is homeostatically controlled, for most wildlife species, 40 K values are very similar to mean values for organisms sampled in the United Kingdom (Beresford et al., 2008b); pine tree wood 40 K activity concentrations were just below the  Journal of Environmental Radioactivity xxx (xxxx) xxx-xxx lower end of the range reported for a relatively limited number of samples in the United Kingdom (Beresford et al., 2008b;Barnett et al., 2014). Similarly, 238 U concentrations were comparable to those in UK wildlife.
On the basis of the data reported here, dose rates to wildlife from natural background radionuclides within the CEZ will be similar to those in many areas of Europe (admittedly this does not include any consideration of exposure to 222 Rn and daughter products which will dominate natural exposure to burrowing animals in some areas ).

Pb concentrations in soil
As noted above there has been some suggestion that there may be high concentrations of Pb in the CEZ as a consequence of emergency measures taken in 1986 (Jagoe et al. 1998). However, at the current sampling site this does not appear to be the case with Pb concentrations in soil samples being below the median of 22.6 mg kg −1 for top soil collected from semi-natural ecosystems across the European Union (Salminen et al., 2005).

Conclusions
This paper and the associated data set  provide transfer parameter values for a range of organisms and elements, including those required for the ICRPs environmental protection framework for which there were previously few data available. Hence, our results will contribute to the development of the international radiological environmental protection framework along with studies conducted elsewhere using similar protocols Thørring et al., 2016;Guillén et al., 2018).
Estimated dose rates at the study site were sufficiently high for most organisms that we would anticipate the potential for some form of effect. It is possible that some radiation induced effects may impact on radionuclide transfer. However, transfer parameter values derived within this study are generally in the range of those reported within international databases. Where this is not the case, the differences can be explained.
Our results raise the question of whether or not it is appropriate to use stable element data within the derivation of radionuclide transfer parameter values. Reasons why there are differences in the transfer of stable and radio-elements need to be investigated; sequential extraction to compare potential differences in available fractions between stable and radio-isotopes, and consideration of comparative distributions within the soil profile may be useful within such investigations.
The study has made a useful contribution in the interpretation of radiation effects studies undertaken within the CEZ. The common use of ambient dose rate may underestimate the absorbed dose rate of organisms by over an order of magnitude. This needs to be taken into account when considering reported studies from the CEZ in relation to the suggested benchmark dose rates used in environmental assessments. It has been suggested that there may be additional stressors in the CEZ, such as Pb contamination from emergency measures conducted in 1986, but our results show that this is not the case at our study site. When interpreting results of studies of radiation effects, dose rates need to be put in the context of natural background dose rates. On the basis of the results presented here, natural background dose rates in the CEZ are comparable to those in many other areas of Europe.
We have published all the underlying data associated with this study , which will hopefully help the development of the CEZ as a long-term observatory site.