Vertical distribution of radiocesium in coniferous forest soil after the Fukushima nuclear power plant accident

Journal of Environmental Radioactivity 137 (2014) 37e45

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Vertical distribution of radiocesium in coniferous forest soil after the Fukushima nuclear power plant accident Mengistu T. Teramage a, *, Yuichi Onda a, Jeremy Patin a, Hiroaki Kato a, Takashi Gomi b, Sooyoun Nam b a b

Center for Research in Isotopes and Environmental Dynamic, University of Tsukuba, Tennodai 1-1-1, Tsukuba shi, Ibaraki 305-8572, Japan Department of International Environmental and Agriculture Science, Tokyo University of Agriculture and Technology, Fuchuu, Tokyo 183-8509, Japan

a r t i c l e i n f o

a b s t r a c t

Article history: Received 31 January 2014 Received in revised form 6 June 2014 Accepted 17 June 2014 Available online

This study deals with the description of the vertical distribution of radiocaesium (137Cs and 134Cs) in a representative coniferous forest soil, investigated 10 months after the Fukushima radioactive fallout. During soil sampling, the forest floor components (understory plants, litter (Ol-) and fermented layers (Of)) were collected and treated separately. The results indicate that radiocesium is concentrated in the forest floor, and high radiocesium transfer factor observed in the undergrowth plants (3.3). This made the forest floor an active exchanging interphase for radiocesium. The raw organic layer (Ol þ Of) holds 52% (5.3 kBq m2) of the Fukushima-derived and 25% (0.7 kBq m2) of the pre-Fukushima 137Cs at the time of the soil sampling. Including the pre-Fukushima 137Cs, 99% of the total soil inventory was in the upper 10 cm, in which the organic matter (OM) content was greater than 10%, suggesting the subsequent distribution most likely depends on the OM turnover. However, the small fraction of the Fukushimaderived 137Cs at a depth of 16 cm is most likely due to the infiltration of radiocesium-circumscribed rainwater during the fallout before that selective adsorption prevails and reduces the migration of soluble 137Cs. The values of the depth distribution parameters revealed that the distribution of the Fukushima-derived 137Cs was somewhat rapid. © 2014 Elsevier Ltd. All rights reserved.

Keywords: Deposition Forest floor Migration Radiocesium 137 Cs 134 Cs

1. Introduction Radiocesium (134Cs, t1/2 ¼ 2.1 y, and 137Cs, t1/2 ¼ 30.2 y) derived from the Fukushima Dai-ichi Nuclear Power Plant (hereinafter FDNPP) accident has contaminated a wide range of environments, including forest areas in Fukushima and neighboring Prefectures. Note that greater than 70% of Japan's archipelago is covered by forests, of which most are evergreen coniferous forests (Onda et al., 2010). However, processes ruling the local dynamics of radiocesium transfer from the forest canopy to the forest floor and its further distribution in soil are still few documented. Numerous studies dealing with the early redistribution of radiocesium deposited onto the forest canopy, including canopy interception and subsequent transfer from the canopy to forest floor were carried out in the aftermath of Chernobyl accident (Bunzl et al., 1989; Tikhomirov et al., 1993) and most of them focused on the dry climatic conditions of Chernobyl-affected

* Corresponding author. Tel.: þ81 90 8569 3775; fax: þ81 29 853 4226. E-mail address: [email protected] (M.T. Teramage). http://dx.doi.org/10.1016/j.jenvrad.2014.06.017 0265-931X/© 2014 Elsevier Ltd. All rights reserved.

regions. However, such information is lacking in the case of the Fukushima reactor accident and humid climatic regions, which might cause a different behavior of radiocesium migration and distribution (Teramage et al., 2014). The deposited fraction of radionuclides onto the forest floor through hydrological pathways (i.e throughfall) and fallen canopy components (i.e. litterfall) will undergo horizontal and vertical migration (Teramage et al., 2013). On flat, undisturbed sites, the highest radiocesium concentration is in the uppermost soil layer and decreases exponentially with the depth (e.g He and Walling, 1997). However, the litter layer on the forest floor can play a unique role in the distribution of radiocesium that is typically lacking in most other land use types. The litter layer accumulates both throughfall and litterfall-derived radiocesium deposits, and the variation of radiocesium distribution on the forest floor can reflect the resultant effects of these two mechanisms of accumulation. As most of the heads of water resources are located in forested watersheds and are intimately linked to downstream ecosystems (Gomi et al., 2002; Sun et al., 2013), the remobilization of radiocesium accumulated in surface organic layers of forest soil may result in contamination of the soil and rivers for a long time.

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

2. Materials and methods 2.1. Study site The study was conducted in a 30-y old stand of Japanese cypress (Chamaecyparis obtusa Endl.) located on Karasawayama (139 44' E; 36 23' N) in the Tochigi Prefecture of central Japan (Fig. 1). The area is located 180 km southwest of the FDNPP. The size of the catchment is 0.8 ha. The climate of the area is humid temperate, with mean annual rainfall of 1259 mm and mean annual temperature of 14.1  C (as

1600

500

1400

400

1200 1000

300

800 200

600 Soil sampling

100

400

200 0

Cumulative precipitation (mm)

Therefore, understanding the early distribution and subsequent migration of radiocesium in the soil profile is essential. Such knowledge is helpful in decisions making about possible countermeasures, to set up environmental baselines, and to establish parameters to predict radiocesium transfer in forested ecosystems. Despite the importance of understanding the movement of radiocesium in forest soils, most of the recent studies on soils affected by radioactive deposits focused on agricultural soil and undisturbed sites in non-forested environments such as pasture land (e.g., He and Walling, 1997; Kato et al., 2012); therefore, little is known about forest soil. Among the few studies on Chernobylaffected forest soils, Karadeniz and Yaprak (2008) reported a relaxation length (i.e., the inverse of the rate of change in the concentration that represents the depth of migration due to change in concentration) of 4e15 cm, and Poreba et al. (2003) indicated a relaxation length of 2.1 cm several years after the accident. However, these findings cannot be simply applied to the Japanese environment due to differences in the climatic conditions and various physiographic features. Recently, Koarashi et al. (2012) reported the vertical distribution of Fukushima-derived 137Cs in different land uses and identified major controlling factors for its migration using the core soil sampling method with coarse soil section of 1e5 cm. However, fine depth sections are expected to better describe the distribution of radiocesium (Kato et al., 2012; Loughran et al., 2002). In addition, the composition and contamination density of radiocesium are expected to vary with the distance from the source. This is mainly because of the difference in particle composition of the radiocesium debris, i.e. heavier composition likely deposits near to source area. Therefore, this study investigated the radiocesium depth distribution in the forest soil profile based on fine depth resolution. The radiocesium migration and soil-to-plant transfer rates were also estimated.

Monthly precipitation (mm)

38

0 Date (Month/Year)

Fig. 2. Monthly (bars) and cumulative (bold line) precipitation for one year at the study site. ¼ indicates arrival of the radionuclide plume at the study site in midMarch (2011). The wet deposition was expected due to the precipitation from March 15 to the first two weeks in April 2011.

obtained from the Karasawa mountain metrological station from 2010 to 2011). The soil type was classified as an orthic cambisol according to the World Reference Base for Soil Resource (IUSS Working group WRB, 1998). The estimated stand density is approximately 2500 trees per hectare. The dominating understory vegetation is composed of sparsely grown understory plants (marlberry (Ardisia japonica (Thunb.) Blume) in addition to various herbs. According to the MEXT (2011) report, the plume extended to the study site was at the origin of 137Cs deposits of approximately 10 kBq m2. The low temperature at the time of accident maintained the plume movement near to the ground surface driven by the local wind. In that context, it is assumed that the deposition primarily occurred through wet deposition. Fig. 2 shows the monthly and cumulative total rainfall around the time of the accident and the sampling date. 2.2. Soil sampling The soil samples were collected from an undisturbed flat area to minimize the effect of the subsequent lateral movement of radionuclides after fallout. The sampling point was purposely selected at the midpoints between tree lines to make the sampling more representatives. We used a rectangular metal-framed scraper plate

Fig. 1. Map of the study area and location of the sampling point. Solid circles in the map represent the capital and Prefectures cites.

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

(internal dimensions of 15 cm  30 cm) with an adjustable depth increment of 5 and 10 mm intervals. This sampling method allows for the collection of numerous and voluminous soil samples that encompass and represent the relatively wider microtopographic variability (Kato et al., 2012; Loughran et al., 2002). The sampling was performed on 16 January 2012 (about 10 month after the accident). The understory vegetation, Ol- (~3 cm thick) and Of- (~4.5 cm thick) organic horizons were carefully separated by hand and scissors (when necessary) to represent the forest floor. The understory vegetation with in the sampling plot which was dominated by sparsely grown marlberry and annual herbs were carefully mowed. Subsequently, their radiocesium content in dry basis was determined after crushing into powder. The Ol-horizon, the original shape of the components was easily recognizable, was composed of periodically falling raw litter. The Of- horizon, which was located under the Ol-horizon, was composed of an early fermented and fragmented litter component in which the original shapes of the litter were difficult to identify. The radiocesium uptake by the understory plants is usually estimated based on the soil-to-plant transfer factor on a dry weight basis (TFw) or aggregated transfer factor (Tagg) coefficients, which are defined as the ratio of average radionuclide concentration in plants (Bq kg1) to that in the soil (Bq kg1) for the TFw or the total soil inventory (Bq m2) for the Tagg (Calmon et al., 2009). The Tagg is often used for the medium-to-long term measurements after deposition, when radiocesium have been redistributed between the different surface layers of forest soils (IAEA, 2010), as radiocesium requires a long time to reach the tree root zone. In our study site, we assumed that the roots of the understory vegetation mainly explore the Of- horizon and the upper 2 cm of soil and it is reasonable to use the TFw instead of the Tagg. The soil below the Of- horizon was scraped layer-by-layer in three major depth resolutions of 5 mm (for upper 5 cm), 10 mm (for 5e10 cm) and 20 mm (for 10e30 cm). We employed a depth increment of 0.5 cm for the upper 5.0 cm to reduce the possible effects of soil thickness on defining the shape of the profile, unlike that used by Koarashi et al. (2012). In the sampling depth, neither cracks nor large roots were encountered. 2.3. Laboratory analysis 2.3.1. Measurement of radiocesium activity All of the samples were dried at 110  C for 24 h to determine the dry weight. The samples from the understory vegetation and from the Ol- and Of- horizons were then milled and mixed to ensure homogenous sample material for each respective sampling unit. The soil samples were disaggregated by gentle grinding and were then passed through a 2-mm sieve. Then, the milled samples from the understory vegetation and from the Ol- and Of- horizons and the <2 mm soil fraction from each soil layer were placed in plastic containers (U-8, As ONE, Tokyo) and sealed for analysis. The analysis was conducted in the laboratory of the University of Tsukuba, which was authorized for independent calibration checks during the worldwide open proficiency test in 2006 (IAEA/AL/171). The activities of 137Cs and 134Cs were determined using gamma ray spectrometry from a high purity n-type germanium coaxial gamma ray detector (EGC 25e195-R, Canberra-Eurysis, Meriden, USA) connected to an amplifier (PSC822, Canberra, Meriden, USA) and a multichannel analyzer (DSA1000, Canberra, Meriden, USA) using the counts at the 662 keV and 605 keV peaks, respectively. The absolute counting efficiency of the detector was calibrated using various weights of IAEA-2006-03 standard soil samples with background correction. In this study, the measured radiocesium concentrations (Bq kg1) were converted to inventories (Bq m2) using the volume and apparent bulk density on the basis of dry

39

weight of each sampling layer. All of the measured activities were decay-corrected to May 20, 2011. 2.3.2. Soil physicochemical property analysis The physicochemical properties of the soil were determined for the <2 mm soil samples. The particle size distribution (sand: >50 mm, silt: 2e50 mm, and clay: <2 mm) in each soil layer was analyzed using a laser diffraction particle size analyzer (SALD-3100, Shimadzu Co., Ltd., Kyoto, Japan). The bulk density was determined from the dry weight and volume of the soil in each layer. The organic matter (OM) content of the samples was determined by the weight lost after the incineration of a known dry weight sample in a muffle furnace at 450  C for 4 h. The pH was determined using a pH meter by mixing 5 g dry soil with 50 ml distilled water to a final 1:5 soil:water ratio. 2.4. Parameters for the radiocesium depth profile To investigate the vertical distribution of radiocesium, we applied a negative exponential profile function, which includes a number of simplified assumptions (Karadeniz and Yaprak, 2008). This method was used to estimate the radiocesium movement in early stages after the fallout. Importantly, in forest soils, the organic-rich upper layers could affect the theoretical exponential function. Therefore, we discussed the radiocesium concentration in the forest floor (hereafter refers to the composition of the understory vegetation and of the Ol- and Of- horizons), while the depth penetration of radiocesium in the soil below the Of- layer was assumed to follow the general formula:

CðzÞ ¼ Cð0ÞeaZ

p

(1)

where z is the depth from the soil surface (cm), C (z) is the concentration of radiocesium at depth z (Bq kg1); C (0) is the radiocesium concentration (Bq kg1) of initial deposit on the upper surface soil; a (cm1) is the reciprocal of the relaxation length; and p (unit-less) is an experimentally determinable parameter depending on the upper soil surface condition and form of transport. The parameter a depends on the characteristics of the radionuclides, the soil type and the physicochemical characteristics, land use type, time elapsed after deposits and climatic conditions (Kato et al., 2012). The reciprocal value of a represents the relaxation length of the radiocesium in the vertical profile (Karadeniz and Yaprak, 2008). The depth can be expressed either in linear (cm or m) or mass depth (kg m2). The linear relaxation length, 1/a, represents the shape of the tail of the depth distribution, whereas the relaxation mass depth describes its penetration strength into soil mass. When p ¼ 1, the following function (Porto et al., 2001) can be used as an alternative to directly estimate h0 by fitting the model to the empirical data of the reference site as: 0

Cðz0 Þ ¼ Cð0Þez h0

(2)

where z0 is the mass depth from soil surface (kg m2); C(z0 ) is the concentration of radiocesium at depth z0 (Bq kg1); C(0) is the radiocesium concentration (Bq kg1) of the surface soil; and h0 can be easily determined using the linearized least square regression method. 2.5. Characterizing the pre-Fukushima

137

Cs

The observed total 137Cs concentration in the soil includes remnants of the pre-Fukushima episodes, whereas due to its short half-life, the observed 134Cs in the environment is exclusively originates from Fukushima. To determine the pre-Fukushima 137Cs concentration, the 134Cs/137Cs ratio was used (Livens et al., 1992).

40

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

The ratio of Fukushima-derived radiocesium deposits was determined from litter collected immediately after the fallout and was found to be 1 (Teramage et al., 2014), and the pre-Fukushima 137Cs residue was determined accordingly. Note that in this study we describe 137Cs as pre-Fukushima137Cs, Fukushima-derived 137Cs (~134Cs) or total 137Cs (representing pre- and Fukushima-derived 137 Cs altogether); otherwise radiocesium is used to represent the general and common features of both radioisotopes.

where t is the sampling year; t0 is the maximum fallout year (2011 for Fukushima, 1986 for Chernobyl and 1963 for bomb fallout); Nz is the mass depth (kg m2) or linear depth (cm) of the pre- and Fukushima-derived 137Cs concentration reduced to 1/e of the maximum concentration; and Wz is the mass depth (kg m2) or linear depth (cm) where the maximum pre- and Fukushimaderived 137Cs concentration is located at the time of measurement. 3. Results and discussion

2.6. Diffusion and migration rates of radiocesium in the mineral soil

3.1. Radiocesium in forest soils

Although the depth distribution of fresh fallout is often described using an exponential function as indicated in Eq. (1), the function fails to describe the downward transport that occurs from the first instant of deposition and is unable to illustrate long-term distribution (Almgren and Isaksson, 2006). The redistribution of radiocesium within the soil profile is the result of a complex set of processes that should be considered by the time-dependent model (Walling et al., 2002). Such transport mechanisms have been characterized by a diffusion coefficient (D, kg2 m4 y1) and a migration or convection rate (V, kg m2 y1) that lump together all of the redistribution processes in the soil column (Walling et al., 2002). These transport coefficients can also be represented in the dimension of D in cm2 y1 and V in cm y1. In this study, we estimated these two parameters based on the formula proposed by Walling et al. (2002):

3.1.1. Soil physicochemical properties The soil type along the profile was found to be silt loam with a mean particle size distribution of 29% (Standard deviation (SD): 9.5) sand, 61% (SD: 8.6) silt and 10% (SD: 1.9) clay. The silt fraction generally dominated the soil texture composition in each examined layer, while the clay, sand contents and bulk density showed no definite pattern (Table 1). The organic matter contents of the Ol- and Of- layers were 87.2 and 74.4%, respectively. Below these layers, the OM content decreased sharply and continuously with increasing depth. However, the OM content was greater than 10% in all of the soil layers in the upper 10 cm, with the highest proportion (27%) in the upper 0e0.5 cm. The pH of the soil was acidic that ranged from 5.10 to 5.92 and showed a slight general increase below the 5 cm soil depth (Table 1).

Dz

ðNz  Wz Þ2 2ðt  t0 Þ

(3)

Wz Vz t  t0

(4)

3.1.2. Radiocesium on the forest floor 3.1.2.1. Understory vegetation. The biomass of the understory vegetation was 0.4 kg m2 which contains 1.53 ± 0.07 kBq kg1 of Fukushima-derived 137Cs. The TFw ratio (Bq kg1 in the plant/ Bq kg1 in the upper soil) was estimated to be 3.3. These values imply that the concentration of Fukushima-derived 137Cs in a given dry-weight understory vegetation exceeds the average

Table 1 Physiochemical properties and radiocesium concentration in soil profile. Depth (cm)

Mass depth (kg m2)

Bulk density (g cm3)

Sand (%)

Silt (%)

Clay (%)

OM (%)

pH

137

Cs (Bq kg1)

134

Cs (Bq kg1)

Inventory proportion 137

Ol Of 0.0e0.5 0.5e1.0 1.0e1.5 1.5e2.0 2.0e2.5 2.5e3.0 3.0e3.5 3.5e4.0 4.0e4.5 4.5e5.0 5.0e6.0 6.0e7.0 7.0e8.0 8.0e9.0 9.0e10 10e12 12e14 14e16 16e18 18e20 20e22 22e24 24e26 26e28 28e30

0.7 6.5 3.8 1.8 2.3 2.7 2.1 4.0 2.6 3.7 3.1 5.9 8.4 6.7 3.8 9.4 8.7 11.6 12.5 13.5 14.2 10.2 12.7 11.7 7.6 11.4 13.1

0.02 0.1 0.8 0.4 0.5 0.5 0.4 0.8 0.5 0.7 0.6 1.2 0.8 0.7 0.4 0.9 0.9 0.6 0.6 0.7 0.7 0.5 0.6 0.6 0.4 0.6 0.7

e e 43 36 31 17 32 41 39 36 36 34 36 36 27 21 17 27 31 34 21 20 17 21 16 18 50

e e 51 56 62 73 58 48 53 57 53 55 56 53 63 68 72 61 57 60 69 68 72 69 72 71 42

e e 6 8 7 10 10 11 8 7 11 11 9 11 10 11 11 11 11 6 10 11 11 10 12 11 8

87.2 74.4 26.7 20 18 17 16 15 14 13 12 13 12 12 12 10 10 9 7 5 5 5 5 5 4 5 4

Error (±) value shows the statistical counting error at the time of measurement; nd: not detected.

4.93 4.96 5.12 5.49 5.19 5.27 5.14 5.39 5.10 5.17 5.17 5.37 5.52 5.36 5.64 5.39 5.71 5.92 5.39 5.83 5.67 5.74 5.73 5.77 5.70 5.71 5.70

669 ± 33 840 ± 43 422 ± 15 362 ± 18 349 ± 18 215 ± 9 168 ± 12 118 ± 7 93 ± 6 92 ± 4 61 ± 2 55 ± 4 45 ± 3 45 ± 3 32 ± 3 29 ± 3 9±1 6±1 Nd 5±1 Nd Nd Nd Nd Nd Nd Nd

666 ± 37 739 ± 43 389 ± 15 285 ± 17 323 ± 19 158 ± 9 139 ± 11 88 ± 7 58 ± 6 56 ± 4 30 ± 3 29 ± 3 15 ± 3 18 ± 3 11 ± 3 9±3 7±1 Nd Nd 5±1 Nd Nd Nd Nd Nd Nd Nd

4 43 13 5 6 5 3 4 2 3 1 3 3 2 1 2 1 1 e 1 e e e e e e e

Cs %

134

Cs %

5 47 15 5 7 4 3 3 1 2 1 2 1 1 0.4 1 1 e e 1 e e e e e e e

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

concentration found in the soil under consideration by three-fold, indicating uptake by the plants. Our result of TFw ratio obtained from bulked understory plants was slightly higher than the values reported by Zibold et al. (2009) for ferns (1.2e3.2) and considerably different from that of blackberries (0.3e0.6). Note that because the forest floor was a secondary receiver of radiocesium next to the phytomass at the time of fallout, the observed Fukushima-derived 137 Cs in the understory vegetation could be from two possible originate, i.e., via root uptake from the forest floor and via the direct deposition on plant organs. Given the possibility of renewal of very short-lived understory plants, the effect of direct interception by understory could be still predominant which may attributed to high measured TFw. Another reason could be the acidic nature and low clay contents in Ol-,Of- and the upper 2 cm soil layers (Table 1) that conceivably favor radiocesium bioavailability to plant uptake (Delvaux et al., 2000; Konopleva et al., 2009; Thorring et al., 2012). 3.1.2.2. Ol- and Of- horizons. The densities of the Ol- and the Ofhorizons were 0.7 kg m2 and 6.5 kg m2, respectively and the radiocesium concentrations were greatest in both layers (Table 1). The retained fraction, defined as the ratio of inventory at each depth section to that of the total inventory of the soil, showed that the Of- horizon retained 47% of the Fukushima-derived 137Cs inventory (Table 1). The raw organic layer (Ol þ Of) holds 52% (5.3 kBq m2) of the Fukushima-derived and 25% (0.7 kBq m2) of pre-Fukushima 137Cs at the time of the soil sampling (16 January 2012), and the remaining is distributed in the soil below the Oflayer. Recent study demonstrated that litterfall tends to continue depositing radiocesium on the forest floor (Teramage et al., 2014). Therefore, the radioactivity on the forest floor is expected to increase in subsequent periods, as considerable proportion of initial atmospheric radiocesium deposits is possibly held in the canopy. In spite of this likelihood, only 5% of the Fukushima-derived 137Cs content was found in the Ol- horizon at the time of observation. Table 2 The retention

41

This result is most likely due to the transfer of the Ol- horizon to the Of- horizon through mechanical and biological breakdown (Rafferty et al., 2000). Comparing our results with those of Chernobyl highlights the differences in the migration of radiocesium in a forested environment. For this reason, we compiled pre-Fukushima 137Cs retention in a layer that combined Ol- and Of- horizons reported in previous studies from diverse locations, forest types and different time periods (Table 2) and were plotted them against a time elapsed between the accidents and sampling periods (Fig. 3). It shows a general decrease over time and the retained 137Cs proportion varied with the forest and soil types. Relatively, the migration of the Fukushima-derived 137Cs in our study seems rapid (Fig. 3), with approximately half of the total inventories were already transported below the Of- horizon in less than a year. Despite the site difference, the tendency of the reported retained pre-Fukushima 137 Cs fraction in litter layer seems to follow that of the deposition density. This can be clearly observed in Fig. 3 (shaded) on the retained fraction values obtained in a similar time distance after the Chernobyl accident. The difference might be attributed partly to the variation in the depositional density, and it likely masked the downward migration rate in highly contaminated areas possibly due to the input load of radiocesium into the litter layer largely exceeds to that leaving the layer. This implies that fallout density should be taken in to account when comparing the migration rate in forested environment. 3.1.3. Distribution of the radiocesium inventory in the soil The total 137Cs inventory of the soil profile was 13 kBq m2, to which the Fukushima accident contributed approximately 77% (10 kBq m2). The pre-Fukushima 137Cs in the studied forest was 2.6 kBq m2, which is in close agreement with the reported values, which range from 2 to 5 Bq m2 (Sakaguchi et al., 2010). The inventory peaks of 134Cs(representing Fukushima-derived 137 Cs) and total 137Cs were located in the Of- horizon and then

137

Cs in the litter layer in different locations affected by Chernobyl nuclear power plant accident (% of the total soil inventory).

Reference

Location

Tree /forest type

Fallout density (kBq md2)

Sampling year

Retained 137Cs in litter horizon Soil type (%)

Witkamp and Frank (1964) Schimmack and Bunzl (1992)

USA

Tulip poplar

e

1963

64

Germany

Norway Spruce forest

30

April, 1989

85

Tobler et al.(1988)

Switzerland

Scots pine forest Norway spruce forest

9 8

Tikhomirov et al. (1993)

Ukraine & Belarus

Birch-oak-pine mixed forest

210e250

April, 1989 80 October, 56 1986 August, 1987 95

Slighly acid colluvial silt loam Parabrown Sandy podsol Regosol Soddy podzolic sandy

Strebl et al.(1996) Strandberg (1994)

Austrian Denmark

Norway spruce forest Scots pine forest

e 0.9

Melin et al.(1994) Raitio and Rantavaara (1994)

Sweden Finland

Pine forest Norway spruce forest

180 e

August, 1988 August,1989 August, 1990 August, 1991 1993 October, 1991 October,1990 1991

Bunzl et al.(1998) Fesenko et al.(2001)

Germany Russia

Scots pine forest Norway spruce forest Pine Pine and Birch mixed forest Pine and Birch mixed forest Japanese cypress forest

e 20 800e2300 890e2850

1991 1990 1996 1996

45.5 35 8.8 12.5

1600

1996

11.5

e Podzolic parabrown Soddy podzolic loamy sand Humic podozol gley loamy sand Soddy podzolic loamy sand

10

January, 2012

52

Orthic brown silt loam

This study

Japan

In this comparison litter layer refers a combination of raw and partially fragmented litter layers.

89 85.3 80.8 76.6 46 20

Soddy podzolic sandy Soddy podzolic sandy Soddy podzolic sandy Soddy podzolic sandy Dystric cambisol Podsolic

40 21

Podsol e

42

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

that the persistence of radiocesium in forest floor and pointed out this may be important, especially in thick humus layers which then acts as a reservoir for plant uptake (Thiry et al., 2000; Goor et al., 2007) and retards the vertical migration in the soil profile of coniferous forest ecosystems (Fawaris and Johanson, 1994; Thiry and Myttenaere, 1993). From a physicochemical point of view, despite its high sorbing capacity, organic matter fraction of soil is not a source of irreversible radiocesium adsorption (Valcke and Cremers, 1994). Radiocesium selective adsorption is rather under the dependence of highly fixing sites born by weathered micaceous clay particles (Maes et al., 1998). Radiocesium-fixing clay minerals are more or less diluted in forest floor according to local bioturbation and were shown to highly influence its mobility and bioavailability in the upper organic layers of forest soils (Kruyts and Delvaux, 2002). In thick and acid humus in particular, characterized by a high percentage of organic matter (>80%) resulting from a low biological activity, the mobility of radiocesium is in general mainly governed by the large pool of reversible adsorption sites. In deeper layers of the forest floor, at the transition with mineral layers, a lower dilution of radiocesium-fixing clay minerals may however promote the radiocesium retention. Therefore, the observed high proportion of radiocesium in the Of- horizon indicates that radiocesium barely leaves the organic horizons. This tendency most likely determines the subsequent downward movement and bioavailability of radiocesium to the plant roots that explore this soil section.

Tobler et al., 1988 Tikhomirov et al., 1993

10 – 2

Fesenko et al., 2001

This study Whitkamp and Frank,1964 Melin et al., 1994

80 kBq k ) B m- ) (200 kBq

(0.99 kBq m- )

Fig. 3. Retained 137Cs in the litter layer in different forest ecosystems over time. The shaded region represents the samples which were collected the same period after the Chernobyl accident but demonstrated different % of retained radiocesium, and the values in parentheses are their corresponding contamination densities.

both inventories sharply decreased in the lower depths, although some of Fukushima-derived 137Cs appeared at a depth of 16 cm (Table 1). The radiocesium inventory in the 0e0.5 cm soil layer (i.e., just below the peak) was approximately three times lower than that of the Of- horizon, indicating radiocesium seems hardly leave Of-horizon. In agreement, Rafferty et al. (2000) evaluated the preFukushima 137Cs migration in Pinus contorta forest under the influence of a full year and demonstrated that only 1% of it migrates into the mineral soil. Brouwer et al. (1994) also demonstrated in sequential extraction experiment that on pure mineral substrate approximately 40% of the pre-Fukushima 137Cs was removed, while the same procedure resulted only 8% on organic horizons. Furthermore, Sombre et al. (1994) also reported comparable trends in Spruce and Oak forests. Similarly, several authors acknowledged 137Cs Concentration

0

200

400

134Cs Concentration (Bq kg -1)

(Bq kg-1) 600

800

0

0

0

1

1

2

2

3

3

4

4

5

5

6

Depth(cm)

3.1.4. Distribution of the pre-Fukushima 137Cs in the soil profile The pre-Fukushima 137Cs activities were obtained by deducting the Fukushima-derived 137Cs estimated from measurements of 134 Cs (average ratio of 134Cs/137Cs equal to 1). Fig. 4a illustrates the depth profile distribution of the total and pre-Fukushima 137Cs. The pre-Fukushima 137Cs tended to dominate in the deeper soil layers. Its depth distribution exhibited a general decreasing pattern with an irregular profile shape and two disincentive peaks. The peaks

7 8 9

200

400

600

800

6

C(0) = 613 α = 0.6 h0 = 11.1 r2 = 0.96

7 8 9

10

10

11

11

12

12

13

13

14

14

15

15

C(0) = 628 α = 0.73 h0 = 8.73 r2 = 0.97

Fig. 4. Depth distribution of the Fukushima-derived radiocesium activity concentration below the Of- horizon of the forest soil; (a) 137Cs and (b) 134Cs concentrations profile. The black solid line shows the measured radiocesium, the gray solid line shows pre-Fukushima radiocesium and the broken line indicates the results fitted by Eq. (1).

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

appeared at a depth of approximately 0.75 cm and 1.75 cm with a 137 Cs concentration of 87 Bq kg1 and 57 Bq kg1, respectively. Such heterogeneous profile could be the result of many complex processes including the redistribution of radionuclides by percolating water and bioturbation processes (Fujiyoshi and Sawmura, 2004). Although it is difficult to clearly distinguish the Chernobyl 137Cs from the atomic bomb 137Cs, the two distinct peaks generally represent the migration distance of pre-Fukushima 137Cs fallout events. Considering the displacement of the highest preFukushima-derived 137Cs peak (at depth of 0.75 cm) in the soil and the time elapsed since its peak deposition on the earth surface (1963/1964), the downward velocity of per-Fukushima 137Cs in the studied soil can be estimated at approximately 0.15 mm y1. This migration rate is far lower than that of previous studies. For example, Schimmack et al. (1989) reported the migration of bombderived 137Cs ranging from 0.7 to 10 mm y1 while Dorr and Munnich (1991) reported as 1.8 mm y1. Despite the likely differences in the study sites, the pre-Fukushima 137Cs peak in our study site has been expected to be relatively deeper but did not show a significant change. This can be explained by efficient progressive fixation of radiocesium by clay minerals even if present in organicrich soils (Konopleva et al., 2009) over time that retards the migration and increases its residence half time (Bunzl et al., 1995). More interestingly, almost all of the pre-Fukushima 137Cs was contained in a soil layer in which the OM content constituted more than 10% (Table 1 and Fig. 4a). Indeed, several studies have demonstrated that forest soils have a strong tendency to retain radiocesium and that the majority of activities are distributed in organic-rich surface layers. For example, Schimmack and Bunzl (1992) reported that only 15e20% of the radiocesium activity in spruce and pine forest is found in the upper 10 cm mineral soil, while most of the activity remains in the organic layer. This implies that the long term downward migration of radiocesium may largely depends on the processes acting in OM-rich soil layers, such as microbial activity and OM turnover. In forest soils, that involves a slow transport rate that likely lasts from a year to several decades, depending on a set of factors governing the decomposition process. However, this transport might have some associated risks if radiocesium is available in mobile form in the uppermost soil of forested hill that the radiocesium-rich surface materials including organo-mineral particles can be easily transported to the surrounding environment by runoff, which could be the potential contaminant for downstream agricultural soils and water resources. 3.1.5. The distribution of Fukushima-derived 137Cs in the soil profile The activities of 134Cs and total 137Cs in the uppermost soil layer below the Of- horizon were 389 ± 15 Bq kgd1and 422 ± 15 Bq kg1, respectively, and the concentrations decreased with increasing depth (Table 1 and Fig. 4a). This distribution pattern typically reflects the nature of radiocesium adsorption along the depth in which the surface organic material filters and retains the majority of the radiocesium. The remaining radiocesium gradually attaches to soil particles along the profile as the infiltration advances that eventually create such a profile shape. In fact, small quantities of Fukushima-derived 137Cs were observed in the 14e16 cm layer (Table 1). This is unexpected considering the common behavior of radiocesium migration along the soil profile and the time elapsed since the accident, particularly in the context of the FDNPP accident, where the fallout occurred during an ecological phase of rest and in limited biogeochemical processes. Schimmack et al. (1989) showed the unchanged depth distribution of Chernobyl's radionuclide fallout between two consecutive soil sampling periods with different precipitation rate, demonstrating the fraction of radionuclides observed in deeper

43

layers is essentially the result of rain shower during the fallout period. Therefore, the most likely reason for the observed fraction at the specified soil depth in our study is the infiltration of radiocesium-circumscribed rainwater at the time of fallout. In fact, this stage can contribute to a rapid initial downward migration (Rafferty et al., 2000) that likely depends on the intensity and duration of precipitation, the soil structure, radiocesium composition (soluble vs insoluble forms) and the pre-existing soil moisture. Nevertheless, this stage cannot be used to describe the long-term radionuclide migration because the initial migration lasts for only a short period of time after fallout. 3.1.6. Characteristics of the vertical distribution of radiocesium in the soil profile The vertical distribution behavior of radiocesium was examined based on the estimated coefficient of the parameters in Eq. (1) and Eq. (2) for soil below the Of- horizon. The value of p in Eq. (1) is often used as 1 in most studies. However, Isaksson and Erlandsson (1995) have indicated that the value of p can also be 0.75 for a lichen carpet, 2.00 in the case of purely diffusional transport, or even lower for active transport processes. As we dealt with the forest floor separately in the preceding sections, p was used as 1, and the empirical data were fitted using Eq. (1). The parameter that represents the inverse of the relaxation length (a) of 134Cs that presents Fukushima-derived 137Cs was 0.7 cmd1 (a 1.4 cm relaxation length) (Fig. 4a and b). Considering the total 137Cs activity in the soil profile, the value of a was closer to 0.6 cmd1 (a 1.7 cm relaxation length). Obviously, this difference can be attributed to the presence of pre-Fukushima 137Cs. The relaxation mass depths in the study soil profile (h0, kg m2) were determined using Eq. (2) and were 8.7 and 11.1 kg m2 for 134Cs and the total 137Cs, respectively. The observed small difference in both estimated parameters between the total and Fukushima-derived 137 Cs might indirectly indicate that the migration of Fukushimaderived 137Cs is somewhat rapid. A direct comparison of our results of radiocesium migration with those of previous studies seems difficult because the previously reported values are either from agricultural fields or are from studies that were conducted several years after the Chernobyl accident. However, the general migration trend can be evaluated. For example, Poreba et al. (2003) observed a relaxation length of 2.1 cm (a ¼ 0.481 cmd1). More recently, Kato et al. (2012) found relatively higher values of h0 (9.1 kg m2) and a (1.2 cmd1) for total 137Cs from an untilled home garden. Koarashi et al. (2012) also investigated the distribution of Fukushima-derived 137Cs in different land uses and reported a relaxation length (1.43e2.9 cm) and relaxation mass depth (7.4e10.9 kg m2) for forest land uses. Given the difference in sampling techniques mentioned earlier, our results are almost consistent with those of Koarashi et al. (2012). However, the observed slight difference in the profile distribution parameters can be attributed to the differences in the sampling date and techniques, land uses and cover type, soil type and precipitation. 3.2. Diffusion and migration rates of radiocesium in the soil profile Unlike the diffusion coefficient (D), which considers the expected depth displacement of the maximum concentration by a factor of 1/e, the migration rate (V) for young fallout is not noticeable. The values of D and V in this study were approximated based on Eqs. (3) and (4), and the results are illustrated in Table 3. The peak of the pre-Fukushima-derived 137Cs occurred deeper (Fig. 5a) than did that of the Fukushima-derived 137Cs, which was on the soil surface (Fig. 5b). As indicated, the calculated values of D and V were 1.5 cm2 y1 and ~0 cm y1 for Fukushima-derived 137Cs, and 0.24 cm2 y1 and 0.03 cm y1 for pre-Fukushima 137Cs,

44

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

Table 3 Values of diffusion and migration coefficients of137Cs in soil. Source of

137

Cs Wz

Nz

V

D

Reference

Mass depth (kg m2) Depth (cm) Mass depth (kg m2) Depth (cm) (kg m2 y1) (cm y1) (kg2 m4 y1) (cm2 y1) Pre-Fukushima Pre-Fukushima Pre-Fukushima Pre-Fukushima Fukushima

e e e 5.7 3.8

e e e 0.75 0

e e e 26.1 12.8

3.6 e e 4.25 2

respectively (Table 3). Specifically, the value of D reflects the lump sum effects of at least three major processes: molecular diffusion, hydrodynamic dispersion and physical mixing and its value generally reduces over time (Schimmack and Marquez, 2006). In strong agreement, the value of D for Fukushima-derived 137Cs was greater than that of the pre-Fukushima 137Cs, suggesting that a more rapid diffusion-like downward transportation tends to dominate in our study site. Rosen et al. (1999) have also reported similar trends for the V-value of 0.5e1.0 cm y1 for the first year and 0.2e0.6 cm y1 thereafter. Moreover, regardless of the influential set factors that emerge from climatic and site differences, our results are also consistent with the range of values reported by Almgren and Isaksson (2006) (Table 3) and most of the references therein. These results imply that the general trends of radiocesium dispersion are similar in that it is rapid during the first years following the emission and slows down thereafter. Combining all of the parameters that were used to describe the depth distribution behavior, it is possible to provide an insight into the vertical migration of radiocesium in the forest environment. However, the radiocesium trapped in the canopy during the fallout will eventually reach the forest floor via, for example, litter fall (Teramage et al., 2014). Therefore, the concentration of radiocesium in the forest floor and soil are expected to increase and affect its subsequent distribution. Schimmack and Marquez (2006) have concluded from a time-series of observations that the convection

0.2e1 e e 0.2 0

e 0e0.35 0e0.52 0.03 0

20e50 e e 8 4.5

e 0.06e2.63 0e2.7 0.24 1.5

Walling et al. (2002) Almgren and Isaksson (2006) Schimmack and Marquez (2006) This study This study

and dispersion coefficients determined from the profiles during the first years after a nuclear accident could mislead the long-term radiation predictions, as these coefficients depend on time and depth. These difficulties are clearly expected to be complicated in a forest ecosystem, as more complex factors are involved in this ecosystem than in that of grassland soils. Therefore, continuous monitoring is required, and precise prediction models should be developed for a clear and better understanding of the timedependent migration of radiocesium in forested environments. 4. Conclusions On the forest floor, the understory plants and the upper few centimeters of soil can be considered as an active radiocesium remobilization interphase, primarily due to its acidic nature and low clay content that made radiocesium bioavailable particularly for short-lived understory plants. Almost all of the radiocesium activity (99% of the total soil inventory) was found in the upper ~10 cm in which the OM content was greater than 10%. The raw organic layer (Ol þ Of) holds 52% of the Fukushima-derived 137Cs at the time of soil sampling, and the remaining 137Cs is distributed in the soil below the Of- layer. Specifically, the Of- horizon seems to accumulate Fukushima-derived 137Cs, which accounts for 47% of the total inventory at the time of sampling, and retards its subsequent migration, indicating that the OM turnover characterizing

Fig. 5. Comparative position of the input parameters used to determine the Diffusion (D) and Migration rate (V) coefficients for the pre- and Fukushima-derived radiocesium distribution.

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45

the Of-layer dynamics and its periodical changes will likely determine the migration, the residence time and the bioavailability of radiocesium. Most downward migration models consider a single phase of migration rate that assume a gradual and slow adsorptionedesorption processes of radiocesium movement in the soil profile. Nevertheless, we observed some fraction of Fukushimaderived 137Cs at a depth 16 cm which most likely due to the infiltration of radiocesium-circumscribed rainwater during the fallout before selective adsorption started. This implies that in forest soil there is additional and quick phase of radiocesium migration. In fact this phase cannot represent the long-term migration of Fukushima-derived 137Cs but still it is worth important to understand the speed of contamination at the initial fallout period. Acknowledgments This study is part of the Core Research for Evolutional Science and Technology (CREST) research project “Development of Innovative Technologies for Increasing in Watershed Runoff and Improving River Environment by the Management Practice of Devastated Forest Plantation”. This work was also partially supported by Grant-in-Aid for Scientific Research on Innovative Areas Grant Number 24110006 of the Ministry of Education, Culture, Sports, Science and Technology of Japan. References Almgren, S., Isaksson, M., 2006. Vertical migration studies of 137Cs from nuclear weapons fallout and Chernobyl accident. J. Environ. Radioact. 91, 90e102. Bunzl, K., Kracke, W., Schimmack, W., Zelles, L., 1998. Forms of fallout137Cs and 239þ240 Pu in successive horizons of forest soil. J. Environ. Radioact. 39 (1), 58e68. Bunzl, K., Schimmack, W., Kreutzer, K., Schierl, R., 1989. Interception and retention of Chernobyl-derived 134Cs, 137Cs and 106Ru in Spruce stand. Sci. Total Environ. 78, 77e87. Bunzl, K., Schimmack, W., Krouglov, S.V., Alexakhin, R.M., 1995. Change with time in the migration of radiocesium in soil, as observed near Chernobyl and Germany, 1986e1994. Sci. Total Environ. 175, 49e56. Calmon, P., Thiry, Y., Zibold, G., Rantavaara, A., Fesenko, S., 2009. Transfer parameter values in temperate forest ecosystems: a review. J. Environ. Radioact. 100, 757e766. de Brouwer, S., Thiry, Y., Myttenaere, C., 1994. Availability and fixation of radiocesium in forest brown acid soil. Sci. Total Environ. 143, 183e191. Delvaux, B., Kruyts, N., Cremers, A., 2000. Rhizospheric mobilization of radiocesium in soils. Environ. Sci. Technol. 34, 1489e1493. Dorr, H., Munnich, O.K., 1991. Lead and Cesium transport in European Forest soils. Water Air Soil. Pollut. 57e58, 809e818. Fawaris, B.H., Johanson, K.J., 1994. Radiocesium in soil and plants in a forest in central Sweden. Sci. Total Environ. 157, 133e138. Fesenko, S.V., Soukhova, N.V., Sanzharova, N.I., Avila, R., Spiridonov, S.I., Klein, D., et al., 2001. Identification of processes governing long-term accumulation of 137Cs by forest trees following the Chernobyl accident. Radiat. Environ. Biophys. 40,105e113. Fujiyoshi, R., Sawamura, S., 2004. Mesoscale variability of vertical profiles of environmental radionuclides (40K, 226Ra, 210Pb and 137Cs) in temperate forest soils in Germany. Sci. Total Environ. 320, 177e188. Gomi, T., Sidle, R.C., Richardson, J.S., 2002. Understanding processes and downstream linkages of head water systems. J. Biosci. 52 (10), 905e916. Goor, F., Thiry, Y., Delvaux, B., 2007. Radiocaesium accumulation in stemwood: integrated approach at the scale of forest stands for contaminated Scots pine in Belarus. J. Environ. Manag. 85, 129e136. He, Q., Walling, D.E., 1997. The distribution of fallout 137Cs and 210Pb in undisturbed and cultivated soils. Appl. Radioact. Isot. 48, 677e690. IAEA, 2010. Handbook of parameter values for the prediction of radionuclide transfer in terrestrial and freshwater environments. Tech. Reports Ser. 472, 1e194. Isaksson, M., Erlandsson, B., 1995. Experimental determination of the vertical and horizontal distribution of 137Cs in the ground. J. Environ. Radioact. 27 (2), 141e160. IUSS Working group WRB, 1998. World Reference Base for Soil Resources. World Soil Resources Reports No. 84, FAO, Rome. Karadeniz, O., Yaprak, G., 2008. Vertical distribution and gamma dose rates of 40K, 232 Th, 238U and 137Cs in the selected forest soils in Izmir, Turkey. J. Radiat. Prot. Dosim., 1e10. http://dx.doi.org/10.1093/rpd/ncn185. Kato, H., Onda, Y., Teramage, M., 2012. Depth distribution of 137Cs, 134Cs and 131I in soil profile after Fukushima Dai-ichi Nuclear power plant accident. J. Environ. Radioact. 111, 59e64. Koarashi, J., Atarashi-Andoh, M., Matsunaga, T., Sato, T., Nagao, S., Nagai, H., 2012. Factors affecting distribution of Fukushima accident-derived radiocesium in soil under different land-use conditions. Sci. Total Environ. 431, 392e401.

45

Konopleva, I., Klemt, E., Konoplev, A., Zibold, G., 2009. Migration and bioavailability of 137Cs in forest soil of southern Germany. J. Environ. Radioact. 100, 315e321. Kruyts, N., Delvaux, H., 2002. Soil organic horizons as a major source for radiocesium biorecycling in forest ecosystems. J. Environ. Radioact. 58, 175e190. Livens, F.R., Fowler, D., Horrill, A.D., 1992. Wet and dry deposition of 131I, 134Cs and 137 Cs at an upland site in northern England. J. Environ. Radioact. 16, 243e254. Loughran, R.J., Wallbrink, P.J., Walling, D.E., Appleby, P.G., 2002. Chapter 3: sampling method. In: Zapata, F. (Ed.), Handbook for the Assessment of Soil Erosion and Sedimentation Using Environmental Radionuclides. Kluwer Academic Publishers, The Netherlands, pp. 41e57. Maes, E., Delvaux, B., Thiry, Y., 1998. Fixation of radiocaesium in an acid brown forest soil. Eur. J. Soil. Sci. 49, 133e140. Melin, J., Wallberg, L., Suomela, J., 1994. Distribution and retention of cesium and strontium in Swedish boreal forest ecosystems. Sci. Total Environ. 157, 93e105. MEXT, 2011. Ministry of Education, Culture, Sports, Science and Technology, Japan (MEXT). Corrections to the Readings of Airborne Monitoring Surveys (Soil Concentration Map) Based on the Prepared Distribution Map of Radiation Doses, etc. By MEXT. http://radioactivity.mext.go.jp/old/en/1270/2011/08/1270_ 083014-2.pdf (accessed 29.11.11.). Onda, Y., Gomi, T., Mizugaki, S., Nonoda, T., Sidle, R., 2010. An overview of the field and modeling studies on effects of forest devastation on flooding and environmental issues. Hydrol. Process. 24, 527e534. Poreba, G., Bluszcz, A., Snieszko, Z., 2003. Concentration and vertical distribution of 137 Cs in agricultural and undistributed soil from Chechlo and Czarnocin areas. J. Methods Appl. Absol. Chronology 22, 67e72. Porto, P., Walling, D.E., Ferro, V., 2001. Validating the use of caesium -137 measurements to estimate soil erosion rates in a small drainage basin in Calabria, Southern Italy. J. Hydrol. 248, 93e108. Rafferty, B., Brenna, M., Dawson, M., Dowding, D., 2000. Mechanism of 137Cs migration in coniferous forest soils. J. Environ. Radioact. 48, 131e143. Raitio, H., Rrantavaara, A., 1994. Airborn radiocesium in Scots pine and Norway spruce needles. Sci. Total Environ. 157, 171e180. Rosen, K., Oborn, I., Lonsjo, H., 1999. Migration of radiocesium in Swedish soil profiles after Chernobyl accident, 1986e1995. J. Environ. Radioact. 46, 45e66. Sakaguchi, A., Kawai, K., Steier, P., Imanaka, T., Hoshi, M., Endo, S., Zhumadilov, K., Yamamoto, M., 2010. Feasibility of using 236U to reconstruct close-in fallout deposition from the Hiroshima Atomic Bomb. Sci. Total Environ. 408, 5392e5398. Schimmack, W., Bunzl, K., 1992. Migration of radiocesium in two forest soils as obtained from field and column investigations. Sci. Total Environ. 116, 93e107. Schimmack, W., Bunzl, K., Zelles, L., 1989. Initial rates of migration of radionuclides from the Chernobyl fallout in undisturbed soils. Geoderma 44, 211e218. Schimmack, W., Marquez, F.F., 2006. Migration of fallout radiocesium in grassland soil from 1986 to 2001; Part II: evaluation of activity-depth profiles by transport models. Sci. Total Environ. 368, 863e874. Sombre, L., Vanhouche, M., de Brouwer, S., Ronneau, C., Lambotte, J.M., Myttenaere, C., 1994. Long-term radiocesium behavior in spruce and Oak forests. Sci. Total Environ. 157, 59e71. Strandberg, M., 1994. Radiocesium in a Danish pine forest ecosystem. Sci. Total Environ. 157, 125e132. Strebl, F., Gerzabek, M.H., Karg, V., Tataruch, F., 1996. 137Cs-migration in soil and its transfer to roe deer in an Austrian forest stand. Sci. Total Environ. 181, 237e247. Sun, X., Onda, Y., Kato, H., Otsuki, K., Gomi, T., 2013. Partitioning of the total evapotranspiration in a Japanese cypress plantation during the growing season. Ecoydrology. http://dx.doi.org/10.1002/eco.1428. Teramage,., T.M., Onda, Y., Kato, H., Gomi, T., 2014. The role of litterfall in transferring Fukushima-derived radiocesium to a coniferous forest floor. Sci. Total Environ. http://dx.doi.org/10.1016/j.scitotenv.2014.05.034. Teramage, T.M., Onda, Y., Kato, H., Wakiyama, Y., Mizugaki, S., Hiramatsu, S., 2013. The relationship of soil organic carbon to 210Pbex and 137Cs during surface soil erosion in hillslope forested environment. Geoderma 192, 59e67. Thiry, Y., Kruytz, N., Delvaux, B., 2000. Respective horizon contributions to Cesium137 soil-to-plant transfer: a pot experiment approach. J. Environ. Qual. 29, 1194e1199. Thiry, Y., Myttenaere, C., 1993. Behavior of radiocaesium in forest multilayered soils. J. Environ. Radioact. 18, 247e257. Thorring, H., Skuterud, L., Steinnes, E., 2012. Distribution and turnover of 137Cs in birch forest ecosystem: influence of precipitation chemistry. J. Environ. Radioact. 110, 69e77. Tikhomirov, F.A., Shcheglov, A.I., Sidorov, V.P., 1993. Forests and forestry: radiation protection measures with special reference to Chernobyl accident zone. Sci. Total Environ. 137, 289e305. Tobler, L., Bajo, S., Wyttenbach, A., 1988. Deposition of 134,137Cs from Chernobyl fallout on Norway spruce and forest soil and its incorporation in to spruce twigs. J. Radioact. 6, 225e245. Valcke, E., Cremers, A., 1994. Sorption desorption dynamics of radiocaesium in organic matter soils. Sci. Total Environ. 157, 275e283. Walling, D.E., He, Q., Appleby, P.G., 2002. Conversion models for use in soil-erosion, soil-redistribution and sedimentation investigations. In: Zapata, F. (Ed.), Handbook for the Assessment of Soil Erosion and Sedimentation Using Environmental Radionuclides. Kluwer Academic Publishers, Dordrecht, Netherlands, pp.111e164. Witkamp, M., Frank, M.L., 1964. First year movement, distribution and availability of 137 Cs in the forest floor under tagged Tulip Poplars. Radiat. Bot. 4, 485e495. Zibold, G., Klemt, E., Konopleva, I., Konoplev, A., 2009. Influence of fertilizing on the 137 Cs soil-plant transfer in a spruce forest of Southern Germany. J. Environ. Radioact. 100, 489e496.