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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 Center for Research in Isotopes and Environmental Dynamic, University of Tsukuba, Tennodai 1-1-1, Tsukuba shi, Ibaraki 305-8572, Japan b Department of International Environmental and Agriculture Science, Tokyo University of Agriculture and Technology, Fuchuu, Tokyo 183-8509, Japan article info Article history: Received 31 January 2014 Received in revised form 6 June 2014 Accepted 17 June 2014 Available online Keywords: Deposition Forest oor Migration Radiocesium 137 Cs 134 Cs abstract This study deals with the description of the vertical distribution of radiocaesium ( 137 Cs and 134 Cs) in a representative coniferous forest soil, investigated 10 months after the Fukushima radioactive fallout. During soil sampling, the forest oor components (understory plants, litter (Ol-) and fermented layers (Of)) were collected and treated separately. The results indicate that radiocesium is concentrated in the forest oor, and high radiocesium transfer factor observed in the undergrowth plants (3.3). This made the forest oor an active exchanging interphase for radiocesium. The raw organic layer (Ol þ Of) holds 52% (5.3 kBq m 2 ) of the Fukushima-derived and 25% (0.7 kBq m 2 ) of the pre-Fukushima 137 Cs at the time of the soil sampling. Including the pre-Fukushima 137 Cs, 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 Fukushima- derived 137 Cs at a depth of 16 cm is most likely due to the inltration of radiocesium-circumscribed rainwater during the fallout before that selective adsorption prevails and reduces the migration of sol- uble 137 Cs. The values of the depth distribution parameters revealed that the distribution of the Fukushima-derived 137 Cs was somewhat rapid. © 2014 Elsevier Ltd. All rights reserved. 1. Introduction Radiocesium ( 134 Cs, t 1/2 ¼ 2.1 y, and 137 Cs, t 1/2 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 oor 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 oor 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 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 oor through hydrological pathways (i.e throughfall) and fallen canopy components (i.e. litterfall) will undergo horizontal and vertical migration (Teramage et al., 2013). On at, 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 oor 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 oor can reect 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 radio- cesium accumulated in surface organic layers of forest soil may result in contamination of the soil and rivers for a long time. * Corresponding author. Tel.: þ81 90 8569 3775; fax: þ81 29 853 4226. E-mail address: [email protected] (M.T. Teramage). Contents lists available at ScienceDirect Journal of Environmental Radioactivity journal homepage: www.elsevier.com/locate/jenvrad http://dx.doi.org/10.1016/j.jenvrad.2014.06.017 0265-931X/© 2014 Elsevier Ltd. All rights reserved. Journal of Environmental Radioactivity 137 (2014) 37e45

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Journal of Environmental Radioactivity 137 (2014) 37e45

Contents lists avai

Journal of Environmental Radioactivity

journal homepage: www.elsevier .com/locate/ jenvrad

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

Mengistu T. Teramage a, *, Yuichi Onda a, Jeremy Patin a, Hiroaki Kato a, Takashi Gomi b,Sooyoun Nam b

a Center for Research in Isotopes and Environmental Dynamic, University of Tsukuba, Tennodai 1-1-1, Tsukuba shi, Ibaraki 305-8572, Japanb 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

Article history:Received 31 January 2014Received in revised form6 June 2014Accepted 17 June 2014Available online

Keywords:DepositionForest floorMigrationRadiocesium137Cs134Cs

* Corresponding author. Tel.: þ81 90 8569 3775; faE-mail address: [email protected] (M.T. Teram

http://dx.doi.org/10.1016/j.jenvrad.2014.06.0170265-931X/© 2014 Elsevier Ltd. All rights reserved.

a b s t r a c t

This study deals with the description of the vertical distribution of radiocaesium (137Cs and 134Cs) in arepresentative 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 theforest floor, and high radiocesium transfer factor observed in the undergrowth plants (3.3). This made theforest floor an active exchanging interphase for radiocesium. The raw organic layer (Ol þ Of) holds 52%(5.3 kBq m�2) of the Fukushima-derived and 25% (0.7 kBq m�2) of the pre-Fukushima 137Cs at the time ofthe soil sampling. Including the pre-Fukushima 137Cs, 99% of the total soil inventory was in the upper10 cm, in which the organic matter (OM) content was greater than 10%, suggesting the subsequentdistribution most likely depends on the OM turnover. However, the small fraction of the Fukushima-derived 137Cs at a depth of 16 cm is most likely due to the infiltration of radiocesium-circumscribedrainwater during the fallout before that selective adsorption prevails and reduces the migration of sol-uble 137Cs. The values of the depth distribution parameters revealed that the distribution of theFukushima-derived 137Cs was somewhat rapid.

© 2014 Elsevier Ltd. All rights reserved.

1. Introduction

Radiocesium (134Cs, t1/2 ¼ 2.1 y, and 137Cs, t1/2 ¼ 30.2 y) derivedfrom the Fukushima Dai-ichi Nuclear Power Plant (hereinafterFDNPP) 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 byforests, of which most are evergreen coniferous forests (Onda et al.,2010). However, processes ruling the local dynamics of radiocesiumtransfer from the forest canopy to the forest floor and its furtherdistribution in soil are still few documented.

Numerous studies dealing with the early redistribution ofradiocesium deposited onto the forest canopy, including canopyinterception and subsequent transfer from the canopy to forestfloor were carried out in the aftermath of Chernobyl accident(Bunzl et al., 1989; Tikhomirov et al., 1993) and most of themfocused on the dry climatic conditions of Chernobyl-affected

x: þ81 29 853 4226.age).

regions. However, such information is lacking in the case of theFukushima reactor accident and humid climatic regions, whichmight cause a different behavior of radiocesium migration anddistribution (Teramage et al., 2014).

The deposited fraction of radionuclides onto the forest floorthrough hydrological pathways (i.e throughfall) and fallen canopycomponents (i.e. litterfall) will undergo horizontal and verticalmigration (Teramage et al., 2013). On flat, undisturbed sites, thehighest radiocesium concentration is in the uppermost soil layerand decreases exponentially with the depth (e.g He and Walling,1997). However, the litter layer on the forest floor can play aunique role in the distribution of radiocesium that is typicallylacking in most other land use types. The litter layer accumulatesboth throughfall and litterfall-derived radiocesiumdeposits, and thevariation of radiocesium distribution on the forest floor can reflectthe resultant effects of these two mechanisms of accumulation.

As most of the heads of water resources are located in forestedwatersheds and are intimately linked to downstream ecosystems(Gomi et al., 2002; Sun et al., 2013), the remobilization of radio-cesium accumulated in surface organic layers of forest soil mayresult in contamination of the soil and rivers for a long time.

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Fig. 2. Monthly (bars) and cumulative (bold line) precipitation for one year at thestudy site. ¼ indicates arrival of the radionuclide plume at the study site in mid-March (2011). The wet deposition was expected due to the precipitation from March15 to the first two weeks in April 2011.

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e4538

Therefore, understanding the early distribution and subsequentmigration of radiocesium in the soil profile is essential. Suchknowledge is helpful in decisions making about possible counter-measures, to set up environmental baselines, and to establish pa-rameters to predict radiocesium transfer in forested ecosystems.

Despite the importance of understanding the movement ofradiocesium in forest soils, most of the recent studies on soilsaffected by radioactive deposits focused on agricultural soil andundisturbed sites in non-forested environments such as pastureland (e.g., He andWalling, 1997; Kato et al., 2012); therefore, little isknown about forest soil. Among the few studies on Chernobyl-affected forest soils, Karadeniz and Yaprak (2008) reported arelaxation length (i.e., the inverse of the rate of change in theconcentration that represents the depth of migration due to changein concentration) of 4e15 cm, and Poreba et al. (2003) indicated arelaxation length of 2.1 cm several years after the accident. How-ever, these findings cannot be simply applied to the Japaneseenvironment due to differences in the climatic conditions andvarious physiographic features. Recently, Koarashi et al. (2012) re-ported the vertical distribution of Fukushima-derived 137Cs indifferent land uses and identified major controlling factors for itsmigration using the core soil sampling method with coarse soilsection of 1e5 cm. However, fine depth sections are expected tobetter describe the distribution of radiocesium (Kato et al., 2012;Loughran et al., 2002). In addition, the composition and contami-nation density of radiocesium are expected to vary with the dis-tance from the source. This is mainly because of the difference inparticle composition of the radiocesium debris, i.e. heaviercomposition likely deposits near to source area. Therefore, thisstudy investigated the radiocesium depth distribution in the forestsoil profile based on fine depth resolution. The radiocesiummigration and soil-to-plant transfer rates were also estimated.

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 areais located 180 km southwest of the FDNPP. The size of the catch-ment is 0.8 ha.

The climate of the area is humid temperate, with mean annualrainfall of 1259 mm and mean annual temperature of 14.1 �C (as

Fig. 1. Map of the study area and location of the sampling point. Sol

obtained from the Karasawa mountain metrological station from2010 to 2011). The soil type was classified as an orthic cambisolaccording to the World Reference Base for Soil Resource (IUSSWorking group WRB, 1998). The estimated stand density isapproximately 2500 trees per hectare. The dominating understoryvegetation is composed of sparsely grown understory plants(marlberry (Ardisia japonica (Thunb.) Blume) in addition to variousherbs.

According to the MEXT (2011) report, the plume extended to thestudy site was at the origin of 137Cs deposits of approximately10 kBq m�2. The low temperature at the time of accident main-tained the plume movement near to the ground surface driven bythe local wind. In that context, it is assumed that the depositionprimarily occurred through wet deposition. Fig. 2 shows themonthly and cumulative total rainfall around the time of the acci-dent and the sampling date.

2.2. Soil sampling

The soil samples were collected from an undisturbed flat area tominimize the effect of the subsequent lateral movement of radio-nuclides after fallout. The sampling point was purposely selected atthe midpoints between tree lines to make the sampling morerepresentatives. We used a rectangular metal-framed scraper plate

id circles in the map represent the capital and Prefectures cites.

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45 39

(internal dimensions of 15 cm � 30 cm) with an adjustable depthincrement of 5 and 10 mm intervals. This sampling method allowsfor the collection of numerous and voluminous soil samples thatencompass and represent the relatively wider microtopographicvariability (Kato et al., 2012; Loughran et al., 2002).

The sampling was performed on 16 January 2012 (about 10month after the accident). The understory vegetation, Ol- (~3 cmthick) and Of- (~4.5 cm thick) organic horizons were carefullyseparated by hand and scissors (when necessary) to represent theforest floor. The understory vegetation with in the sampling plotwhich was dominated by sparsely grown marlberry and annualherbs were carefully mowed. Subsequently, their radiocesiumcontent in dry basis was determined after crushing into powder.The Ol-horizon, the original shape of the components was easilyrecognizable, was composed of periodically falling raw litter. TheOf- horizon, which was located under the Ol-horizon, wascomposed of an early fermented and fragmented litter componentin which the original shapes of the litter were difficult to identify.

The radiocesium uptake by the understory plants is usuallyestimated based on the soil-to-plant transfer factor on a dry weightbasis (TFw) or aggregated transfer factor (Tagg) coefficients, whichare defined as the ratio of average radionuclide concentration inplants (Bq kg�1) to that in the soil (Bq kg�1) for the TFw or the totalsoil inventory (Bq m�2) for the Tagg (Calmon et al., 2009). The Tagg isoften used for the medium-to-long term measurements afterdeposition, when radiocesium have been redistributed betweenthe different surface layers of forest soils (IAEA, 2010), as radio-cesium requires a long time to reach the tree root zone. In our studysite, we assumed that the roots of the understory vegetationmainlyexplore the Of- horizon and the upper 2 cm of soil and it isreasonable to use the TFw instead of the Tagg.

The soil below the Of- horizon was scraped layer-by-layer inthree major depth resolutions of 5 mm (for upper 5 cm), 10mm (for5e10 cm) and 20 mm (for 10e30 cm). We employed a depthincrement of 0.5 cm for the upper 5.0 cm to reduce the possibleeffects of soil thickness on defining the shape of the profile, unlikethat used by Koarashi et al. (2012). In the sampling depth, neithercracks nor large roots were encountered.

2.3. Laboratory analysis

2.3.1. Measurement of radiocesium activityAll of the samples were dried at 110 �C for 24 h to determine the

dry weight. The samples from the understory vegetation and fromthe Ol- and Of- horizons were then milled and mixed to ensurehomogenous sample material for each respective sampling unit.The soil samples were disaggregated by gentle grinding and werethen passed through a 2-mm sieve. Then, the milled samples fromthe understory vegetation and from the Ol- and Of- horizons andthe <2 mm soil fraction from each soil layer were placed in plasticcontainers (U-8, As ONE, Tokyo) and sealed for analysis. The anal-ysis was conducted in the laboratory of the University of Tsukuba,which was authorized for independent calibration checks duringthe worldwide open proficiency test in 2006 (IAEA/AL/171). Theactivities of 137Cs and 134Cs were determined using gamma rayspectrometry from a high purity n-type germanium coaxial gammaray detector (EGC 25e195-R, Canberra-Eurysis, Meriden, USA)connected to an amplifier (PSC822, Canberra, Meriden, USA) and amultichannel analyzer (DSA1000, Canberra, Meriden, USA) usingthe counts at the 662 keV and 605 keV peaks, respectively. Theabsolute counting efficiency of the detector was calibrated usingvarious weights of IAEA-2006-03 standard soil samples withbackground correction. In this study, the measured radiocesiumconcentrations (Bq kg�1) were converted to inventories (Bq m�2)using the volume and apparent bulk density on the basis of dry

weight of each sampling layer. All of the measured activities weredecay-corrected to May 20, 2011.

2.3.2. Soil physicochemical property analysisThe 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 wasanalyzed using a laser diffraction particle size analyzer (SALD-3100,Shimadzu Co., Ltd., Kyoto, Japan). The bulk density was determinedfrom the dry weight and volume of the soil in each layer. Theorganic matter (OM) content of the samples was determined by theweight lost after the incineration of a known dryweight sample in amuffle furnace at 450 �C for 4 h. The pH was determined using a pHmeter by mixing 5 g dry soil with 50 ml distilled water to a final 1:5soil:water ratio.

2.4. Parameters for the radiocesium depth profile

To investigate the vertical distribution of radiocesium, weapplied a negative exponential profile function, which includes anumber of simplified assumptions (Karadeniz and Yaprak, 2008).This method was used to estimate the radiocesium movement inearly stages after the fallout. Importantly, in forest soils, theorganic-rich upper layers could affect the theoretical exponentialfunction. Therefore, we discussed the radiocesium concentration inthe forest floor (hereafter refers to the composition of the under-story vegetation and of the Ol- and Of- horizons), while the depthpenetration of radiocesium in the soil below the Of- layer wasassumed to follow the general formula:

CðzÞ ¼ Cð0Þe�aZp(1)

where z is the depth from the soil surface (cm), C (z) is the concen-tration of radiocesium at depth z (Bq kg�1); C (0) is the radiocesiumconcentration (Bq kg�1) of initial deposit on the upper surface soil; a(cm�1) is the reciprocal of the relaxation length; and p (unit-less) isan experimentally determinable parameter depending on the uppersoil surface condition and form of transport.

The parameter a depends on the characteristics of the radio-nuclides, the soil type and the physicochemical characteristics, landuse type, time elapsed after deposits and climatic conditions (Katoet al., 2012). The reciprocal value of a represents the relaxationlength of the radiocesium in the vertical profile (Karadeniz andYaprak, 2008). The depth can be expressed either in linear (cm orm) or mass depth (kg m�2). The linear relaxation length, 1/a, rep-resents the shape of the tail of the depth distribution, whereas therelaxation mass depth describes its penetration strength into soilmass. When p ¼ 1, the following function (Porto et al., 2001) can beused as an alternative to directly estimate h0 by fitting the model tothe empirical data of the reference site as:

Cðz0Þ ¼ Cð0Þe�z0h0 (2)

where z0 is the mass depth from soil surface (kg m�2); C(z0) is theconcentration of radiocesium at depth z0 (Bq kg�1); C(0) is the radio-cesiumconcentration (Bqkg�1) of the surface soil; andh0 canbeeasilydetermined using the linearized least square regression method.

2.5. Characterizing the pre-Fukushima 137Cs

The observed total 137Cs concentration in the soil includesremnants of the pre-Fukushima episodes, whereas due to its shorthalf-life, the observed 134Cs in the environment is exclusivelyoriginates from Fukushima. To determine the pre-Fukushima 137Csconcentration, the 134Cs/137Cs ratio was used (Livens et al., 1992).

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e4540

The ratio of Fukushima-derived radiocesium deposits was deter-mined from litter collected immediately after the fallout and wasfound to be 1 (Teramage et al., 2014), and the pre-Fukushima 137Csresidue was determined accordingly. Note that in this study wedescribe 137Cs as pre-Fukushima137Cs, Fukushima-derived 137Cs(~134Cs) or total 137Cs (representing pre- and Fukushima-derived137Cs altogether); otherwise radiocesium is used to represent thegeneral and common features of both radioisotopes.

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

Although the depth distribution of fresh fallout is oftendescribed using an exponential function as indicated in Eq. (1), thefunction fails to describe the downward transport that occurs fromthe first instant of deposition and is unable to illustrate long-termdistribution (Almgren and Isaksson, 2006). The redistribution ofradiocesium within the soil profile is the result of a complex set ofprocesses that should be considered by the time-dependent model(Walling et al., 2002). Such transport mechanisms have beencharacterized by a diffusion coefficient (D, kg2 m�4 y�1) and amigration or convection rate (V, kg m�2 y�1) that lump together allof the redistribution processes in the soil column (Walling et al.,2002). These transport coefficients can also be represented in thedimension of D in cm2 y�1 and V in cm y�1. In this study, we esti-mated these two parameters based on the formula proposed byWalling et al. (2002):

DzðNz �WzÞ22ðt � t0Þ

(3)

VzWz

t � t0(4)

Table 1Physiochemical properties and radiocesium concentration in soil profile.

Depth (cm) Mass depth (kg m�2) Bulk density (g cm�3) Sand (%) Silt (%) C

Ol 0.7 0.02 e e e

Of 6.5 0.1 e e e

0.0e0.5 3.8 0.8 43 510.5e1.0 1.8 0.4 36 561.0e1.5 2.3 0.5 31 621.5e2.0 2.7 0.5 17 73 12.0e2.5 2.1 0.4 32 58 12.5e3.0 4.0 0.8 41 48 13.0e3.5 2.6 0.5 39 533.5e4.0 3.7 0.7 36 574.0e4.5 3.1 0.6 36 53 14.5e5.0 5.9 1.2 34 55 15.0e6.0 8.4 0.8 36 566.0e7.0 6.7 0.7 36 53 17.0e8.0 3.8 0.4 27 63 18.0e9.0 9.4 0.9 21 68 19.0e10 8.7 0.9 17 72 110e12 11.6 0.6 27 61 112e14 12.5 0.6 31 57 114e16 13.5 0.7 34 6016e18 14.2 0.7 21 69 118e20 10.2 0.5 20 68 120e22 12.7 0.6 17 72 122e24 11.7 0.6 21 69 124e26 7.6 0.4 16 72 126e28 11.4 0.6 18 71 128e30 13.1 0.7 50 42

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

where t is the sampling year; t0 is the maximum fallout year (2011for Fukushima, 1986 for Chernobyl and 1963 for bomb fallout); Nz isthe mass depth (kg m�2) or linear depth (cm) of the pre- andFukushima-derived 137Cs concentration reduced to 1/e of themaximum concentration; and Wz is the mass depth (kg m�2) orlinear depth (cm) where the maximum pre- and Fukushima-derived 137Cs concentration is located at the time of measurement.

3. Results and discussion

3.1. Radiocesium in forest soils

3.1.1. Soil physicochemical propertiesThe 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 fractiongenerally dominated the soil texture composition in each examinedlayer, while the clay, sand contents and bulk density showed nodefinite pattern (Table 1).

The organic matter contents of the Ol- and Of- layers were 87.2and 74.4%, respectively. Below these layers, the OM contentdecreased sharply and continuously with increasing depth. How-ever, the OM content was greater than 10% in all of the soil layers inthe upper 10 cm, with the highest proportion (27%) in the upper0e0.5 cm. The pH of the soil was acidic that ranged from 5.10 to5.92 and showed a slight general increase below the 5 cm soil depth(Table 1).

3.1.2. Radiocesium on the forest floor3.1.2.1. Understory vegetation. The biomass of the understoryvegetation was 0.4 kg m�2 which contains 1.53 ± 0.07 kBq kg�1 ofFukushima-derived 137Cs. The TFw ratio (Bq kg�1 in the plant/Bq kg�1 in the upper soil) was estimated to be 3.3. These valuesimply that the concentration of Fukushima-derived 137Cs in a givendry-weight understory vegetation exceeds the average

lay (%) OM (%) pH 137Cs (Bq kg�1) 134Cs (Bq kg�1) Inventoryproportion

137Cs % 134Cs %

87.2 4.93 669 ± 33 666 ± 37 4 574.4 4.96 840 ± 43 739 ± 43 43 47

6 26.7 5.12 422 ± 15 389 ± 15 13 158 20 5.49 362 ± 18 285 ± 17 5 57 18 5.19 349 ± 18 323 ± 19 6 70 17 5.27 215 ± 9 158 ± 9 5 40 16 5.14 168 ± 12 139 ± 11 3 31 15 5.39 118 ± 7 88 ± 7 4 38 14 5.10 93 ± 6 58 ± 6 2 17 13 5.17 92 ± 4 56 ± 4 3 21 12 5.17 61 ± 2 30 ± 3 1 11 13 5.37 55 ± 4 29 ± 3 3 29 12 5.52 45 ± 3 15 ± 3 3 11 12 5.36 45 ± 3 18 ± 3 2 10 12 5.64 32 ± 3 11 ± 3 1 0.41 10 5.39 29 ± 3 9 ± 3 2 11 10 5.71 9 ± 1 7 ± 1 1 11 9 5.92 6 ± 1 Nd 1 e

1 7 5.39 Nd Nd e e

6 5 5.83 5 ± 1 5 ± 1 1 10 5 5.67 Nd Nd e e

1 5 5.74 Nd Nd e e

1 5 5.73 Nd Nd e e

0 5 5.77 Nd Nd e e

2 4 5.70 Nd Nd e e

1 5 5.71 Nd Nd e e

8 4 5.70 Nd Nd e e

detected.

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45 41

concentration found in the soil under consideration by three-fold,indicating uptake by the plants. Our result of TFw ratio obtainedfrom bulked understory plants was slightly higher than the valuesreported by Zibold et al. (2009) for ferns (1.2e3.2) and considerablydifferent from that of blackberries (0.3e0.6). Note that because theforest floor was a secondary receiver of radiocesium next to thephytomass at the time of fallout, the observed Fukushima-derived137Cs in the understory vegetation could be from two possibleoriginate, i.e., via root uptake from the forest floor and via the directdeposition on plant organs. Given the possibility of renewal of veryshort-lived understory plants, the effect of direct interception byunderstory could be still predominant which may attributed tohigh measured TFw. Another reason could be the acidic nature andlow 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 Of-horizons were 0.7 kg m�2 and 6.5 kg m�2, respectively and theradiocesium concentrations were greatest in both layers (Table 1).The retained fraction, defined as the ratio of inventory at eachdepth section to that of the total inventory of the soil, showed thatthe Of- horizon retained 47% of the Fukushima-derived 137Cs in-ventory (Table 1). The raw organic layer (Ol þ Of) holds 52%(5.3 kBq m�2) of the Fukushima-derived and 25% (0.7 kBq m�2) ofpre-Fukushima 137Cs at the time of the soil sampling (16 January2012), and the remaining is distributed in the soil below the Of-layer.

Recent study demonstrated that litterfall tends to continuedepositing radiocesium on the forest floor (Teramage et al., 2014).Therefore, the radioactivity on the forest floor is expected to in-crease in subsequent periods, as considerable proportion of initialatmospheric radiocesium deposits is possibly held in the canopy. Inspite of this likelihood, only 5% of the Fukushima-derived 137Cscontent was found in the Ol- horizon at the time of observation.

Table 2The retention 137Cs in the litter layer in different locations affected by Chernobyl nuclear

Reference Location Tree /forest type Fallout density(kBq md2)

Witkamp and Frank(1964)

USA Tulip poplar e

Schimmack and Bunzl(1992)

Germany Norway Spruce forest 30

Scots pine forest 9Tobler et al.(1988) Switzerland Norway spruce forest 8

Tikhomirov et al. (1993) Ukraine &Belarus

Birch-oak-pine mixedforest

210e250

Strebl et al.(1996) Austrian Norway spruce forest e

Strandberg (1994) Denmark Scots pine forest 0.9

Melin et al.(1994) Sweden Pine forest 180Raitio and Rantavaara

(1994)Finland Norway spruce forest e

Scots pine forest e

Bunzl et al.(1998) Germany Norway spruce forest 20Fesenko et al.(2001) Russia Pine 800e2300

Pine and Birch mixedforest

890e2850

Pine and Birch mixedforest

1600

This study Japan Japanese cypress forest 10

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

This result is most likely due to the transfer of the Ol- horizon to theOf- horizon through mechanical and biological breakdown(Rafferty et al., 2000).

Comparing our results with those of Chernobyl highlights thedifferences in the migration of radiocesium in a forested environ-ment. For this reason, we compiled pre-Fukushima 137Cs retentionin a layer that combined Ol- and Of- horizons reported in previousstudies from diverse locations, forest types and different time pe-riods (Table 2) and were plotted them against a time elapsed be-tween the accidents and sampling periods (Fig. 3). It shows ageneral decrease over time and the retained 137Cs proportion variedwith the forest and soil types. Relatively, the migration of theFukushima-derived 137Cs in our study seems rapid (Fig. 3), withapproximately half of the total inventories were already trans-ported below the Of- horizon in less than a year. Despite the sitedifference, the tendency of the reported retained pre-Fukushima137Cs fraction in litter layer seems to follow that of the depositiondensity. This can be clearly observed in Fig. 3 (shaded) on theretained fraction values obtained in a similar time distance after theChernobyl accident. The differencemight be attributed partly to thevariation in the depositional density, and it likely masked thedownward migration rate in highly contaminated areas possiblydue to the input load of radiocesium into the litter layer largelyexceeds to that leaving the layer. This implies that fallout densityshould be taken in to account when comparing the migration ratein forested environment.

3.1.3. Distribution of the radiocesium inventory in the soilThe total 137Cs inventory of the soil profile was 13 kBq m�2, to

which the Fukushima accident contributed approximately 77%(10 kBq m�2). The pre-Fukushima 137Cs in the studied forest was2.6 kBq m�2, which is in close agreement with the reported values,which range from 2 to 5 Bq m�2 (Sakaguchi et al., 2010).

The inventory peaks of 134Cs(representing Fukushima-derived137Cs) and total 137Cs were located in the Of- horizon and then

power plant accident (% of the total soil inventory).

Samplingyear

Retained 137Cs in litter horizon(%)

Soil type

1963 64 Slighly acid colluvial siltloam

April, 1989 85 Parabrown

April, 1989 80 Sandy podsolOctober,1986

56 Regosol

August, 1987 95 Soddy podzolic sandy

August, 1988 89 Soddy podzolic sandyAugust,1989 85.3 Soddy podzolic sandyAugust, 1990 80.8 Soddy podzolic sandyAugust, 1991 76.6 Soddy podzolic sandy1993 46 Dystric cambisolOctober,1991

20 Podsolic

October,1990 40 Podsol1991 21 e

1991 45.5 e

1990 35 Podzolic parabrown1996 8.8 Soddy podzolic loamy sand1996 12.5 Humic podozol gley loamy

sand1996 11.5 Soddy podzolic loamy sand

January,2012

52 Orthic brown silt loam

r layers.

Tobler et al., 1988

Tikhomirov et al., 1993

Fesenko et al., 2001

This study

Whitkamp and Frank,1964

Melin et al., 1994

- )(20 kBq m- )

(0.9 kBq m- )

10 – 2

80 kBqk0 kBqB

9 kBq

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

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e4542

both inventories sharply decreased in the lower depths, althoughsome 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 thanthat of the Of- horizon, indicating radiocesium seems hardly leaveOf-horizon. In agreement, Rafferty et al. (2000) evaluated the pre-Fukushima 137Cs migration in Pinus contorta forest under the in-fluence of a full year and demonstrated that only 1% of it migratesinto the mineral soil. Brouwer et al. (1994) also demonstrated insequential extraction experiment that on pure mineral substrateapproximately 40% of the pre-Fukushima 137Cs was removed, whilethe same procedure resulted only 8% on organic horizons.Furthermore, Sombre et al. (1994) also reported comparable trendsin Spruce and Oak forests. Similarly, several authors acknowledged

0

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

0

1

2

3

4

5

6

7

8

9

10

11

12

13

14

15

0 200 400 600 800

Dep

th(c

m)

137Cs Concentration (Bq kg-1)

C(0) = 613α = 0.6h0 = 11.1r2 = 0.96

Fig. 4. Depth distribution of the Fukushima-derived radiocesium activity concentration belblack solid line shows the measured radiocesium, the gray solid line shows pre-Fukushima

that the persistence of radiocesium in forest floor and pointed outthis may be important, especially in thick humus layers which thenacts as a reservoir for plant uptake (Thiry et al., 2000; Goor et al.,2007) and retards the vertical migration in the soil profile ofconiferous forest ecosystems (Fawaris and Johanson, 1994; Thiryand Myttenaere, 1993). From a physicochemical point of view,despite its high sorbing capacity, organic matter fraction of soil isnot a source of irreversible radiocesium adsorption (Valcke andCremers, 1994). Radiocesium selective adsorption is rather underthe dependence of highly fixing sites born by weatheredmicaceousclay particles (Maes et al., 1998). Radiocesium-fixing clay mineralsare more or less diluted in forest floor according to local bio-turbation and were shown to highly influence its mobility andbioavailability in the upper organic layers of forest soils (Kruyts andDelvaux, 2002). In thick and acid humus in particular, characterizedby a high percentage of organic matter (>80%) resulting from a lowbiological activity, the mobility of radiocesium is in general mainlygoverned by the large pool of reversible adsorption sites. In deeperlayers of the forest floor, at the transition with mineral layers, alower dilution of radiocesium-fixing clay minerals may howeverpromote the radiocesium retention. Therefore, the observed highproportion of radiocesium in the Of- horizon indicates that radio-cesium barely leaves the organic horizons. This tendency mostlikely determines the subsequent downward movement andbioavailability of radiocesium to the plant roots that explore thissoil section.

3.1.4. Distribution of the pre-Fukushima 137Cs in the soil profileThe pre-Fukushima 137Cs activities were obtained by deducting

the Fukushima-derived 137Cs estimated from measurements of134Cs (average ratio of 134Cs/137Cs equal to 1). Fig. 4a illustrates thedepth profile distribution of the total and pre-Fukushima 137Cs. Thepre-Fukushima 137Cs tended to dominate in the deeper soil layers.Its depth distribution exhibited a general decreasing pattern withan irregular profile shape and two disincentive peaks. The peaks

0 200 400 600 800

134Cs Concentration (Bq kg-1)

C(0) = 628α = 0.73

h0 = 8.73r2 = 0.97

ow the Of- horizon of the forest soil; (a) 137Cs and (b) 134Cs concentrations profile. Theradiocesium and the broken line indicates the results fitted by Eq. (1).

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45 43

appeared at a depth of approximately 0.75 cm and 1.75 cm with a137Cs concentration of 87 Bq kg�1 and 57 Bq kg�1, respectively. Suchheterogeneous profile could be the result of many complex pro-cesses including the redistribution of radionuclides by percolatingwater and bioturbation processes (Fujiyoshi and Sawmura, 2004).Although it is difficult to clearly distinguish the Chernobyl 137Csfrom the atomic bomb 137Cs, the two distinct peaks generallyrepresent the migration distance of pre-Fukushima 137Cs falloutevents. Considering the displacement of the highest pre-Fukushima-derived 137Cs peak (at depth of 0.75 cm) in the soiland the time elapsed since its peak deposition on the earth surface(1963/1964), the downward velocity of per-Fukushima 137Cs in thestudied soil can be estimated at approximately 0.15 mm y�1. Thismigration rate is far lower than that of previous studies. Forexample, Schimmack et al. (1989) reported the migration of bomb-derived 137Cs ranging from 0.7 to 10 mm y�1 while Dorr andMunnich (1991) reported as 1.8 mm y�1. Despite the likely differ-ences in the study sites, the pre-Fukushima 137Cs peak in our studysite has been expected to be relatively deeper but did not show asignificant change. This can be explained by efficient progressivefixation of radiocesium by clay minerals even if present in organic-rich soils (Konopleva et al., 2009) over time that retards themigration and increases its residence half time (Bunzl et al., 1995).

More interestingly, almost all of the pre-Fukushima 137Cs wascontained in a soil layer in which the OM content constituted morethan 10% (Table 1 and Fig. 4a). Indeed, several studies havedemonstrated that forest soils have a strong tendency to retainradiocesium and that the majority of activities are distributed inorganic-rich surface layers. For example, Schimmack and Bunzl(1992) reported that only 15e20% of the radiocesium activity inspruce and pine forest is found in the upper 10 cm mineral soil,while most of the activity remains in the organic layer. This impliesthat the long term downwardmigration of radiocesiummay largelydepends on the processes acting in OM-rich soil layers, such asmicrobial activity and OM turnover. In forest soils, that involves aslow 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 radi-ocesium is available in mobile form in the uppermost soil offorested hill that the radiocesium-rich surface materials includingorgano-mineral particles can be easily transported to the sur-rounding environment by runoff, which could be the potentialcontaminant for downstream agricultural soils and waterresources.

3.1.5. The distribution of Fukushima-derived 137Cs in the soil profileThe activities of 134Cs and total 137Cs in the uppermost soil layer

below the Of- horizonwere 389 ± 15 Bq kgd1and 422 ± 15 Bq kg�1,respectively, and the concentrations decreased with increasingdepth (Table 1 and Fig. 4a). This distribution pattern typically re-flects the nature of radiocesium adsorption along the depth inwhich the surface organic material filters and retains the majorityof the radiocesium. The remaining radiocesium gradually attachesto soil particles along the profile as the infiltration advances thateventually create such a profile shape.

In fact, small quantities of Fukushima-derived 137Cs wereobserved in the 14e16 cm layer (Table 1). This is unexpectedconsidering the common behavior of radiocesium migration alongthe soil profile and the time elapsed since the accident, particularlyin the context of the FDNPP accident, where the fallout occurredduring an ecological phase of rest and in limited biogeochemicalprocesses. Schimmack et al. (1989) showed the unchanged depthdistribution of Chernobyl's radionuclide fallout between twoconsecutive soil sampling periods with different precipitation rate,demonstrating the fraction of radionuclides observed in deeper

layers is essentially the result of rain shower during the falloutperiod. Therefore, the most likely reason for the observed fractionat the specified soil depth in our study is the infiltration ofradiocesium-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 andduration of precipitation, the soil structure, radiocesium composi-tion (soluble vs insoluble forms) and the pre-existing soil moisture.Nevertheless, this stage cannot be used to describe the long-termradionuclide migration because the initial migration lasts for onlya short period of time after fallout.

3.1.6. Characteristics of the vertical distribution of radiocesium inthe soil profile

The vertical distribution behavior of radiocesiumwas examinedbased on the estimated coefficient of the parameters in Eq. (1) andEq. (2) for soil below the Of- horizon. The value of p in Eq. (1) isoften used as 1 in most studies. However, Isaksson and Erlandsson(1995) have indicated that the value of p can also be 0.75 for alichen carpet, 2.00 in the case of purely diffusional transport, oreven lower for active transport processes. As we dealt with theforest 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 relaxationlength (a) of 134Cs that presents Fukushima-derived 137Cs was0.7 cmd1 (a 1.4 cm relaxation length) (Fig. 4a and b). Consideringthe total 137Cs activity in the soil profile, the value of awas closer to0.6 cmd1 (a 1.7 cm relaxation length). Obviously, this difference canbe attributed to the presence of pre-Fukushima 137Cs. The relaxa-tion mass depths in the study soil profile (h0, kg m�2) were deter-mined using Eq. (2) and were 8.7 and 11.1 kg m�2 for 134Cs and thetotal 137Cs, respectively. The observed small difference in bothestimated parameters between the total and Fukushima-derived137Cs might indirectly indicate that the migration of Fukushima-derived 137Cs is somewhat rapid.

A direct comparison of our results of radiocesium migrationwith those of previous studies seems difficult because the previ-ously reported values are either from agricultural fields or are fromstudies that were conducted several years after the Chernobyl ac-cident. However, the general migration trend can be evaluated. Forexample, Poreba et al. (2003) observed a relaxation length of 2.1 cm(a ¼ 0.481 cmd1). More recently, Kato et al. (2012) found relativelyhigher values of h0 (9.1 kgm�2) and a (1.2 cmd1) for total 137Cs froman untilled home garden. Koarashi et al. (2012) also investigated thedistribution of Fukushima-derived 137Cs in different land uses andreported a relaxation length (1.43e2.9 cm) and relaxation massdepth (7.4e10.9 kgm�2) for forest land uses. Given the difference insampling techniques mentioned earlier, our results are almostconsistent with those of Koarashi et al. (2012). However, theobserved slight difference in the profile distribution parameters canbe attributed to the differences in the sampling date and tech-niques, 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 ex-pected depth displacement of the maximum concentration by afactor of 1/e, the migration rate (V) for young fallout is notnoticeable. The values of D and V in this study were approximatedbased 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 wason the soil surface (Fig. 5b). As indicated, the calculated values of Dand Vwere 1.5 cm2 y�1 and ~0 cm y�1 for Fukushima-derived 137Cs,and 0.24 cm2 y�1 and 0.03 cm y�1 for pre-Fukushima 137Cs,

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

Source of 137Cs Wz Nz V D Reference

Mass depth (kg m�2) Depth (cm) Mass depth (kg m�2) Depth (cm) (kg m�2 y�1) (cm y�1) (kg2 m�4 y�1) (cm2 y�1)

Pre-Fukushima e e e 3.6 0.2e1 e 20e50 e Walling et al. (2002)Pre-Fukushima e e e e e 0e0.35 e 0.06e2.63 Almgren and Isaksson (2006)Pre-Fukushima e e e e e 0e0.52 e 0e2.7 Schimmack and Marquez (2006)Pre-Fukushima 5.7 0.75 26.1 4.25 0.2 0.03 8 0.24 This studyFukushima 3.8 0 12.8 2 0 0 4.5 1.5 This study

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e4544

respectively (Table 3). Specifically, the value of D reflects the lumpsum effects of at least three major processes: molecular diffusion,hydrodynamic dispersion and physical mixing and its valuegenerally reduces over time (Schimmack and Marquez, 2006). Instrong agreement, the value of D for Fukushima-derived 137Cs wasgreater than that of the pre-Fukushima 137Cs, suggesting that amore rapid diffusion-like downward transportation tends todominate in our study site. Rosen et al. (1999) have also reportedsimilar trends for the V-value of 0.5e1.0 cmy�1 for the first year and0.2e0.6 cm y�1 thereafter. Moreover, regardless of the influentialset factors that emerge from climatic and site differences, our re-sults are also consistent with the range of values reported byAlmgren and Isaksson (2006) (Table 3) and most of the referencestherein. These results imply that the general trends of radiocesiumdispersion are similar in that it is rapid during the first yearsfollowing the emission and slows down thereafter.

Combining all of the parameters that were used to describe thedepth distribution behavior, it is possible to provide an insight intothe vertical migration of radiocesium in the forest environment.However, the radiocesium trapped in the canopy during the falloutwill eventually reach the forest floor via, for example, litter fall(Teramage et al., 2014). Therefore, the concentration of radiocesiumin the forest floor and soil are expected to increase and affect itssubsequent distribution. Schimmack and Marquez (2006) haveconcluded from a time-series of observations that the convection

Fig. 5. Comparative position of the input parameters used to determine the Diffusion (D)distribution.

and dispersion coefficients determined from the profiles during thefirst years after a nuclear accident could mislead the long-termradiation predictions, as these coefficients depend on time anddepth. These difficulties are clearly expected to be complicated in aforest ecosystem, as more complex factors are involved in thisecosystem than in that of grassland soils. Therefore, continuousmonitoring is required, and precise prediction models should bedeveloped for a clear and better understanding of the time-dependent migration of radiocesium in forested environments.

4. Conclusions

On the forest floor, the understory plants and the upper fewcentimeters of soil can be considered as an active radiocesiumremobilization interphase, primarily due to its acidic nature andlow clay content that made radiocesium bioavailable particularlyfor short-lived understory plants. Almost all of the radiocesiumactivity (99% of the total soil inventory) was found in the upper~10 cm in which the OM content was greater than 10%. The raworganic layer (Olþ Of) holds 52% of the Fukushima-derived 137Cs atthe time of soil sampling, and the remaining 137Cs is distributed inthe soil below the Of- layer. Specifically, the Of- horizon seems toaccumulate Fukushima-derived 137Cs, which accounts for 47% ofthe total inventory at the time of sampling, and retards its subse-quent migration, indicating that the OM turnover characterizing

and Migration rate (V) coefficients for the pre- and Fukushima-derived radiocesium

M.T. Teramage et al. / Journal of Environmental Radioactivity 137 (2014) 37e45 45

the Of-layer dynamics and its periodical changes will likely deter-mine the migration, the residence time and the bioavailability ofradiocesium. Most downward migration models consider a singlephase of migration rate that assume a gradual and slow adsorp-tionedesorption processes of radiocesium movement in the soilprofile. Nevertheless, we observed some fraction of Fukushima-derived 137Cs at a depth 16 cm which most likely due to the infil-tration of radiocesium-circumscribed rainwater during the falloutbefore selective adsorption started. This implies that in forest soilthere is additional and quick phase of radiocesium migration. Infact this phase cannot represent the long-term migration ofFukushima-derived 137Cs but still it is worth important to under-stand the speed of contamination at the initial fallout period.

Acknowledgments

This study is part of the Core Research for Evolutional Scienceand Technology (CREST) research project “Development of Inno-vative Technologies for Increasing in Watershed Runoff andImproving River Environment by the Management Practice ofDevastated Forest Plantation”. This work was also partially sup-ported by Grant-in-Aid for Scientific Research on Innovative AreasGrant 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 nuclearweapons fallout and Chernobyl accident. J. Environ. Radioact. 91, 90e102.

Bunzl, K., Kracke, W., Schimmack, W., Zelles, L., 1998. Forms of fallout137Cs and239þ240Pu in successive horizons of forest soil. J. Environ. Radioact. 39 (1), 58e68.

Bunzl, K., Schimmack, W., Kreutzer, K., Schierl, R., 1989. Interception and retentionof 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 inthe 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 parametervalues in temperate forest ecosystems: a review. J. Environ. Radioact. 100,757e766.

de Brouwer, S., Thiry, Y., Myttenaere, C., 1994. Availability and fixation of radio-cesium in forest brown acid soil. Sci. Total Environ. 143, 183e191.

Delvaux, B., Kruyts, N., Cremers, A., 2000. Rhizospheric mobilization of radiocesiumin 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 incentral 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 byforest trees following the Chernobyl accident. Radiat. Environ. Biophys. 40,105e113.

Fujiyoshi, R., Sawamura, S., 2004. Mesoscale variability of vertical profiles of envi-ronmental radionuclides (40K, 226Ra, 210Pb and 137Cs) in temperate forest soils inGermany. Sci. Total Environ. 320, 177e188.

Gomi, T., Sidle, R.C., Richardson, J.S., 2002. Understanding processes and down-stream linkages of head water systems. J. Biosci. 52 (10), 905e916.

Goor, F., Thiry, Y., Delvaux, B., 2007. Radiocaesium accumulation in stemwood: in-tegrated approach at the scale of forest stands for contaminated Scots pine inBelarus. J. Environ. Manag. 85, 129e136.

He, Q., Walling, D.E., 1997. The distribution of fallout 137Cs and 210Pb in undisturbedand cultivated soils. Appl. Radioact. Isot. 48, 677e690.

IAEA, 2010. Handbook of parameter values for the prediction of radionuclide transferin terrestrial and freshwater environments. Tech. Reports Ser. 472, 1e194.

Isaksson, M., Erlandsson, B., 1995. Experimental determination of the vertical andhorizontal distribution of 137Cs in the ground. J. Environ. Radioact. 27 (2),141e160.

IUSS Working group WRB, 1998. World Reference Base for Soil Resources. WorldSoil Resources Reports No. 84, FAO, Rome.

Karadeniz, O., Yaprak, G., 2008. Vertical distribution and gamma dose rates of 40K,232Th, 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 insoil 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 soilunder different land-use conditions. Sci. Total Environ. 431, 392e401.

Konopleva, I., Klemt, E., Konoplev, A., Zibold, G., 2009. Migration and bioavailabilityof 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 radio-cesium 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 and137Cs 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: samplingmethod. In: Zapata, F. (Ed.), Handbook for the Assessment of Soil Erosion andSedimentation Using Environmental Radionuclides. Kluwer Academic Pub-lishers, The Netherlands, pp. 41e57.

Maes, E., Delvaux, B., Thiry, Y., 1998. Fixation of radiocaesium in an acid brownforest soil. Eur. J. Soil. Sci. 49, 133e140.

Melin, J., Wallberg, L., Suomela, J., 1994. Distribution and retention of cesium andstrontium 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 (SoilConcentration Map) Based on the Prepared Distribution Map of RadiationDoses, 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 fieldand modeling studies on effects of forest devastation on flooding and envi-ronmental issues. Hydrol. Process. 24, 527e534.

Poreba, G., Bluszcz, A., Snieszko, Z., 2003. Concentration and vertical distribution of137Cs 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 mea-surements 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 137Csmigration in coniferous forest soils. J. Environ. Radioact. 48, 131e143.

Raitio, H., Rrantavaara, A., 1994. Airborn radiocesium in Scots pine and Norwayspruce needles. Sci. Total Environ. 157, 171e180.

Rosen, K., Oborn, I., Lonsjo, H., 1999. Migration of radiocesium in Swedish soilprofiles 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 falloutdeposition from theHiroshimaAtomic Bomb. Sci. Total Environ. 408, 5392e5398.

Schimmack, W., Bunzl, K., 1992. Migration of radiocesium in two forest soils asobtained from field and column investigations. Sci. Total Environ. 116, 93e107.

Schimmack, W., Bunzl, K., Zelles, L., 1989. Initial rates of migration of radionuclidesfrom the Chernobyl fallout in undisturbed soils. Geoderma 44, 211e218.

Schimmack, W., Marquez, F.F., 2006. Migration of fallout radiocesium in grasslandsoil from 1986 to 2001; Part II: evaluation of activity-depth profiles by transportmodels. 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 for-ests. Sci. Total Environ. 157, 59e71.

Strandberg, M., 1994. Radiocesium in a Danish pine forest ecosystem. Sci. TotalEnviron. 157, 125e132.

Strebl, F., Gerzabek, M.H., Karg, V., Tataruch, F., 1996. 137Cs-migration in soil and itstransfer 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 totalevapotranspiration 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 trans-ferring Fukushima-derived radiocesium to a coniferous forest floor. Sci. TotalEnviron. 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 soilerosion in hillslope forested environment. Geoderma 192, 59e67.

Thiry, Y., Kruytz, N., Delvaux, B., 2000. Respective horizon contributions to Cesium-137 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 inbirch forest ecosystem: influence of precipitation chemistry. J. Environ. Radio-act. 110, 69e77.

Tikhomirov, F.A., Shcheglov, A.I., Sidorov, V.P., 1993. Forests and forestry: radiationprotection 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 Chernobylfallout on Norway spruce and forest soil and its incorporation in to sprucetwigs. J. Radioact. 6, 225e245.

Valcke, E., Cremers, A., 1994. Sorption desorption dynamics of radiocaesium inorganic 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.), Hand-book for the Assessment of Soil Erosion and Sedimentation Using EnvironmentalRadionuclides. Kluwer Academic Publishers, Dordrecht, Netherlands, pp.111e164.

Witkamp, M., Frank, M.L., 1964. First year movement, distribution and availability of137Cs 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 the137Cs soil-plant transfer in a spruce forest of Southern Germany. J. Environ.Radioact. 100, 489e496.