Processes affecting land-surface dynamics of 129I impacted by atmospheric 129I releases from a spent nuclear fuel reprocessing plant

Science of the Total Environment 704 (2020) 135319

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Processes affecting land-surface dynamics of 129I impacted by atmospheric 129I releases from a spent nuclear fuel reprocessing plant Masakazu Ota a,⇑, Hiroaki Terada a, Hidenao Hasegawa b, Hideki Kakiuchi b a b

Research Group for Environmental Science, Japan Atomic Energy Agency, 2-4 Shirakata, Tokai, Ibaraki 319-1195, Japan Department of Radioecology, Institute for Environmental Sciences, 1-7 Ienomae, Obuchi, Rokkasho, Kamikita, Aomori 039-3212, Japan

h i g h l i g h t s

g r a p h i c a l a b s t r a c t


A field near a spent nuclear fuel

reprocessing plant was contaminated with 129I.
Leaf contamination was caused by foliar adsorption of 129I due to wet deposition. 129
Foliar uptake of I2 via stomata and root uptake of soil 129I had less impact. 129
For soil I, root uptake was more influential than volatilization by methylation.

a r t i c l e

i n f o

Article history: Received 20 August 2019 Received in revised form 28 October 2019 Accepted 30 October 2019 Available online 20 November 2019 Editor: Mae Sexauer Gustin Keywords: Radioiodine Foliar adsorption Foliar uptake Root uptake Methylation

a b s t r a c t Terrestrial environments impacted by atmospheric releases of 129I from nuclear plants become contaminated with 129I; however, the relative importance of each land-surface 129I-transfer pathway in the process of the contamination is not well understood. In this study, transfers of 129I in an atmospherevegetation-soil system are modeled and incorporated into an existing land-surface model (SOLVEG-II). The model was also applied to the observed transfer of 129I at a vegetated field impacted by atmospheric releases of 129I (as gaseous I2 and CH3I) from the Rokkasho reprocessing plant, Japan, during 2007. Results from the model calculation and inter-comparison of the results with the measured environmental samples provide insights into the relative importance of each 129I-transfer pathway in the processes of 129I contamination of leaves and soil. The model calculation revealed that contamination of leaves of wild bamboo grasses was mostly caused by foliar adsorption of inorganic 129I (81%) following wet deposition of 129I. In contrast, accumulation of 129I in the leaf due to foliar uptake of atmospheric 129I2 (2%) was lesser. Root uptake of soil 129I was low, accounted for 17% of the 129I of the leaf. The low root-uptake of 129I in spite of the 129I contained in the soil was ascribed to the fact that the most fraction (over 90%) of the soil 129 I existed in ‘‘soil-fixed” (not plant-available) form. Regarding the 129I-transfer to the soil, wet deposition of 129I was ten-fold more effective than dry deposition of atmospheric 129I2; however, the deposition of 129I during the year represented only 2% of the model-assumed 129I that pre-existed in the soil; indicating the importance of long-term accumulation of 129I in terrestrial environments. The model

⇑ Corresponding author. E-mail address: [email protected] (M. Ota). https://doi.org/10.1016/j.scitotenv.2019.135319 0048-9697/Ó 2019 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).

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calculation also revealed that root uptake of inorganic 129I can be more influential than volatilization by methylation in exportation of 129I from soil. Ó 2019 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http:// creativecommons.org/licenses/by/4.0/).

1. Introduction Radiation dose from the long-lived radioisotope of iodine (129I, with the half-life of 1.6  107 years) that is released from spent nuclear fuel reprocessing plants constitutes a significant portion of the public dose due to nuclear facilities (Müller and Pröhl, 1993; UNSCEAR, 2000; Shinohara, 2004). Terrestrial environments impacted by atmospheric releases of 129I become contaminated with 129I (Hauschild and Aumann, 1985; Robens et al., 1989), and 129 I transferred to agricultural food products and forage crops may cause radiation exposures to human consumers and grazing animals (Marter, 1963; Soldat, 1976; Kirchner 1994). Radiological impacts of 129I releases would last longer, because of its slow remobilization (followed by loss by volatilization) from soils (Kocher, 1981; Sheppard et al., 1994) —long-lasting high levels of 129 I were observed in soils decades later in those areas affected by the Chernobyl’s fallout (Mironov et al., 2002; Hou et al., 2003; Michel et al., 2005). Therefore, dynamics of 129I, in particular, transfer to vegetation and turnover in soil, in the impacted areas must be well understood to assess the dose due to 129I releases from reprocessing plants (Muramatsu et al., 1983, 1993; Robens et al., 1989; Roulier et al., 2019). Terrestrial environments affected by atmospheric 129I become contaminated with 129I via various 129I transfer pathways. During operations of a reprocessing plant, gaseous 129I is released to the atmosphere — note; particulate 129I is removed from the off-gas by filtering treatments (Sasaki et al., 2000; Mikami and Koarashi, 2003; Shinohara, 2004). Gaseous 129I contains inorganic (mostly elemental iodine, 129I2) and organic (primarily methyl-iodide, CH129 forms of iodine (Nakamura and Ohmomo, 1980; 3 I) Wershofen and Aumann, 1989; Moran et al., 1999). Notably, inorganic 129I mixes well in terrestrial environments because of its high solubility in water (Hunter-Smith et al., 1983; Whitehead, 1984; Moore et al., 1995). Atmospheric 129I2 deposits dissolves to freewater inside leaves (leaf cellular water) via stomatal openings (foliar uptake) (Chamberlain, 1959; Chamberlain and Chadwick, 1966; Clark and Smith, 1988). When leaves are wet, atmospheric 129 I2 may also dissolve in leaf surface water (liquid water retained on the leaf) (Allen and Neff, 1975; Garland and Cox, 1984), and the water-dissolved inorganic 129I is adsorbed into cuticle waxes covering the leaf surface (foliar adsorption) (Shaw et al., 2007; Hurtevent et al., 2013). Inorganic 129I contained in rainwater is also directly intercepted by aerial part of vegetation (Soldat, 1976; Moran et al., 1999; Pröhl, 2009). Furthermore, 129I deposited to soil can be transferred to plants by root-water uptake (Robens et al., 1989; Rao and Fehn, 1999; Kashparov et al., 2005; Weng et al., 2009; Humphrey et al., 2019). Although CH129 3 I is less bioavailable, i.e., its water-solubility is low, a certain fraction of CH129 3 I in the atmosphere and soil may be transferred to vegetation via foliar uptake (Nakamura and Ohmomo, 1980) and root uptake (Muramatsu et al., 1995). Assessing contributions of these 129Itransfer pathways to the contamination of vegetation is necessary to accurately estimate ingestion doses of 129I (Tikhomirov and Ryzhova, 1981; Whitehead, 1984; Roulier et al., 2018, 2019); however, since these 129I-transfers co-occur, it is difficult to ascertain the impact of the individual 129I-transfer process from results of field observations (Thiessen et al., 1999). Model simulation is a practical approach to investigate the role and importance of each transfer-pathway in land-surface dynamics

of deposited radionuclides (Boone et al., 1985; Hinton, 1994; Ota et al., 2012, 2016a, 2017; Ota and Tanaka, 2019; Schell et al., 1996). Regarding 129I, individual processes affecting its transfer to vegetation has been studied experimentally; for example, interception of rainwater-dissolved 129I by leaves (Chamberlain, 1970; Hoffman et al., 1992; Kinnersley and Scott, 2001; Pröhl, 2009), foliar adsorption of inorganic 129I (Shaw et al., 2007; Hurtevent et al., 2013; Humphrey et al., 2019), foliar uptake of 129I2 (Barry and Chamberlain, 1963; Chamberlain and Chadwick, 1966; Adams and Voillequé, 1971), and root uptake of inorganic 129I (Whitehead, 1973; Muramatsu et al., 1983; Sheppard and Evenden, 1988; Amiro and Johnston, 1989; Sheppard et al., 1993; Ashworth et al., 2003; Hong et al., 2008; Humphrey et al., 2019) and CH129 3 I (Nakamura and Ohmomo, 1980). Soil-related processes that could affect the 129I transfer to vegetation has been studied. These include, adsorption and desorption of 129I (Yoshida et al., 1995; Fukui et al., 1996; Dai et al., 2004, 2009; Shetaya et al., 2012), downward transport of water-dissolved 129I (Boone et al., 1985; Robens et al., 1989; Ashworth et al., 2003; Kashparov et al., 2005; Yuita et al., 2005; Weng et al., 2009), and volatilization of 129I due to methylation (Sheppard et al., 1994; Muramatsu and Yoshida, 1995; Amachi et al., 2003). Based on the knowledge obtained from these process-based researches, models to estimate land-surface dynamics of 129I have been developed (USNRC, 1977; Whicker and Kirchner, 1987; Davis et al., 1993; Müller and Pröhl, 1993; Abbott and Rood, 1994; Schell et al., 1996; Katagiri et al., 1997; Weng et al., 2009). The model by USNRC (1977) calculates transfers of 129I from the atmosphere and soil to vegetation by using deposition velocity and soil-to-plant transfer factor. The models named ECOSYS-87 (Müller and Pröhl, 1993) and FORESTPATH (Schell et al., 1996) include interception of wet-deposited 129 I by aerial part of vegetation. Furthermore, the BIOTRAC model (Davis et al., 1993) considers loss of soil 129I by methylation. However, there is yet no model that evaluates behavior of 129I under field conditions, where the transport of 129I is influenced by various influential hydrological and physiological processes (Thiessen et al., 1999). For instance, rain-interception and foliar adsorption of inorganic 129I affected by water budget on leaves (Garland and Cox, 1984; Hoffman et al., 1992; Kinnersley and Scott, 2001), foliar uptake of 129I2 and CH129 3 I controlled by the opening and closure of stomata (Barry and Chamberlain, 1963; Chamberlain and Chadwick, 1966; Adams and Voillequé, 1971; Nakamura and Ohmomo, 1980), root uptake of 129I driven by leaf transpiration (Amiro and Johnston, 1989; Sheppard et al., 1993; Weng et al., 2009), and redistribution of inorganic 129I in soils caused by soil– water flow (Boone et al., 1985; Robens et al., 1989; Bundt et al., 2000; Ashworth et al., 2003; Weng et al., 2009). An application of such enhanced models to an actually-observed environmental transfer of 129I will clarify the process(s) and its/their importance for land-surface dynamics of deposited 129I. This study aims to clarify the role of each 129I-transfer process in the contamination of vegetation and soil that are impacted by atmospheric 129I-releases from spent nuclear fuel reprocessing plant. To achieve this, a 129I-model that integrates all the processes mentioned earlier is herein developed and incorporated into an existing land-surface model (SOLVEG-II). This modeling approach links hydrological and physiological processes to 129I transfers, allowing quantification of the contribution of each transfer pathway to the dynamics of 129I under actual meteorological condi-

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tions. Furthermore, the model was applied to the observed transfer of 129I in a vegetated field impacted by atmospheric releases of 129I from the Rokkasho reprocessing plant (RRP) of Japan Nuclear Fuel Ltd. (JNFL). Based on the results of the model calculation and comparison with measured environmental samples, the performance of the developed model is evaluated, and the influence of each 129 I-transfer process is discussed.

2. Model In this section, a land-surface model, SOLVEG-II, was used to develop a 129I model in order to evaluate 129I-transfers under field conditions. SOLVEG-II is a one-dimensional model that computes the transportation of heat, water, CO2, and other trace gases in a multi-layered atmosphere-vegetation-soil system (Yamazawa, 2001; Nagai, 2005; Katata et al., 2011). SOLVEG-II is run with meteorological data as input to the upper atmospheric boundary, below which a vegetation canopy and a soil exist. In the atmospheric model, transports of heat, water (vapor), and CO2 by turbulent diffusion are evaluated (Nagai, 2005). Leaf photosynthesis and transpiration are evaluated by the vegetation-canopy model (Nagai, 2005) using Farquhar’s photochemistry (Farquhar et al., 1980) and Ball-Berry and Leuning’s stomata-photosynthesis interdependence (Collatz et al., 1991). Liquid water interaction with leaves was evaluated by accounting for properties of canopy structure (leaf area density, leaf angle, maximum water storage on leaf, etc.) and vertical distribution of throughfall intensity inside the layered canopy (Nagai, 2005). The soil model calculates temperature, water content, and flux throughout the soil profile by considering heat conduction and Richard-type water transport (Yamazawa, 2001). Modeled processes for the transport of heat, water, CO2, and other trace gases have been proposed for several land-use types such as grassland and forest (Katata et al., 2011; Nagai, 2005; Ota et al., 2012, 2013, 2016a, 2016b, 2017; Ota and Tanaka, 2019). The 129I model (see, Fig. 1) is developed to evaluate land-surface transfer of 129I influenced by wet deposition of 129I by rain and dry deposition of gaseous 129I that is the likely form of 129I to be released to the atmosphere from reprocessing plants (see,

Fig. 1. Transport processes of inorganic

129

Section 1); i.e., the developed model ignores particulate 129I. The I model consists of two submodels, one of which evaluates the dynamics of inorganic 129I (see, Appendix A) and the other evaluates the dynamics of organic 129I, as CH129 3 I, (see, Appendix B). The two submodels are simultaneously run with the input of concentrations of inorganic 129I and CH129 3 I in the atmosphere and the rainwater at the upper atmospheric boundary, and linked by below-ground methylation process (see, Fig. 1). In the inorganic 129I model, 129I2 (gas) in the atmosphere and inorganic 129I dissolved in plants (leaf) and soil water—representing iodide (129I) and iodate (129IO 3 )—are considered (see, Fig. 1). The transport of 129I2 by turbulent diffusion is calculated in the atmospheric model, while the transfer of 129I to leaf is modeled by interception of rainwater-dissolved inorganic 129I by the leaf, dissolution of atmospheric 129I2 to leaf-surface water (liquid water retained on plant leaf), foliar adsorption of 129I, foliar uptake of 129 I2 via stomata, and soil-to-leaf transfer of inorganic 129I by root uptake. The input of 129I to the soil is considered by infiltration of rainwater-dissolved 129I and dissolution of atmospheric 129I2 into the soil water via the ground surface. Further, the inorganic 129I in each soil-layer is distributed to the following three pools: 129I dissolved in soil water, 129I adsorbed to the soil, and 129I fixed in the soil. The dissolved 129I is mobile in the soil by vertical waterflow, and available for plants through root-water uptake; whereas the soil-adsorbed 129I is reversely exchangeable with the 129I in soil water, where the partition of 129I between the two pools is determined by distribution coefficient (K d ) that depends on organic matter content of the soil. The soil-fixed 129I is the nonexchangeable fraction that is strongly bound to soil constituents owing to strong physicochemical interaction with the soil organic matter or self-diffusion of 129I into the inner structure of the soil constituents, and thus it is neither mobile in the soil nor available for plants. 129 The CH129 3 I model (see, Fig. 1) considers gaseous CH3 I in the 129 atmosphere and the soil, and water-dissolved CH3 I in the rain, soil, and leaves. The transport of CH129 3 I by turbulent diffusion is calculated in the atmosphere, while its transfer to the leaf is modeled by reversible exchanges of CH129 3 I between canopy air and leaf surface water, and between canopy air and leaf cellular water via stomata, and transfer of soil CH129 3 I to the leaf cellular water by 129

I and CH129 3 I considered in the developed

129

I model.

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root uptake. The soil model calculates generation of CH129 3 I due to microbial methylation of inorganic 129I dissolved in soil water, and gaseous and aqueous transport of the generated CH129 3 I.

3. Model application 3.1. Site, period, and environmental samples In this section, we apply the 129I model to a vegetated field site impacted by atmospheric releases of 129I from the RRP. The RRP is located in Rokkasho village in Aomori Prefecture, north end of the main island of Japan. JNFL conducted test operations of the RRP with atmospheric discharges of 129I from April 2006 through April 2009. The release amount of gaseous 129I during this period was the greatest in the year 2007: 1.5  108 Bq (2006), 3.6  108 Bq

(2007), 2.4  108 Bq (2008), and 1.8  106 Bq (2009) (JNFL, 2019). Therefore, we apply the model simulation to approximate one-year period (January 3 – December 31) of the year 2007 (see, Table 1). In the year 2007, the 129I discharges occurred from February to April, and from September to December, as shown in Fig. 2a. The study site (the Obuchi site, in Table 1) is an undisturbed area forested with wild bamboo grasses, and located 2.5 km south-east from the RRP’s main stack (Aomori Prefecture, 2008; Kakiuchi et al., in preparation). The mean temperature and the annual precipitation in this region are 9 °C and 1300 mm, respectively (record from 1982 to 2010 at the nearest observatory, AMeDAS-Rokkasho of the Japan Meteorological Agency; http:// www.jma.go.jp/jp/amedas). At the Obuchi site, leaves of the bamboo grasses and surface litter-materials (O-horizon) were collected on June 12 and August 10, 2007, and top 12 cm soil (below the Ohorizon) was collected on August 10, 2007 (Aomori Prefecture,

Table 1 Settings of model simulation. Parameters or inputs

Values

References

Period Site Distribution coefficient (Kd)

January 3, 2007 – December 31, 2007 The Obuchi site (141.3515°E, 40.9471°N) Control case: 0.023–0.035a m3 kg1 Increased adsorption case: 3-fold of the control caseReduced adsorption case: 1/3-fold of the control case Bamboo grass (Sasa senanensis var. senanensis) 1m 0.1 kg m2c 0.2–0.5 m2 m3d

This study Aomori Prefecture (2008), Kakiuchi et al. (in preparation) Calculated by Eq. A.20b Uncertainty in the modeled Kd (Fig. A.1).

Plant type Thickness of canopy Initial leaf biomass Leaf area density e-folding depth of root-water uptake (zr) Initial leaf 129I concentration Initial soil 129I concentration (vss1 þ vss2 ) Hourly input of meteorological dataf Hourly input of 129I concentration in air 129 I2:CH129 3 I in air

Hourly input of 129I concentration in rain Inorganic 129I:CH129 3 I in rain Time constant of 129I adsorption to leaf surface (tads) Time constant of soil (tfix)

129

I fixation in

Time constant of 129I mobilization in soil (t mob ) Time constant of in soil (t met )

129

I methylation

0.25 m

Aomori Prefecture (2008), Kakiuchi et al. (in preparation) Shibata (1992), Yabe et al. (2012) Yokoyama and Shibata (1998) Shibata (1992), Yokoyama and Shibata (1998), Yabe et al. (2012), Fukuzawa et al. (2013) Shibata (1987)

Control case: 1.2 mBq kg1 Zero leaf-129I case: 0 Bq kg1 0–1.5 mBq kg1e (Fig. 3)

Assumed (Subsection 3.2) Assumed (Subsection 3.3) Assumed (Subsection 3.2)

Calculation by a mesoscale model MM5

Ota et al. (2016b)

Shown in Fig. 2c (observation at the IES)

Aomori Prefecture (2008), Hasegawa et al. (in preparation) Mikami and Koarashi (2003) Shinohara (2004) Wershofen and Aumann (1989) Hasegawa et al. (2017)

Control case: 50%:50% 129 I2-90% case: 90%:10% 129 I2-10% case: 10%:90% Shown in Fig. 2b (observation at the IES) 100%:0% Control case: 22 min Rapid adsorption case: 7 min Slow adsorption case: 60 min Control case: 830 daysg Rapid fixation case: 500 daysg Slow fixation case: 1400 daysg Control case: 32 yearsg Rapid mobilization case: 10 yearsg Slow mobilization case: 490 yearsg Control case: 48 years Rapid methylation case: 0.32 years Slow methylation case: 5600 years

Assumed (Subsection 3.2) Shaw et al. (2007)

Takeda et al. (2015) Kashparov et al. (2005) Assumed (Subsection 3.3) Raich and Schlesinger (1992)

Sheppard et al. (1994)

a Values are 0.035, 0.030, 0.028, 0.023, 0.025, and 0.025 m3 kg1 for the soil layers at the depths of 0–2, 2–4, 4–6, 6–8, 8–12, and 12–100 cm, respectively. In the soil below the depth of 12 cm, the same value as the depth of 8–12 cm was assumed. b Organic matter content (qOM ) was determined from the measured organic carbon content of the soil sampled at the Obuchi site during the year 2007 (Aomori Prefecture, 2008; Kakiuchi et al., in preparation); the values are 0.135, 0.110, 0.076, 0.077, and 0.086 kg-C kg1 for the soil layers at the depths of 0–2, 2–4, 4–6, 6–8, and 8–12 cm, respectively. These organic carbon contents were converted to the organic matter content using the ratio of 1.724 kg-OM kg-C1 (Pribyl, 2010). c The minima (observed in April) of the aboveground biomass of bamboo grass’s understory in a cool temperate (the same climate as the Rokkasho region) forest in Japan (Yokoyama and Shibata, 1998). d Leaf area density is constant (0.2 m2 m3) from January 3 to March 31, increases to the maximum of 0.5 m2 m3 until September 1, and then decreases to the minimum of 0.2 m2 m3 by December 31. These seasonal changes were taken from the observed results of leaf area densities of bamboo grasses growing in Japanese forests (Shibata, 1992; Yokoyama and Shibata, 1998; Yabe et al., 2012). The maximal value (0.5 m2 m3) was determined so that the SOLVEG-II’s calculation of the leaf biomass at the end of the year 2007 is approximately 0.7 kg m2, which is an observed annual production of aboveground biomass of bamboo grasses’ understory (Fukuzawa et al., 2013). e Corresponds to the 129I inventory of 0.23 Bq m2 in the one-meter depth soil (calculated using model-assumed dry bulk density of Eq. A.18). f Air pressure, air temperature, solar radiation, longwave radiation, wind speed, humidity, precipitation intensity, and CO2 concentration in the air. g Fractions of soil-adsorbed and soil-fixed 129I in the 129I contained in the soil were calculated by Eq. (1) for each case: 7% and 93% for the control case, 4% and 96% for the rapid fixation case, 11% and 89% for the slow fixation case, 19% and 81% for the rapid mobilization case, and less than 1% and greater than 99% for the slow mobilization case.

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Fig. 2. (a) Monthly atmospheric discharges of 129I and 14C during the RRP’s test operation in the year 2007 (JNFL, 2019), and hourly model-inputs of (b) precipitation intensity, concentration of inorganic 129I in rainwater, and (c) concentrations of 129I2 and CH3129I in the atmosphere. In (c), monthly observations of gaseous and particulate 129I in the air at the Institute for Environmental Sciences (IES, locating 2.2 km north-northeast of the Obuchi site) are also plotted (Aomori Prefecture, 2008; Hasegawa et al., in preparation).

2008; Kakiuchi et al., in preparation). In addition, pine needles were collected on June 12 and August 10, 2007, at the adjacent area forested with wild pine trees (Sasa senanensis var. senanensis) (Aomori Prefecture, 2008; Kakiuchi et al., in preparation). These plant and soil samples were taken to laboratory and the activity of 129I was analyzed. The detailed procedure for the determination of 129I has been described elsewhere (Ueda et al., 2018). Briefly, the dried sample was ignited under a flow of oxygen in a quartz tube furnace after addition of an iodine carrier (Woodward Iodine Co., USA). And then the iodine volatilized by the ignition was trapped in tetramethyl-ammonium hydroxide (TMAH) solution. The iodine in the solution was precipitated as AgI. The 129I/127I atom ratio in the AgI sample was determined with an accelerator mass spectrometer (AMS) at the PRIME Laboratory of Purdue University, USA. The 127I concentration in the sample was measured using an inductively coupled plasma mass spectrometer (Agilent-7700, Agilent Technologies, Inc., USA) after extraction of iodine with TMAH solution. A standard reference material (IAEA-375) with the reference value of 1.7 mBq kg1-dry weight of 129I concentration was also analyzed to maintain high accuracy of the measured results.

5

SOLVEG-II’s default parameters for loam (Yamazawa, 2001) were used in the calculations of heat and water transport in the soil. The organic matter content (qOM ) that is needed to calculate the distribution coefficient (K d ) using Eq. A.20 was determined from the measured organic carbon content of the soil sampled at the site (see, footnote (b) of Table 1). Table 1 present the calculated K d values as ‘‘control case”—this refers to the model calculation that used the model’s default parameters; however, test calculations that used tuned-parameters are described in Subsection 3.3. A bamboo grasses’ canopy was established to the lower five atmospheric layers of the model. Vegetation parameters that are required to run SOLVEG-II were obtained in the literature: thickness of the canopy (Shibata, 1992; Yabe et al., 2012), initial leafbiomass (Yokoyama and Shibata, 1998), leaf area density (Shibata, 1992; Yokoyama and Shibata, 1998; Yabe et al., 2012; Fukuzawa et al., 2013), and e-folding depth of root-water uptake (Shibata, 1987). It was assumed that the organic matter produced by photosynthesis is accumulated in a single leaf compartment. Furthermore, 129I adsorbed to the leaf surface, and those transferred to leaf cellular water were also assumed to accumulate in this compartment. Possibly the bamboo grasses’ leaf contained 129 I at the beginning of the year 2007 since 129I was released from the RRP in the previous year, 2006 (see, Subsection 3.1), and the average lifetime of bamboo grass’s leaf is one to two years (Shibata, 1992; Yokoyama and Shibata, 1998). The leaf 129I concentration at the beginning of the year 2007 was set to 1.2 mBq kg1, which is the concentration of 129I in the litter sampled at the site on 12 June 2007 (see, Fig. 3). Here, we assumed that the litter consisted of the bamboo grasses’ leaves that had fallen immediately before the year 2007, given that litter-fall of bamboo grasses occurs from autumn to early spring (Shibata, 1992). The effect of the initial setting of the leaf 129I concentration on the model calculation is analyzed in the later part of this paper (Subsection 4.3). For the soil 129I, there was no measurement immediately before the year 2007. However, we used the measurement of 129I in the soil (top 50 cm horizon) sampled at the site during the year 2003 as the soil 129I concentrations at the beginning of the year 2007, which is represented in Fig. 3 by ‘‘Soil-Obs. (2003)”. The fractions of soil-adsorbed and soil-fixed 129I, i.e., vss1 and vss2 of Eq. A.17 and Eq. A.21, respectively, in this soil were estimated. We assumed that a semi-equilibration was achieved for the dynamics of 129I in this soil; i.e., the fixation and the mobilization of 129I defined by Eq. 1 A.21 balanced, as: t 1 fix qb vss1 ¼ t mob qb vss2 . Therefore, the ratio of the

3.2. Models’ settings and inputs A 12 m-high atmosphere (including a vegetation canopy) and a 1 m-thick soil were represented in the SOLVEG-II model. The atmosphere was divided into ten layers (with boundary heights of 0.1, 0.3, 0.5, 0.7, 1, 1.5, 3, 5, 8, and 12 m above the ground) and the soil was divided into 19 layers (with boundary depths of 1, 2, 3, 4, 5, 6, 7, 8, 10, 12, 15, 20, 25, 30, 40, 50, 60, 80, and 100 cm of the soil). The soil texture was set to loam (Tsukada et al., 2008) and the

Fig. 3. Observed (Aomori Prefecture, 2008; Kakiuchi et al., in preparation) and model-calculated concentrations of 129I in the top 12 cm soil (sectioned to the analyzed depths of 0–2, 2–4, 4–6, 6–8, and 8–12 cm). Observations of 129I in the litter are also plotted on the dotted region. Error bars attached to the observed data represent counting errors of 129I analysis. In the model calculation, the observation of soil 129I during the year 2003 (‘‘Soil-Obs. (2003)”) was used as the initial value of the soil 129I concentration (see, Subsection 3.2).

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concentration of soil-fixed (vss2 ) to soil-adsorbed obtained as:

vss2 =vss1 ¼ tmob =tfix ¼ 14:1

I (vss1 ) was

129

ð1Þ

where t mob = 32 years and tfix = 830 days (see, Table 1). The model calculation was run with the time increment of 5 s, while the input meteorological data were prepared from a calculation of a mesoscale model, MM5 (Grell et al., 1994). The MM5 was applied to the Rokkasho region for the year 2007 (Ota et al., 2016b). Hourly-interval calculation results obtained at the location of the Obuchi site were used as the SOLVEG-II’s input meteorological data. To precisely simulate the water input to the Obuchi site, and the resulting wet deposition of 129I, the precipitation intensity of the input data were modified as follows. A factor was multiplied to the MM5-calculated hourly precipitation intensity in each month so that the modified precipitation (shown in Fig. 2b) agrees with the monthly precipitation observed at the Institute for Environmental Sciences (IES) (Hasegawa et al., 2017), which is located 2.2 km north-northeast of the Obuchi site. Measurements of 129I in the atmosphere and rainwater were not performed at the Obuchi site. Therefore, the model’s inputs (hourly interval) of concentrations of 129I in the atmosphere (Fig. 2c) and the rain (Fig. 2b) were prepared from measurement obtained at the IES. The possibility of using this data as the 129I input to the Obuchi site is discussed in Subsection 5.1. At the IES, ambient air was sampled continuously during the year 2007, and the concentrations of gaseous and particulate 129I were analyzed at a one-month interval (Aomori Prefecture, 2008; Hasegawa et al., in preparation). The gaseous 129I was analyzed as the sum of the inorganic and the organic 129I; thus, the ratio of 129I2 to CH129 was unknown. However, we assumed that 129I2: 3 I 129 CH3 I = 50% : 50% (‘‘control case” of Table 1), provided that the off-gas of a reprocessing plant (Tokai reprocessing plant, Japan) contained approximately the same activity of 129I2 and CH129 3 I (Mikami and Koarashi, 2003). Meanwhile, a greater fraction of 129 CH129 I in the 3 I (at most ~ 90%) was observed for the gaseous atmosphere near a reprocessing plant, the Karlsruhe reprocessing plant (Wershofen and Aumann, 1989); therefore, additional calculations that applied altered-ratios of 129I2 to CH129 3 I to the input data of 129I were performed (see, Subsection 3.3). At the IES, rainwater was also collected, and the concentration of 129I (all dissolved forms) was analyzed at a one-month interval (Hasegawa et al., 2017). We assumed that all the 129I contained in the rainwater were inorganic 129I (i.e., the input data of CH129 3 I concentration in the rainwater were set to 0 Bq m3) because the solubility of inorganic iodine (I2) in water is 30 to 40 times that of CH3I, as detailed by Henry’s law (Hunter-Smith et al., 1983; Whitehead, 1984; Moore et al., 1995). 3.3. Sensitivity analysis To check the model’s sensitivity to the parameters that could have uncertainties and investigate the impact of each 129I transfer process on the dynamics of 129I at the site, we performed several test calculations by varying the values of the model’s parameters or inputs in the expected range. The followings are considered as the variables: time constant (s) (i.e. an inverse of reaction rate (s1)) of foliar adsorption of 129I (t ads ); distribution coefficient of soil 129I (K d ); time constants of mobilization of soil 129I (t mob ); fixation of soil 129I (t fix ); and methylation of soil 129I (tmet ); ratio of 129 I2 to CH129 3 I in the atmospheric input term; efficiency of soilto-leaf transfer for root-absorbed 129I, and initial concentration of 129 I in the leaf. The developed model utilized the time constant t ads = 22 min for the foliar adsorption of water-dissolved inorganic 129I (see, Eq.

A.10). Meanwhile, slower adsorption of 129I to leaf surfaces has been observed: time constant of 60 min (Shaw et al., 2007). Hence, we assumed an uncertainty at a factor of three (=60 min/22 min) for the model-assumed value of t ads and performed test calculations varying t ads in this range: t ads = 60 min—a relatively slow adsorption case, or 7 min—a considerably rapid adsorption case (see, Table 1). The calculation of distribution coefficient K d using Eq. A.20 has uncertainty at a factor of three (see, Fig. A.1). Thus, the test calculations were performed by setting the K d value to three-fold (for the increased adsorption case) or one-third (for the reduced adsorption case) of the value calculated by Eq. A.20. Regarding the modeled fixation of 129I in the soil (with the time constant, tfix = 830 days, see Eq. A.21), a sharp decrease of waterexchangeable (i.e., soil-adsorbed) iodine have been observed (Sheppard and Evenden, 1988; Kashparov et al., 2005). Precisely, the time constant of 500 days was obtained (Kashparov et al., 2005). Thus, we assumed uncertainty at a factor of 1.7 (=830 day s/500 days) for this value of t fix , and performed test calculations varying tfix between (see, Table 1), t fix = 500 days (for considerably rapid fixation case) and t fix = 1400 days (for relatively slow fixation case). For the mobilization process, we assumed that microbial decomposition of soil organic matter causes the mobilization of soil-fixed 129I (see, Eq. A.21), and used a global average value—obtained in the literature—of turnover time of soil organic matter as the time constant of this process (tmob = 32 years). The turnover of soil organic matter, however, is site-specific, revealing turnover time on the order of a decade to century (Parton et al., 1987); for instance, 10 years to 490 years (Raich and Schlesinger, 1992). Therefore, the test calculations were performed assuming a relatively small (10 years for rapid mobilization case) or considerably high value (490 years for slow mobilization case) for t mob . In the calculations where the revised values of tfix or t mob is used, the initial concentrations of soil-adsorbed and soil-fixed 129I were reset following Eq. (1) (see footnote (g) of Table 1). Methylation of 129I in soils is site-specific (Muramatsu and Yoshida, 1995; Muramatsu et al., 1995; Ashworth, 2009), revealing time constants on the order of months (0.32 years) to thousand years (5600 years) (Sheppard et al., 1994). Therefore, the test calculations were performed by setting the time constant tmet (see, Eq. A.14) to a smaller (0.32 years, for rapid methylation case) or greater value (5600 years, for slow methylation case), compared with the default value of 48 years (see, Table 1). When preparing the atmospheric input term of 129I, we assumed a ratio of 129I2:CH129 3 I = 50%:50% (see, Subsection 3.2). However, the inorganic 129I may represent ~ 90% of gaseous 129I contained in offgas of a reprocessing plant (Shinohara, 2004). Meanwhile, the most likely form of gaseous 129I to be released from a reprocessing plant is the organic one (Nakamura and Ohmomo, 1980; Moran et al., 1999); for example, the organic 129I (including CH129 3 I) was found to represent ~ 90% of the gaseous 129I in the atmosphere around Karlsruhe reprocessing plant (Wershofen and Aumann, 1989). Therefore, the test calculations were performed by applying the ratio 129I2: 129 CH129 I2-90% case, or 10%:90%, 129I2-10% case, to the 3 I = 90%: 10%, input data of atmospheric 129I concentrations. The developed model assumes that all the 129I taken-up by root is immediately transferred to leaves by transpiration flow (see, Eq. A.11 and Eq. A.13). However, it has been indicated that inorganic iodine taken-up by roots are stored in the roots, and not transported to other plant parts (Whitehead, 1973; Muramatsu et al., 1983; Hong et al., 2008). To quantify the effect of this phenomenon on the calculation of leaf 129I concentration, we carried out an experiment assuming the flux of soil-to-leaf 129I transfer to be zero (EIroot = 0 Bq m2 s1) when using Eq. A.11 for the calculation, i.e., the zero root-uptake case.

M. Ota et al. / Science of the Total Environment 704 (2020) 135319

7

As mentioned earlier, the modeled bamboo grass’s leaf contained 129I at the beginning of the year 2007, which implies that the leaf had developed before the year 2007 and so contained the 129I that had accumulated in the previous year(s) (see, Subsection 3.2). However, since bamboo grasses form new leaves every year (Shibata, 1992), it is expected that the leaves formed during the year 2007 were not affected by the previous accumulation of 129 I. To quantify the effect of this phenomenon on the model calculation of leaf 129I concentration, we experimented by setting the initial concentration of 129I in the leaf to 0 Bq kg1, i.e., zero leaf-129I case (see Table 1).

4. Results 4.1. Leaf

129

I

Here, we first present the result of model calculation of 129I transfer to leaf surface, i.e., foliar adsorption, which revealed that foliar adsorption of 129I is mainly caused by wet deposition of inorganic 129I by rain, rather than dry deposition of 129I2. When it rained, a fraction of rainwater precipitating inside the bamboo grasses’ canopy was intercepted and retained on the leaves. As a result, inorganic 129I dissolved in the rainwater was trapped on the leaves (the results are not shown here). Dissolution of 129I2 from canopy air to the retained leaf-surface water co-occurred. A comparison of the monthly values of ‘‘rain interception” and ‘‘dissolution” in Fig. 4a showed that the input of 129I to the leaf surface water by the rain interception exceeds the dissolution for most months. Hence, noticeable foliar adsorption of 129I was obtained from September to December (see, Fig. 4a) when the concentrations of 129I in the rainwater were high (Fig. 2b). The model calculation revealed that the accumulation of 129I in the leaf, i.e., the leaf cellular water (see Fig. 1), was less than the

Fig. 4. Model-calculated transfers of 129I to (a) leaf surface, (b) leaf cellular water, and (c) biomass and concentration of 129I in the bamboo grasses’ leaf. In (a) and (b), the results are shown by 129I-transfer per unit area of the ground surface, and integrated values of the adsorption and accumulation, are also plotted. In (c), the observed results of leaf 129I concentration (Aomori Prefecture, 2008; Kakiuchi et al., in preparation) are plotted by triangles; error bars attached represent counting errors of 129I analysis. Note: the scale of the left axis in (a) is ten times that in (b).

Fig. 5. Model-calculated transfers of 129I (per unit area of the ground surface) at the Obuchi site during the year 2007.

adsorption of 129I on the leaf surface. It was found that soil-toleaf 129I transfer by root-water uptake occurred (see, ‘‘root uptake” in Fig. 4b) due to the 129I contained in the soil (footnote (e) of Table 1), with month-to-month variations affected by seasonal changes of leaf area density in the canopy (footnote (d) of Table 1) and leaf transpiration (thus, root-water uptake), which are evaluated by SOLVEG-II (result not shown here). Foliar uptake of atmospheric 129I2 via stomata was the highest in April (see, Fig. 4b), which was influenced by the high atmospheric concentration of 129 I2 in this month (see ‘‘input of 129I2” in Fig. 2c). A study of the 129 I (7.0  105 Bq m2, see Fig. 4b) transferred to the leaf cellular water during the year revealed that 87% were delivered by the root uptake as presented by annual transfers of 129I in Fig. 5. This 129I accumulation in the leaf was 1/4 of the 129I adsorption on the leaf (see, Fig. 4a). The evaluation of the CH129 3 I model showed that a fraction (0.4%) of CH129 3 I generated in the soil was taken-up by roots and transferred to the leaf (see, Fig. 5). Further, the soil at the site act 129 as a CH129 I initially 3 I source, due to methylation of inorganic contained in the soil (noted in Subsection 4.2). As a result, CH129 3 I was released from the soil to the canopy air and dissolution of the released CH129 3 I to the leaf surface water occurred (Fig. 5). However, these CH129 3 I that were transferred to the leaf were released to the atmosphere and did not contribute to the leaf 129I concentration. The model calculations of leaf biomass and leaf 129I concentration, i.e., the sum of the 129I in and on the leaf, were plotted on Fig. 4c. The leaf biomass was observed to increased significantly (by 4.6 times) from May to October, majorly due to the high values of leaf area density and photosynthesis (organic matter production) calculated by the SOLVEG-II in this season (result not shown here). The concentration of 129I in the leaf revealed that there were slight increases in these quantities during April and the period after September, mainly caused by the foliar adsorption of 129I (see, Fig. 4a). Furthermore, noticeable decrease (by 60%) of the 129 I concentration was recorded from May to August (see, Fig. 4c), which was due to the accumulation of organic matter in the leaf; in other words, the leaf 129I activity was diluted. However, the model underestimated the 129I concentration on June 12 by a factor of two and overestimated it on August 10 by a factor of 14.

8

4.2. Soil

M. Ota et al. / Science of the Total Environment 704 (2020) 135319 129

I

The evaluation of the soil 129I model revealed that the model calculated hourly flux of wet deposition of inorganic 129I during rainfall was found to be several orders of magnitude greater than the flux of dry deposition (dissolution) of atmospheric 129I2 to the soil (results not shown here). On a monthly basis, the wet deposition was observed to exceed the dry deposition (see, Fig. 6a). Of the 129I deposited (5.3  103 Bq m2, see Fig. 6a) during the year 2007, 91% were delivered by wet deposition, see ‘‘Infiltration” and ‘‘Dissolution” to the soil in Fig. 5. This 129I deposition in the year 2007 was 2% of the model-assumed inventory of 129I in the one-meter thick soil (see footnote (e) of Table 1). Furthermore, it was found that 129I depositions affect the concentrations of 129I on the soil surface up to several cm horizons. For instance, in the top 1 cm soil-layer (Fig. 6b), the concentration of inorganic 129I in the soil water was observed to increase during rainfalls in the month of September, October, and December due to the relatively high concentrations of 129I in the rainwater in these months (see Fig. 2b). After these rainfalls, adsorption of the deposited 129I occurred in the soil. Hence, the concentration of soiladsorbed 129I was increased after September (see Fig. 6b). In the soil below the depth of 3 cm, no observable variations in the concentration of soil-adsorbed 129I were calculated (results not shown here); because the upper soil layers retained the deposited 129I. For soil-fixed 129I, the concentrations across the entire soil-profile were unchanged (e.g., see ‘‘soil-fixed 129I” in Fig. 6b for the top 1 cm layer) because the deposition of 129I in the year 2007 was much less than the 129I that existed in the soil (mentioned in the previous paragraph) that consisted of 93% of the soil-fixed 129I (footnote (g) of Table 1). The methylation of inorganic 129I in the soil was evaluated throughout the simulation period (data not shown here); however, only 0.001% of the total content of 129I in the soil column was methylated in the year (see Fig. 5). Therefore, the methylation did not influence the concentration of inorganic 129I in the soil. Fig. 3 shows the vertical profiles of the concentration of 129I, i.e., the sum of the soil-adsorbed and soil-fixed 129I, in the top 12 cm horizon. Compared with the value assumed at the beginning of the year 2007 (‘‘Soil-Obs. (2003)”), no observable variations were calculated at the time of the soil-sampling (‘‘Soil-Cal. (Aug 10, 2007)”). However, the model-calculated 129I concentration underestimated the observation ‘‘Soil-Obs. (Aug 10, 2007)” by a factor of 2 for the surface 2 cm layer, whereas the calculation was about

Fig. 6. Model-calculated results of (a) monthly and integrated depositions of inorganic 129I to the soil, and (b) concentrations of inorganic 129I in the top 1 cm layer of the soil.

two to three times the observation in the layers at depths of 8 to 12 cm. 4.3. The sensitivity of the model calculation of leaf and soil

129

I

Fig. 7 shows a comparison of the concentrations of 129I in the bamboo grass leaf calculated for the 14 test cases (Subsection 3.3). It was found that of the selected parameters or settings, the initial value of leaf 129I concentration had the greatest impact on the model calculation. Precisely, a noticeable decrease of leaf 129I concentration—for example, by 61% of the control case for the result on June 12—was calculated for the zero leaf-129I case. Also, the time constant of foliar adsorption significantly influenced the model evaluation. For instance, the leaf 129I concentrations on June 12 obtained from the rapid and the slow adsorption cases were found to differ by a factor of 1.6. Furthermore, the variabilities of leaf 129I concentration to some extent (less than a factor of 1.4) were observed for the following four pairs: 129I2-90% and 129I210% cases; increased and reduced adsorption cases; rapid and slow fixation cases; and, rapid and slow mobilization cases. The difference of leaf 129I concentrations between the 129I2-90% case and 129 I2-10% case was caused by the altered dissolution of atmospheric 129I2 in the wet leaf surface and the subsequent foliar adsorption of 129I. For the three pairs of increased and reduced adsorption cases, rapid and slow fixation cases, and, rapid and slow mobilization cases, the differences of leaf 129I concentration were ascribed to the altered root-uptake of 129I that was caused either by the revised adsorption of 129I in the soil (for the increased and reduced adsorption cases) or the revised partition of the soil 129I to the soil-adsorbed (i.e., water-soluble) fraction (for rapid and slow fixation cases, and, rapid and slow mobilization cases; see footnote (g) of Table 1). Neither the below-ground methylation nor the 129I-retention by roots significantly impacted the leaf 129I concentration (see Fig. 7). Moreover, gaps from the observation were noticeable for all results obtained from these test calculations. It was also found that the model calculation of soil 129I is less sensitive to the selected parameters or settings. In particular, of the soil 129I concentrations obtained from the test calculations (data not shown here), those from the two pairs—increased and reduced adsorption cases; and, rapid and slow mobilization cases—exhibited observable variabilities. Compared with the slow mobilization case, the 129I concentrations calculated at the rapid mobilization case were observed to be smaller (by less than 5%) in the upper soil layers (of depth 0–8 cm) and greater (by less than 9%) below these horizons. A similar difference was observed

Fig. 7. Comparison of the observed (Aomori Prefecture, 2008; Kakiuchi et al., in preparation) and model-calculated concentrations of 129I in the bamboo grasses’ leaf on the sampling dates. Calculation results are shown for the control case and the 14 test cases (Subsection 3.3). Error bars attached to the observed data represent counting errors of 129I analysis.

M. Ota et al. / Science of the Total Environment 704 (2020) 135319

between the reduced and increased adsorption cases. These differences were attributed to the altered transport of 129I in the soil by the revised partition of the soil 129I to soil-adsorbed (i.e., the watersoluble) fraction, discussed earlier. However, compared with the control case, the 129I concentrations obtained from these test cases differed by no more than 5%.

5. Discussion 5.1. Processes affecting the transfer of

129

I to leaves

The model-calculated transfers of 129I to the bamboo grasses’ leaf presented in Subsections 4.1 and 4.3 qualitatively agree with results presented in the literature on the dynamics of 129I in vegetated fields and clarify the relative importance of each 129I-transfer process considered in the model as a contributor to the leaf 129I (see Table S1). Notably, the transfer of 129I to the leaf was mainly caused by foliar adsorption of inorganic 129I (81%, see Fig. 5), while the accumulation of 129I in the leaf by foliar uptake of 129I2 (2%) or root uptake of 129I (17%) was lesser, which is in agreement with literature data (Soldat, 1976; Tikhomirov and Ryzhova, 1981; Whitehead, 1984; Whicker and Kirchner, 1987; Hinton, 1994; Pröhl, 2009; Hurtevent et al., 2013). Further, foliar adsorption was mostly (90%, see Fig. 5) induced by the retention of rainwater-dissolved inorganic 129I to leaf surface following raininterception, whereas dissolution of 129I2 from the canopy air to the wet leaf-surface had minor impact (10%) (Clark and Smith, 1988; Moran et al., 1999; Kinnersley and Scott, 2001; Shaw et al., 2007). Despite the 129I contained in the soil, relatively-low root uptake of 129I was ascribed to the fact that most (93%) of the soil 129 I occurred in the soil-fixed form—not plant-available, which demonstrates the importance of fractionation of soil 129I by adsorption, fixation, and mobilization processes (Sheppard and Evenden, 1988; Muramatsu et al., 1993, 1995; Yoshida et al., 1995; Kashparov et al., 2005). Furthermore, leaf 129I concentration was decreased significantly (by 60%) by accumulation of organic matter in the leaf (Marter, 1963; Whitehead, 1984; Whicker and Kirchner, 1987; Müller and Pröhl, 1993; Thiessen et al., 1999). CH129 3 I transferred to the leaf via atmosphere-leaf exchange or soil-to-leaf transfer was released to the atmosphere, and thus these 129 CH129 I concen3 I-transport processes had no impact on the leaf tration (Nakamura and Ohmomo, 1980; Amiro and Johnston, 1989). Clearly, the developed 129I model, which considers some 129 I-transfer processes, is useful to quantify the contribution of each 129I-transfer process on the contamination of leaves impacted by atmospheric releases of 129I by a reprocessing plant. Despite these indications of 129I-transfers by the developed model, there are still gaps between the model calculation and the field observation of leaf 129I concentration (see Fig. 4c). However, to obtain information for better modeling of land-surface dynamics of deposited 129I, and investigate the importance of the processes not considered in the developed model, we analyze possible causes of these gaps. Precisely, the following three main causes are identified: errors in the model calculation due to uncertainties in i) the input data of 129I (see Subsection 3.2); ii) the model’s parameters (see Subsection 3.3), and iii) the limitation of the model’s conceptual design due to the lack of other processes possibly affecting the leaf 129I concentration. For the uncertainty in the input data of 129I, we note that the concentrations of 129I in the atmosphere and the rainwater at the Obuchi site could differ from those measured at the IES (which is used in Subsection 3.2 as input data for the model), and that the temporal resolution of the prepared input 129I data was insufficient. Concerning the difference of the 129I concentrations between the Obuchi site and IES, the dispersion of the air-borne radioactiv-

9

ity released from the RRP is analyzed. In Ota et al. (2016b), a dispersion model’s calculation was applied to the atmospheric discharge of 14C (half-daily interval data) from the RRP during the year 2007. The model calculation (horizontal resolution of 300 m; cf. the distance between the Obuchi site and the IES is 2.2 km, see Subsection 3.2) exhibited similar values (differing only by 12%) for the annual mean of the atmospheric concentration of 14 C at the two locations. This suggests that the atmospheric concentrations of 129I at the two locations are similar, at least, on the annual mean basis, given the similar temporal patterns of the release rates of 14C and 129I as might be expected from the monthly values shown in Fig. 2a. The concentrations of 129I in the rainwater at the Obuchi site and the IES were also expected to be similar because 129I in rainwater is provided by rain-scavenging of atmospheric 129I (Moran et al., 1999; Kadowaki et al., 2018); in other words, the concentration of 129I in rainwater depends on that in the atmosphere (Muramatsu et al., 1987; Clark and Smith, 1988). On the other hand, the low temporal resolution of input 129I data—that is on the monthly interval—is considered as an explanation for the gaps between the model calculation and the field observation of leaf 129I (Fig. 4c). In the area around a reprocessing plant, the atmospheric concentration of discharged-radioactivity fluctuates rapidly—for example on an hourly time-scale—due to local meteorology (atmospheric dispersion) (Maro et al., 2017; Ota et al., 2016b; Terada et al., 2013). For 129I, due to rainscavenging of atmospheric 129I, the concentration in rainwater is variable even during an event of single rainfall (Muramatsu et al., 1987; Moran et al., 1999). Dry and wet depositions of 129I influenced by such temporal changes of 129I concentrations in the atmosphere and rain could not be identified by the model calculation when monthly-defined input data of 129I are used (see, Fig. 2b and c); moreover, input data of 129I at time-scales shorter than one month were not available (see Subsection 3.2). Regarding ii) the uncertainty in model parameters, a number of test calculations were performed (see Subsection 3.3) to check the effect of the modeled processes with parameters having uncertainties, including time constants of foliar adsorption of 129I, mobilization, fixation of soil 129I, methylation of soil 129I, and distribution coefficient in the soil. As shown in Fig. 7, it was found that the gaps in the observed leaf 129I concentration were noticeable for all the results obtained from these test calculations, suggesting that the modeling or parameter values associated with these processes are not significantly responsible for the gaps. Concerning iii) the limitation of the model’s concept and transfers representation, we identified some unconsidered sources and transport processes for 129I that could affect the leaf 129I concentration at the site. In particular, the model’s underestimation on June 12 (Fig. 4c) could be due to contribution from dry deposition of particulate 129I (Hungate et al., 1963; Martin, 1963; Perkins, 1963; Chamberlain, 1970); resuspension of 129I (Whicker and Kirchner, 1987; Nicholson, 1988; Thiessen et al., 1999), and emission of 129I2 from the soil (Whitehead, 1984; Fuge, 1990; Sheppard et al., 1993; Fukui et al., 1996). Whereas for the model overestimation on August 10, we considered influence from the presence or absence of previous accumulation of 129I in the leaf; weathering of leaf 129I (Martin, 1963; Chamberlain and Chadwick, 1966; Chamberlain, 1970; Garland and Cox, 1984; Muramatsu et al., 1987); methylation of 129I on leaf (Chamberlain, 1970; Hurtevent et al., 2013); methylation of 129I in leaf (Amiro and Johnston, 1989; Muramatsu et al., 1995; Muramatsu and Yoshida, 1995; Amachi et al., 2003); and retention of root-absorbed 129I by the roots (Whitehead, 1973, 1984; Muramatsu et al., 1983; Ashworth et al., 2003). In the year 2007, particulate 129I was observed at the IES (shown by ‘Particulate 129I’ in Fig. 2c); that indicated that the Obuchi site could be affected by depositions of particulate 129I. Meanwhile, in the model calculation, wet deposition of particulate 129I (if any)

10

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was accounted for because the sampled-rainwater (i.e., the measurement used for the model’s input, see Fig. 2b) was analyzed for all dissolved forms of 129I (Hasegawa et al., 2017). However, dry deposition of particulate 129I was not accounted for in the model calculation. Given the similar atmospheric concentrations for the particulate and gaseous 129I (Fig. 2c), and the deposition velocity of particulate 129I being less than that of 129I2—for instance, typical values over vegetated surfaces are 1 mm s1 for particulate 129I (submicron particles) and 1 cm s1 for gaseous I2 (Tikhomirov and Ryzhova, 1981; Müller and Pröhl, 1993; Hinton, 1994)—dry deposition of particulate 129I at the Obuchi site was expected to be lesser than that of 129I2. This shows that the dry deposition of particulate 129I might have minor impact on the leaf 129 I concentration since the leaf contamination was much more affected by wet deposition than the dry deposition of 129I2 (see Fig. 5); thus, the particulate 129I. Resuspended particles such as particulate organic matter and fine soil particles could also adhere to the aerial part of plants (Whicker and Kirchner, 1987; Nicholson, 1988; Thiessen et al., 1999). The litter material that contained 129I (see Fig. 3) could be the source of resuspended 129I at the site, which may explain the model’s underestimation of leaf 129I concentration on June 12 (Fig. 4c). However, it has been shown that leaf contamination due to resuspension of deposited radionuclides is, in general, less significant than the other transport processes (Malek et al., 2002; Sauras-Yera et al., 2003); for example, foliar deposition of resuspended soil-particles contributed less than 2% of the leaf contamination with 137Cs—mostly caused by root uptake—for vegetations cultivated on a 137Cs-contaminated soil (Malek et al., 2002). In forest, the resuspension pathway is rarely evoked. In our case, the following analysis confirmed the insignificance of the resuspension process. It is known that deposition of resuspended particles on leaves decreases with height above the ground due to dilution of atmospheric concentration of resuspended particles by turbulent diffusion (Lindberg et al., 1986; Lovett and Lindberg, 1992). By contrast, the measured concentrations of 129I in the pine needles growing at a site adjacent the Obuchi site (1. 5 ± 0.1 mBq kg1 on June 12 and 0.5 ± 0.1 mBq kg1 on August 10 (Aomori Prefecture, 2008; Kakiuchi et al., in preparation)) were greater than, or comparable to the measured concentrations of 129I in the bamboo grasses’ leaf (see Fig. 4c), despite the higher level of the pine trees’ canopy at the adjacent site than the bamboo grasses’ canopy at the Obuchi site. Given the similar deposition velocities of fine particles to pine needles and leaves (Dasch, 1987), this result indicates that the 129I-transfer processes other than the dry deposition of resuspended 129I were responsible for the contaminations of these leaves. Also, it is well known that iodide (I ) in soils can be oxidized to I2, i.e., not methylated (Fukui et al., 1996), that is a process not considered in the developed 129I model. The I2 generated in soils would be released from the soil (Whitehead, 1984; Fuge, 1990) and may deposit to leaves (Sheppard et al., 1993). However, the following analysis indicated the insignificance of such soil-atmosphere-leaf transfer of 129I2. In a complementary simulation, it was assumed 129 that all the CH129 I2. However, it 3 I released from the soil were should be noted that this is a strong assumption because the methylation of inorganic iodine is, in general, more readily than oxidation to I2 (Muramatsu and Yoshida, 1995; Moran et al., 1999; Amachi et al., 2003; Ashworth, 2009). Then, we find an input of 129I2 from the soil to the bamboo grasses’ canopy at 2.1  106 Bq m2 yr1, i.e., ‘‘Release from soil” of CH129 3 I in Fig. 5. This hypothetical input of 129I2 to the canopy is much less than the 129I transferred to the bamboo grasses’ leaves via other pathways, which is 3.7  104 Bq m2, see Fig. 4a and b. Concerning the previous accumulation of 129I in the leaf, we assumed that the analyzed leaf had developed before the year

2007 and so contained the 129I at the beginning of the year 2007 (see Subsection 3.2). Nevertheless, the leaf 129I concentration calculated for zero leaf-129I case well-agreed with the observation on August 10 (see Fig. 7), compared with the control case. This suggests that the previous accumulation of 129I is insignificant; i.e., the analyzed leaf could have been formed during the year 2007. On the other hand, the model’s underestimation of leaf 129I concentration on June 12 became significant for the zero leaf-129I case (see Fig. 7), suggesting that the leaf sampled on this day could have been formed before the year 2007. These comparisons indicate that both the previous accumulation of 129I and development of individual leaf should be considered and measured to predict 129I concentrations of leaves exposed to atmospheric 129I for a period exceeding the life-span of the leaf—one to two years for the bamboo grasses’ leaf (Shibata, 1992; Yokoyama and Shibata, 1998). The effect of weathering of leaf 129I partly explains the model’s overestimation of leaf 129I concentration on August 10 (Fig. 4c). It is well known that iodine adsorbed on leaf surfaces are lost due to the erosion of cuticle waxes by wind and rainfall (Martin, 1963; Chamberlain and Chadwick, 1966; Chamberlain, 1970; Kinnersley and Scott, 2001; Hurtevent et al., 2013). Relatively rapid losses of 129 I adsorbed on leaves have been observed; for example, time constants of decrease of leaf 129I concentration at six days for pine needles (Muramatsu et al., 1987), and 5- to 20- days for grass leaves (Chamberlain and Chadwick, 1966; Chamberlain, 1970; Adams and Voillequé, 1971; Kirchner 1994). A sensitivity analysis using the ECOSYS-87 model (Hinton, 1994) showed that the loss of leaf 129 I by weathering more readily decreased leaf 129I concentration than growth dilution. Therefore, probably our model overvalued the retention of adsorbed 129I by leaves. The effect of weathering would be particularly significant for the leaf sampled on August 10; i.e., the leaf that could have been formed during the year 2007, as mentioned earlier, because turnover (and thus erosion) of cuticle waxes occurs more readily in young (growing) leaves than mature leaves (Baker and Hunt, 1986). Another process that may affect the leaf 129I concentration is volatilization (methylation) of 129I on the leaf (Chamberlain, 1970; Amiro and Johnston, 1989; Muramatsu and Yoshida, 1995; Muramatsu et al., 1995; Ashworth, 2009). It has been observed that about 20% of I added on leaves of radish plants volatilized in four days (Hasegawa et al., 2015). Furthermore, Hurtevent et al. (2013) observed a more rapid loss; 20% of iodine applied as a solution on leaves of spring wheats volatilized in 16 h. Again, our model could overvalue the retention of adsorbed 129I by the leaf. As for the methylation of 129I in the leaf, it has been observed that inorganic 129I taken-up by plants could be methylated and released via stomatal openings as in the form of CH129 3 I (Amiro and Johnston, 1989; Muramatsu et al., 1995; Muramatsu and Yoshida, 1995; Amachi et al., 2003). However, it is supposed that this process did not significantly influence the concentration of 129 I in the bamboo grasses’ leaves, because the 129I transferred in the leaf (thus, could be released as CH129 3 I) was found to represent a minor part (19%) of the total 129I transferred to the leaf (see Fig. 5). As regard to the retention of 129I by roots, experiments have shown that the concentration of iodine in roots is 10- to 20- times higher than those in leaves or shoots for plants, such as grass, timothy, clover, spinach, and cabbage grown on iodinecontaminated soils (Weng et al., 2009) or iodine-solution cultures (Whitehead, 1973; Muramatsu et al., 1983; Hong et al., 2008). It has also been observed that inorganic iodine taken-up by roots adsorbs to root tissues, and is not transported to the other plant parts (Adams and Voillequé, 1971; Whitehead, 1984; Weng et al., 2009). In the present study, however, the leaf 129I concentration calculated for zero-root uptake case did not significantly differ (at most by 13%, see Fig. 7) from the one obtained at the control

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case. This indicates that even if the retention of 129I in the roots occurred, which leads to a low 129I translocation to the upper part of the plant, there would be no significant change in the concentration of 129I in the leaf. Overall, as summarized in Table S1, it was found that the foliar adsorption due to wet deposition of inorganic 129I by rain is the most effective process for the contamination of leaves with 129I among several 129I-transfer processes assessed. Further, previous accumulation of 129I in the leaf also needed to be quantified; this would be particularly essential for plants growing around a spent nuclear fuel reprocessing plant because they are exposed to atmospheric 129I over the years due to the operation of the reprocessing plant (Wershofen and Aumann, 1989; Katagiri et al., 1997; Shinohara, 2004). Regarding processes decreasing the leaf 129I concentration, growth dilution by the accumulation of organic matter in leaves, and on-leaf processes of weathering and methylation were found to be highly significant (see Table S1). Nevertheless, only little is known about the on-leaf processes (Thiessen et al., 1999; Kinnersley and Scott, 2001); besides, the current status of modeling of weathering and methylation are low (see Table S1). Future research should, therefore, be focused on the further study of dynamics of 129I adsorbed on leaf surfaces and how to incorporate it into environmental 129I models. 5.2. Processes affecting the below-ground dynamics of

129

I

In this subsection, we discuss processes affecting below-ground dynamics of 129I. The model calculation of soil 129I presented in Subsections 4.2 and 4.3, qualitatively well-agreed with results obtained ondynamics of 129I in soils, and provides insights into the role of each below-ground process on the retention and turnover of 129I deposited to soils (see Table S2). Wet deposition of inorganic 129I by rain was—by ten times—more significant than dry deposition (dissolution) of atmospheric 129I2, see Fig. 5 (Clark and Smith, 1988; Moran et al., 1999; Englund et al., 2010). Besides, the deposited 129I adsorbed to the top several-cm layer of the soil and did not migrated to the deeper part during the one-year period (Boone et al., 1985; Muramatsu and Ohmomo, 1986; Muramatsu et al., 1987, 1993; Robens et al., 1989; Sheppard et al., 1993, 1994; Rao and Fehn, 1999; Mironov et al., 2002). A large fraction (over 90%) of the soil 129I occurred in immobile soil-fixed form during the year (see Fig. 6b) as assumed by the initial condition (represented by Eq. (1)) and this was achieved by slow mobilization of soil-fixed 129I (Schmitz and Aumann, 1995; Hou et al., 2003; Hansen et al., 2011; Roulier et al., 2018); consequently, the vertical profile of 129I in the soil remained unchanged in the year (see Fig. 3) (Kashparov et al., 2005; Weng et al., 2009). Furthermore, only trace fractions (0.001%, see Fig. 5) of the soil 129I were methylated in the year—hence, methylation did not influence the concentration of inorganic 129I in the soil (Bostock et al., 2003; Sheppard et al., 2006; Ashworth, 2009; Roulier et al., 2019). Moreover, neither root uptake of inorganic 129I nor CH129 3 I influenced the soil 129 I concentration (Tikhomirov and Ryzhova, 1981; Roulier et al., 2018). Obviously, simulation by the developed 129I model is useful to investigate the role and importance of each below-ground processes in controlling dynamics of 129I in soils, which are often difficult to be assessed from field observations. However, there are gaps between the model calculation and the field observation of soil 129I concentrations, as shown in Fig. 3. The gaps could be attributed to i) the errors in the model calculation due to inadequate setting of soil 129I concentration at the beginning of the year 2007 (Subsection 3.2); ii) uncertainty in the model’s parameters (Subsection 3.3); iii) difficulty in obtaining soil 129I concentrations that are representative of those of a site; and iv) limitation of the model concept due to insufficient knowledge on the processes affecting below-ground dynamics of 129I.

11

Regarding the setting of soil 129I, we used the measurement of soil 129I obtained during the year 2003 as the soil 129I concentration at the beginning of the year 2007 (Subsection 3.2). Therefore, the difference between the two data of ‘‘Soil-Obs. (2003)” and ‘‘SoilObs. (Aug 10, 2007)” shown in Fig. 3 would reflect the deposition of 129I from the year 2003 till 2007, i.e., not only the simulation period of the year 2007. In fact, the model-calculated deposition—5.3  103 Bq m2, see Fig. 6a—of 129I during the year 2007 represented only 18% of the increase of 129I (by 2.9  102 Bq m2) in the top 12 cm horizon (i.e., the horizon where the 129I measurements were available) determined from the two observed concentrations of ‘‘Soil-Obs. (2003)” and ‘‘Soil-Obs. (Aug 10, 2007)” shown in Fig. 3. For the 129I deposition from the year 2003 till 2006, we consider depositions of the RRP-released 129I during the year 2006 and the non-RRP-derived 129I from the year 2003 to 2006. During the RRP’s test operation in the year 2006, bulk (wet and dry) deposition of 129 I at 4.0 ± 0.2 mBq m2 yr1 was observed at the IES (Hasegawa et al., 2017). Assuming the same 129I deposition for the Obuchi site, this 129I deposition represents 14% of the increase of 129I in the top 12 cm soil, as mentioned earlier. For the deposition of non-RRP-derived 129I, it is known that the 129I discharges from the European reprocessing plants affect the 129I levels in the northern hemisphere (Rao and Fehn, 1999; Moran et al., 1999; Hou et al., 2003; Kadowaki et al., 2018). It was also indicated that 129 I derived from the past bomb-tests and 129I released from the Tokai reprocessing plant (locating in Ibaraki Prefecture, about 500 km south of the study site) which operated from 1977 till 2007, have influences on the levels of atmospheric 129I in Japan (Toyama et al., 2012). These non-RRP derived 129I could have influences not only on the 129I contained in the soil sampled in 2003 (Fig. 3), but also the atmospheric deposition of 129I from 2003 to 2006. Hasegawa et al. (2017) estimated the deposition of the non-RRP-derived 129I in Rokkasho at 0.5 ± 0.2 mBq m2 yr1 that calculates a four-year (2003–2006) deposition of 2.1 ± 0.6 mBq m 2 —representing 7% of the increase of 129I in the top 12 cm soil. Totally, the 129I deposition from the year 2003 to 2006 accounts for approximately 20% of the increase of 129I in the top 12 cm soil. Concerning ii) the uncertainty in the model’s parameters, some soil parameters such as distribution coefficient, time constants of mobilization, fixation, and methylation could have uncertainties, as discussed in Subsection 3.3. However, the sensitivity analysis, performed in Subsection 4.3, showed that the soil 129I concentrations obtained from the test calculations were similar to those of the control case—deferring by less than 5%. This indicates that modeling or parameter values of the associated processes are not responsible for the gaps between the observed and modelcalculated soil 129I concentrations. Regarding iii) the difficulty in obtaining the soil 129I concentrations representative of those of a site, we identify spatial heterogeneities of deposition of airborne radionuclides and solution transport in soils in a site. In sites impacted by the Chernobyl or Fukushima’s fallout, factors of 2- to 3 differences were observed for the deposition of airborne radionuclides—129I and 137Cs (Mironov et al., 2002; Tagami et al., 2011). For the below-ground transport, individual soil in a site may reveal different capacity of 129 I transport (Mironov et al., 2002; Moran et al., 2002), given that factors affecting iodine transport, such as contents of organic matter and Fe-Al oxides (Sheppard et al., 1994; Evans and Hammad, 1995; Yoshida et al., 1995; Dai et al., 2004), reveal spatial distributions in a site (Hook et al., 1991; John et al., 2007). Moreover, heterogeneous water flow—such as preferential flow through large macropores—which is unpredictable by the model may cause different depth-distribution patterns for radionuclides deposited to soils (Whitehead, 1984; Bundt et al., 2000). The different depthdistributions of 129I observed in the soil cores sampled in the year

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2003 and 2007 (see Fig. 3) may reflect these heterogeneous natures of atmospheric deposition and below-ground transport. Considering iv) the limitation of the model’s concept, there are several sources and transfer processes of soil 129I not considered in the developed model. For the model’s underestimation of soil 129I concentration in the top 2 cm layer (Fig. 3), we analyze influences from dry deposition of particulate 129I (McMahon and Denison, 1979; Sehmel, 1980); deposition of leaf 129I due to weathering (Hauschild and Aumann, 1985; Whicker and Kirchner, 1987; Ghuman et al., 1993; Abbott and Rood, 1994; Muramatsu et al., 2004; Hurtevent et al., 2013); transfer of 129I contained in the litter to the soil (Hauschild and Aumann, 1985; Ghuman et al., 1993; Roulier et al., 2018); and reappearance of root-absorbed 129I in the soil (Thiessen et al., 1999). On the other hand, concerning the model’s overestimation of soil 129I concentrations observed at depths from 8 to 12 cm (Fig. 3), we consider the effects due to emission of 129I2 (loss of 129I) from soil (Whitehead, 1984; Fuge, 1990; Sheppard et al., 1993; Fukui et al., 1996), and retention of atmospheric deposits of 129I by above-ground litter (Roulier et al., 2018, 2019). As noted in Subsection 5.1, dry deposition of particulate 129I is not accounted for in the model calculation and may explain the model’s underestimation of soil 129I concentration observed in the surface 2 cm layer (Fig. 3). However, the following analysis indicates the insignificance of this process as a 129I input to the soil. Given the similar concentrations of particulate and gaseous 129I in the atmosphere (see Fig. 2c) and the deposition velocity of particulate 129I being less than that of 129I2—typical values for deposition to the ground surface are 0.5 mm s1 for particulate iodine (submicron particles) and 3 mm s1 for I2 (Müller and Pröhl, 1993; Hinton, 1994)—it is expected that dry deposition of particulate 129I to the soil is less than that of 129I2, and thus, much lesser than the wet deposition of inorganic 129I (see Fig. 5). It is worth noting that 129I removed from leaf surfaces by weathering deposits onto ground surface (the litter layer at the Obuchi site) (Chamberlain and Chadwick, 1966; Abbott and Rood, 1994; Kinnersley and Scott, 2001; Hurtevent et al., 2013) and can be sent to soil by precipitation wash-off (Perkins, 1963; Whicker and Kirchner, 1987; Hoffman et al., 1992). However, such leaf-tosoil transfer of 129I was probably insignificant, because foliar adsorption of 129I, i.e., the leaf 129I that could be sent to the soil, was much less (by 18 folds) than the direct input of 129I to the soil by wet and dry depositions (see Fig. 5). Rainfall-leaching of radionuclides due to litter decomposition may occur (Ghuman et al., 1993; Hauschild and Aumann, 1985; Ota et al., 2016a; Roulier et al., 2018, 2019). Therefore, the 129I contained in the litter (Fig. 3) could have been transferred to the soil during the year 2007. However, such litter-to-soil 129I transfer was probably insignificant for the following reason. The turnover of litter by microbial decomposition is a slow process with a time constant of several years (Parton et al., 1987; Berg and McClaugherty, 2003). Therefore, the concentration of 129I in the litter was thought to be unchanged during the one-year simulation period; indeed no significant difference was observed between the 129I concentrations in the litter sampled at the timings spanning by two months: 1.2 ± 0.1 mBq kg1 on June 12 and 1.0 ± 0.5 mBq kg1 on August 10, 2007 (see Fig. 3). These concentrations of 129I in the litter were significantly lower than that of the top 2 cm soil (3.2 ± 0.4 mBq kg1, see Fig. 3), suggesting that the litter-to-soil 129I transfer did not increase the 129I concentration in the soil. As for the root-absorbed 129I, 129I taken-up by plant roots might be a source of soil 129I. As noted in Subsection 5.1, root-absorbed 129 I could be stored in roots (Whitehead, 1973, 1984; Muramatsu et al., 1983; Hong et al., 2008). Fine roots of bamboo grasses which are responsible for uptake of soil water, and thus, the dissolved

129

I, are replaced several times every year (Shibata, 1987; Fukuzawa et al., 2013). Dead fine roots are rapidly (by months) decomposed (Silver and Miya, 2001; Trumbore et al., 2006), which would result in a release of the root-stored radionuclides to soil (Thiessen et al., 1999). However, 129I taken-up by the bamboo grass roots that could be released to the soil, was much less than the input of 129I to the soil by wet and dry depositions (see Fig. 5); indicating that the root-absorbed 129I could never be a significant source of soil 129I. Below-ground oxidation of inorganic 129I to 129I2 and subsequent emission of the generated 129I2 from soil that was noted in Subsection 5.1, might explain the model’s overestimation of concentrations of 129I observed in the soil layer at depths from 8 to 12 cm (Fig. 3). However, it is expected that this process did not significantly influence the soil 129I concentration because methylation of inorganic 129I in soils is, in general, more readily than oxidation to 129I2 (Muramatsu and Yoshida, 1995; Moran et al., 1999; Amachi et al., 2003; Ashworth, 2009); besides, the methylation had no observable impact on the soil 129I concentration at the site during the year. Another process unconsidered in the model that may affect the below-ground dynamics of 129I is the retention of atmospheric deposits of 129I by the litter layer. It has been shown that humus litter materials act as temporary storage of atmospheric deposits of iodine (Roulier et al., 2018, 2019). This would reduce the transfer of deposited 129I to soil; i.e., the model calculation could have overvalued the input of 129I to the soil. However, 129I retention by the litter was unlikely. For instance, if we assume that all of the 129I deposited to the ground surface so far (including the 129I derived from European reprocessing plant, Tokai reprocessing plant and the past bomb-tests, as mentioned in the fourth paragraph in Subsection 5.2) were retained and stored in the litter, we estimated an integrated 129I deposition at 0.8 ± 0.6 mBq m2 that is the 129I stock in the litter calculated from the average of the measured 129I concentrations in the litter (1.1 ± 0.5 mBq kg1, Fig. 3) and the average stock of litter (0.7 ± 0.4 kg m2) in forested areas in the Aomori Prefecture (Ugawa et al., 2012). This estimates of integrated 129I deposition are much less than the 129I stock in the soil (model-assumed value of 0.23 Bq m2, Table 1), suggesting that most of the 129I deposited so far had been sent to the soil—not being stored in the litter. Overall, due to the limited availability of measurement of soil 129 I at the beginning of the year 2007, and the spatial variabilities expected for the atmospheric deposition and the below-ground transport of 129I at the site, there may be possibilities that the model calculation did not perfectly simulate the dynamics of soil 129 I during the year. However, we believe that the model calculation that accounted for a number of below-ground processes is useful to obtain insights into the long-term dynamics of 129I in the soil that is important for contaminations of terrestrial environments impacted by long-term operation (thus 129I deposition) of a reprocessing plant (Tikhomirov and Ryzhova, 1981; Shinohara, 2004). From the model calculation, it was indicated that root uptake can export more 129I out of the rooting-zone of the soil than methylation. It is believed that methylation is responsible for the turnover (loss) of soil 129I (Sheppard et al., 1994, 2006; Muramatsu and Yoshida, 1995; Muramatsu et al., 1995; Amachi et al., 2003; Ashworth, 2009); however, the results obtained in this study revealed that root uptake of 129I is much greater (by 24 folds) than the methylation (see Fig. 5). Weng et al. (2009) observed similar dynamics of soil 129I—uptake of soil iodine by cabbage roots (10% of the 125I added to the soil) was greater than the amount of iodine lost from the soil due to volatilization or leaching (0.4%) during a 15 days’ cultivation after a pulsed input of 125I solution to the soil. As a reference of the long-term turnover, we calculated the effective turnover-time (i.e. residence time) of 129I in the root-

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ing zone of the soil (see Appendix C), being approximately 6000 years for the methylation and approximately 300 years for the root uptake. However, it should be noted that an organic-rich soil with slow methylation is assumed in the present application (Section A.3.2). Therefore, more rapid methylation has been observed often; for example, effective turnover-time of several years to decades was observed (Boone et al., 1985; Robens et al., 1989; Kocher, 1991; Sheppard et al., 1994; Muramatsu and Yoshida, 1995). Additionally, as stated earlier, root-absorbed 129I may reappear in soils due to root death, although this phenomenon is not experimentally confirmed. In these respects, further enhancement of the modeling of dynamics of root-absorbed 129I (i.e. sequestration of 129I by root level and reappearance of root 129 I to soil) and evaluation of root uptake of 129I for soils revealing wider range of rate of methylation, are required to well document the role of root uptake on the long-term dynamics of soil 129I. 6. Conclusions In this study, a model to calculate transfers of 129I in an atmosphere-vegetation-soil system was developed and applied to the observed transfer of 129I at a vegetated field impacted by the atmospheric release of 129I from RRP during the year 2007. The relative importance of each 129I-transfer pathway assessed by the model (Tables S1 and S2) will be useful information to conduct field monitoring campaigns for planned-releases of 129I from a reprocessing plant. Overall, it was found that leaf contamination with 129I was mainly caused by foliar adsorption of 129I induced by the wet deposition of inorganic 129I by rain. The wet deposition of 129I was the main input pathway of 129I to the soil as well. However, the deposition of 129I in the year 2007 represented only 2% of 129 I that pre-existed in the studied soil, indicating the effect of long-term accumulation and turnover of 129I in soils. Additionally, it was suggested that future research should be focused on the dynamics of 129I adsorbed on leaves; in particular, weathering and methylation, and the fate of root-absorbed 129I in soil-root system for better prediction of short-term (months to years) dynamics of deposited 129I and long-term (decades to centuries) turnover of 129 I transferred to soils, respectively. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgments A part of this work was supported by JNFL [the JNFL’s Contract Research 2017]. We gratefully thank to Mr. K. Sasaki and Mr. H. Kasai of JNFL for providing valuable comments on the results of model calculations. We also gratefully acknowledge Dr. H. Nagai of JAEA for having helpful discussions on the model development, and Dr. S. Ueda of IES and Dr. T. Kobayashi of JAEA for their supports to conduct this work. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2019.135319.

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