Soil column experiments for iodate and iodide using K-edge XANES and HPLC–ICP-MS

Journal of Geochemical Exploration 107 (2010) 117–123

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Journal of Geochemical Exploration j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / j g e o ex p

Soil column experiments for iodate and iodide using K-edge XANES and HPLC–ICP-MS Yoko S. Shimamoto a,⁎, Takaaki Itai a, Yoshio Takahashi a,b a b

Department of Earth and Planetary Systems Science, Faculty of Science, Hiroshima University, 1-3-1 Kagamiyama, Higashi-Hiroshima, Hiroshima 739-8526, Japan Laboratory for Multiple Isotope Research for Astro- and Geochemical Evolution (MIRAGE), Hiroshima University, 1-3-1 Kagamiyama, Higashi-Hiroshima, Hiroshima 739-8526, Japan

a r t i c l e

i n f o

Article history: Received 13 March 2009 Accepted 7 November 2009 Available online 17 November 2009 Keywords: Iodide Iodate HPLC–ICP-MS XANES Soil column experiment

a b s t r a c t Radioactive iodine is one of the most problematic radionuclides because of its long half life and high mobility. Mobility of iodine depends on the chemical form to a great extent. This paper reports the results of soil − column experiments we conducted to evaluate the mobilities of IO− 3 and I . In order to determine the − on soil, adsorption isotherms were obtained by batch experiments. mechanisms of adsorption of IO− 3 and I − Both adsorption isotherms of IO− 3 and I are well explained by Langmuir model. The adsorption maximum of − IO− 3 is about five times larger than that of I . In the column experiments, iodine distributions between soil and pore water in the soil column were determined at various depths. Chemical forms of iodine in soil and pore water were determined by X-ray absorption near-edge structure (XANES), and high performance liquid chromatography connected to inductively coupled plasma mass spectrometry (HPLC–ICP-MS), respectively. Vertical profiles of iodine in pore water were simulated using Visual MODFLOW. Our results showed, upon I− infiltration through the column, that a small amount of I− adsorbed on soil, and its mobility is mainly controlled by advection and dispersion. The profile of iodine concentration in pore water was well simulated by assuming equilibrium-controlled Langmuir type adsorption without considering any chemical − transformations. For the IO− 3 addition system into the column, however, IO3 adsorbed to soil to a larger − was also degree, which causes a much larger retardation effect than I−. In addition, reduction of IO− 3 to I confirmed in both soil and pore water by XANES and HPLC–ICP-MS, respectively. The fraction of I− increased − toward the deeper end in both phases because of its lower affinity for soil than IO− 3 , where the reduced I was released to the pore water and transported by the water flow. In this study, such reduction effect was clearly demonstrated by the speciation analyses of iodine in both soil and water phases, which confirmed that the mobility of I− is a dominant factor that controls the fate of iodine in the surface environment. © 2009 Elsevier B.V. All rights reserved.

1. Introduction Iodine is an essential micronutrient for human beings as well as animals for their production of thyroid hormones, and also for the proper functioning of the thyroid gland. Iodine deficiency, which can lead to severe metabolic disorders, is a worldwide public health problem. It is also the case that isotopes of radioiodine (129I and 131I), which are emitted by the processing of nuclear fuel, nuclear testing, and nuclear accidents are of serious environmental concern (Moran et al., 1999; Taghipour and Evans, 2000; Oktay et al., 2000; Buraglio et al., 2001). In addition, 129I, which has a long half life (1.6× 107 years), is disposed along with other radioactive wastes in potential deep underground geological repositories. When radioiodine is emitted to the air, the fraction of IO− 3 may increase during transportation under oxic conditions (Truesdale and Jones, 1996). Iodide constitutes the main

⁎ Corresponding author. Tel.: +81 82 424 7460; fax: +81 82 424 0735. E-mail address: [email protected] (Y.S. Shimamoto). 0375-6742/$ – see front matter © 2009 Elsevier B.V. All rights reserved. doi:10.1016/j.gexplo.2009.11.001

chemical form for the assessment of the migration of 129I released from underground nuclear-waste repository, because iodide is stable under reductive condition. Van Loon et al. (2003a,b) and Glaus et al. (2008) studied diffusion of iodide and iodate through Opalinus Clay, which is under investigation as a potential host rock for a high-level radioactive waste repository in Switzerland. Um and Serne (2005) performed a series of batch sorption and column experiments to investigate sorption and transport behavior of 99Tc, 129I, 79Se, and 90Sr on and through borehole sediments collected at the Hanford Site in the U.S. Major chemical forms of iodine in soil include iodide (I−), iodate (IO− 3 ), and organically bound iodine (Yamada et al., 1999, 2002; Yuita, 1992; Yamaguchi et al., 2006; Shimamoto and Takahashi, 2008), and the environmental behavior of iodine can be different among these species. In earlier research, Hu et al. (2005) conducted integrated column and batch experiments to investigate the inter-conversion, sorption and transport of various iodine species (iodide, iodate, and 4-iodoaniline). They used different types of sediments which exhibited a wide variation in organic matter, clay mineralogy, soil pH, and texture. They confirmed the occurrences of iodate reduction, irreversible retention or mass loss

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of iodide, and rate limited and non-linear sorption. They identified that iodate reduction to iodide was presumably mediated by structural Fe(II) in some clay minerals in batch experiments. In order to demonstrate the interaction between iodine and soil components more precisely, investigation of the speciation of iodine in soil is necessary. X-ray absorption fine structure (XAFS) has been recently used as an in-situ speciation method for iodine in solid materials (Fuhrmann et al., 1998; Feiters et al., 2005; Kodama et al., 2006; Schlegel et al., 2006; Yamaguchi et al., 2006, 2008; Shimamoto and Takahashi, 2008). Use of XAFS enables us to measure the samples without any pretreatments, and it can be applied to both solid and liquid phases. In our previous study, we compared characteristics of K-edge and LIII-edge X-ray absorption near-edge structure (XANES) (Shimamoto and Takahashi, 2008). The spectrum of LIII-edge XANES is more effective in distinguishing the chemical forms of iodine, but the detection limit is high due to the interference of Ca in fluorescence mode. Meanwhile, K-edge XANES of iodine is featureless among iodine species, though it has a lower detection limit than LIII-edge because of no interferences from other major elements. The speciation of iodine in solution is also established using high performance liquid chromatography connected to inductively coupled plasma mass spectrometry (HPLC–ICP-MS). The combination of XANES and HPLC–ICP-MS is a powerful tool to interpret detailed chemical reactions of iodine between solid and liquid phases. Most of the earlier studies on column experiments discussed breakthrough curves that analyze solution eluted from the column. There are few papers which directly measured the distribution of iodine in both soil and pore water in the column. In order to ascertain the chemical reactions during migration of iodine in soil, determination of the chemical forms of iodine in soil and pore water is required, which is more readily assessed by the analyses on water and soil in the column. This paper reports the column experiments we performed using IO3− and I− as initial chemical forms of iodine, and the determination of iodine species in both soil and pore water samples by iodine K-edge XANES and HPLC–ICP-MS, respectively. Adsorption isotherms of IO− 3 and I− were also obtained by batch experiments to estimate the mechanisms of adsorption. 2. Experiments 2.1. Reagents Iodine reference compounds for XANES experiments, such as KI and KIO3 (purity N 99%) were purchased from Wako Pure Chemicals (Osaka, Japan). For solid compounds, pellet samples were prepared from mixtures of each iodine reference sample and powdered boron nitride (BN; Wako Pure Chem.). The concentrations of standard samples were 5 wt.%, and the thickness of the pellets was 4 to 5 mm. CH3I was diluted to 0.4 wt.% by ethanol. One percent of KI and KIO3 solutions was also prepared as references. All the samples were packed into airtight polyethylene bags (thickness: 75 μm). Suwannee River Humic Acid (SRHA) was purchased from International Humic Substances Society as the reference of iodine in humic substances (Ritchie and Perdue, 2003). 2.2. XANES Iodine K-edge XANES was measured at beamline BL01B1 using a bending magnet at SPring-8 (Hyogo, Japan). In order to obtain the incident X-rays, an Si(311) double-crystal monochromator with two mirrors was used. The beam size was varied from 0.2 (vertical) × 0.3 (horizontal) to 0.2 (v) × 6 (h) mm2 depending on the iodine concentrations in the samples. XANES spectra were measured in fluorescence mode using a 19-element Ge semiconductor detector, except for reference materials that were measured in transmission mode. Analysis of the data, including background subtraction, normalization, and linear combination fitting of XANES spectra, were performed with REX2000

Ver. 2.5 software (Rigaku). The contribution of background was subtracted from the original spectra by extrapolation of linear absorption obtained by least squares analysis in the pre-edge region. The background-subtracted spectra were normalized by defining the absorption at 33.300 keV as unity. The sample spectra were fitted in the energy range from 33.170 to 33.210 keV by the linear combination of − reference solution XANES spectra of two components, IO− 3 and I samples, since the simulation can fit the spectra in a better manner rather than using solid KIO3 and KI as reference samples, except for the spectra collected at 5 to 6 cm in the IO− 3 column.

2.3. HPLC–ICP-MS The I−/IO− 3 ratio in a solution was determined by HPLC–ICP-MS. The pump and oven used were a Pu-2089 Plus (JASCO) and a Co-2065 Plus (JASCO), respectively. An anion exchange column (TSK-gel Super IC-AP; 7.5 cm, Tosoh) was used at a constant temperature of 40 °C. The mobile phase was a mixture of 0.25% tetramethylammonium hydroxide (TMAH) and 0.3% methanol. The flow rate was 1.0 ml/min. The I− peak was observed at around 128 s, while IO− 3 was observed at 220 s.

2.4. Column experiments We obtained Gleysols from plowed layer of a paddy field of the National Institute for Agro-Environmental Sciences (NIAES) at Tsukuba, Japan. Characteristics of this soil are reported in Takahashi et al. (2003) and Yamaguchi et al. (2006), and this soil is classified as a fine textured gray lowland soil, having a light clay (LiC) texture. This soil contains SiO2, Al2O3, Fe2O3, and organic carbon at concentrations of 59.6, 17.8, 7.19, and 1.68 wt.%, respectively. The sample was dried in an oven at 50 °C and sieved (b0.5 mm). The soil density was 2.46 mg/cm3. The soil was incrementally packed with the air-dried sediment in an acrylic column (4.0 cm inner diameter, 30.0 cm height). Small amounts of soil were gently placed in the column by a spatula and pushed by 2.0 kg weight to ensure a uniform bulk density, which was repeated until the height of the soil column reached 10.7 cm. Calculated porosity and pore space of the soil column were 54.6% and 73.4 cm3, respectively. Eighty grams of Milli-Q water was added to the column from the top to moisten the soil. Fifty milliliter each of 39.4 mmol/L KI or KIO3 solution was added to the column from the top. Each solution was drained out from the column within 12 h. After the entire volume of the solution was drained off, soil was scooped up at 1 cm intervals. The wet soil samples were centrifuged and separated into pore water and soil. Collected pore water was immediately filtered by a 0.45 μm membrane filter (PTFE; Advantec, Tokyo). Species of iodine were determined by K-edge XANES for soil, and by HPLC–ICP-MS for pore water. All the samples of solutions for HPLC–ICP-MS were diluted by MQ water to appropriate concentrations after filtration, and measurements were conducted within 2 h after the entire volume of solution was drained out. Iodine concentrations in solutions were determined by ICP-MS (Agilent 7500cx). Each solution was diluted by 1.25% TMAH, and 1 μg/kg Ge was added as an internal standard. Iodine concentration in soil was determined by a method reported by Yamada et al. (1996) as follows. Each soil sample was dried at 50 °C, and the dried soil was homogenized by an agate mortar. The mixture of 0.1 g of soil and 5% TMAH was heated at 70 °C in an oven for 3 h. After filtration by a 0.45 μm membrane filter, iodine concentration in the filtrate was determined by ICP-MS. Iodine concentration eluted by MQ water was 0.12 μmol/L determined by batch experiment following next section. The sterilized soil, which was irradiated by a dose of 5 mrad of γ-ray, was also used for column experiment of IO3− in order to examine the effect of bacterial activities.

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2.5. Iodine adsorption isotherm − In order to determine the mechanism of adsorption of IO− 3 and I on or I− soil, adsorption isotherms were examined. Three milliliter of IO− 3 solution was added to 5.0 g of dried soil in a polystyrene vessel. This solid to solution ratio is close to the column experiment condition (weight ratio; solid:solution ≈ 2:1). Initial iodine concentrations in solution were adjusted to 0, 0.33, 0.66, 6.74, 13.48, 20.20, 27.05, and 40.31 mmol/L for IO− 3 , and 0.77, 3.93, 7.80, 23.40, 39.26, and 55.25 mmol/L for I−. The samples were then placed on a reciprocating shaker (100 rpm) at 25 °C for 10 h determined by preliminary experiments on the equilibrium time. Following the period of shaking, solution was separated by centrifugation, and filtered by 0.45 μm membrane filters. The pH values of supernatant were measured by a pH-glass electrode (F-52, Horiba). Each sample of solution was diluted by TMAH to an appropriate concentration for measurements by ICP-MS.

3. Results 3.1. Adsorption isotherm Initial concentration of iodine in the soil was 1.47 mg/kg. Kaolinite was identified as the main clay mineral by X-ray diffraction analysis using preferred orientation method for the clay fractions (b2 μm) separated by sedimentation method. The pH values of the supernatant − after 10 h were within 5.1 to 5.2 for IO− 3 , and 4.9 to 5.3 for I . Both − and I can be fitted by Langmuir (Fig. 1) and isotherms of IO− 3 Freundlich isotherms. Freundlich isotherm is an empirical equation, which can be applied to solids with heterogeneous surface properties (Stumm, 1992). The Freundlich equation is described as follows, n

½Iads = KF ½Iaq

ð1Þ

where [I]ads and [I]aq are concentrations of iodine in solid and in solution, respectively. KF is the Freundlich constant, and n is the Freundlich exponent. The KF and n values determined to fit the experimental data were (KF, n) = (2.15, 0.64) and (0.22, 0.63) for IO− 3 and I−, respectively. The correlation coefficients (R2) between fit and − experimental data are 0.998 and 0.991 for IO− 3 and I , respectively. When the exponent n is equal to unity, the Freundlich isotherm is equivalent to the linear isotherm. Our fitting results showed that both − to the soil was non-linear adsorption. adsorption of IO− 3 and I Meanwhile, Langmuir equation is a theoretical formula, which can be explained as a specific case of Freundlich equation. The Langmuir model assumes a monolayer adsorption, which is typical for anion adsorption where a finite number of adsorption sites react until all the sites are occupied (Stumm, 1992). The Langmuir equation is described as follows, ½Iads =

½Imax × A × ½Iaq 1 + A × ½Iaq

:

− Fig. 1. Adsorption isotherm for (a) IO− 3 , and for (b) I . Closed squares show experimental data. Broken line shows fitting curve by Langmuir model.

remaining in the solution from the measured value for each soil using the water content. Upon IO− 3 infiltration through the column, henceforth abbreviated as IO− 3 column, concentration of iodine in pore water linearly decreased with increasing depth from 0 to 6 cm (Fig. 2a). Concentrations of iodine in soil were almost constant around 25 to 29 mmol/kg at 0 to 3 cm depths (Fig. 3a). The adsorption maximum of IO− 3 determined by Langmuir isotherm is similar to these values,

ð2Þ

The [I]max is adsorption maximum, and A is the Langmuir adsorption constant. The estimated adsorption maxima and Langmuir adsorption constants ([I]max, A) are (24.79 mmol/kg, 0.065 L/mmol) − and (4.60 mmol/kg, 0.025 L/mmol) for IO− 3 and I , respectively. The − adsorption maximum of IO3 is 5.4 times larger than that of I−. The values of R2 between fit and experimental data are 0.994 and 0.995 for − IO− 3 and I , respectively. 3.2. Column experiments 3.2.1. Iodine concentrations in pore water and soil The vertical profiles of concentration of iodine in pore water and in soil are shown in Figs. 2 and 3, respectively. Net concentration of iodine in soil was calculated by subtracting concentration of iodine

Fig. 2. Vertical profile of iodine concentration in pore water for (a) IO− 3 column, and for (b) I− column. The inserted figure shows the horizontal axis in log scale.

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reaching 78% at depths of 10 to 10.7 cm. The total amount of reduced I− in the pore water was about 0.2% of the total iodine in the column.

3.2.3. Iodine species in soil XANES spectra of reference materials and all the samples are shown in Figs. 5 and 6. In the I− column, the spectra at 0 to 9 cm depths are identical to that of I− solution (Fig. 6) suggesting the adsorption of I− to the soil components without any chemical transformation. At 9 to 10 cm depth, the spectrum is more rectilinear after absorption edge when compared with that of I− solution. The reference spectrum of KI is more featureless than that of I− solution, and similar to the spectra at 9 to 11 cm. At 9 to 11 cm depths, concentration of iodine in soil is less than that of the upper part. The adsorption of I− at the deeper end might be relatively inner-sphere like, where iodine loading is low. There is also a possibility of chemical transformation of I− to other iodine species such as IO− 3 , I2, or organic I at the depth, but it is difficult to discuss in more details due to the poor quality of the spectrum.

− Fig. 3. Vertical profile of iodine concentration in soil for (a) IO− 3 column, and for (b) I column.

showing that concentration of iodine in soil at 0 to 3 cm depths reached a saturation value of IO− 3 to the soil. At 4 to 6 cm depths, concentration of iodine in soil drastically decreased. The rate of decrease of iodine concentration in the upper part of the column is different between soil and pore water, which means that the adsorption of iodine on soil is non-linear. Upon I− infiltration through the column (=I− column), most of I− remained in pore water and was transported by water flow toward the deeper end (Fig. 2b). Concentrations of iodine in soil and pore water were almost constant at 0 to 5 cm depths (Figs. 2b and 3b). Adsorption capacity of I− to the soil was estimated at 4.1 mmol/kg from the column experiment. This value is also close to the adsorption maximum determined by Langmuir model. 3.2.2. Species of iodine in pore water In the I− column, only I− was detected from all the depths, showing that I− is stable under this column's condition. In the IO− 3 − column, IO− was confirmed (Fig. 4). At 0 to 3 cm 3 reduction to I depths, almost 100% of iodine existed as IO− 3 . However, below the depth of 4 cm, I− fraction increased toward the deeper end, finally

Fig. 5. Iodine K-edge XANES spectra of reference materials. (SRHA: Suwannee River Humic Acid).

− Fig. 4. Fraction of IO− in IO− 3 and I 3 column determined by HPLC–ICP-MS.

− Fig. 6. Iodine K-edge XANES spectra of soil samples in IO− column. 3 and I

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In the IO− 3 column, the spectrum collected at 0 to 5 cm was similar to − that of IO− 3 solution (Fig. 6). Because of the featureless spectrum of I , it − is difficult to distinguish the presence of a small amount of I species in the IO− 3 column especially from 0 to 6 cm depths. At 5 to 6 cm depth, the peak height of around 33.18 keV is higher than that at 0 to 4 cm depth. Although the reason is not clear at present, solid KIO3 reference spectrum is nearly identical to that at 5 to 6 cm (Fig. 7). This might have been caused by the different mechanisms of adsorption between high iodine loading and low iodine loading at the adsorption sites at the surface of soil. For example, it has been speculated that the outer-sphere complex of IO− 3 to the soil is important at a higher loading, while the inner-sphere complex can be dominant at a lower loading. The spectrum collected at 8 to 11 cm depths did not show any absorption at the energy corresponding to the edge, since iodine concentration in the soil was too low. At 6 to 7 cm depth, however, the spectrum lost the clear peak shape of IO− 3 around 33.18 keV, which shows existence of I−. This is consistent with the fact that the reduction − of IO− 3 to I was also found in the pore water analysis. The speciation in both phases is effective to fully understand the migration of IO− 3 including such chemical transformations.

Table 1 Parameters for simulation of column experiments. Conductivitya Kx [cm/s] Ky [cm/s] Kz [cm/s]

1.0 × 10− 10 1.0 × 10− 10 6.881 × 10− 5

Storage Specific storage [1/m] Specific yield Effective porosity Total porosity

1.0 × 10− 5 0.20 0.50 0.55

Longitudinal dispersivity [m]

1.0 × 10− 4

Drain Conductance [m2/h]

1.920 × 10− 3

Zone budget Function Time [h] Flow rate [m3/h] a

3.2.4. Simulation of iodine concentration in pore water Concentration of iodine in pore water was simulated by Visual MODFLOW Premium (Waterloo Hydrogeologic, Inc.; WHI). The equation describing the fate and transport of species k in three-dimensional, transient water flow systems can be written as follows: " # ∂ðθC k Þ ∂ ∂C k ∂ k k − θDij ðθvi C Þ + qs Cs + ∑Rn = ∂t ∂xi ∂xi ∂xj

ð2Þ

where, Ck is the dissolved concentration of species k [ML− 3]; θ is the porosity of the subsurface medium [dimensionless]; t is time [T]; xi is the distance along the respective Cartesian coordinate axis [L]; Dij is the hydrodynamic dispersion coefficient tensor [L2T− 1]; vi is the seepage or linear pore water velocity [LT− 1]; qs is the volumetric flow rate per unit volume of aquifer representing fluid sources (positive) and sinks (negative) [T− 1]; Cks is the concentration of the source or sink flux for species k [ML− 3]; ∑Rn is the chemical reaction term [ML− 3T− 1]. The numeric engine was USGS MODFLOW 2000 from WHI. Transport simulation was performed by DoD MT3D. V. 1.5 (Public Domain) that assumed Langmuir isotherm to describe the equilibrium iodine sorption. The simulation presumed a flow at steady state. In order to create a constant flow velocity, fictitious empty spaces were established above and below the soil column. The upper empty space was filled with

Fig. 7. Iodine K-edge XANES spectra of soil samples at 0 to 1 and 5 to 6 cm depths of IO− 3 column with the spectra of IO− 3 solution and KIO3, respectively.

121

(Constant head-In) × (1) + (Constant Head-Out) × (− 1) 12 5.33 × 10− 6

Hydraulic conductivity in the direction of the model X, Y, and Z-axes.

39.4 mmol/L of iodine solution, which was set at a constant concentration for the entire duration of the simulation of 12 h. Other tabulated parameters of the soil column are shown in Table 1. The parameters of hydraulic conductivity (Kz), drain, and zone budget were determined properly in order to design the average flow rate for the column experiment. Langmuir adsorption constant and the total concentration of available sorption sites determined by the adsorption isotherms were applied to the simulation. The results of the simulation with the experimental data are shown in Fig. 8. The experimental data are average values of triplicate and − columns, respectively. The large duplicate repetitions for IO− 3 and I error around the front was caused possibly by the degree of soil compaction, which can affect the effective porosity and water flow. The longitudinal dispersivity and effective porosity were fitting parameters optimized by the simulation of experimental profile of I− column. Subsequently, these parameters were also applied to the simulation of IO− 3 column, since such kind of physical properties must be similar among the two systems. In the I− column, distribution of I− is controlled

− Fig. 8. Concentration of iodine in pore water in (a) IO− 3 column and (b) I column with simulation curve calculated by Visual MODFLOW Premium. The solid line shows simulation curve, dots show experimental data. The experimental data are average − column, values of triplicate and duplicate repetitions for IO− 3 column and I respectively.

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by advection and dispersion because adsorption of I− on soil is small. However, in the IO− 3 column, the retardation effect by adsorption is very large when compared with the water advection. The simulation of IO− 3 column did not perfectly fit the experimental data, especially at 0 to 3 cm depths. 4. Discussion − In order to determine the mechanism of adsorption of IO− 3 and I on soil, we performed batch experiments at various concentrations. − fitted well by Langmuir equation, Adsorption isotherms of IO− 3 and I which assumes monolayer adsorption and availability of maximum number of adsorption sites (Fig. 1). The column experiments demonstrated that the concentration of iodine in soil is almost constant at 0 to − 3 cm for IO− 3 , and at 0 to 5 cm for I (Fig. 3). These concentration values are similar to the adsorption maximum obtained by Langmuir fitting. − These observations suggest that both IO− 3 and I amounts adsorbed on soil can be explained by Langmuir type adsorption. The value of adsorption constant (A) of IO− 3 to the soil is 2.6 times larger than that of I−. Such low reactivity of I− is well known in earlier studies (Yuita, 1992; Hu et al., 2005; Yamaguchi et al., 2006). Ticknor and Cho (1990) performed various adsorption experiments of iodate and iodide on calcite, chlorite, epidote, goethite, gypsum, hematite, kaolinite, muscovite, and quartz. Iodide was not removed from solution by sorption interactions with any of common granitic fracture-filling minerals. Kaplan et al. (2000) revealed that a significant amount of I− sorbed to illite, but sorbed I− could be easily desorbed with halide salts. I− adsorbed to illite predominantly by a reversible physical adsorption to the pH dependent edge sites. Nagata et al. (2009) also showed that the adsorption on ferrihydrite is mainly due to the outer-sphere complexation on the surface. − to soil The different affinity and adsorption maxima of IO− 3 and I generated different vertical profiles of iodine concentration in pore water (Fig. 2). As IO− 3 adsorbed on soil to a large degree, the profile of IO− 3 in pore water showed a linear decrease with increasing depth at 0 to 6 cm, and only a small amount of IO− 3 reached below 6 cm depth. On the other hand, I− was almost constant at 0 to 5 cm depth, since most of I− continued to remain in solution and was transferred by water flow. The vertical profiles of iodine in solution were simulated by Visual MODFLOW (Fig. 8). The profile of I− was well simulated by assuming that the equilibrium was controlled by Langmuir adsorption. The migration of I− is mainly controlled by advection and dispersion, and not by adsorption as the adsorption capacity of I− is very small. However, the vertical profile of IO− 3 did not fit well by the present model especially at 1 to 3 cm, the reason of which is not clear at present. However, taking account of the fact that the parameters of longitudinal dispersivity and effective porosity were fixed to the values determined for the I− column, the depth where the concentration of iodine dramatically decreased in the IO− 3 column (=3–4 cm) was relatively well produced by the simulation. We determined the chemical species of iodine in pore water and soil by HPLC–ICP-MS and iodine K-edge XANES, respectively. In the I− column, almost all the I− existed as I− in both phases (Fig. 6b). This means that I− is a stable chemical form under this column's condition. − − By contrast, in the IO− 3 column, reduction of IO3 to I was confirmed in both phases. The I− fraction in pore water gradually increased toward the deeper end from 3 cm depth onward. In addition, the presence of I− fraction in soil was only indicated at 6 to 7 cm depth, while IO− 3 was confirmed from 0 to 5 cm depth. This gap might be caused by the different detection limits between HPLC–ICP-MS and XANES. The detection limit for I− determined by HPLC–ICP-MS and K-edge XANES is about 0.3 wt.% and 10 wt.%, respectively. The small amount of I− in soil was not confirmed using XANES. However, this gap in depth for the presence of I− between soil and pore water can also be explained by the fact that most of the reduced I− is readily released to the pore water due to the much lower affinity of I− to the

− soil than that of IO− was transferred to the deeper 3 . The released I part by water flow and advection, and adsorbed to the soil only to some extent. Such low affinity of I− to the soil created the larger fraction of I− at the deeper part in the pore water of the IO− 3 column. In order to verify the adequacy of the simulation at the deeper end of the column, the profiles in log scale are shown in Fig. 9. Although the profile was reasonably simulated between 0 and 5 cm of the IO− 3 column, there is a large and systematic discrepancy between experimental data and simulations below 5 cm depth in the IO− 3 column. The − results can be explained by the reduction of IO− 3 to I , since we did not − consider the chemical transformation of IO− 3 to I in the simulation. The cannot reach beyond a depth of 7.3 cm. simulation predicted that IO− 3 The concentration of iodine of N0.01 mmol/L at the deeper end is not caused by the leaching effect from the soil, because only 0.12 μmol/L of iodine was leached from the soil by MQ water under the same condition with the batch experiment. Identification of the species of iodine in both soil and water phases enabled us to clearly demonstrate that the reduction induced the migration of iodine to the depth of 10.7 cm in the IO− 3 column. The simulation based on reactive transport model must be applied to this phenomenon in future in consideration of the reduction − effect of IO− 3 to I . According to Hu et al. (2005), the reduction of IO− 3 can be mediated by the presence of structural Fe(II) in some clay minerals by batch experiments. They also stated that there is a limited capacity for reduction for each medium. Since the soil used in our experiments is a clay rich soil, similar reactions might have occurred in our column − experiments. Bacterial activity can also reduce IO− 3 to I (Amachi et al., − 2005). The reduction of IO− 3 to I in our soil columns could have been caused by bacterial activities too. However, we had used sterilized soil for use in the IO− 3 column (data not shown), which rules out any − bacterial reduction of IO− 3 to I in our soil columns. This means that the in our column experiments occurred without bacterial reduction of IO− 3 activities. The addition of a high amount of iodine was needed to couple the speciation analyses with the column experiments. However, in the natural environment, concentration of iodine is not so high when compared with that of our column experiments. Based on the speciation analyses, it is possible to show clearly that the reduction of iodate to iodide cannot be negligible for the evaluation of the migration of iodine in surface soil–water systems. Similar effects can prove important in the migration of iodine at lower concentrations in a soil–water system, since − can be more significant at the lower the reduction of IO− 3 to I concentration level.

Fig. 9. Concentration of iodine in pore water with simulation curve. The horizontal axis of Fig. 8 was changed into log scale.

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5. Conclusions This paper reports the results of our soil column experiments and − batch experiments of IO− 3 and I as initial chemical forms of iodine using surface soil sample from a paddy field. Adsorption isotherms of IO− 3 and I− fitted well by Langmuir model. The estimated adsorption maxima of − IO− 3 and I to the soil were 24.82 and 4.61 mmol/kg, respectively. With the addition of I− in the column, a large amount of I− continued to remain in a stable state in the solution, which was transported by advection and dispersion with a lesser degree of retardation to soil by adsorption. The vertical profile of I− column was well simulated under assumption of equilibrium controlled by Langmuir type adsorption. On the other hand, IO− 3 adsorbed on soil to a larger degree in the column − − was injected with IO− 3 solution. In addition, reduction of IO3 to I confirmed in both soil and pore water by XANES and HPLC–ICP-MS, respectively. The reduced I− was adsorbed to soil components to some extent, but most of the I− was released to pore water and was transported to the deeper end by water flow. The reduction effect in IO− 3 column experiment is shown clearly in our study by coupling the speciation analyses such as XANES and HPLC–ICP-MS applied to soil and water phases, respectively. A more elaborate model that takes the reduction effect into consideration must be developed to simulate the depth profile of IO− 3 in the soil column. Acknowledgements The authors thank Dr. S. Endo for operating gamma radiation. This research project was conducted under a research contract with the Japan Nuclear Energy Safety Organization (JNES). This work was performed with the approval of JASRI (2003B0385, 2007B1175, and 2008A1463). The English language of this paper was refined by Dr. Nachiketa Das. References Amachi, S., Fujii, T., Shinoyama, H., Muramatsu, Y., 2005. Microbial influences on the mobility and transformation of radioactive iodine in the environment. Journal of Nuclear and Radiochemical Sciences 6, 21–24. Buraglio, N., Aldahan, A., Possnert, G., Vintersved, I., 2001. 129I from the nuclear reprocessing facilities traced in precipitation and runoff in northern Europe. Environmental Science and Technology 35, 1579–1586. Feiters, M.C., Küpper, F.C., Meyer-Klaucke, W., 2005. X-ray absorption spectroscopic studies on model compounds for biological iodine and bromine. Journal of Synchrotron Radiation 12, 85–93. Fuhrmann, M., Bajt, S., Schoonen, M.A.A., 1998. Sorption of iodine on minerals investigated by X-ray absorption near edge structure (XANES) and 125I tracer sorption experiments. Applied Geochemistry 13, 127–141. Glaus, M.A., Müller, W., Van Loon, L.R., 2008. Diffusion of iodide and iodate through Opalinus Clay: monitoring of the redox state using an anion chromatographic technique. Applied Geochemistry 23, 3612–3619.

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