Remobilization of radiocesium on riverine particles in seawater: The contribution of desorption to the export flux to the marine environment

Marine Chemistry 176 (2015) 51–63

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Remobilization of radiocesium on riverine particles in seawater: The contribution of desorption to the export flux to the marine environment Hyoe Takata ⁎, Kazuyuki Hasegawa, Shinji Oikawa 1, Natsumi Kudo, Takahito Ikenoue, Ryosuke S. Isono, Masashi Kusakabe Central Laboratory, Marine Ecology Research Institute, 300 Iwawada, Onjuku-machi, Isumi-gun, Chiba 299-5105, Japan

a r t i c l e

i n f o

Article history: Received 22 May 2015 Received in revised form 30 June 2015 Accepted 9 July 2015 Available online 14 July 2015 Keywords: Fukushima-derived radiocesium Riverine particles Desorption Export flux

a b s t r a c t We evaluated the influence of radiocesium (137Cs) desorbed from riverine particles on the export flux of dissolved 137Cs from rivers to the marine environment, where it may accumulate and become concentrated in marine biota. The samples were collected from stations along the Abukuma (December 2013), Kuji (November 2013), Naka (November 2013), and Tone (October 2013) rivers, in the catchments of which large quantities of radionuclides from the Fukushima Dai-ichi Nuclear Power Plant (FNPP) accident were deposited. We estimated the extent of the desorbed fraction from riverine particles by conducting a particle–seawater partitioning and desorption experiment in which sieved particles (b 74 μm) from the riverbank soils and river-bottom sediments were added to filtered (b0.45 μm) seawater. After the addition of the particles, the dissolved 137Cs concentration increased rapidly as a result of desorption of 137Cs from the particles, and then became approximately constant. The desorbed 137Cs fraction (i.e., the amount desorbed in seawater relative to the initial total amount in the particles) after 7 d ranged from 0.75 to 6.6%. This variability in the desorbed 137Cs fraction reflects the varied composition of the particles (clay minerals, Fe/Mn oxides/hydrous oxides, and organic matter), which controls the adsorption–desorption reaction. We calculated the export flux from each river to the coastal area in two ways: (1) by multiplying the sum of the desorbed fraction and dissolved concentration in the river water by the river discharge (i.e., the flux with the desorbable fraction), and (2) by multiplying just the dissolved concentration in the river water by the discharge (the flux without the desorbable fraction). The estimated 137Cs flux under normal flow conditions when the desorbed fraction was included was 72–86 MBq/d from the Abukuma River, and 4.4–4.6 MBq/d from the Kuji River. These fluxes are respectively 2.5–21.9% and 1.1–9.9% larger than the fluxes calculated without the desorbed fraction (Abukuma, 71 MBq/d; Kuji, 4.2 MBq/d). Under high-flow conditions, the rivers would likely carry a greater suspended particle load with high 137Cs (N 1000 Bq/kg-dry) to the marine environment, thereby increasing the contribution of desorption to the export flux of dissolved 137 Cs. These results show that reactive particulate Cs (i.e., the desorbable fraction) must be taken into account when estimating the radiocesium fluxes from rivers to coastal regions near the FNPP. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Damage to the Fukushima Dai-ichi Nuclear Power Plant (FNPP) caused by the Tohoku-oki earthquake and subsequent tsunami on 11 March 2011 resulted in the release of large amounts of radionuclides to the surrounding environment. Radiocesium (134Cs and 137Cs) derived from the damaged FNPP contaminated surrounding areas of the Japanese islands and the North Pacific Ocean (Chino et al., 2011; Kawamura et al., 2011; Yoshida and Kanda, 2012). Estimates of the total amount of 137Cs released directly into seawater range from 3 to 30 PBq (Bailly du Bois et al., 2012; Charette et al., 2013; Rypina et al., ⁎ Corresponding author. E-mail address: [email protected] (H. Takata). 1 Present address: Nuclear Regulation Authority, 1-9-9 Roppongi, Minato-ku, Tokyo, Japan.

http://dx.doi.org/10.1016/j.marchem.2015.07.004 0304-4203/© 2015 Elsevier B.V. All rights reserved.

2013; Tsumune et al., 2012), and estimates of the amount released to the atmosphere between 12 March and 30 April 2011 range from 11 to 15 PBq (Chino et al., 2011; Kawamura et al., 2011 and references therein). Of this atmospheric 137Cs, 20–50% (2–6 PBq) was deposited on land (Kawamura et al., 2011; Stohl et al., 2012; Morino et al., 2011, 2013). Most of the Fukushima-derived 137Cs deposited on land is thought to remain in the soil, with about 70% estimated to be in the surface layer (2–5 cm depth) (Fujiwara et al., 2012; Koarashi et al., 2012). Soil with 137 Cs of N100 kBq/km2 covers an area of about 3000 km2 on eastern Honshu Island (MEXT, 2011). As a result of processes such as erosion and resuspension of river-bottom sediment, much of this contaminated soil is likely to be eventually transported as suspended particles in rivers to the marine environment (Knighton, 1998; Matsunaga et al., 1991; Ritter et al., 2006). For example, Nagao et al. (2013) investigated the effect of a heavy rain event associated with a typhoon in September 2011

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on 137Cs export from two Fukushima rivers and reported a 137Cs concentration in the river water (both dissolved and particulate phases) of ~ 0.85 Bq/L, with the particulate phase (i.e., suspended particles) accounting for close to 100% of the total. Ueda et al. (2013), however, estimated that the amount of 137Cs discharged in 2011 after the accident in the Hiso and Wariki rivers (50–60 km northwest of the FNPP) accounted for only 0.5% and 0.3%, respectively, of the total amount of radiocesium deposited in their catchments. Therefore, riverine export of FNPP-derived radiocesium to the marine environment is likely to continue for many years. Riverine input of radiocesium has important effects on the biogeochemistry of radiocesium in river–sea systems (Tagami and Uchida, 2013), and radiocesium in the dissolved phase in particular is easily taken up by marine biota (IAEA, 2004). Although as noted above much of the radiocesium carried by rivers is in particulate form (i.e., adsorbed onto the surfaces of suspended particles), the salinity increase in estuaries results in the desorption of this radiocesium, thus increasing the concentration of radiocesium in the dissolved phase in coastal seawaters (Turner, 1996; Fuhrmann et al., 1997; Garnier et al., 1996; Li et al., 1984; Staunton and Roubaud, 1997; Tagami and Uchida, 2013; Turner et al., 1993). However, the contribution of the desorbed 137Cs fraction to the export flux of radiocesium in rivers to the marine environment has not yet been quantified, as these studies mainly focus on the desorption processes during the estuarine mixing. The main goal of this study was to evaluate the contribution of desorption of radiocesium from suspended particles to the riverine export flux of FNPP-derived 137Cs to the marine environment. To achieve this goal, we determined experimentally the desorbable fraction of radiocesium in soil and sediment samples collected along four rivers near the FNPP. We also examined how the desorbed fraction in seawater changed with time and with the particle concentration in the seawater. We also determined dissolved Cs activities in the water and suspended particles in water samples from each river. Then we calculated the contribution of desorption to the export flux by multiplying the desorbed fraction by the river discharge, and the sum of the desorbed fraction and the dissolved concentration by the river discharge, and compared the results. In addition, we examined how the composition of the suspended particles affected the desorbed Cs fraction by measuring the Si and Al (indicators of the clay mineral content), Fe and Mn, and organic carbon and organic nitrogen in the collected particles. Clay minerals, Fe and Mn oxides and hydrous oxides, and organic matter (i.e., organic carbon plus organic nitrogen) are important substrates for the sorption of elements, and the relative contribution of these substrates to element sorption depends on how an element interacts with the substrate. We also compared the desorptive behavior between stable Cs and 137Cs. 2. Methods and materials 2.1. Study area In 2013, we collected samples from the Abukuma (sampled on 5 December), Kuji (20 November), Naka (20 November), and Tone (29 October) rivers in eastern Honshu, Japan (Fig. 1). Detailed sampling points in each river were listed in Table 1. The estimated amounts of 137 Cs derived from the FNPP accident deposited in the catchments of these rivers range from 6000 to ~ 3,000,000 Bq/m2 (Tsuji et al., 2014; Saito et al., 2015). The Abukuma River (catchment area, ~ 5400 km2; total length, 239 km; mean annual freshwater discharge, 52 m3/s) flows through Fukushima and Miyagi prefectures and discharges into northeastern Sendai Bay located in the northern area from FNPP. It has the largest catchment area in Fukushima Prefecture of any river. The Kuji River (catchment area, 1490 km2; total length, 124 km; mean annual freshwater discharge, 26 m3/s) and the Naka River (3270 km2; 150 km; 74 m3/s) flow through Fukushima, Tochigi, and Ibaraki prefectures. The Tone River flows through the Kanto region and has the largest

catchment area (16,840 km2) among these four rivers. Its annual freshwater discharge is 237 m3/s. All of these rivers have “salt-wedge” or “partially mixed” estuaries (water depth, ~ 10 m), in which a layer of freshwater floating on top of seawater gradually thins in the seaward direction and a wedge of seawater on the bottom of the estuary thins in the landward direction. Moreover, when these rivers flow into the Pacific Ocean, mixing occurs suddenly when the rivers enter the ocean itself, and then the suspended particle load is immediately exposed to seawater with high salinity. 2.2. Sampling We collected 60 L of river water in a clean 20-L high-density polyethylene bucket from one station in each river (Fig. 1). The water was filtered within 24 h of collection through a 0.45 μm Millipore membrane filter, and the filtrates were stored in the dark in 20-L polyethylene bottles until analysis (see Section 2.6). The filters, which contained suspended particles from the water samples, were air-dried at room temperature and then each was stored in a 100-mL polyethylene bottle until analysis (Section 2.5). Samples of surface soil or sediment (to ~5 cm depth) from the riverbank and the river bottom were collected with a shovel at four stations along each river (except for the Kuji River) and stored in polyethylene bottles (Fig. 1, Table 1). At each station we also measured the conductivity of the water in contact with the sediment to confirm that it was not affected by seawater. In the laboratory, the collected soil and sediment samples were air-dried and then sieved through a 74-μm mesh sieve. We considered these sieved particles to be potential sources of the suspended particle load in each river (Stumm and Morgan, 1996; Martino et al., 2004). The sieved particles were stored in the dark for later analysis (Sections 2.3 and 2.5) and use in the particle–seawater partitioning and desorption (P&D) experiment (Section 2.4). We collected seawater samples from the Pacific Ocean off Onjuku-machi, Chiba Prefecture, near our laboratory. A total of about 400 L of seawater (salinity, 31) was collected in clean 20-L highdensity polyethylene buckets. The seawater was filtered through a 0.45-μm Millipore membrane filter, and the filtrate was stored in 20-L polyethylene bottles at room temperature (~ 20 °C) in the dark until the P&D experiment (Section 2.4) and seawater analyses (Sections 2.6 and 2.7). 2.3. Particle characterization Concentrations of Al, Mn, Fe, and stable Cs in the sieved particles were determined by inductively coupled plasma mass spectrometry (ICP-MS: 8800 Triple Quadrupole ICP-MS, Agilent Japan, Tokyo) after full digestion by HNO3 and HF as described by Takata et al. (2010). To measure the organic carbon (OC) and organic nitrogen (ON) contents, 1 g of sieved particles was first treated with 2 mL of 1 M HCl to remove inorganic carbonates, and then air-dried at room temperature. Then, the OC and ON concentrations were measured with an elemental analyzer (EuroEA 3000, EuroVector). To estimate the concentration of lithogenic Si, we first estimated the biogenic Si content, as described by DeMaster (1981), by digesting about 1 g of sieved particles in a 1 M Na2CO3 solution for 5 h at 85 °C. Then we determined the lithogenic Si concentration by weighing the sieved particles before and after HF extraction. 2.4. Particle–seawater partitioning and desorption experiment The protocol of the P&D experiment is outlined in Fig. 2. Sieved particles (~10 g) were transferred to 10-L bottles filled with 5 L of filtered seawater to achieve a particle concentration of about 2000 mg/L, which is close to the reported suspended particle concentrations in Japanese rivers after heavy rainfall events (e.g., Tama River, ~2000 mg/L, Nihei et al., 2008; Ara River, 100–2100 mg/L, Toda et al., 2000). Before

H. Takata et al. / Marine Chemistry 176 (2015) 51–63

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Fig. 1. (A) Map of eastern Japan showing the locations of the four rivers. Satellite views of each river showing the locations of the sampling stations: (B) Abukuma, (C) Kuji, (D) Naka, and (E) Tone rivers. River water and suspended particles were collected at the stations shown in red.

the experiment, the concentration of 137Cs was 3.2 ± 0.4 mBq/L and that of stable Cs (i.e., 133Cs) was 2.72 ± 0.038 nmol/L in the seawater (Sections 2.6 and 2.7). Five subsamples from each station were used in the experiment so that we could examine temporal differences in the dissolved concentrations of the two isotopes. Thus, the amount of dissolved Cs in the seawater was measured 30 min, 6 h, 1 d, 3 d, and 7 d after the addition of the sieved particles. During the experiment, the bottles containing the filtered seawater and sieved particles were shaken in the dark at 20 °C. The pH was nearly constant at 8.0 ± 0.2 during the experimental period. At the end of the measurement period, the

particulate and dissolved phases were separated by filtration through a pre-weighed 0.45-μm pore size Millipore membrane filter. The filtrates were then placed into 5-L polyethylene bags and acidified to pH b 2 by adding 10 mL of 15 M HNO3, and then the bags were stored at room temperature in the dark until analysis for stable Cs, 134Cs, and 137Cs. Before this experiment, we measured the desorption of 137Cs from the sieved particles in river water with a salinity of b0.1 (2000 mg of sieved particles per liter of river water). We found that the percentage of desorbed radiocesium in this case was negligibly low (b 0.1%), consistent with previously published reports (e.g., Li et al., 1984).

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size Millipore membrane filters. It is important to consider the effect of dilution, which might lead to the continued release of radiocesium by sieved particles. Although the initial sieved particle concentration might be 2000 mg/L, once it enters the sea the dilution is greater; thus, to reach the same equilibrium, a greater amount of radiocesium would be desorbed from the sieved particles. Decreases of the partition coefficients of Cu, Zn, Ag, and Cs associated with elevated particle concentrations in estuarine environments are well documented (Turner, 1996; Benoit et al., 1994; Li et al., 1984; Santschi et al., 1997; Takata et al., 2012; Tang et al., 2002).

Table 1 Sampling station locations and sampling dates. Study site

Stationa

North latitude

East longitude

Date

Abukuma River

A-1 A-2 A-3 A-4 K-1 K-2 K-3 K-4 N-1 N-2 N-3 N-4 T-1 T-2 T-3 T-4

38°4′49.95″ 38°4′27.61″ 38°5′48.98″ 38°5′39.80″ 36°28′55.01″ 36°29′22.60″ 36°29′45.43″ 36°29′36.92″ 36°20′19.10″ 36°21′4.02″ 36°21′51.12″ 36°21′9.61″ 35°44′44.32″ 35°45′18.72″ 35°45′14.81″ 35°46′24.14″

140°54′25.21″ 140°54′28.30″ 140°52′28.12″ 140°52′46.73″ 140°36′6.63″ 140°35′19.03″ 140°34′41.10″ 140°34′5.32″ 140°34′54.62″ 140°33′48.57″ 140°32′42.59″ 140°33′33.19″ 140°49′05.46″ 140°48′01.45″ 140°47′16.92″ 140°45′12.58″

5 Dec. 13 5 Dec. 13 5 Dec. 13 5 Dec. 13 20 Nov. 13 20 Nov. 13 20 Nov. 13 20 Nov. 13 20 Nov. 13 20 Nov. 13 20 Nov. 13 20 Nov. 13 29 Oct. 13 29 Oct. 13 29 Oct. 13 29 Oct. 13

Kuji River

Naka River

Tone River

2.5. Determination of 134Cs and 137Cs in the sieved particles and in the suspended particles from river water Before the P&D experiment, an aliquot (10 g) of sieved particles was placed in a plastic container and 134Cs and 137Cs activities of the particles were analyzed by nondestructive gamma-ray spectrometry with a coaxial high-purity Ge detector (most were analyzed with the Canberra GC4018 detector, which has a relative efficiency of 40%). The counting system was calibrated against a mixed standard volume source containing 10 radionuclides. The detection limits of 134Cs and 137Cs (calculated as three times the background fluctuation) were approximately 3–10 Bq/kg-dry for a counting period of several tens of thousands of seconds. The 134Cs and 137Cs activities of the suspended particles remaining on the filters after filtration of the river water samples were measured in the same manner.

a At the stations shown in boldface, soil samples were collected from the riverbank. The others were river sediments.

We also examined the effect of the particle concentration in the seawater on the experimental results by performing the experiment with four different concentrations of sieved particles from station A-4 in seawater (20, 50, 100, 500, and 2000 mg/L). The initial concentration of 137Cs in the sieved particles was 3700 ± 375 Bq/kg-dry. The particles were added to 5 L of filtered seawater, and then after 7 d, the particulate and dissolved phases were separated by filtration through 0.45-μm pore

Sieved particles (SP) (<74 µm) Determination of stable Cs by ICP-MS

10g

10g

10g

10g

Filtered seawater (<0.45 µm)

10g

Determination of radiocesiums by gamma-ray spectrometry

5L seawater

5L seawater

5L seawater

5L seawater

5L seawater

5L seawater Control

30 m of shaking

6 h of shaking

1 d of shaking

7 d of shaking

3 d of shaking

7 d of shaking

Filtration by 0.45 µm filter Determination of radiocesiums and stable Cs

Symbols Equation Desorbed fraction (%) =

( Cdt

Cd0 ) × Vsw Cp0 × Vp

× 100

Cdt : Dissolved Cs concentration at a time interval after the addition (Bq/L or nmol/L) Cd0 : Dissolved Cs concentration in control seawater (Bq/L or nmol/L) Vsw : seawater volume (5L) Cp0 : Cs concentration in SP before the addition (Bq/kg-dry or µmol/kg-dry) Vp : Weight of SP added (0.01 kg)

Fig. 2. Schematic diagram of the protocol of the partitioning and desorption experiment and the equation used to calculate the desorbed fraction. SP, sieved particles.

H. Takata et al. / Marine Chemistry 176 (2015) 51–63

2.6. Determination of 134Cs and 137Cs in river water and seawater

55

We analyzed river water samples (10 L) and seawater samples (5–10 L) for 134Cs and 137Cs activities in the same manner. Seawater samples were analyzed both before and after the P&D experiment. The pH of the water sample was adjusted to about 1 with 15 M HNO3 , 0.26 g of CsCl was added as a carrier, and the sample was stirred for at least 1 h. Then, Cs was co-precipitated with ammonium phosphomolybdate (AMP, KANSO Technos Co. Ltd.) (2.5 g-AMP/5 L) while the mixture was stirred for 1–2 h. Radiocesium present in the AMP as an impurity was 0.05 mBq/g-AMP (KANSO Technos Co, Ltd.). The Cs-AMP precipitate was allowed to settle (6–16 h) and then was collected onto a glass fiber filter (5C, ADVANTEC Co. Ltd.) and washed with nitric acid. The AMP-Cs compound was air-dried for at least 1 week, and its weight yield was determined gravimetrically. In all samples the yield exceeded 95%. Dried AMP/Cs (2.5–5 g) was put into a 4-mL Teflon tube, and 134Cs and 137Cs activities were counted in a germanium well detector (ORTEC GWL-90-15) for several tens of hours. The detection limits for 134Cs and 137Cs were calculated as three times the value of the statistical counting error. The minimum detection limits of 134Cs and 137Cs were 2.9 and 1.3 mBq/L, respectively, in the water sample from the Kuji River (134Cs, below the detection limit; 137Cs, 1.9 mBq/L), with a counting time of 25,000 s. The maximum detection limits of 134Cs and 137Cs were 14 and 7.9 mBq/L, respectively, in the seawater sample used for the P&D experiment (134Cs, 157 mBq/L; 137 Cs, 330 mBq/L). The concentrations of 134Cs and 137Cs were decaycorrected to the sampling date.

[concentrated HNO3 diluted with deionized water (18 MΩ) from a Milli-Q water (MQW) purification system (Millipore Corporation, Tokyo, Japan) to pH 1.5]. The rinse solution was also filtered. The residue on the filter was dissolved in 20 mL of 0.15 M NH4OH eluent, which had first been added to the 100-mL polyethylene bottle to dissolve any AMP adhering to the bottle walls. The eluate was passed through ~2 mL of ion exchange resin (AG50W-8X, Bio-Rad, Tokyo, Japan), which was packed in a polypropylene syringe mounted in a column stand (Muromachi Technos Co. Ltd.), at a gravity flow rate of 1 mL/min. Next, the resin was washed with 20 mL of MQW to remove any metals (e.g., Mo). Cs adsorbed onto the resin was extracted with 4 mL of 5 M HNO3 eluent and collected in a 20-mL high-density polyethylene vial. Then the solution was diluted with MQW to make a 1 M HNO3 solution (final volume about 20 mL). Cs in the solution was measured by ICP-MS. The instrumental detection limit was defined as three standard deviations of the blank solution value (0.0022–0.0030 nmol/L, n = 5). We also determined a procedural blank value based on the separation of 20 mL of aqueous AMP solution (n = 5), made by adding 3 mg of AMP to a 20-mL aliquot of HNO3-acidified MQW adjusted to a pH of 1.5. This aliquot was then processed in the same manner as the water samples. The procedural blank value for our method was about 0.022 nmol/L, which is less than 10% of the concentration range of Cs in coastal waters (about 2.2–3.0 nmol/L). The detection limit for the procedural blank, defined as three times the standard deviation of the blanks, was 0.0075 nmol/L (sample volume, 20 mL). The error was calculated as the standard deviation of three replicate samples measured by ICP-MS.

2.7. Determination of stable Cs in seawater

3. Results and discussion

The stable Cs concentration was determined in a subsample (~20 mL) of each seawater sample before and after the P&D experiment following the analytical procedure of Takata et al. (2013). In brief, 3 mg of AMP (Kishida Kagaku Co.) was added to 20 mL of seawater that had been acidified with 15.3 M HNO3 (pH ~ 1) in a 100-mL polyethylene bottle. Then the mixture was filtered through a membrane filter (b0.2 μm), and the bottle was washed with 20 mL of rinse solution

3.1. Concentrations of 137Cs and stable Cs in the sieved particles The 137Cs concentration in the sieved particles varied significantly with respect to sampling location, ranging from 16 to 3700 Bq/kg-dry (Fig. 3, top). The northernmost stations along the Abukuma River and the southernmost stations along the Tone River showed the greatest influence by the accident, and 137Cs concentrations were not necessarily

137

Cs (Bq/kg-dry)

5000 4000

Abukuma Tone

3000

Kuji

2000

Naka

1000

Stable Cs (µmol/kg-dry)

0 A-1 A-2 A-3 A-4

K-1 K-2 K-3 K-4

N-1 N-2 N-3 N-4

T-1 T-2 T-3 T-4

A-1 A-2 A-3 A-4

K-1 K-2 K-3 K-4

N-1 N-2 N-3 N-4

T-1 T-2 T-3 T-4

40 30 20 10 0

Fig. 3. Activities of 137Cs and stable Cs concentrations in sediments collected from along the Abukuma (stations A-1–A-4), Kuji (stations K-1–K-4), Naka (stations N-1–N-4), and Tone (stations T-1–T-4) rivers.

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higher at stations closer to the FNPP. Stable Cs concentrations in the sieved particles, which ranged from 9.0 to 26 μmol/kg-dry (average, 17 ± 5.1 μmol/kg-dry), were smaller than the 137Cs concentrations by a factor of 230. We considered three factors as possible explanations for the observed variability in the 137Cs concentration of the sieved particles. First, the airborne radionuclides deposited onto the land might vary spatially (Yasunari et al., 2011). For example, after the Chernobyl Nuclear Power Plant accident, the behavior of radiocesium changed with deposition distance from the power plant, suggesting heterogeneity of deposition of the radioactive aerosols on the soil surface (Krouglov et al., 1998). After the FNPP accident, the airborne 137Cs-bearing particles were mainly sulfate aerosols (Kaneyasu et al., 2012), although some were spherical, water-insoluble particles containing Fe, Zn, and Cs (Adachi et al., 2013). In remote areas, different physical and chemical properties of the radioactive aerosols would affect the initial distribution of radionuclides in the environment and their particular behavior at local sites. Second, the observed variability might reflect differences in the penetration depth of 137Cs into the surface soil layer at different sites, or radiocesium-bearing surface soil may have been eroded at some sites by surface runoff during heavy rainfall events that occurred in the sampling area after the accident (Japan Japan Meteorological Agency, 2013). Third, the variability might be due to the adsorbed radiocesium having different chemical and physical characteristics in the soils at the different stations (Cremers et al., 1988; Hird et al., 1996; Valcke and Cremers, 1994).

3.2. Desorption of Cs from sieved particles added to seawater (P&D experiment) The half-life of 134Cs is relatively short (2 y). Therefore, all detectable Cs in the environment after the FNPP accident was derived from the accident. In several of the seawater samples, 134Cs was already below the detection limit. As a result, it is difficult to evaluate the behavior of this isotope. Fortunately, there was a significant relationship between desorbed fraction of 134Cs and that of 137Cs for the P&D experiments (Fig. 4), suggesting that 137Cs have a similar behavior with 134Cs. In addition, before the accident, concentrations of 137Cs in soils, river water and seawater in Japan are thought to be negligible low, so detected 137Cs in this study was mainly derived from the FNPP accident. Therefore, we can discuss the desorptive behavior of FNPP-derived radiocesium on the basis of our measured 137Cs data.

200 134

137

Cs = 0.45x Cs - 0.95 r = 0.99, n = 51, p < 0.0001

180 160

134

Cs (mBq/L)

140 120 100 80

Abukuma River Kuji River Naka River Tone river

60 40 20 0 0

100

200

300

400

3.2.2. Time-series variation of the desorbed Cs fraction Time series of the desorbed 137Cs fraction (Fig. 6) in the experimental results show that 137Cs was rapidly desorbed when the sieved particles were added to the filtered seawater, and the desorbed fraction reached a constant value within 30 min. Rapid desorption has been well documented previously (e.g., McKinley et al., 2004; Liu et al., 2004; Wang et al., 2010; Zachara et al., 2002). As shown by complexation models described in these studies, there are two distinct populations of sorption sites on particle surfaces, and the population with the faster desorption rate is higher than that with the slower rate; as a result, the overall desorption rate is fast. The desorbed 137Cs fraction for the entire experimental period ranged from 1.4% to 4.2% for Abukuma River, from 0.75% to 4.5% for Kuji River, from 2.3% to 4.5% for Naka River, and from 3.8% to 6.6% for Tone River. By comparison, salinity-induced desorption of 2% to 30% of radiocesium within one day has been reported for Baltic Sea sediments (Knapinska-Skiba et al., 2001) and Japanese soils collected

Percentage of desorbed137Cs from SP (%)

134

3.2.1. Influence of particle concentration on desorption of Cs from sieved particles in the P&D experiment The desorbed 137Cs fraction remained nearly constant (within 1.5–2.5%) for sieved particle concentrations between 20 and 2000 mg/L (Fig. 5). This result suggests that, even with a sieved particle concentration of 2000 mg/L, desorption of all of the desorbable 137Cs occurred when the particles were added to the seawater. To investigate desorption occurring during the transport of suspended particles in the rivers, it would be necessary to consider how changes in the suspended particle concentration affected the desorbed 137Cs fraction. However, the results of our P&D experiment provide convincing evidence that the desorbed fraction was not related to the particle concentration in seawater. We concluded, therefore, that we could disregard changes in the percentage of 137Cs desorption with increasing particle concentrations, at least up to concentrations of 2000 mg/L, which is the maximum concentration that has been observed around the study area after heavy rain (Toda et al., 2000). Therefore, we inferred that P&D experiments could be carried out with a sieved particle concentration of 2000 mg/L-seawater for more accurate quantification of the desorbed fraction. The minimum 137 Cs concentration in the sieved particles used in the experiment was 154 Bq/kg-dry (sample K-3); thus, if a low particle concentration had been used in the experiment, accurate measurement of the radioactivity of the desorbed fraction would not be possible.

500

5

4

3

2

1

0 10

100

1000

5000

SP (mg/L)

137

Cs (mBq/L)

Fig. 4. Relationship between desorbed 134Cs and desorbed 137Cs in the results of the P&D experiment.

Fig. 5. Relationship between the desorbed 137Cs fraction and the concentration of sieved particles (SP) in seawater in the P&D experiment with different sieved particle concentrations. The sieved particles used in this experiment were from station A-4 (Abukuma River).

H. Takata et al. / Marine Chemistry 176 (2015) 51–63

57

10

A-1

8

A-4

A-2 137

6

Cs Stable Cs

Abukuma

4

Percentage of desorped Cs from SP (%)

2 0 0

30

60

90 120 150 1800

30

60

90 120 150 1800

30

60

90 120 150 180

10

K-1

8

K-3

K-2

Kuji

6 4 2 0 0

30

60

90 120 150 1800

30

60

90 120 150 1800

30

60

90 120 150 180

10

N-1

8

N-4

N-3

Naka

6 4 2 0 0

30

60

90 120 150 1800

30

60

90 120 150 1800

30

60

90 120 150 180

10

T-4

T-2

T-1

8

Tone

6 4 2 0 0

30

60

90 120 150 1800

Time (h)

30

60

90 120 150 1800

Time (h)

30

60

90 120 150 180

Time (h)

Fig. 6. Temporal changes in the desorbed fractions of 137Cs and stable Cs.

near the FNPP (Tagami and Uchida, 2013). Desorption of Cs occurs in seawater because of the large amounts of K+ and NH+ 4 found in seawater replace adsorbed Cs in an ion exchange reaction (Evans et al., 1983; Pardue et al., 1989; Staunton and Roubaud, 1997). Desorption losses of ~15% of 137Cs bound to river sediments after seawater contact have been reported near the mouth of the Columbia River (Robertson et al., 1973), in the Hudson River estuary (Jinks and Wrenn, 1976), and in Bombay Harbor (Patel et al., 1978). The K content of seawater (8.1–11.2 mmol/L) when salinity is N 30 (e.g., Takata et al., 2010) is two orders of magnitude higher in seawater than in most river water. The high concentration of K in seawater can account for the release (desorption) of 137Cs from the sieved particles. Although our P&D experiments were conducted with aliquots of seawater with a uniform salinity (31) and K concentration, the desorbed fraction of the sieved particle samples varied from 0.75% to 6.6% (average, 3.3%). This variability may reflect differences in the characteristics of the sieved particles used in the experiment.

3.3. Relationship between desorptive behaviors of stable Cs and radiocesium in river and marine systems We compared the relationships between desorbed stable Cs and desorbed 137Cs from sieved particles in the four river systems (Fig. 7). If the desorption behavior of the two isotopes was similar, we would expect the desorbed fractions of 137Cs and of stable Cs to be linearly correlated. Although the desorbed fractions of both Stable Cs (range: 0–8.9%) and 137Cs (range: 0.75%–6.6%) varied in the similar range, we found a weak relationship between them, indicating that the mobility of stable Cs is not necessarily a natural analogue for the mobility of FNPP-derived radiocesium. The lack of a significant relationship can be attributed to the different origins of the two isotopes in the sieved particles. The 137Cs was likely only adsorbed on the particle surfaces, whereas stable Cs was likely present not only on the surface but also as structural Cs, that is, within the crystal lattices of the various minerals composing the particles. Structural Cs is not readily exchanged.

58

H. Takata et al. / Marine Chemistry 176 (2015) 51–63

3.4. Factors controlling the extent of desorption

Desorbed stable Cs (%)

10

The extent of desorption depends on the kinetics of the reaction relative to the flushing time of the coastal region where the river meets the ocean and the reversibility of the sorption reaction. The flushing time of riverine particles in seawater is controlled by the interplay between river flow, tidal amplitude, and the suspended particle size (Officer and Lynch, 1981; Turner and Millward, 2002), but the kinetics of the chemical reaction are apparently fast (Fig. 6), thereby allowing chemical equilibrium (i.e., the desorption process) to be achieved within the flushing time. With regard to the reversibility of the sorption reaction, among substrates, 137Cs is preferentially adsorbed onto clay minerals, and 137Cs may be strongly immobilized within their matrices (Absalom et al., 1996; Wauters and Cremers, 1996). However, our measurements of desorbed 137Cs fractions ranging from 0.75% to 6.6% of the total 137Cs suggest that the sieved particles had some exchangeable sites. Moreover, because we collected a large volume of seawater (400 L) for evaluating desorption from the sieved particles, ion concentrations in the seawater (i.e., Na+, Mg2+, K+, Ca2+) and the seawater pH should not have differed during the experiment. In contrast, the compositions of the sieved particles used in this study were heterogeneous: in addition to clay minerals, they contained total Fe (0.48–0.80 mol/kg-dry) and total Mn (0.011–0.022 mol/kg-dry) and organic matter (organic carbon plus organic nitrogen, 0.74–8.9 mol/kg-dry) (Appendix, Table A), and differing relative proportions of these components would lead to variability in the availability of exchangeable sites for radiocesium, and thus in the desorbable fraction (Wen et al., 2008; Takata et al., 2012). To evaluate the role of clay minerals, we plotted the desorbed Cs fraction against Si (excluding biogenic silica), Al, Mn, Fe, and OM in the sieved particles used for the P&D experiment (Fig. 8). Here, Al and Si are indicators of clay minerals; Cs can be taken up by these layered minerals only if the aluminosilicate layers are splayed to form interlayer wedges. We expected to find a negative linear relationship

8

6

4

2

Y = 1.69+ 0.68X r = 0.49, n = 59, p < 0.0001 0 0

2

4

Desorbed

6

137

8

Cs (%)

Fig. 7. Relationship between the desorbed fractions of 137Cs and stable Cs obtained from 1 day to 1 week after the addition of sieved particles to seawater during the P&D experiment. The error for 137Cs is one sigma, based on the counting statistics, and that for stable Cs was calculated as the standard deviation of three replicate samples measured by ICP-MS. The red line was fit by the weighted least squares method.

Furthermore, there is a well-known aging effect on Cs desorption (Nyffeler et al., 1984; Rani and Sasidhar, 2012); thus, because the stable Cs had been associated with the particles for a much longer time than the 137Cs, which was derived mainly or entirely from the FNPP accident, it had stronger affinity for the particles.

10

r = 0.72, n = 12, p = 0.0071

r = -0.51, n = 12, p=0.091

r = -0.29, n = 12, p=0.35 8

Desorbed

137

Cs (%)

6 4 2 0 4

6

8

Si (mol/kg-dry)

10

2.0

2.5

3.0

3.5

4.0

Al (mol/kg-dry)

10

0.4

0.5

0.6

0.7

0.8

0.9

1.0

Fe (mol/kg-dry)

r = 0.004, n = 12, p = 35

r = 0.31, n = 12, p = 0.32 8 6 4 2 0 0

2

4

6

8

OM (mol/kg-dry)

10 0.00

0.01

0.02

0.03

0.04

0.05

Mn (mol/kg-dry)

Fig. 8. Relationship between the desorbed fractions of 137Cs and various chemical constituents of the sieved particles. Note that Si does not include biogenic Si, and organic matter (OM) is the sum of organic carbon and organic nitrogen. The dotted lines indicate the 95% confidence limits of the fitted regression line. The error for 137Cs is one sigma, based on counting statistics, and that for Al, Mn, and Fe was calculated as the standard deviation of three replicate samples measured by ICP-MS.

36 74.0 65.0 76.9 30.0 ~100 40.0 99.0 26.6 60.1 28.1 24.9 Nov. 20, 2013 Nov. 20, 2013 Oct. 29, 2013 Kuji River Naka River Tone River

Activities were calculated from the total 137Cs activity and the percentage of 137Cs in the particulate phase.

Jul.–Dec. 2011 Same River

Ohta River Natsui River

a

(%) (mBq/L)

41.3 170.0 66.0 29.2 284.0 49.0 89.0 25.0 673.0 4.7 5.2 8.3 26.3 126 ± 2.0 43 ± 4.0 22.4 ± 0.6 86.0 ± 6.0 – 35.6a 25.0a 179a 2.8 ± 0.2 1.5 ± 0.2 2.1 ± 0.2

(mBq/L) (mBq/L)

15.0 ± 1.1 44.0 ± 3.0 23.0 ± 2.0 6.7 ± 0.6 198 ± 7.0 – 53.4a 0.25a 494a 1.9 ± 0.4 3.8 ± 0.5 6.3 ± 0.6 2307 21,000 9773 2241 ± 59 57,333 – – – – 1413 ± 111 39 ± 5 89 ± 11

(Bq/kg-dry) (mg/L)

11.4 6.0 4.4 10.0 1.5 – – – – 2.0 37.4 23.2 140°52′11″ 140°38′36.96″ 140°38′36.96″ 140°52′28.12″ 140°59′25.44″ – – – – 140°34′5.32″ 140°33′33.19″ 140°45′12.58″

(E) (N)

Aug. 5, 2012 Sep. 2012 Jan. 2013 Dec. 5, 2013 Jun. 2013 Jun.–Dec. 2011 Abukuma River

38°5′45″ 37°53′34.08″ 37°53′34.08″ 38°5′48.98″ 37°36′11.16″ – – – – 36°29′36.92″ 36°21′9.61″ 35°46′24.14″

Percentage of 137Cs in particulate phase Total 137Cs in river water Particulate 137Cs in river water Dissolved 137Cs in river water 137 Cs in suspended particle

Suspended particle Longitude Latitude Sampling date Study site

3.5.1. Contribution of desorption to the dissolved phase The desorbed 137Cs fraction relative to total 137Cs in the sieved particles ranged from 0.75% to 6.6%; thus, most of the 137Cs remained in the sieved particles. However, during the P&D experiments, the dissolvedphase 137Cs activity in the seawater 7 d after the addition of the sieved particles increased from 3 mBq/L to 122–141 mBq/L (Abukuma samples), 6.8–56 mBq/L (Kuji samples), 31–68 mBq/L (Naka samples), and 26–414 mBq/L (Tone samples). Even though the desorbed 137Cs fraction comprised only a few percent of the total 137Cs in the sieved particles, under the same experimental conditions the contribution of desorption to dissolved 137Cs in seawater increased by a factor of 3 to 100, depending on the 137Cs activity of the sieved particles (Fig. 3). The mean 137Cs activity (1055 Bq/kg-dry) in the sieved particles measured in this study was significantly higher than 137Cs activities in coastal sediments (Otosaka and Kobayashi, 2013; Otosaka and Kato, 2014; Kusakabe et al., 2013). It can be assumed that post-depositional remobilization (desorption) of FNPP-derived 137Cs in the relatively more contaminated suspended particles after they reach the sea is a

Table 2 Concentrations of 137Cs in river water and suspended particles and related data.

3.5. Effect of desorption on the export flux of 137Cs from rivers to the marine environment

Reference

between the desorbed 137Cs fraction and the Al or Si content because the exchangeability of the clay minerals declines when 137Cs is strongly adsorbed onto them. We did not find a significant negative relationship between the desorbed 137Cs fraction and the Si concentration, but we found weak negative correlations between 137Cs and the Al concentration (|r| = 0.51 P = 0.091, Fig. 8). Organic matter and oxides/hydrous oxides of Fe and Mn are also important components of sediment with respect to sorption–desorption interactions with 137Cs (Thørring et al., 2014; Valcke and Cremers, 1994). We found no relationship between the desorbed 137Cs isotope fractions and Mn, but the desorbed 137Cs fraction increased with an increasing Fe content (r = 0.72 for 137Cs). The positive relationship between 137Cs desorption and the Fe content suggests that Fe oxides/ hydrous oxides may control to some extent the exchangeability of 137 Cs. Dumat et al. (1997) reported that iron oxide does not play a role in the adsorption of Cs onto clay minerals in soil. However, they conducted their adsorption experiment with synthetic clay minerals coated with iron oxides or organic materials (e.g., fulvic acid), whereas we used natural soils and sediments with heterogeneous compositions. These differences in the nature of the iron oxides might explain our different results. The organic matter content was not significantly related to the desorbed Cs fraction of either isotope (Fig. 8). The affinity of Cs for organic matter and interactions between them depend primarily on the molecular weight of humic substances, their cation exchange capacity, and their pH (e.g., Dunigan and Francis, 1972; Rigol et al., 2002; Sheppard et al., 1980; Valcke and Cremers, 1994), and the presence of organic matter has even been reported to inhibit Cs adsorption (Bunzl and Schultz, 1985; Maguire et al., 1992). In general, organic matter can have different origins, such as terrestrial, riverine, or autochthonous (e.g., Das et al., 2008; Graham et al., 2001), and its properties can also be modified by biological and microbial activities. In this study, we observed relatively high organic matter contents at stations K-1 and T-2, but the desorbed Cs fractions (e.g., 137Cs) differ considerably between those two stations (K-1, 1.4%; T-2, 6.6%; Fig. 8). This difference may reflect different compositions (or sources) of the organic matter. Further investigation is necessary to evaluate the exchangeability of Cs in relation to the organic matter composition. We found no statistically significant correlations between the desorbed 137Cs fraction and the measured element concentrations (except Fe contents). It is probable that the chemical heterogeneity of the sieved particles increased the affinity of 137Cs for sorption sites on their surfaces, but independent evaluation of the importance of each component is difficult.

59 Sakaguchi et al. (2015) Tsuji et al. (2014) Tsuji et al. (2014) This study Tsuji et al. (2014) Nagao et al. (2013) Nagao et al. (2013) Nagao et al. (2013) Nagao et al. (2013) This study This study This study

H. Takata et al. / Marine Chemistry 176 (2015) 51–63

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H. Takata et al. / Marine Chemistry 176 (2015) 51–63

secondary source of radionuclides in coastal seawater near the FNPP and in marine biota. Therefore, we consider that, in addition to dissolved 137 Cs in fresh river waters, a main pathway of dissolved 137Cs input from rivers to the coastal ocean is desorption from the suspended particles with accumulated 137Cs in river flow. The higher the 137Cs concentration is in a river's suspended particle load, the larger will be their contribution of remobilized 137Cs, and the more they will contribute to seawater contamination in coastal regions. Similarly, in the Danube–Black Sea mixing zone (Gulin et al., 2013) and in the southern Baltic Sea (Knapinska-Skiba et al., 2001), the total (particulate plus dissolved phase) 137Cs concentration increases continuously from river to marine waters, but the fraction of particulate 137Cs decreases along the salinity gradient. 3.5.2. Evaluation of the export flux of 137Cs from rivers to the marine environment It is worthwhile to evaluate the impact of 137Cs desorption on the 137 Cs budget in the coastal area, because significant desorption would add to the radiological risk to the environment. We calculated the export flux of 137Cs via the rivers (i.e., the dissolved phase in river water plus the desorbable Cs fraction in riverine suspended particles) from the riverine water discharge of dissolved 137 Cs activity in the river water, and the experimentally estimated desorbed 137Cs fraction (0.75–6.6%) from particulate 137Cs in the river water (Table 2). The estimated 137Cs fluxes are listed in Table 3, together with results for the rivers near the estuarine and coastal areas after the FNPP accident reported by previous studies (Nagao et al., 2013; Sakaguchi et al., 2015; Tsuji et al., 2014). The estimated 137Cs fluxes (desorbed fraction plus that dissolved in the river water) in the Abukuma and Kuji rivers in this study were 72–86 MBq/d and 4.4–4.6 MBq/d, respectively. These ranges are respectively 2.5–21.9% and 1.1–9.9% larger than 137Cs fluxes calculated by simply multiplying the dissolved 137Cs activity in the river water by the discharge of the two rivers (Abukuma, 71 MBq/d; Kuji, 4.2 MBq/d) (Table 3). In contrast, although the Naka and Tone rivers had relatively large amounts of suspended particles, the increase rate of desorbed 137 Cs to the 137Cs flux in these rivers was less than 0.3%. The differences in the contribution of desorbed 137Cs among the four rivers reflect the differences in the suspended particle 137Cs concentration among them. The dissolved 137Cs concentration in the four rivers varied within a factor of 3 (i.e., Abukuma, 6.7 mBq/L; Kuji, 1.9 mBq/L; Naka, 3.8 mBq/L; and Tone, 6.3 mBq/L), but the 137Cs concentration in the suspended particles was about 2 orders of magnitude larger in the Abukuma (2241 Bq/kg-dry) and Kuji (1413 Bq/kg-dry) rivers than in the Naka

(39 Bq/kg-dry) and Tone (89 Bq/kg-dry) rivers (Table 2). The concentrations of suspended particles in the Abukuma and Kuji rivers (10 and 2.0 mg/L, respectively) were significantly lower than those in the Naka and Tone rivers; as a result the percentage of 137Cs in the particulate fraction of the Abukuma and Kuji rivers was high relative to their total 137Cs activity (76.9% and 60.1%, respectively; Table 2). We obtained similar trends for the export flux and desorbable 137Cs fraction in the Abukuma River when we used previously published data to calculate them (Sakaguchi et al., 2015; Tsuji et al., 2014); that is, although the suspended particle concentrations were low (4.4–11.4 mg/L), the 137Cs concentrations in the particles were remarkably high (2307–21,000 Bq/kg-dry), resulting in a high percentage contribution of desorbable 137Cs to the export flux. For a desorbable 137Cs fraction of 6.6%, the estimated 137Cs export flux in the dissolved phase and the desorbable fraction from the Abukuma River was 175–547 MBq/d, and the increase rate of the desorbed fraction to the export flux ranged from 11.6% to 18.9% (Table 3). These results indicate that in rivers whose catchments are highly contaminated with FNPP-derived 137Cs, such as the Abukuma and Kuji rivers, a considerable amount of 137Cs is remobilized by desorption from suspended particles. According to data collected by the Japan Meteorological Agency (http://www.data.jma.go.jp/obd/stats/etrn/index.php), no heavy rains occurred in the catchment areas of the four rivers during the three days before the sampling date. Thus, the river water sampling was conducted under normal flow conditions. The contribution of the desorbable fraction to the export flux may be less intense under normal flow conditions than under high-flow conditions. After heavy rains, the percentage of 137Cs in the particulate fraction of rivers was high relative to their total 137Cs activity was higher, and thus the contribution of river-transported 137Cs to the marine environment would likely be higher as well. Nagao et al. (2013) proposed that the direct input of suspended solids eroded from the ground surface by heavy rains might explain the higher values measured in the Natsui River (~ 20 km south of FNPP) and the Same River (~ 40 km southwest of FNPP) measured immediately after the FNPP accident. Although there are no data on the suspended particle concentrations in these two rivers, particulate phase 137Cs accounted for more than 99% of the total 137Cs concentration under high-flow conditions in both rivers (Table 2). Although we calculated the contribution of desorption to be a few percent under normal flow conditions in the rivers, after a heavy rain event the contribution increased to more than 50% in the Same River (Table 3). Considering these results for high-flow conditions, the transport of suspended particles with high 137Cs (N1000 Bq/kg-dry) in the Abukuma and Kuji rivers to the marine environment could increase markedly under flood conditions, which

Table 3 Summary of fluvial transport of 137Cs. Study site

Abukuma River

Sampling date

Ohta River Natsui River

Aug. 5, 2012 Sep. 2012 Jan. 2013 Dec. 5, 2013 Jun. 2013 Jun.–Dec. 2011

Same River

Jul.–Dec. 2011

Kuji River Naka River Tone River

Nov. 20, 2013 Nov. 20, 2013 Oct. 29, 2013

a b c d

River discharge

137 Cs export flux for dissolved phasea

Estimated 137Cs export flux for dissolved phase and desorbable fraction b

Increase rate of desorbable 137 Cs for the export flux

(m3/s)

(MBq/d)

(MBq/d)

(%)

121 121 121 121 0.15 238–350c 11–23.7d 238–350c 11–23.7d 26 79 499

157 460 240 71 2.57 – 51–109 5–8 469–1011 4.2 25.6 269.8

159–175 470–547 244–270 72–86 2.6–3.0 – 52–114 9–58 471–1036 4.4–4.6 26.0–26.3 273–276

Flux was calculated by multiplying dissolved 137Cs in the river water by the river discharge. Estimated by assuming that 0.75–6.6% of the 137Cs in the suspended particles would be desorbed in seawater. High flow. Normal flow.

1.3–11.6 2.1–18.9 1.4–12.3 2.5–21.9 0.3–2.9 – N3 N50 N1.0 1.1–9.9 0.3–2.6 0.2–2.2

Reference

Sakaguchi et al. (2015) Tsuji et al. (2014) Tsuji et al. (2014) This study Tsuji et al. (2014) Nagao et al. (2013) Nagao et al. (2013) Nagao et al. (2013) Nagao et al. (2013) This study This study This study

(molN/kg-dry)

0.11 0.22 0.13 0.17 0.48 0.17 0.11 0.12 0.13 0.31 0.16 0.14 0.18 0.69 0.06 0.21 1.5 2.8 1.6 2.2 6.2 2.5 1.4 1.4 1.7 4.0 2.2 1.8 2.2 8.3 0.7 2.9 1828 ± 249 3526 ± 258 331 ± 10 3700 ± 345 1020 ± 95 220 ± 19 154 ± 20 92 ± 7 438 ± 32 276 ± 6 1211 ± 103 344 ± 63 303 ± 16 3221 ± 77 16 ± 3 213 ± 15 780 ± 105 1551 ± 135 137 ± 7 1622 ± 154 464 ± 34 96 ± 12 70 ± 6 44 ± 15 199 ± 11 122 ± 1 538 ± 42 157 ± 24 140 ± 3 1475 ± 48 8±3 101 ± 9

(mol/kg-dry)

0.61 0.58 0.63 0.57 0.52 0.51 0.50 0.49 0.49 0.48 0.52 0.54 0.70 0.61 0.76 0.80

(mol/kg-dry)

0.022 0.019 0.014 0.020 0.011 0.014 0.016 0.016 0.012 0.014 0.016 0.011 0.012 0.015 0.012 0.018

(mol/kg-dry)

3.0 3.4 3.3 3.5 2.9 2.8 2.7 2.8 2.6 2.9 2.9 2.6 2.8 2.1 2.9 2.9

(mol/kg-dry)

7.4 7.9 7.4 7.7 6.8 8.0 8.5 8.3 8.3 7.5 7.3 7.7 8.2 5.9 8.8 7.5

11 17 9 11 18 23 20 17 14 13 26 12 18 16 18 25

(molC/kg-dry)

OC Cs

(Bq/kg-dry) (Bq/kg-dry)

137

Cs 134

Mn Al

Stable Cs

would most likely raise the contribution of desorption to the export flux of 137Cs in the dissolved phase to the sea.

Silicon

Fe

61

4. Conclusions In this study, we showed that the desorption of radiocesium from sieved particles from riverbank soil and river sediment samples, as presumed sources of suspended particles in the river water, contributed to the export flux of dissolved radiocesium, which in the marine environment can easily accumulate in marine biota. Our P&D experiments showed that although the desorbed fractions of both stable Cs and 137Cs varied in almost the same range, they did not necessarily co-vary. The variation in the desorbed fraction among the rivers might be due to the different reactivity of the sorbed Cs (i.e., different sorption sites, such as frayed edge sites), and to the presence of Fe and Mn oxides and hydrous oxides and organic matter in the sieved particles. Our investigation of the factors controlling the desorbed 137Cs fraction indicated that a slight increase in Al (and/or Si) was associated with reduced exchangeability, owing to strong adsorption of 137Cs onto clay minerals. In contrast, increased amounts of Fe seemed to increase the exchangeability of 137Cs. We could not determine, however, the effect of the organic matter content of the particles on their reactivity, because the reactivity would depend on the composition (sources) of the organic matter. Hence, further characterization of the organic matter is necessary. Finally, we estimated the export flux of 137Cs in the dissolved phase from rivers to the coastal marine environment. We found that the higher the 137Cs concentration is in suspended particles, or the higher percentage of 137Cs in the particulate phase relative to the total amount in the river, the larger the contribution of 137Cs desorption to the export flux becomes. Thus, to assess the effect of riverine radiocesium on coastal regions near the FNPP, it is necessary to take into account the flux of the desorbable 137Cs fraction in riverine suspended particles. Acknowledgments We thank Shinichi Yamano (KANSO Co. Ltd.) for technical assistance. We are also grateful for helpful comments on the manuscript from two anonymous reviewers. The marine environmental radioactivity survey is part of a research project carried out for the Nuclear Regulation Authority.

Dec. 5, 2013 Dec. 5, 2013 Dec. 5, 2013 Dec. 5, 2013 Nov. 20, 2013 Nov. 20, 2013 Nov. 20, 2013 Nov. 20, 2013 Nov. 20, 2013 Nov. 20, 2013 Nov. 20, 2013 Nov. 20, 2013 Oct. 29, 2013 Oct. 29, 2013 Oct. 29. 2013 Oct. 29. 2013 (E)

140°54′25.21″ 140°54′28.30″ 140°52′28.12″ 140°52′46.73″ 140°36′6.63″ 140°35′19.03″ 140°34′41.10″ 140°34′5.32″ 140°34′54.62″ 140°33′48.57″ 140°32′42.59″ 140°33′33.19″ 140°49′05.46″ 140°48′01.45″ 140°47′16.92″ 140°45′12.58″

(N)

Tone River

Naka River

Kuji River

Abukuma River

A-1 A-2 A-3 A-4 K-1 K-2 K-3 K-4 N-1 N-2 N-3 N-4 T-1 T-2 T-3 T-4

38°4′49.95″ 38°4′27.61″ 38°5′48.98″ 38°5′39.80″ 36°28′55.01″ 36°29′22.60″ 36°29′45.43″ 36°29′36.92″ 36°20′19.10″ 36°21′4.02″ 36°21′51.12″ 36°21′9.61″ 35°44′44.32″ 35°45′18.72″ 35°45′14.81″ 35°46′24.14″

Date Longitude Station

Latitude

Appendix A. Supplementary data

Study site

Table A The chemical composition and Cs activities of the sieved riverine particles from each sampling station.

(μmol/kg-dry)

ON

H. Takata et al. / Marine Chemistry 176 (2015) 51–63

Table A lists the activities of the Cs radionuclides and the concentrations of Si, Al, Mn, Fe, stable Cs, and organic C and N in the sieved particles from each of the sampling stations. References Absalom, J.P., Crout, N.M.J., Young, S.D., 1996. Modeling radiocesium fixation in upland organic soils of Northwest England. Environ. Sci. Technol. 30 (9), 2735–2741. Adachi, K., Kajino, M., Zaizen, Y., Igarashi, Y., 2013. Emission of spherical cesium-bearing particles from an early stage of the Fukushima nuclear accident. Sci. Rep. 3. http:// dx.doi.org/10.1038/srep02554 (Art. No. 2554.). Bailly du Bois, P., Laguionie, P., Boust, D., Korsakissok, I., Didier, D., 2012. Estimation of marine source-term following Fukushima Dai-ichi accident. J. Environ. Radioact. 114, 2–9. Benoit, G., Oktay-Marshall, S.D., Cantu II, A., Hood, E.M., Coleman, C.H., Corapcioglu, M.O., Santschi, P.H., 1994. Partitioning of Cu, Pb, Ag, Zn, Fe, Al, and Mn between filterretained particles, colloids and solution in six Texas estuaries. Mar. Chem. 45, 307–336. Bunzl, K., Schultz, W., 1985. Distribution coefficients of 137Cs and 85Sr by mixtures of clay and humic material. J. Radioanal. Nucl. Chem. 90, 23–37. Charette, M.A., Breier, C.F., Henderson, P.B., Pike, S.M., Rypina, I.I., Jayne, S.R., Buesseler, K.O., 2013. Radium-based estimates of cesium isotope transport and total direct ocean discharges from the Fukushima Nuclear Power Plant accident. Biogeosciences 10, 2159–2167. Chino, M., Nakayama, H., Nagai, H., Terada, H., Katata, G., Yamazawa, H., 2011. Preliminary estimation of release amounts of 131I and 137Cs accidentally discharged from

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