Behavior of Radionuclides in the Environment III: Fukushima (Behavior of Radionuclides in the Environment, 3) 9811667985, 9789811667985

Citation preview

Kenji Nanba Alexei Konoplev Toshihiro Wada Editors

Behavior of Radionuclides in the Environment III Fukushima

Behavior of Radionuclides in the Environment III

Kenji Nanba • Alexei Konoplev • Toshihiro Wada Editors

Behavior of Radionuclides in the Environment III Fukushima

Editors Kenji Nanba Fukushima University Institute of Environmental Radioactivity Fukushima, Japan

Alexei Konoplev Fukushima University Institute of Environmental Radioactivity Fukushima, Japan

Toshihiro Wada Fukushima University Institute of Environmental Radioactivity Fukushima, Japan

ISBN 978-981-16-6798-5 ISBN 978-981-16-6799-2 https://doi.org/10.1007/978-981-16-6799-2

(eBook)

© Springer Nature Singapore Pte Ltd. 2022 This work is subject to copyright. All rights are reserved by the Publisher, whether the whole or part of the material is concerned, specifically the rights of translation, reprinting, reuse of illustrations, recitation, broadcasting, reproduction on microfilms or in any other physical way, and transmission or information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed. The use of general descriptive names, registered names, trademarks, service marks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. The publisher, the authors, and the editors are safe to assume that the advice and information in this book are believed to be true and accurate at the date of publication. Neither the publisher nor the authors or the editors give a warranty, expressed or implied, with respect to the material contained herein or for any errors or omissions that may have been made. The publisher remains neutral with regard to jurisdictional claims in published maps and institutional affiliations. This Springer imprint is published by the registered company Springer Nature Singapore Pte Ltd. The registered company address is: 152 Beach Road, #21-01/04 Gateway East, Singapore 189721, Singapore

Preface

This is Volume III comprising a part of the set of three volumes Behavior of Radionuclides in the Environment, focusing on Fukushima. The series was initiated by the enthusiasm of Professor Kenji Kato from Shizuoka University whose tremendous contribution was a key to the successful publication of the first two volumes: Volume I—Function of Particles in Aquatic Systems devoted mostly to the behavior of radionuclides in the environment surrounding PA Mayak (Kato et al. 2020), and Volume II—Chernobyl is dedicated to the long-term behavior of Chernobyl-derived radionuclides in the environment (Konoplev et al. 2020). The magnitude 9.0 Great East Japan Earthquake occurred at 14:46 on 11 March 2011 and the following devastating tsunami caused the accident at the Fukushima Dai-ichi Nuclear Power Plant (FDNPP), which led to extensive environmental contamination by several radionuclides, particularly 134Cs (half-life Т1/2 ¼ 2.06 years) and 137 Cs (Т1/2 ¼ 30.17 years). Radiocesium deposition north-west of the FDNPP resulted in a trace of contamination 50–70 km long and 20 km wide (МЕХТ 2012). This has rekindled the interest in the behavior of radiocesium, particularly given the geoclimatic conditions of Japan. During the 10 years since the accident, a number of large-scale national and international programs and projects have been implemented to investigate in detail and gather data about Fukushima-derived radionuclides in environmental compartments. Soon after the Fukushima accident in July 2013 the Institute of Environmental Radioactivity (IER) at Fukushima University was established to meet the challenges associated with environmental radioactivity after the FDNPP accident. Now 10 years after the accident and 8 years after the establishment of the IER, it is time to share achievements and progress. This Volume III—Fukushima is composed of contributions prepared mostly by IER staff members and their collaborators and is designed to review their major achievements in research on the behavior of Fukushima-derived radionuclides, ranging from air transport, mobility and bioavailability in soil-water environment, vertical and lateral migration in soils and sediments, soil-to-plant and soil-to-animal transfer, water-to-aqueous biota transfer, and radiation effects. v

vi

Preface

The studies after the Fukushima accident and comparison with Chernobyl research outcomes have clearly demonstrated that the behavior of accidentally released radionuclides in the environment is governed by radionuclide speciation in fallout or liquid discharges to the ocean and site-specific environmental characteristics. With this in view, Volume III is composed of six parts: Part I—Atmospheric Transport and Fallout of Radionuclides; Part II—Fate and Transport of Radionuclides in Soil-Water Environment; Part III—Radionuclide Behavior in Freshwater Environment; Part IV—Radionuclide Behavior in Coastal and Marine Environment; Part V—Radionuclide Transfer in Agricultural Environment; Part VI—Radionuclide Transfer in Terrestrial Environment. Overall twenty-three individual chapters are included in the Volume prepared by 47 scientists representing the IER and its partnering institutions. The book may be of interest to researchers of environmental contamination, radioecology and radiation safety, as well as students and all those majoring in environmental radioactivity and radioecology. We would be gratified if the main messages of this Volume could be used for the benefit of future generations of scientists. Fukushima, Japan April 30, 2021

Kenji Nanba Alexei Konoplev Toshihiro Wada

References Kato K, Konoplev A, Kalmykov SN (eds) (2020) Behavior of radionuclides in the environment I: function of particles in aquatic system. Springer, 225 p. ISBN 978-981-15-0679-6 Konoplev A, Kato K, Kalmykov SN (eds) (2020) Behavior of radionuclides in the environment II: Chernobyl. Springer, 295 p. ISBN 978-981-15-3567-3 MEXT (Ministry of Education, Culture, Sports, Science and Technology of Japan) (2012) Results of the (i) fifth airborne monitoring survey and (ii) airborne monitoring survey outside 80 km from the Fukushima Dai-ichi NPP. Tokyo. http:// radioactivity.nsr.go.jp/en/contents/6000/5790/24/ 203_0928 14e.pdf

Contents

Part I 1

2

3

Air Dose Rate in Fukushima Prefecture Measured during March 2011: The First Organized Measurement after Fukushima Daiichi Nuclear Power Plant Accident . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Kenji Nanba, Katsuhiko Yamaguchi, Naoaki Shibasaki, Yoshitaka Nagahashi, Kotaro Hirose, Takahide Kurosawa, Katsuhiko Kimura, Tsugiko Takase, Nobuo Shinoda, Akira Tanaka, Hiromasa Ikuta, Dai Oyama, Yoshimasa Koyama, Kencho Kawatsu, Takayuki Takahashi, and Hitoshi Kanazawa Atmospheric Transport and Deposition of Fukushima-Derived Radionuclides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Shigekazu Hirao Airborne Radiation Survey after the Accident . . . . . . . . . . . . . . . . . Tatsuo Torii and Yukihisa Sanada

Part II 4

5

6

Atmospheric Transport and Fallout of Radionuclides

3

9 17

Fate and Transport of Radionuclides in Soil-Water Environment

Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment: Review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Alexei Konoplev, Yoshifumi Wakiyama, Toshihiro Wada, Yasunori Igarashi, Volodymyr Kanivets, and Kenji Nanba Dynamics of Radiocesium Solid–Liquid Distribution Coefficient in Soil and Water Environments . . . . . . . . . . . . . . . . . . . . . . . . . . . Sadao Eguchi and Noriko Yamaguchi An Overview of Fukushima-Derived Strontium Radioisotopes . . . . Ismail M. M. Rahman, Hikaru Sawai, M. Ferdous Alam, and Zinnat A. Begum

33

69 79

vii

viii

7

Contents

Erosion and Redeposition of Sediments and Sediment-Associated Radiocesium on River Floodplains (the Niida River Basin and the Abukuma River as an Example) . . . . . . . . . . . . . . . . . . . . . . . . . . . Valentin Golosov, Alexei Konoplev, Yoshifumi Wakiyama, Maxim Ivanov, and Mikhail Komissarov

Part III

97

Radionuclide Behavior in Freshwater Environment

8

Dynamics of Radiocesium in Urban River in Fukushima City . . . . . 137 Kenji Nanba, Shota Moritaka, and Yasunori Igarashi

9

Temporal Variations in Particulate and Dissolved 137Cs Activity Concentrations in the Abukuma River During Two High-Flow Events in 2018 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 153 Yoshifumi Wakiyama, Alexei Konoplev, Nguyễn Thoa, Takuya Niida, Hirofumi Tsukada, Tsugiko Takase, Kenji Nanba, Valentin Golosov, and Mark Zheleznyak

10

The Effect of Groundwater Bypass at the Fukushima Daiichi Nuclear Power Plant in 2014 by Detailed Facies Analysis and 3D Groundwater Simulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 177 Hikaru Sato

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 199 Mark Zheleznyak, Sergii Kivva, Oleksandr Pylypenko, and Maksim Sorokin

Part IV

Radionuclide Behavior in Coastal and Marine Environment

12

Spatiotemporal Variation of Radiocesium in Coastal and Oceanic Seawater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 255 Hyoe Takata and Yuichiro Kumamoto

13

Spatiotemporal Variation of Radiocesium in Coastal Marine Sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 285 Masashi Kusakabe

14

Cesium Radioactivity in Marine and Freshwater Products and Its Relation to the Restoration of Fisheries in Fukushima: A Decade Review . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 313 Toshihiro Wada, Yoshiharu Nemoto, Tsuneo Fujita, Gyo Kawata, Kyoichi Kamiyama, Tadahiro Sohtome, Kaoru Narita, Masato Watanabe, Shinya Shimamura, Masahiro Enomoto, Shotaro Suzuki, Yosuke Amano, Daigo Morishita, Akira Matsumoto, Yoshiaki Morioka, Atsushi Tomiya, Toshiyuki Sato, Kouji Niizeki, Takashi Iwasaki, Michio Sato, Takuji Mizuno, and Kenji Nanba

Contents

Part V

ix

Radionuclide Transfer in Agricultural Environment

15

Spatial Distribution and Temporal Change of 137Cs Activity Concentration in Dissolved and Suspended Fractions of Irrigation Waters Collected from Fukushima . . . . . . . . . . . . . . . . . 355 Hirofumi Tsukada

16

Mineralogical Factors Controlling the Ability to Retain 137Cs in Andosols in Japan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 365 Atsushi Nakao, Shiori Uno, Ryoji Tanaka, and Junta Yanai

17

Soil Properties Affecting Soil-to-Crop Transfer of FukushimaDerived Radiocesium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 377 Akira Takeda

18

Model of Radionuclide Uptake by Plants via Foliar Pathway: Kyshtym, Сhernobyl, Fukushima . . . . . . . . . . . . . . . . . . . . . . . . . . 389 Boris Prister

Part VI

Radionuclide Transfer in Terrestrial Environment

19

Behavior of Fukushima-Derived Radiocesium in Forest Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 427 Vasyl Yoschenko, Kenji Nanba, Tatsuhiro Ohkubo, and Hiroaki Kato

20

Radiocesium Contamination in Wild Rodents Inhabiting Forested Areas Inside the Evacuated Area in Fukushima, Japan . . 463 Hiroko Ishiniwa, Manabu Onuma, and Masanori Tamaoki

21

Concentrations and Transfer Parameters of 137Cs for Wild Boar . . 473 Donovan Anderson, Hirofumi Tsukada, and Thomas G. Hinton

22

Variation of Cesium-137 Concentration in Wild Boar and Asian Black Bear . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 481 Yui Nemoto

23

Physicochemical Fractions of Radiocesium in the Stomach Contents of Wild Boar and Its Transfer to Muscle Tissue . . . . . . . . 495 Rie Saito and Hirofumi Tsukada

Index . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 507

Part I

Atmospheric Transport and Fallout of Radionuclides

Chapter 1

Air Dose Rate in Fukushima Prefecture Measured during March 2011: The First Organized Measurement after Fukushima Daiichi Nuclear Power Plant Accident Kenji Nanba, Katsuhiko Yamaguchi, Naoaki Shibasaki, Yoshitaka Nagahashi, Kotaro Hirose, Takahide Kurosawa, Katsuhiko Kimura, Tsugiko Takase, Nobuo Shinoda, Akira Tanaka, Hiromasa Ikuta, Dai Oyama, Yoshimasa Koyama, Kencho Kawatsu, Takayuki Takahashi, and Hitoshi Kanazawa

Abstract Extensive survey of dose rate on the surface of the ground in Fukushima Prefecture was carried out from 25 to 31 March 2011, immediately after the accident. The map based on this first organized measurement on the surface of the ground has identified an area of high radiation dose rates in the northwest direction from the Fukushima Daiichi Nuclear Power Plant. This was the one of the responses the local university made during the emergency period after the accident. Keywords Fukushima · Dose rate · Deposition · Local emergency response

K. Nanba (*) · K. Yamaguchi · N. Shibasaki · Y. Nagahashi · T. Kurosawa · K. Kimura · T. Takase · N. Shinoda · A. Tanaka · H. Ikuta · D. Oyama · Y. Koyama · K. Kawatsu · T. Takahashi Faculty of Symbiotic System Science, Fukushima University, Fukushima, Japan e-mail: [email protected] K. Hirose Faculty of Symbiotic System Science, Fukushima University, Fukushima, Japan Faculty of Science and Engineering, Waseda University, Tokyo, Japan H. Kanazawa Faculty of Symbiotic System Science, Fukushima University, Fukushima, Japan Faculty of Engineering, Yamagata University, Yamagata, Japan © The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2022 K. Nanba et al. (eds.), Behavior of Radionuclides in the Environment III, https://doi.org/10.1007/978-981-16-6799-2_1

3

4

1.1

K. Nanba et al.

Introduction

The Fukushima Daiichi Nuclear Power Plant (FDNPP) was hit by the Great East Japan Earthquake at 14:46 and resulting tsunami on 11 March 2011, which caused the loss of external power supply and the inability to fully operate the cooling function for the reactors. The Prime Minister issued a declaration of a nuclear emergency at 19:03 on the same day, based on the Act on Special Measures Concerning Nuclear Emergency Preparedness, due to the increased risk of damage to the reactors and the release of radioactive materials. As a result, at 20:50 on 11 March 2011, the prefectural government issued an evacuation order within a radius of 2 km, and at 21:23 on the same day, the national government issued an evacuation order within a radius of 3 km and an indoor stay order within a radius of 3–10 km from the FDNPP. The evacuation order was expanded by the evening of 12 March to include a 20-km radius of the FDNPP and evacuation orders within a 10-km radius of the Fukushima Daini Nuclear Power Plant. The indoor stay order was also expanded within a radius of 20–30 km from the FDNPP. These orders continued until 22 April when the “deliberate” evacuation was ordered in some area beyond 20 km radius (Fukushima Prefecture 2013). On 16 March, the U.S. Nuclear Regulatory Commission advised the American people to evacuate the area beyond 50 miles (80 km) from the Fukushima Daiichi Nuclear Power Plant. By this time, a series of accidents had occurred at Fukushima Daiichi, including hydrogen explosions in Units 1, 3, and 4 and a failure of the pressurized containment vessel (PCV) in Unit 2 (National Diet of Japan Fukushima Nuclear Accident Independent Investigation Commission [NAIIC] 2012). An increase in the value of dose rate from the monitoring posts set up by Fukushima Prefecture in Fukushima City was recorded from the evening of 15 March, and the value rapidly rose to 24 μSv/h (Fukushima Prefecture 2011). After that, Ministry of Education, Culture, Sports, Science and Technology (MEXT) started monitoring of radiation levels at various points in the prefecture and began to report along with the concentration of radionuclides in the soil; on 18 March, high levels were reported from some points, but due to the limited number of observation points, it was difficult to determine whether each of them was an independent spot with high radiation levels or an extensive area of high radiation levels continuous with other high points. In this context, the first survey of the state of deposition in the prefecture after the accident was an airborne survey. The survey was conducted jointly by MEXT and US Department of Energy (US DOE) from 17 to 19 March and published on 22 March, covering an area of about 30 km from the nuclear power plant (US DOE 2011). The results showed that the dose rate on the ground surface was higher in the northwest direction. However, the resolution of both the dose rate and its spatial distribution was not sufficient to identify the areas that needed to be evacuated.

1 Air Dose Rate in Fukushima Prefecture Measured during March 2011: The. . .

1.2

5

Method

Extensive survey of dose rate on the surface of the ground in Fukushima Prefecture was carried out by the authors, the members of Fukushima University Radiation Measurement Team, which is a group of voluntary faculty members of the Faculty of Symbiotic Systems Science of Fukushima University from 25 to 31 March 2011. We used a NaI scintillation survey meter (TCS-171, ALOKA) or an ionization chamber survey meter (ICS-311, ALOKA) borrowed from the Association for the Promotion of Radiation Application, Tohoku Radiation Science Center, and other institutions. The ionization chamber survey meter was used when the air dose rate exceeded 30 μSv/h. The measurements were carried out at uncovered points at a height of 1 m above the ground surface. The sites to be measured were set at a mesh of 2 km in the Nakadori and Hamadori areas where the dose rate was relatively high, and at a mesh of approximately 10 km in the Aizu and other areas in Fukushima Prefecture. A contour map of dose rate distribution was prepared based on the measurements at 372 locations in the prefecture. The digital maps provided by the Geospatial Information Authority of Japan and a software Surfer 9 (Golden Software Inc.) were used. The radiation dose rate was compensated to the value on 30 March by using the parameter of exponential decrease of the value of the monitoring post set up by Fukushima Prefecture in Fukushima City (Fukushima Prefecture 2011).

1.3

Results and Discussion

The map (Fig. 1.1) based on the first organized measurement on the ground has identified an area of high radiation dose rates in the northwest direction from the FDNPP. At points farther than 20 km from the plant, where no evacuation order had been issued but elevated radiation doses had been reported, it was confirmed that there were elevated radiation doses with an extent sufficient to warrant evacuation orders as a continuous area. These areas were later evacuated as a deliberate evacuation area (Fukushima Prefecture 2013). Maximum dose rate measured was 70 μSv/h. The distribution of dose rate showed that the highlands in 30–50 km from the NPP in northwest direction were boundaries of high deposition, which was explained by deposition process (Sanada et al. 2018). The patches of hotspot deposition were also found between 50 and 60 km from the NPP in west to northwest direction. The radiation measurement team of Fukushima University communicated the measurement results and the resulting dose distribution maps to the municipalities concerned in March. This was because we wanted to prioritize the delivery of the measurement results close to the residents as the parties concerned, and also to inform them of the expected future instructions from the government.

6

K. Nanba et al.

Fig. 1.1 A counter map of the distribution of dose rate measured at 1 m above the ground. The values are calibrated as of 30 March 2011(see text). Counter is shown in white line and values of dose rates (μSv/h) are shown. Measured points are blue dots. Red cross shows the location of the Fukushima Daiichi Nuclear Power Plant and concentric circular lines show the distance from there. Coordinate system is UTM (WGS84, Zone 54)

Fukushima University has started the postnuclear accident emergency responses immediately after the accident. The mapping of the dose rate in this report is one of those responses. Subsequently, the Institute for Environmental Radioactivity (most of members of which contributed to this book) was established in 2013, and the Department of Food and Agricultural Sciences was established in 2019. With the cooperation of the local community, Fukushima University has been engaged in research and education including reconstruction from the nuclear disaster and

1 Air Dose Rate in Fukushima Prefecture Measured during March 2011: The. . .

7

dissemination of lessons learned from the accident to the world. Even now, 10 years after the accident, there are still many issues to be addressed regarding the reconstruction of the region after the Fukushima nuclear accident and the decommissioning of FDNPP. As an university and research institute located in Fukushima Prefecture, we are committed to addressing these issues.

References Fukushima Prefecture (2011) Results of air dose rate monitoring. https://www.pref.fukushima.lg.jp/ sec_file/monitoring/m-1/7houbu0311-0331.pdf Fukushima Prefecture (2013) Record of Great East Japan Earthquake and Steps of Reconstruction . https://www.pref.fukushima.lg.jp/sec_file/koho/e-book/index.html National Diet of Japan Fukushima Nuclear Accident Independent Investigation Commission [NAIIC] (2012) Official report . Tokuma Shoten, 592p Sanada Y, Katata G, Kaneyasu N, Nakanishi C, Urabe Y, Nishizawa Y (2018) Altitudinal characteristics of atmospheric deposition of aerosols in mountainous regions: lessons from the Fukushima Daiichi nuclear Power Station accident. Sci Total Environ 618:881–890 US DOE (2011) Releases radiation monitoring data from Fukushima area. https://www.energy.gov/ articles/us-department-energy-releases-radiation-monitoring-data-fukushima-area. Accessed 22 Mar 2011

Chapter 2

Atmospheric Transport and Deposition of Fukushima-Derived Radionuclides Shigekazu Hirao

Abstract The environmental impacts of radioactive materials released into the atmosphere from the Tokyo Electric Power Company (TEPCO) Fukushima Daiichi Nuclear Power Plant as a result of the tsunami caused by the 2011 Tohoku Earthquake are both wide-ranging and long-lasting. More than a decade after the accident, this review provides an overview of the current understanding of the early environmental impacts via the atmosphere and the related findings regarding the release of radionuclides into the atmosphere. Keywords Atmospheric dispersion · Deposition · Source term · Radionuclides · Fukushima Daiichi Nuclear Power Plant · Monitoring data

2.1

Outline of the Fukushima Daiichi Nuclear Power Plant Accident and Related Atmospheric Release

The 2011 Tohoku Earthquake struck at 14:46 on March 11, and 50 min later, a tsunami with a wave height of approximately 13 m reached the Tokyo Electric Power Company (TEPCO) Fukushima Daiichi Nuclear Power Plant (FDNPP). Although the reactors of Units 1, 2, and 3, which were in operation, automatically shut down immediately after the earthquake, the reactor and turbine buildings were flooded by the tsunami. Although external power supplies were lost because of the earthquake, two emergency diesel generators could be used. However, when some equipment with the generators, such as power transmission and receiving facilities, were flooded by the tsunami, all AC power was lost within the power plant resulting in an extremely critical problem. The loss of power resulted in the failure of the reactor cooling systems, which led to a number of severe accidents, including core meltdown, within the reactors of Unit 1 (from March 11), Unit 3 (from March 13),

S. Hirao (*) Institute of Environmental Radioactivity, Fukushima University, Fukushima, Fukushima, Japan e-mail: [email protected] © The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2022 K. Nanba et al. (eds.), Behavior of Radionuclides in the Environment III, https://doi.org/10.1007/978-981-16-6799-2_2

9

10

S. Hirao

and Unit 2 (from March 14). Although a series of venting operations was conducted, hydrogen explosions occurred in Units 1, 3, and 4, with the explosion in Unit 4 caused by a hydrogen leak from Unit 3. The hydrogen explosions occurred as a result of a loss in the confinement function owing to the rise in pressure of the primary containment vessels (PCVs) and the hydrogen gas leakage from the PCVs. As a result, large quantities of radionuclides were released from the reactors into the atmosphere. Even a decade after the accident, the details of the damage to the reactors are still unclear, with limited information on the amount of radionuclides released into the atmosphere from the reactors, as well as their physical and chemical properties.

2.2

Monitoring of Environmental Radiation

Because of the earthquake, a number of monitoring posts and exhaust stack monitors ceased functioning owing to the direct damage caused by the tsunami, as well as power outages in the vicinity of the plant. Therefore, the only on-site monitoring of radiation took place at the boundary of the site by an operator in a monitoring car. Even if the exhaust stack monitors had been operational, data would still have been limited, as atmospheric releases may have occurred from other locations in the plant. Although roads and other infrastructure beyond the site had also been damaged, the Fukushima Prefecture began emergency radiation monitoring within a 10-km radius on March 12. From approximately 8–12 am on March 12, radioactive iodine and cesium were measured in airborne dust collected from the towns of Okuma and Namie, located in the vicinity of the nuclear power plant. From March 13, the Ministry of Education, Culture, Sports, Science and Technology (MEXT) began to support emergency monitoring, allowing the scope of the monitoring to be expanded to a 30-km radius. Environmental monitoring was performed systematically from March 15 onward, with large-scale contamination recorded to the northwest of the FDNPP. From April onward, the ambient air dose rates were measured at existing monitoring posts on the boundary of the site. However, considering that the majority of atmospheric release occurred in March, there was a lack of information to clarify the state and amount of atmospheric release and dispersion. In addition, as hardly any atmospheric concentration data were available, estimating the source term and the impacts of radiation exposure had large uncertainty. In the decade since the accident, useful data regarding the atmospheric dispersion of radionuclides have come to the fore, including a review of atmospheric monitoring and related atmospheric effects (Hirose 2020). One such data source is the radioactivity analysis of filter tape for suspended particulate matter (SPM), obtained from various SPM monitoring stations. For example, data on cesium-137 and cesium-134 concentrations at 99 points in the Kanto and Tohoku regions from March 12 to 23, 2011, which were not accessible via emergency monitoring, have now become available (Oura et al. 2015; Tsuruta et al. 2018). For both Fukushima and Ibaraki prefectures, for which no SPM data were available, atmospheric

2 Atmospheric Transport and Deposition of Fukushima-Derived Radionuclides

11

concentrations of radioactive cesium, iodine, and noble gases have now been evaluated using the pulse height distribution of NaI(Tl) detectors from air dose rate measurements at monitoring posts (Hirayama et al. 2017; Terasaka et al. 2016; Moriizumi et al. 2019). Furthermore, data on highly radioactive cesium-rich microparticles (CsMPs) in the environment (Adachi et al. 2013; Igarashi et al. 2019) and their spatial distribution (Ikehara et al. 2020) are now becoming available.

2.3

Estimation of Atmospheric Release

Although source term information is essential for analyzing the atmospheric dispersion of radioactive materials, a large amount of uncertainty remains regarding source terms at the time of the Fukushima accident, which hinders important analyses and impact assessments. The following two methods can be used to estimate atmospheric release in the event of a nuclear emergency: (1) estimation by analyzing accident progression in the reactor using severe accident analysis codes or (2) inverse estimation by combining atmospheric dispersion calculations and environmental monitoring data. As data from within the reactors during the FDNPP accident are not available, the second method has been used to estimate the amount of radionuclides released into the atmosphere (Chino et al. 2011, 2016; Terada et al. 2012; Hirao et al. 2013; Katata et al. 2015), based on atmospheric and fallout data. Release rates were obtained by dividing the measured concentrations by the calculated values obtained from atmospheric dispersion simulations, assuming set unit release rates. For periods when nuclide concentrations are not available, release rates are estimated using dose rates due to ground shine from deposited nuclides on ground surfaces. The above studies indicate that the release of radioactive materials into the atmosphere varied greatly over the first half of March 2011, from several minutes to several hours, and decreased from April onward. Within this context, temporal changes occurred in the release rates of both radioactive cesium and iodine (Fig. 2.1) (Katata and Chino 2017). Specifically, large releases occurred due to venting and a hydrogen explosion in reactor of Unit 1 on March 12, followed by a hydrogen explosion in reactor of Unit 3 on March 14. Large releases related to events inside reactor of Units 2 and 3 continued through to March 20. The total amounts of I-131 and Cs-137 released in March 2011 were 151 and 14 PBq, respectively. An inverse estimation from ocean inventories using Cs-137 concentrations in Pacific Ocean water estimated the total amount of released Cs-137 as between 15 and 21 PBq (Aoyama et al. 2020), which is consistent with the results from Katata and Chino (2017). However, the results of a recent inverse estimation using newly obtained environmental monitoring data suggest a lower release amount (Terada et al. 2020). Thus, to summarize, current estimates of the total amount of radionuclides released into the atmosphere include a relative uncertainty of a few tens of percent, while current estimates of release rates over shorter time periods (several hours) may have a relative uncertainty to an order of magnitude of one.

12

S. Hirao

Fig. 2.1 Temporal changes in release rates of radioactive iodine (I-131) and radioactive cesium (Cs-137) (Katata et al. (2015) and Chino et al. (2016), cited in Katata and Chino (2017)

2.4

Atmospheric Dispersion and Deposition

The atmospheric dispersion of radioactive material during the FDNPP accident has been investigated by a number of researchers, both in Japan and internationally, through the use of various atmospheric dispersion models (Yamazawa and Hirao 2011; Morino et al. 2011; Povinec et al. 2013; Mathieu et al. 2018). Currently,

2 Atmospheric Transport and Deposition of Fukushima-Derived Radionuclides

13

Fig. 2.2 Schematic diagrams of analyzed transport routes related to plumes (a) P2, P3, and P4 and (b) P7, P8, and P9. Thick arrows indicate the general trend in the movement of each plume (Nakajima et al. 2017)

regardless of the model used, similar results regarding the advection and dispersion of radioactive plumes are coming to the fore (Yamazawa and Igarashi 2020). In summary, the venting and hydrogen explosion at Unit 1 on March 12 resulted in the dispersion of radionuclides to the northwest of the reactor and contamination by dry deposition. From the evening of March 14 to the morning of March 15, a plume transported in a southerly direction dispersed over a wide area, and after passing through Gunma and Tochigi prefectures via Ibaraki Prefecture, caused contamination in these areas by both dry and wet deposition (including fog deposition). Thereafter, from the morning to the evening of March 15, a radionuclide plume that had dispersed to the southeast passed through the Nakadori region of Fukushima Prefecture. From the afternoon of March 15 until the morning of March 16, the plume then dispersed in a northwesterly direction. On March 15, wet deposition occurred due to rain and snowfall, causing the most widespread contamination of the FDNPP accident in the Nakadori region and northwest of the FDNPP. In total, nine plumes were identified based on their transport route characteristics (Fig. 2.2) (Nakajima et al. 2017). The movement of plumes obtained from atmospheric dispersion models corresponds to and was validated by temporal changes in atmospheric concentrations, deposition amounts, and air dose rates observed at various locations. However, when the measured values are quantitatively compared with the calculated values, a number of questions arise, such as the calculation of wet deposition processes (including in-cloud deposition) within the model and the influence of the spatial resolution of the model.

14

2.5

S. Hirao

Overview of Resuspension

We conducted a review of the atmospheric behavior of radioactive materials in the early stages of the FDNPP accident to determine the distribution of radioactive contamination. Notably, radioactive cesium is still being detected in the atmosphere around the FDNPP by the regulatory authority of Fukushima Prefecture. Although the observed atmospheric concentrations reflect the effects of radionuclide resuspension, concentration levels of radioactive cesium, which affects the impacts of radiation exposure, have not been reported. As the primary mechanisms of resuspension include wind and decontamination activities (Akimoto 2014; Hirose 2020), these can be investigated using a resuspension factor, which is the airborne radioactivity concentration divided by the soil concentration (Calmon et al. 2009). Resuspension factors have been reported as between 10 6 m 1 and 10 5 m 1 in the early stages of an accident (over the first days and months following the accident), subsequently decreasing to between 10 10 m 1 and 10 9 m 1 after 3–4 years. Notably, resuspension factors with the similar values were obtained since the FDNPP accident (Ochiai et al. 2016), and these results are consistent with those of previous studies. The primary objective of these observations is to monitor the release of radionuclides into the atmosphere associated with the decommissioning of reactors, and thus, should continue into the future.

2.6

Conclusion

In this study, we reviewed of the radioactive materials released into the atmosphere during the Fukushima Daiichi Nuclear Power Plant accident, including the related dispersion and deposition processes. As environmental monitoring data are extremely limited, the amount of released radionuclides and the atmospheric transport processes have had to be estimated using atmospheric dispersion models. Thus, it will be necessary to reevaluate the estimated results to reflect updated information on the atmospheric concentrations of radionuclides, nuclide composition, as well as related physical and chemical properties.

References 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 3:2554. https://doi.org/10.1038/ srep02554 Akimoto K (2014) Resuspension and transfer of radioactive particulate matter and subsequent secondary contamination. Japanese J Heal Phys 49:17–28. https://doi.org/10.5453/jhps.49.17

2 Atmospheric Transport and Deposition of Fukushima-Derived Radionuclides

15

Aoyama M, Tsumune D, Inomata Y, Tateda Y (2020) Mass balance and latest fluxes of radiocesium derived from the Fukushima accident in the western North Pacific Ocean and coastal regions of Japan. J Environ Radioact 217:106206. https://doi.org/10.1016/j.jenvrad.2020.106206 Calmon P, Fesenko S, Voigt G, Linsley G (2009) Quantification of radionuclide transfer in terrestrial and freshwater environments (IAEA-TECDOC-1616). International Atomic Energy Agency, Vienna Chino M, Nakayama H, Nagai H et al (2011) Preliminary estimation of release amounts of 131 I and 137 Cs accidentally discharged from the Fukushima Daiichi Nuclear Power Plant into the atmosphere. J Nucl Sci Technol 48:1129–1134. https://doi.org/10.1080/18811248.2011. 9711799 Chino M, Terada H, Nagai H et al (2016) Utilization of (134)Cs/(137)Cs in the environment to identify the reactor units that caused atmospheric releases during the Fukushima Daiichi accident. Sci Rep 6:31376. https://doi.org/10.1038/srep31376 Hirao S, Yamazawa H, Nagae T (2013) Estimation of release rate of iodine-131 and cesium-137 from the Fukushima Daiichi nuclear power plant. J Nucl Sci Technol 50:139–147. https://doi. org/10.1080/00223131.2013.757454 Hirayama H, Matsumura H, Namito Y, Sanami T (2017) Estimation of Xe-135, I-131, I-132, I-133 and Te-132 concentrations in plumes at monitoring posts in Fukushima Prefecture using pulse height distribution obtained from NaI(Tl) detector. Trans At Energy Soc Japan 16:1–14. https:// doi.org/10.3327/taesj.J16.014 Hirose K (2020) Atmospheric effects of Fukushima nuclear accident: a review from a sight of atmospheric monitoring. J Environ Radioact 218:106240. https://doi.org/10.1016/j.jenvrad. 2020.106240 Igarashi Y, Kogure T, Kurihara Y et al (2019) A review of Cs-bearing microparticles in the environment emitted by the Fukushima Dai-ichi Nuclear Power Plant accident. J Environ Radioact 205–206:101–118. https://doi.org/10.1016/j.jenvrad.2019.04.011 Ikehara R, Morooka K, Suetake M et al (2020) Abundance and distribution of radioactive cesiumrich microparticles released from the Fukushima Daiichi Nuclear Power Plant into the environment. Chemosphere 241:125019. https://doi.org/10.1016/j.chemosphere.2019.125019 Katata G, Chino M (2017) Source Term, Atmospheric dispersion, and deposition of radionuclides during the Fukushima Daiichi Nuclear Power Station accident. Earozoru Kenkyu 32:237–243. https://doi.org/10.11203/jar.32.237 Katata G, Chino M, Kobayashi T et al (2015) Detailed source term estimation of the atmospheric release for the Fukushima Daiichi Nuclear Power Station accident by coupling simulations of an atmospheric dispersion model with an improved deposition scheme and oceanic dispersion model. Atmos Chem Phys 15:1029–1070. https://doi.org/10.5194/acp-15-1029-2015 Mathieu A, Kajino M, Korsakissok I et al (2018) Fukushima Daiichi-derived radionuclides in the atmosphere, transport and deposition in Japan: A review. Appl Geochemistry 91:122–139. https://doi.org/10.1016/j.apgeochem.2018.01.002 Moriizumi J, Oku A, Yaguchi N et al (2019) Spatial distributions of atmospheric concentrations of radionuclides on 15 March 2011 discharged by the Fukushima Dai-Ichi Nuclear Power Plant Accident estimated from NaI(Tl) pulse height distributions measured in Ibaraki Prefecture. J Nucl Sci Technol 57:495–513. https://doi.org/10.1080/00223131.2019.1699191 Morino Y, Ohara T, Nishizawa M (2011) Atmospheric behavior, deposition, and budget of radioactive materials from the Fukushima Daiichi nuclear power plant in March 2011. Geophys Res Lett 38:12. https://doi.org/10.1029/2011GL048689 Nakajima T, Misawa S, Morino Y et al (2017) Model depiction of the atmospheric flows of radioactive cesium emitted from the Fukushima Daiichi Nuclear Power Station accident. Prog Earth Planet Sci 4:2. https://doi.org/10.1186/s40645-017-0117-x Ochiai S, Hasegawa H, Kakiuchi H et al (2016) Temporal variation of post-accident atmospheric 137Cs in an evacuated area of Fukushima Prefecture: Size-dependent behaviors of 137Csbearing particles. J Environ Radioact 165:131–139. https://doi.org/10.1016/j.jenvrad.2016.09. 014

16

S. Hirao

Oura Y, Ebihara M, Tsuruta H et al (2015) Determination of atmospheric radiocesium on filter tapes used at automated SPM monitoring stations for estimation of transport pathways of radionuclides from Fukushima Dai-ichi Nuclear Power Plant. J Radioanal Nucl Chem 303:1555–1559. https://doi.org/10.1007/s10967-014-3662-4 Povinec PP, Gera M, Holý K et al (2013) Dispersion of Fukushima radionuclides in the global atmosphere and the ocean. Appl Radiat Isot 81:383–392. https://doi.org/10.1016/j.apradiso. 2013.03.058 Terada H, Katata G, Chino M, Nagai H (2012) Atmospheric discharge and dispersion of radionuclides during the Fukushima Dai-ichi Nuclear Power Plant accident. Part II: Verification of the source term and analysis of regional-scale atmospheric dispersion. J Environ Radioact 112:141–154. https://doi.org/10.1016/j.jenvrad.2012.05.023 Terada H, Nagai H, Tsuduki K et al (2020) Refinement of source term and atmospheric dispersion simulations of radionuclides during the Fukushima Daiichi Nuclear Power Station accident. J Environ Radioact 213:106104. https://doi.org/10.1016/j.jenvrad.2019.106104 Terasaka Y, Yamazawa H, Hirouchi J et al (2016) Air concentration estimation of radionuclides discharged from Fukushima Daiichi Nuclear Power Station using NaI(Tl) detector pulse height distribution measured in Ibaraki Prefecture. J Nucl Sci Technol 53:1919–1932. https://doi.org/ 10.1080/00223131.2016.1193453 Tsuruta H, Oura Y, Ebihara M et al (2018) Time-series analysis of atmospheric radiocesium at two SPM monitoring sites near the Fukushima Daiichi Nuclear Power Plant just after the Fukushima accident on March 11, 2011. Geochem J 52:103–121. https://doi.org/10.2343/geochemj.2.0520 Yamazawa H, Hirao S (2011) Impacts of Fukushima Daiichi NPP accident through atmospheric environment. J At Energy Soc Japan 53:479–483. https://doi.org/10.3327/jaesjb.53.7_479 Yamazawa H, Igarashi Y (2020) Recent understanding on the release of radionuclides and their behavior in the atmosphere. Radioisotopes 69:19–30. https://doi.org/10.3769/radioisotopes. 69.19

Chapter 3

Airborne Radiation Survey after the Accident Tatsuo Torii and Yukihisa Sanada

Abstract Airborne radiation survey was conducted to evaluate the influence of radionuclides emitted by the Fukushima Daiichi Nuclear Power Plant (FDNPP) accident. Carrying out airborne radiation surveys using a manned helicopter, the Japan Atomic Energy Agency (JAEA) has developed and established an analysis method concurrently with the development of this survey method. In particular, because the background radiation level differs greatly between the East and West regions of Japan, the JAEA has developed a discrimination method for natural radionuclide and cosmic rays using the gamma energy spectra. The reliability of the airborne radiation monitoring data was confirmed through comparison with large amounts of ground measurement data. Here, we report on the measurement technique and the results. The ambient dose rate has been decreased by the increasing attenuation effect due to radioactive cesium penetration into the soil in addition to the physical decay of radioactive cesium. These results indicate the importance of airborne monitoring to evaluate and predict the radiation exposure of residents. Keywords Airborne radiation survey · Manned helicopter · Radioactive cesium · Iodine-131

3.1

Introduction

In order to evaluate the influence of radionuclides emitted by the accident at the Fukushima Daiichi Nuclear Power Plant (FDNPP) of the Tokyo Electric Power Company Holdings, Inc. caused by the Great East Japan Earthquake, various types T. Torii (*) Institute of Environmental Radioactivity, Fukushima University, Fukushima, Japan e-mail: [email protected] Y. Sanada (*) Collaborative Laboratories for Advanced Decommissioning Science, Japan Atomic Energy Agency, Fukushima, Japan e-mail: [email protected] © The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2022 K. Nanba et al. (eds.), Behavior of Radionuclides in the Environment III, https://doi.org/10.1007/978-981-16-6799-2_3

17

18

T. Torii and Y. Sanada

of environmental radiation monitoring data have been obtained by many governmental institutes and universities. An airborne radiation monitoring technique is suitable to grasp the overall distribution of the ambient dose rate (referred to hereinafter as the air dose rate) and the deposition of radionuclides because it is able to (1) measure widely distributed radionuclides with less manpower and within short periods, (2) display maps that are easy to understand visually, and (3) measure in locations (e.g., forests and mountains) where people cannot easily enter. Analyzing the temporal and spatial changes in the air dose rate based on multiple airborne radiation monitoring datasets is useful for the prediction and evaluation of the radiation exposure of inhabitants. The current airborne radiation monitoring technique was established in the early 2000s. In particular, European scientists conducted pioneering studies of airborne radiation monitoring after the accident at the Chernobyl nuclear power plant. The method of data acquisition, calibration, and mapping developed by Aage et al. (1999) is a current basic technique. Allyson and Sanderson (1998) proposed a calibration technique using a Monte Carlo simulation. Tyler et al. (1996) proposed field-of-view airborne radiation monitoring by comparing airborne data with ground data. Through the European comparison project (Eccomags project), the data acquisition method and analysis methods of airborne radiation monitoring of each European country were unified (Sanderson et al. 2004). In Japan, triggered by the Three Mile Island nuclear power plant accident in 1979, research and development of an airborne radiation monitoring system using a manned helicopter (MRM: manned helicopter radiation monitoring) were initiated by researchers of primarily the Japan Atomic Energy Research Institute (reorganized as the Japan Atomic Energy Agency [JAEA]) (Saito et al. 1988; Nagaoka and Moriuchi 1990). However, methods of data acquisition, data analysis, and mapping of the MRM, which correspond to measurements of wide areas in this case, had not been established.

3.2

National Project after the Accident

After the accident at the FDNPP, the MRM national project was started by the Ministry of Education, Culture, Sports, Science and Technology of Japan (MEXT) and the Department of Energy in the United States (Lyons and Colton 2012; Blumenthal 2012). Even though the MRM initially monitored only the area around the FDNPP, the surveyed areas were gradually expanded and eventually airborne radiation monitoring was used in eastern Japan, excluding Hokkaido, after October 2011 and in western Japan and Hokkaido, after May 2012 (Sanada et al. 2014a). The distributions of the air dose rate at a height of 1 m above ground level (agl.) and the contamination concentration of radioactive cesium (134Cs and 137Cs) on the ground surface were monitored for all areas in Japan via this MRM project. Moreover, Torii et al. (2013) reported to analysis the 131I distribution map using survey data at an early stage after the accident. This survey project is ongoing with regularly repeating

3 Airborne Radiation Survey after the Accident

19

surveys conducted by the Nuclear Regulatory Authority of Japan (NRA) and the JAEA (Nuclear Regulatory Authority [NRA] 2020). While conducting the MRM project, we developed and established an analysis method concurrently with the development of the monitoring method. In particular, because the background radiation level differs greatly between East and West regions of Japan, we developed a discrimination method for the natural radionuclides as the dominant background, a method for setting the parameters for conversion to the air dose rate near the ground level, and an analysis method (Sanada et al. 2017). The effective half-life of the air dose rate was evaluated using MRM data from 2011 to 2016 (Sanada et al. 2018). Several years after the FDNPP accident, evaluation of the temporal change in the air dose rate is required for the evaluation of the radiation exposure of inhabitants and postaccident response. Knowledge of not only the average change in the air dose rate but also the regional characteristics of the change in the air dose rate can be obtained through the analysis of sequential MRM data. This article presents the methods and results of the MRM missions performed around the FDNPP. In addition, the tendency of the temporal change and the characteristics of the change in the air dose rate are discussed based on these results.

3.3 3.3.1

Basic Methodology Data Acquisition and Analysis

The advantage of MRM is that it can collect ground surface γ-ray radiation data with a spatial resolution of 0.25 km over wide areas including mountainous regions that are inaccessible from the ground. The image for data acquisition and analysis method is shown in Fig. 3.1. MRM relied on a dedicated radiation detection system (RSX-3, Radiation solution Inc., Canada) installed on a manned helicopter. It acquires a once-per-second readout of spectrometers to produce a 1024-channel energy spectrum at 3 keV per channel. The readings are synchronized with time from a global positioning system (GPS) receiver. The spectrum and the GPS data (date, time, latitude, longitude, and height above ellipsoid) were recorded every second. The spectrum data were processed using parameters such as an attenuation factor and an air dose rate conversion factor to retrieve the air dose rate at the surface level. The detailed analytical technique was in accordance with the IAEA technical standard (International Atomic Energy Agency [IAEA] 2003). We calculated the altitude distribution of air dose rates with the following procedure. First, we superimposed raster data from a digital elevation model (DEM; Geospatial Information Authority of Japan [GSI] 2015) on raster data from the air dose rate of MRM embedded in Geographic Information System software. Second, we combined the DEM data with the air dose rate data on the nearest grid point. The spatial resolution of the air dose rate and the DEM data is 0.25-km grid and 0.01-km grid, respectively. Mapping was performed by supplementing unmeasured areas via interpolation of the measured results. Even though various methods such as kriging and spline

Fig. 3.1 Image for data acquisition and analysis of MRM

20 T. Torii and Y. Sanada

3 Airborne Radiation Survey after the Accident

21

approaches have been proposed for interpolation, the inverse distance weighted (IDW) method, which assigns weights to the values of the measurement points linearly and in inverse proportion to the distance, was applied to the MRM data. The IDW method is easy to use when analyzing a large amount of data because the parameter setting is simple (International Atomic Energy Agency [IAEA] 2003). This interpolation processing was conducted using ArcGIS software (Environmental Systems Research Institute Inc., CA, USA). The spatial resolution of the resulting contour map for the background absorbed dose rate was 250 m
250 m.

3.3.2

Validation

Confirming the reliability of airborne radiation monitoring data is essential because previous studies have indicated that airborne radiation monitoring data tend to overestimate the air dose rates over complex topography (Tyler et al. 1996; Malins et al. 2015). In this article, we compared the MRM data to a large amount of ground measurement data. These ground measurement data were obtained multiple times at the same positions around the FDNPP (Mikami et al. 2019). To the extent possible, we selected the ground measurement and MRM data at nearly the same times and locations. Both sets of data were corrected for radioactive cesium decay. The error of interpolation can be evaluated by a semivariogram because IDW provides more advanced processing than the NaI survey meter (TCS-172B, Hitachi Inc., Tokyo, Japan). The air dose rate of the ground data (Dg) and the airborne data (Da) were compared by visualizing the unevenness using a scatter diagram. The relative deviation (RD) of each measurement mesh was calculated as follows in order to evaluate the accuracy of the airborne radiation monitoring:
RD ¼ Da  Dg =Dg :

ð3:1Þ

Calculated RDs were used to evaluate the total error and statistical uncertainty, which is shown as a histogram of frequency. Air dose rates retrieved from the MRM were compared to those measured at the ground. The data of the 2019 campaign are shown in Fig. 3.2a,b, respectively, as examples of comparison results. As shown in the figure, the airborne radiation survey and the ground measurement data exhibit positive correlations and agree well. Histograms of the RD of the MRM have shapes similar to Gaussian distribution whose peaks are near zero. However, the average RDs, which indicates the parameter of a Gaussian distribution, was 0.41  1.0. These errors were expressed as the standard deviation (SD) of RD, which indicates the parameter of a Gaussian distribution. These parameters indicated that the MRM tended to be higher than the ground measurements were in relatively good agreement with the ground measurements. These results indicate the accuracy of the MRM.

22

T. Torii and Y. Sanada

Fig. 3.2 Comparison of the airborne survey and the fixed point survey at the campaign of 2019

3.4 3.4.1

Survey Results Early-Stage Survey

For the first time after the accident, the department of energy and air force of the United States conducted the MRM around FDNPP. Torii et al. (2013) made a contour map of 131I by analyzing these data (Fig. 3.3a). In the measurement immediately after the accident, the map was made by combining Monte Carlo simulation and spectroscopic method to compensate for the lack of sufficient calibration data and the influence noise. After that, the Ministry of Education, Culture, Sports, Science and Technology Japan (MEXT) determined to measure the air dose rate which is derived from radiocesium with USA’s cooperation. The contour map shown in Fig. 3.3b is the result of reanalysis of data acquired in March 2011 by taking consider with the calibration data which were not sufficient at that time by Sanada et al. (2018). These data helped to grasp the situation at FDNPP immediately after the accident. Radioactive plume was mainly released to northwest from FDNPP. The deposition process of 131I and radioactive cesium seemed to be different because 131I decayed by the short half-life (8.02 days) at the measurement time. Especially, deposition of 131I was relatively higher in the south region from FDNPP. This tendency does not contradict the atmospheric simulation result (Katata et al. 2014). Since 14 months after the accident, JAEA carried out measurement the whole area in Japan (Sanada et al. 2014a). Figure 3.4 shows maps of the air dose rate and deposition of radiocesium that were constructed from the MRM measured data. The red- and orange-colored area with a high air dose rate exceeding 10 μSv h1 have oriented toward the northwest from the FDNPP to an area approximately 30 km of the FDNPP. For areas from 3 km to 80 km of the FDNPP, areas with 0.5–40 μSv h1 spread gradually in the areas neighboring the red-colored high air dose rate area. The air dose rate is relatively high south of the FDNPP. As shown in Fig. 3.4b, radioactive cesium was spread to 200 km west direction from FDNPP. The region of

3 Airborne Radiation Survey after the Accident

a 131I depotition (kBq m-2 ) at 3 April, 2011

23

b Air dose rate at 1m agl. (µSv h-1) at 29 April, 2011

Fig. 3.3 The results of MRM in early stage after the FDNPS accident. (a) 131I deposition (kBq m2) at April 3, 2011 (Torii et al. 2013) and (b) air dose rate at 1 m agl. (μSv h1) at 29 April, 2011 (Sanada et al. 2018)

relatively high air dose rate areas of north direction from FDNPP was thought to derive from wet deposition (Katata et al. 2014).

3.4.2

Evaluation for Temporal Change

Air dose rateAir dose rates maps from multiple measurement periods are shown in Fig. 3.5a–e). These maps show strong spatially defined areas with a relatively high radiation dose rate and deposited radioactive cesium to the northwest, west–northwest, west, and south of FDNPP. They also show that the high air dose rate areas shrink with time. Figure 3.6a shows the air dose rate at approximately 100 km radius area from FDNPP in all national project campaigns. The effective half-life was calculated using all data. The effective half-life is defined to sum up the physical decay and ecological decrease by such as migration or weathering. The time trend of ambient dose rate (Dr(t)) can be approximated by the following double exponential function with effective half-lives of Tshort and Tlong:

24

T. Torii and Y. Sanada

a Air dose rate at 1m agl. (µSv h-1) at 31 May, 2012

b Radiocesium deposition (kBq m-2) at 31 May, 2012

Fig. 3.4 The results of MRM in whole area in Japan. (a) Air dose rate at 1 m agl. (μSv h1) at 31 May, 2012 and (b) Radiocesium deposition (kBq m2) at 31 May, 2012

Dr ðt Þ ¼ m exp ð ln 2=T short t Þ þ n exp  ln 2=T long t




ð3:2Þ

where t is the time after the FDNPP accident and m and n are the proportion of the short-lived and long-lived component to the total air dose rate at t ¼ 0, respectively. The mean value of each survey campaign and the fitting curve is shown in Fig. 3.6b for clarification of the fitting curve. The fitting curve which sums the two exponential equations reproduced the mean of the air dose rate well. The short and long effective half-life of air dose rate (Tshort and Tlong) were 0.671 (95% confidence interval [CI]: 0.506–0.836) and 4.39 (95% CI: 3.46–5.32), respectively. These values were the important parameters of the prediction in the future. Saito et al. (2019) summarized the evaluated half-lives of air dose rate by some measurement method. These values were almost the same level by comparing with this study. The effective half-life of the air dose rate should be updated by adding a future survey result.

3 Airborne Radiation Survey after the Accident 140o15’E

140o30’E 140o45’E

141oE

140o15’E

25

140o30’E 140o45’E

141oE

37o45’N 37o30’N 37oN 36o45’N

(c)Esri Japan

b 19 November, 2013

140o30’E 140o45’E

141oE

140o15’E

140o30’E 140o45’E

(c)Esri Japan

c 4 November, 2015 141oE

N

N

Air dose rate at 1 m agl. ( Sv h-1)

38oN

38oN

N

37o15’N

37o15’N 37oN 36o45’N

a 5 November, 2011 140o15’E

0.5

1

5

10

50

(c)Esri Japan

d 16 November, 2017

36o45’N

37oN

37oN

37o15’N

37o15’N

37o30’N

37o30’N

37o45’N

37o45’N

0.1

36o45’N

140o30’E 140o45’E 141oE

38oN

38oN 37o30’N

37o45’N 37o30’N 37o15’N 37oN 36o45’N

(c)Esri Japan

140o15’E

N

37o45’N

38oN

N

(c)Esri Japan

e 2 November, 2019

Fig. 3.5 The temporal change of air dose rate. (a) 5 November, 2011. (b) 19 November, 2013. (c) 4 November, 2015. (d) 16 November, 2017. (e) 2 November, 2019

3.5

Advanced Technology Using UAV

A radiation measurement technique that is more detailed than the method using a manned helicopter is required to formulate a decontamination plan and evaluate the effect of decontamination. As shown in Fig. 3.7a, radiation measurement using an unmanned aerial vehicle (UAV) is one solution because an unmanned helicopter can be used to generate detailed air dose rateAir dose rates maps by flying below 150 m, in accordance with Japanese aviation law. Normally, this UAV can operate by

26

T. Torii and Y. Sanada

Fig. 3.6 Evaluation for effective half-life of air dose rate. (a) Shows all the data with the y-axis in log units. (b) Shows only mean value of each measurement campaign with the y-axis in linear units

Fig. 3.7 Unmanned aerial monitoring system

dedicated software in mobile ground base (Fig. 3.7b). The UAV system was developed for monitoring high air dose rate areas and riverbed areas (Sanada et al. 2014b; Sanada and Torii 2015). Analytical methods for converting airborne detector count rates to air dose rates at 1 m agl. Have been established based on the MRM method, and the validity of the MRM method has been demonstrated through comparisons with large amounts of ground measurement data (Sanada and Torii 2015). Several years after the FDNPP accident, evaluation of temporal changes in the air dose rate is required for the evaluation of radiation exposure to inhabitants and for postaccident response. In previous studies, as shown in Fig. 3.8a1–a3, temporal changes in the air dose rate at approximately 5 km from the FDNPP were evaluated regionally based on detailed monitoring results using a UAV (Sanada et al. 2018). Furthermore, Fig. 3.8b shows the cesium-134/137 ratio deposited on the ground. The blue color means that the ratio is small, and the plume diffused in the northwest direction show a different distribution from other plumes (Nishizawa et al. 2015In

3 Airborne Radiation Survey after the Accident

27

Fig. 3.8 Distribution of air dose rates (a) and cesium-134/137 ratio (b) around the FDNPP. Riverbed deposition of radiocesium in the Ukedo River (c)

addition, Fig. 3.8c1 and c2 shows the air dose rate distribution in the Ukedo River basin, which flows around 10 km north of the FDNPP. As is clear from the figure, it can be seen that radioactive cesium is deposited at the confluence of rivers and riverbeds, and the dose rate decreases with the passage of time.

3.6

Conclusion

In order to investigate the influence of artificial radionuclide deposition, the distributions of the air dose rate were measured via MRM. In Japan, no technical experience or example measurements were available for evaluating the air dose rate distribution via widespread MRM prior to the FDNPP accident. Therefore, we developed and established an analysis method concurrently while carrying out actual monitoring. We obtained new knowledge concerning the tendency of the air dose rate to decrease by analyzing past monitoring results. (1) The difference between the deposition situation of 131I and radioactive cesium was depicted to analyze the

28

T. Torii and Y. Sanada

early-stage MRM. (2) The short component of the effective half-life from the FDNPP accident to 0.671 years. This value was effective at the early stage (from 0 to 2 years) from the accident. (3) The long component of the effective half-life from the FDNPP accident to 4.39 years. This value was effective from the accident 2 years later. The effective half-life is an important parameter for the prediction of future air dose rate. Hereafter, it will be important to carefully monitor and investigate the migration of radioactive cesium emitted by the FDNPP accident due to weathering effects. Therefore, the accuracy of the analysis technique as well as the measurement method should be made more precise in the future.

References Aage HK, Korsbech U, Bargholz K, Hovgaard J (1999) A new technique for processing airborne gamma ray spectrometry data for mapping low level contaminations. Appl Rad Isot 51:651–662 Allyson JD, Sanderson DCW (1998) Monte Carlo simulation of environmental airborne gammaspectrometry. J Environ Radioact 38:259–282 Blumenthal DJ (2012) Introduction to the special issue on the U.S. response to the Fukushima accident. Health Phys 102:482–484 Geospatial Information Authority of Japan [GSI] (2015) Maps & Geospatial Information. http:// www.gsi.go.jp/ENGLISH/page_e30031.html. Accessed 31 Aug 2020 International Atomic Energy Agency [IAEA] (2003) Guidelines for radioelement mapping using gamma ray spectrometry data, IAEA-TECDOC-1363 Katata G, Chino M, Kobayashi T, Terada H, Ota M, Nagai H, Kajino M, Draxler R, Hort MC, Malo A, Torii T, Sanada Y (2014) Detailed source term estimation of the atmospheric release for the Fukushima Dai-ichi nuclear Power Station accident by coupling simulations of atmospheric dispersion model with improved deposition scheme and oceanic dispersion model. Atom Chem Phy 14:14725–14832 Lyons C, Colton D (2012) Aerial measuring system in Japan. Health Phys 102:509–515 Malins A, Okumura M, Machida M, Saito K (2015) Topographic effects on ambient dose equivalent rates from radiocesium fallout. Proceedings of joint international conference on mathematics and computation, supercomputing in nuclear applications and the Monte Carlo method (M&C + SNA + MC 2015) Mikami S, Tanaka H, Matsuda H, Sato S, Hoshide Y, Okuda N, Suzuki T, Sakamoto R, Ando M, Saito K (2019) The deposition densities of radiocesium and the air dose rates in undisturbed fields around the Fukushima Dai-ichi nuclear power plant; their temporal changes for five years after the accident. J Environ Radioact 210:105941 https://doi.org/10.1016/j.jenvrad.2019.03. 017 Nagaoka T, Moriuchi S (1990) Aerial radiological survey and assessment system ARSAS. Jpn J Health Phys 25:391–398. In Japanese Nishizawa Y, Yoshida M, Sanada Y, Torii T (2015) Distribution of the 134Cs/137Cs ratio around the Fukushima Daiichi nuclear power plant using an unmanned helicopter radiation monitoring system. J Nucl Sci Tech 53:468–474 Nuclear Regulatory Authority [NRA] (2020) Results of airborne monitoring in Fukushima Prefecture and neighboring prefectures and the Fourteenth Airborne Monitoring in the 80km zone from the Fukushima Daiichi NPP: https://radioactivity.nsr.go.jp/en/contents/14000/13990/24/ 14th_Airborne_eng.pdf. Accessed 30 Aug 2020

3 Airborne Radiation Survey after the Accident

29

Saito K, Sakamoto R, Tsutsumi M, Nagaoka T, Moriuchi S (1988) Prompt estimation of release rates of gaseous radioactivity from a nuclear plant using an aerial survey. Radiat Prot Dosim 22: 77–85 Saito K, Mikami S, Andoh M, Matsuda N, Kinase S, Tsuda S, Yoshida T, Sato T, Seki A, Yamamoto H, Saanda Y, Wainwright-Murakami H, Takemiya H (2019) Summary of temporal changes in air dose rates and radionuclide deposition densities in the 80 km zone over five years after the Fukushima nuclear power plant accident. J Environ Radioact 210:105878 Sanada Y, Sugita T, Nishizawa Y, Kondo A, Torii T (2014a) The airborne radiation monitoring in Japan after the Fukushima Daiichi nuclear power plant accident. Prog Nucl Sci Tech 4:76–80 Sanada Y, Kondo A, Sugita T, Nishizawa Y, Yuuki Y, Ikeda K, Shoji Y, Torii T (2014b) Radiation monitoring using an unmanned helicopter in the evacuation zone around the Fukushima Daiichi nuclear power plant. Explor Geophys 45:3–7 Sanada Y, Torii T (2015) Aerial radiation monitoring around the Fukushima Dai-ichi nuclear power plant using an unmanned helicopter. J Environ Radioact 139:294–299 Sanada Y, Ishizaki A, Nishizawa Y, Urabe Y (2017) Airborne radiation monitoring using a manned helicopter. Bunseki Kagaku 66:149–162. In Japanese Sanada Y, Urabe Y, Sasaki M, Ochi K, Torii T (2018) Evaluation of ecological half-life of dose rate based on airborne radiation monitoring following the Fukushima Dai-ichi nuclear power plant accident. J Environ Radioact 192:417–425 Sanderson DCW, Cresswell AJ, Scott EM, Lang JJ (2004) Demonstrating the European capability for airborne gamma spectrometry: results from the eccomags exercise. Rad Prot Dosimet 109: 119–125 Torii T, Sugita T, Okada CE, Reed MS, Blumenthal DJ (2013) Enhanced analysis methods to derive the spatial distribution of 131I deposition on the ground by airborne surveys at an early stage after the Fukushima Daiichi nuclear power plant accident. Health Phys 105:192–200 Tyler AN, Sanderson DCW, Scott EM, Allyson JD (1996) Accounting for spatial variability and fields of view in environmental gamma ray spectrometry. J Environ Radioact 33:213–315

Part II

Fate and Transport of Radionuclides in Soil-Water Environment

Chapter 4

Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment: Review Alexei Konoplev, Yoshifumi Wakiyama, Toshihiro Wada, Yasunori Igarashi, Volodymyr Kanivets, and Kenji Nanba

Abstract The chapter reviews the studies of the Fukushima-derived radiocesium behavior in the soil–water environment. Quantitative characteristics of radiocesium solid–liquid distribution (apparent and exchangeable distribution coefficients) and radiocesium wash-off from contaminated catchments (dissolved and particulate wash-off ratios) were obtained on the basis of field research and monitoring data. A conceptual model accounting for transformation of Fukushima-derived radiocesium chemical forms in soils and sediments is outlined, and key kinetic and equilibrium parameters of this model were estimated for geoclimatic conditions of Fukushima. Fukushima-derived radiocesium was found to be strongly bound by soil and sediment particles. Radiocesium apparent distribution coefficient Kd in Fukushima Rivers is much higher (at least by an order of magnitude) than that in rivers of the Chernobyl area, which is most likely due to two reasons: high binding ability of soils and sediments in Fukushima-contaminated areas, and the presence of water-insoluble hot glassy microparticles (CsMPs) in the Fukushima accidental fallout. Despite radiocesium in closed and semi-closed ponds being relatively persistent, a decline in both particulate and dissolved 137Cs activity concentrations was revealed. The reduction rate of the particulate 137Cs activity concentrations was much higher than that for dissolved 137Cs. As a result, the apparent distribution coefficient Kd(137Cs) in the sediment–water system decreased with the rate constant 0.12–0.18 yr.
1. Assuming that the decrease in Kd is associated with decomposition of glassy Cs-rich microparticles, the timescale of 137Cs leaching from them in the ponds studied was estimated to be 5–8 years. The obtained estimates are consistent with the findings of recent laboratory experiments. The processes of wash-off, river transport, and radionuclide vertical migration in catchment soils were considered in an integrated way using the semi-empirical diffusional model. This approach enables description of changes in the particulate and dissolved 137Cs wash-off ratios using

A. Konoplev (*) · Y. Wakiyama · T. Wada · Y. Igarashi · K. Nanba Institute of Environmental Radioactivity at Fukushima University, Fukushima, Japan V. Kanivets Ukrainian Hydrometeorological Institute, Kiev, Ukraine © The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2022 K. Nanba et al. (eds.), Behavior of Radionuclides in the Environment III, https://doi.org/10.1007/978-981-16-6799-2_4

33

34

A. Konoplev et al.

only two physically meaningful parameters Deff—the 137Cs effective dispersion coefficient in the topsoil layer—and Kd—the 137Cs apparent distribution coefficient. Particulate 137Cs wash-off ratios from the catchments of the Fukushima area display only minor differences compared to those in the Chernobyl area, being at the lower limit of the Chernobyl values. Somewhat lower values of Np(137Cs) in the Fukushima area are explained by higher values of the effective dispersion coefficient Deff(137Cs) in typical Fukushima soils. Dissolved 137Cs wash-off ratios for Fukushima catchments are at least an order of magnitude lower than those for Chernobyl, mainly due to an order of magnitude difference in the 137Cs distribution coefficients for the Fukushima and Chernobyl Rivers. Keywords Fukushima · Radiocesium · Dissolved · Particulate · Distribution · Leaching · Wash-off

4.1

Introduction

The accident at Fukushima Daiichi Nuclear Power Plant (FDNPP) led to extensive contamination of the soil–water environment by cesium radioactive isotopes 134Cs (half-life Т1/2 ¼ 2.06 years) and 137Cs (Т1/2 ¼ 30.17 years). Radiocesium deposition northwest of the NPP resulted in a trace of contamination 50–70 km long and 20 km wide (Chino et al. 2011; Hirose 2012; Ministry of Education, Culture, Sports, Science and Technology of Japan [MEXT] 2012; Saito et al. 2015). The initial ratio of 134Cs/137Cs isotopes in the Fukushima fallout was about one (Hirose 2012; Chaisan et al. 2013). The contribution of 134Cs to the radioactive contamination of soils, as compared to 137Cs, decreases over time due to its more rapid decay. The behavior of accidentally released radionuclides in the environment is determined by their properties and environmental characteristics which are key for their mobility and bioavailability (Konoplev et al. 2020). The contaminated territory of Fukushima Prefecture is characterized by an expansive and differentiated hydrographic network, dominated by the largest river of the area—Abukuma, with its tributaries the Hirose and Kuchibuto Rivers (Fig. 4.1). Other rivers include Ukedo, with its tributary Takase; the Niida River with its tributaries Hiso, Mizunashi, Iitoi, and Wariki; Uda, Mano, Ota, Odaka, Maeda, Kuma, Tomioka, Ide, Kido, Natsui, and Same Rivers (Yoshimura et al. 2015a; Evrard et al. 2015). All these rivers ultimately end in the Pacific Ocean. Thus, the river catchments contaminated from the FDNPP accident became a long-term source of secondary contamination of water bodies by surface runoff and radiocesium flux to the Ocean (Konoplev 2016; Konoplev et al. 2016a, 2018a). Moreover, surface runoff and river transport result in the transfer of radiocesium from contaminated evacuated areas to cleaner populated regions, and the settling of radiocesium in bottom sediments of river reservoirs and on floodplains (mostly during rainy seasons). After the Chernobyl accident, closed and semi-closed water bodies such as lakes and ponds were found to be more sensitive to radioactive contamination as compared

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

35

Fig. 4.1 River catchments contaminated after the FDNPP accident with a map of 134+137Cs topsoil layer activity concentrations (from Evrard et al. 2015)

36

A. Konoplev et al.

Fig. 4.2 Locations of studied ponds in Okuma town, vicinity of FDNPP

with rivers and dam reservoirs (Comans et al. 1989; Konoplev et al. 1992a, 1998, 2002a). Within Fukushima Prefecture, there are more than 3700 ponds of varying sizes, many of which are used for paddy water supply (Tsukada et al. 2017). Such irrigation ponds have been created over centuries in Japan for rice cultivation. Following the accident, these ponds are also of concern because they are used for fishing and watering of agricultural fields and may cause crop contamination (Fukushima and Arai 2014; Tsukada et al. 2017; Wada et al. 2019; Yoshikawa et al. 2014). The radiocesium behavior in the selected ponds in the vicinity of the FDNPP (Fig. 4.2) has been studied starting from 2014 to enable the identification of mid- and long-term trends and to elucidate mechanisms behind the 137Cs sediment– water distribution and time changes (Konoplev et al. 2018b; Wakiyama et al. 2017, 2019a). Characteristics of the ponds are presented in Table 4.1. After deposition of radionuclides on the ground surface, over time the contamination migrates down through the soil profile. Radionuclides migrate vertically in solution and as colloids with infiltration water flow, or attached to fine soil particles (Konoplev et al. 2016b). Transport of radiocesium in solution by infiltration is slower than the water flow because of sorption–desorption and fixation on soil particles. Fine soil particles containing radiocesium can move by penetrating through pores, cracks, and cavities, as well as with infiltration flow (lessivage), and as a result of vital activity of plants and biota (bioturbation) (Bulgakov et al. 1991a; Konoplev et al. 2016b). Nevertheless, the vertical migration of radionuclides in undisturbed soils can be described by the convection–dispersion equation using the effective values of dispersion coefficient and convective velocity (Konoplev et al. 1992b; Konshin 1992; Ivanov et al. 1997; Bossew and Kirchner 2004; Mamikhin et al. 2016). Vertical migration of radionuclide in soil leads to contamination of deeper soil layers and penetration of radionuclides to groundwater. The processes of vertical and lateral migration lead to gradual reduction in contamination of catchment soil, also referred to as natural attenuation, particularly its top layer (International Atomic Energy Agency [IAEA] 2006; Konoplev et al. 2018b). This, in turn, results in a

a

0.24 Closed, Filled out 2.3  1.3 6500 2.0 Fluviosol 18  3 (0.1  0.04)  10
3

Distance to FDNPP fence, km Pond type 6.9  1.9 4100 1.0 Fluviosol 63  20 (0.1  0.04)  10
3

3.75 Semi-closed, irrigation

Suzuuchi 37 24.950’N 140 58.7910 E 3.50 Open, Recreational 2.4  0.5 10700 2.5 Terrestrial regosol 49  16 (0.1  0.04)  10
3

Funasawa 37 24.363’N 140 59.1730 E

1.0  0.3 8100 5.0 Andosol 27  3 (0.1  0.04)  10–3

Kashiramori 37 22.996’N 140 56.4000 E 7.0 Semi-closed, recreational

Ammonium concentrations are not provided here since they are vastly variable and in many cases were less than detection limit of the methodology applied

137

Cs deposition σ, MBq/m2 Water surface area, m2 Maximum depth, m Catchment soil type K+ in water, μeq/L Stable 133Cs+ in water, μeq/L

Inkyozaka 37 25.499’N 141 01.050 E

Pond Coordinates

Table 4.1 Characteristics of the Okuma town ponds studied in the FDNPP exclusion zone (based on Konoplev et al. 2018b; Konoplev et al. 2021a and Wakiyama et al. 2017)a

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . . 37

38

A. Konoplev et al.

gradual decrease of radionuclide concentrations in both dissolved and particulate forms of surface runoff and river water (Konoplev et al. 2020, 2021b). The areas of Japan contaminated as a result of FDNPP accidents are characterized by wet monsoon climate with highly variable total annual precipitation (1200–1800 mm). Less than 5–7% of precipitation falls as snow. Extreme rains with precipitation amount >200 mm and >100 mm per event occur once every 5–7 years and every 1.5 years, respectively, according to meteorological observations. Several rains >50 mm fall each year, most frequently during the monsoon season between June and October. Some heavy rains can, however, occur in April– May, when soil surface on cultivated fields is less protected by vegetation (Laceby et al. 2016). The geology of the territory is highly heterogeneous, and the mountains serve as a partition between the Abukuma River valley, the largest in the region, and the ocean. The proportion of clays is 20–30%. These mountains are 1200–1300 m high and folded metamorphic and sedimentary rocks with numerous magmatic intrusions. Parent rock materials in the Fukushima-contaminated areas are primarily granites and volcanic ashes that are subject to physicochemical weathering in the humid monsoon climate conditions. In the region there are several active and dormant volcanoes; hot springs of different geochemical compositions are abundant. Soil diversity is great due to the combination of mountain rocks of different lithological compositions, intense weathering and denudation from high seismicity, and the steep inclination of mountain slopes. The interfluve areas include brown soils (under beech forest), ashy-volcanic, rich in humus, acidic allophonic (andosol), and leached brown soils. The valley floors are used mainly for growing rice and are represented by alluvial soils strongly modified as a result of many years’ land use. Undisturbed alluvial soils occur on the leveed parts of river valleys and along the canalized parts of stream channels typical of intermountain depressions. The arable lands, mainly paddy fields, occupy about 12% of the total territory in the region, and occur primarily on extensive depressions and piedmont plain. After the Fukushima accident, cultivation on arable land with high radiocesium deposition levels has been banned and decontamination activities have been implemented to remove contaminated soil layers (Konoplev et al. 2016a). Occurrence of high flow events in rivers especially during typhoon’s season facilitates radionuclide wash-off from contaminated catchments and their lateral migration (Evrard et al. 2014; Konoplev et al. 2018a; Nagao et al. 2013; Yamashiki et al. 2014). The goal of this chapter is to synthesize relevant research on fate and transport of Fukushima-derived radiocesium in soil–freshwater environment including catchment soils, rivers, ponds, and dam reservoirs.

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

4.2

39

Radiocesium Speciation and Solid–Liquid Distribution in Soil–Water Environment

The mobility and bioavailability of radiocesium of accidental origin are governed by the ratio of radiocesium chemical forms in fallout and site-specific environmental characteristics determining the rates of leaching, fixation–remobilization, as well as sorption–desorption of the mobile fraction (its solid–liquid distribution) (Konoplev and Bobovnikova 1991; Konoplev et al. 1992b). Radiocesium in the environment is strongly bound to soil and sediment particles containing micaceous clay minerals (illite, vermiculite, etc.). This is due to two basic processes—high selective reversible sorption and fixation (Konoplev and Konopleva 1999; Beresford et al. 2016; Konoplev 2020). Figure 4.3 presents a conceptual model for transformation of Fukushima-derived radiocesium speciation in soil–water environment.

4.2.1

Particulate Radiocesium

To study suspended sediments and particulate radiocesium in ponds and dam reservoirs, sediment traps were used (Konoplev et al. 2021a) based on the methodology developed by Antsiferov and Kosyan (1986). Each integrated sediment trap was a plastic cylinder 75 mm in diameter and 100 mm in height with a close-fitting lid. Around the circumference of the lid were six equally spaced holes of 7 mm

Fig. 4.3 Conceptual model of transformation processes for Fukushima-derived radiocesium speciation in soil–water environment (updated and modified after Konoplev et al. 2018b and Konoplev 2020)

40

A. Konoplev et al.

Fig. 4.4 Installation of integrated sediment traps in Okuma town ponds and Shinobu Dam reservoir at Abukuma River Table 4.2 Comparison of the 137Cs activity concentrations in the suspended sediments (SS) collected with the integrated traps and SS extracted by filtration of 4 L surface layer water through 0.45 μm filter for three ponds in Okuma town (from Konoplev et al. 2021a) Pond Cs in SS trapped, Bq/kg 137 Cs in SS filtered, Bq/kg(c)

137

Inkyozaka (1.3  0.02)  105a (1.0  0.2)  105

Suzuuchi (1.1  0.3)  105b (1.0  0.4)  105

Funasawa (0.78  0.02)  105a (0.77  0.19)  105

a

Collected from 07.10.2019 to 08.05.2019, gamma spectrometry measurement error Arithmetic mean with the standard deviation for six separate traps exposed from 05.23.2018 to 07.13.2018 c Arithmetic mean with the standard deviation for four surface water sample measurements around the time of SS trapping b

diameter. There was a central sleeve in each trap to accommodate the rod on which five or six of these traps were mounted to form a pillar. The rod was securely fixed on a support to ensure the structure was robust enough to withstand flowing water and to preclude movement and vibrations (Fig. 4.4). In case of relatively shallow Okuma town ponds, we used 5 or 6 traps for an individual cross section while for deeper dam reservoirs we used at least 10 traps for an individual cross section. Table 4.2 presents a comparison of the 137Cs activity concentrations in the integrated samples of suspended sediments and the average activity concentrations in suspended sediments extracted by the filtration of water samples collected during the same period. These measurements were in strong agreement, thereby confirming that the suspended sediments were uniformly distributed across the depths in the ponds and that sampling water from the ponds’ respective surface layers is representative.

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

41

Fig. 4.5 Differential particle size distribution (PSD) for the suspended sediments collected using the integrated traps exposed at the vertical cross section of the three Okuma town ponds (from Konoplev et al. 2021a)

Figure 4.5 shows the particle size distribution (PSD) of the suspended sediments collected using the sediment traps from the ponds. The sediments from all three ponds showed two slight peaks at approximately 1 μm (clay), whereas at 430 μm (medium sand) in Inkyozaka, and at 920 μm (coarse sand) in Funasawa. A further 2000 μm (coarse sand) was trapped in Suzuuchi. Sediments from Suzuuchi and Funasawa had one major peak at approximately 30 μm (coarse silt), whereas sediments from Inkyozaka had two major peaks at approximately 3–5 μm (fine silt) and 50 μm (coarse silt), respectively. A vastly different situation was observed in a deeper water body, Shinobu Dam reservoir on the Abukuma River. Figure 4.6 presents the results of the experiment from December 2018 to February 2019 to study suspended sediments using ten integrated traps installed along the central cross section of Shinobu Dam reservoir on the Abukuma River. During the period of sampling, the water depth at the site varied from 7.0 to 10.5 m and was on average 8.2  1.2 m. It can be seen from Fig. 4.6 that the suspended particle concentration in mg/L is quite uniformly distributed along the water depth since the weight of suspended sediments accumulated by individual traps is almost the same for all distances from the bottom. At the same time, the 137Cs activity concentrations of accumulated sediments differ by about an order of magnitude: from 4.2 kBq/kg on the peak at 230 cm from the bottom to 0.5 kBq/kg at 150 and 330 cm from the bottom. This important finding should be taken into account when sampling water, especially for particulate 137Cs measurements. The PSDs of suspended sediments collected at different depths in Shinobu Dam reservoir (at distances 1.5, 2.7, 2.9, and 3.3 m from the bottom) are presented in

42

A. Konoplev et al.

Fig. 4.6 Dependence of suspended sediments (SS) accumulation in the traps and their 137Cs activity concentrations (Bq/kg) on distance from the bottom at a cross section in Shinobu Dam reservoir in Abukuma River. Distance from the bottom L ¼ D0-D, where D0 is the current total water depth at the sampling site, and D is the current depth for a given sediment trap

Fig. 4.7 PSDs of suspended sediments at four different distances from the bottom in the central part of Shinobu Dam reservoir of Abukuma River (December 2018–February 2019)

Fig. 4.7. The PSDs for all studied depths were quite uniform and characterized by a main maximum at about 30 μm and a slight maximum at about 1 μm. The suspended sediments near the bottom are characterized by the occurrence of sand particles with PSD maximum at 2 mm. The PSDs do not differ much at different depths and do not seem to explain the differences in 137Cs activity concentrations as a function of water

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

43

depth. The differences in sediment composition and 137Cs exchangeability at different depths may play a certain role and should be further studied.

4.2.2

Radiocesium Speciation and its Transformation in Soil– Water Environment

Immediately after the FDNPP accident, the hypothesis was that radiocesium had deposited as part of condensation particles in water-soluble forms. Evidence to support this theory initially appeared in the studies by Kaneyasu et al. (2012) who concluded that radiocesium was transported in the atmosphere by sulfate aerosol particles of 0.5–0.6 μm diameter. Radiocesium in these particles is assumed to be water-soluble and washable by precipitation. However, the later studies (Adachi et al. 2013; Abe et al. 2015) have revealed spherical glassy aerosol particles of a few μm in diameter, as far as 170 km from the FDNPP, containing, apart from radiocesium, uranium and other elements representative of nuclear fuel and reactor materials. Particles of similar properties have also been identified by Niimura et al. (2015) using autoradiography of soils, plants, and mushrooms. Later on, coarser particles (up to hundreds of μm) that have higher radiocesium activity (sometimes more than kBq per particle) with irregular shape have been identified in surface soil and dust samples collected in the vicinity of FDNPP within a few km range (Igarashi et al. 2019). The major component of these radiocesium-bearing microparticles (CsMPs) is SiO2 (Satou et al. 2018). What is important, CsMPs are insoluble in water and persistent in the environment (Igarashi et al. 2019; Miura et al. 2020). In terms of the behavior of Fukushima-derived radiocesium, the critical questions regarding CsMPs are as follows: (1) what was the fraction of CsMPs activity in the total radiocesium release and deposition at different locations? and (2) what is the rate of radiocesium leaching from CsMPs as a result of their decomposition? Ikehara et al. (2018, 2020) succeeded to characterize quantitatively the content of CsMPs in surface soils collected in 20 locations of the contaminated area at different directions from the FDNPP. They found that CsMPs account for a significant fraction of FDNPP-deposited radiocesium. Based on the plume trajectories they concluded that CsMPs were formed during a relatively short period from the late afternoon of March 14, 2011, to the late afternoon of March 15, 2011, and that Unit 3 of the FDNPP was the most plausible source of the CsMPs at the beginning of the release. The number of CsMPs in surface soils at various directions, distances, and deposition levels in the soils was in the range of 0.9–101 particles/g and radioactivity fraction of CsMPs was in the range of 15–80%. Occurrence of CsMPs in soils and sediments can actually substantially impact the radiocesium solid–liquid distribution in the soil–water environment (Konoplev 2015). However, the variability of CsMPs fraction in deposition in a relatively wide range makes it difficult to account for

44

A. Konoplev et al.

CsMPs in assessment of mobility and bioavailability of radiocesium in the soil– water environment and its dynamics. Miura et al. (2018) discovered CsMPs in suspended sediments collected from Kuchibuto River (right bank tributary of the Abukuma River flowing from the most contaminated section of the Abukuma River catchment) during 2011–2016. The fraction of radiocesium incorporated in CsMPs to the total radiocesium in the sediments was found to be up to 67%. Another hypothesis explaining high affinity of radiocesium with soil and sediment particles was also discussed in the literature. Naulier et al. (2017) suggested that particulate organic matter could contribute significantly to the increased particulate concentrations of radiocesium in river sediments. This hypothesis, however, is different from the prevalent and generally accepted view in post-Chernobyl radioecology regarding the role of organic matter in radiocesium dynamics (Comans et al. 1998; Valcke and Cremers 1994). Some of post-Fukushima studies have not confirmed this either (Iwagami et al. 2017; Takahashi et al. 2017, 2020), and therefore this hypothesis requires further investigation. High ability of soils and sediments to selectively sorb radiocesium is believed to be due to the presence of clay minerals with the crystal lattice of group 2:1 (Cremers et al. 1988; Konoplev 2020). At least two types of sorption sites can be distinguished for micaceous clay minerals: regular surface ion exchange sites (RES) which are non-selective for cesium, and highly selective sorption sites for cesium occurring in the area of the frayed edges of neighboring layers of micaceous clay crystal lattice (FES) (Fig. 4.3). Radiocesium is adsorbed on RES non-selectively, that is, the selectivity coefficient of its sorption with respect to one-charge ions K+, Na+, NH4+, and others is close to 1. On the other side, the selectivity coefficient for Cs sorption on FES is about 1000 in respect to K+ and about 200 for NH4+ (Wauters et al. 1996). The selective sorption sites FES constitute a relatively small portion of ion exchange sorption sites (1 to 5%) for most soils and sediments (De Preter 1990). Due to high selectivity of FES for cesium and because radiocesium, and actually even stable cesium, occurs at trace concentrations in the environment, the exchangeable radiocesium is almost entirely adsorbed by FES in most sediments and soils.

4.2.3

Radiocesium Solid–Liquid Distribution in Soil–Water Environment

The partitioning of the radionuclide between the sediment and solution is described by the apparent distribution coefficient Kd (L/kg) defined by the ratio of the particulate radionuclide activity concentration [R]p (Bq/kg) to its dissolved activity concentration [R]d (Bq/L) at equilibrium (International Atomic Energy Agency [IAEA] 2010; Konoplev 2015):

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

45

Table 4.3 Distribution coefficients of 137Cs, Kd (L/kg), averaged over several cross sections of river systems in the contaminated areas after the FDNPP accident at early phase (2012–2013) (based on data from Yoshimura et al. 2015a) River system Abukuma Ukedo Niida Оhtа Маnо

River(s) Suriage, Ara, Aise, Otakine, Hirose, Syakadou Ukedo, Takase Niida, Mizunashi Оhtа Маnо

Kd, L/kg 6.2  105 5.0  105 2.0  105 9.3  105 2.1  105

Fig. 4.8 Time variations in the 137Cs distribution coefficient in the Pripyat River after the Chernobyl accident (Konoplev et al. 2020) and in the Ukedo River after the Fukushima Daiichi Nuclear Power Plant accident based on the data from (Nakanishi and Sakuma 2019; Yoshimura et al. 2015a) (from Konoplev et al. 2021b)

Kd ¼

½Rp ½Rd

ð4:1Þ

Table 4.3 summarizes Kd(137Cs) in the basic river systems of the contaminated areas of Fukushima at early stage after the accident. Soon after the FDNPP accident, it was found that apparent Kd values for Fukushima-derived radiocesium in rivers are at least an order of magnitude higher as compared with Chernobyl-derived radiocesium in the rivers of Ukraine, Belarus, and Russia (Nagao et al. 2013; Konoplev 2015; Konoplev et al. 2016a; Takahashi et al. 2017, 2020; Yoshimura et al. 2015a). This fact suggests high ability of the soils and sediments in the Fukushima zone to bind radiocesium which was confirmed by laboratory measurements (Nakao et al. 2014). Recent data for Fukushima Rivers confirmed this conclusion (Nakanishi and Sakuma 2019; Takahashi et al. 2020).

46

A. Konoplev et al.

Table 4.4 Kd(137Cs) values (L/kg) in rivers of the Fukushima area in comparison with the pre-Fukushima data for Japanese rivers and for the major rivers of the Chernobyl area for the initial period after the Chernobyl accident (modified after Konoplev 2015) Observation References River site period Kd(137Cs), L/kg Fukushima Tsuji et al. (2014) Abukuma2012–2013 6.6  105 Fukushima Hiso 2011 2.3  105 Ueda et al. (2013) Wariki 2011 4.8  105 Ueda et al. (2013) Japan, Chernobyl, and nuclear weapons test global fallout Tone 1985–1986 (2.8  1.9)  104 Hirose et al. (1990) Kuji 1987–1988 (10  6)  104 Matsunaga et al. (1991) Chernobyl Pripyat– 1987–1990 (2.0  1.4)  104 French-German Initiative [FGI] (2006), Chernobyl Konoplev et al. (2002b) Uzh– 1987–1990 (3.1  2.3)  104 French-German Initiative [FGI] (2006), Cherevach Konoplev et al. (2002b) Dnieper– 1987–1990 (1.6  1.0)  104 French-German Initiative [FGI] (2006), Nedanchichi Konoplev et al. (2002b)

Fig. 4.8 compares magnitudes and time changes of Kd(137Cs) in Pripyat River at Chernobyl from 1990 to 2016 (Konoplev et al. 2020) and Ukedo River from 2012 to 2018 (Nakanishi and Sakuma 2019; Yoshimura et al. 2015a). Importantly, that in both cases apparent Kd(137Cs) is quite stable in time and does not demonstrate any trend. It should be pointed out that the radiocesium distribution coefficients observed after the Fukushima Daiichi NPP accident considerably exceed not only the Kd values determined in Chernobyl zone, but also the Kd values determined for radiocesium deposited as a result of nuclear weapon tests and Chernobyl accident in Japanese rivers (Table 4.4). Therefore, another possible cause of so high values of Kd could be related to radiocesium speciation in the fallout after the FDNPP accident (Konoplev 2015). For contaminated areas of Fukushima, [R]p includes the radiocesium embedded in glassy hot microparticles (R1 in Fig. 4.3), the radiocesium embedded in organic debris (R3 in Fig. 4.3), the radiocesium fixed by clay minerals (R4 in Fig. 4.3), and the exchangeably sorbed radiocesium (R2 in Fig. 4.3). The exchangeable form of radiocesium occurs at instantaneous ion exchange equilibrium with the liquid phase, whereas the non-exchangeable form is not involved in radiocesium exchange with the solution in the immediate term. It seems therefore appropriate to use the concept of the exchangeable distribution coefficient K ex d , which is the ratio of the exchangeable radionuclide activity concentration in sediments [R]ex to its activity concentration in solution at equilibrium [R]d (Konoplev 2020; Konoplev and Bulgakov 2000; Konoplev et al. 2002a):

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

K ex d ¼

½Rex ¼ αex K d ½Rd

47

ð4:2Þ

where αex is the fraction of the exchangeable radionuclide in sediments, and Kd is the apparent (also known as total) distribution coefficient. Using the exchangeable distribution, coefficient has certain merits because this parameter is governed by physically meaningful characteristics of the environment such as the capacity of sorption sites in sediments and water cation composition (Konoplev 2020; Konoplev and Bulgakov 2000; Konoplev et al. 2002a). Conversely, the value of the apparent Kd is strongly dependent on the non-exchangeable radionuclide fraction which is determined by the rates of radiocesium leaching from glassy hot particles (kl), organic debris decomposition (kd), fixation (transformation of exchangeable to non-exchangeable fixed form) (kf), and remobilization (transformation of fixed form to exchangeable form) (kr). The post-Chernobyl studies have shown that the timescale of radiocesium fixation is weeks or even months, depending on environmental conditions, whereas the timescale of remobilization can be up to a few years (Konoplev and Bulgakov 1996; Konoplev et al. 1992b). However, the rate of radiocesium leaching from glassy Cs-rich microparticles (CsMP) has not been reliably determined for environmental conditions yet. It can be expected that this process is extremely slow, with a timescale of at least several years or more (Okumura et al. 2019). Applying the equation of ion exchange equilibrium for Cs sorption on FES in the presence of potassium (potassium scenario), which is the major environmental competitor for radiocesium, an equation can be formulated to relate the radiocesium exchangeable distribution coefficient to the sediment FES capacity ([FES]) and K+ concentration in solution ([K+]) (Konoplev and Konopleva 1999; Konoplev et al. 2002a; Wauters et al. 1996): þ RIPex ðK Þ ¼ K c ðCs=K Þ½FES ¼ K ex d ½K 

ð4:3Þ

where Kc(Cs/K) is the selectivity coefficient for ion exchange of Cs+ on FES with respect to cation K+. Equation (4.3) defines the exchangeable radiocesium interception potential RIPex(K), which is an intrinsic property of a given sediment and characterizes its ability to sorb cesium selectively and reversibly. In those cases when the relative contributions of potassium and ammonium competition with Cs are comparable, the following ratios are applicable (Konoplev 2020; Konoplev et al. 2002a; Wauters et al. 1996):  þ
þ  FES RIPex ðK Þ ¼ K ex d ½K  þ K c ðNH 4 =K Þ NH 4

ð4:4Þ

where K FES c ðNH 4 =K Þ is the selectivity coefficient of ammonium with respect to potassium for the FES sites.

48

A. Konoplev et al.

Table 4.5 Characteristics of the 137Cs speciation, distribution, and interception potential in the sediment–water system of the Okuma ponds (from Konoplev et al. 2021a) Pond Inkyozaka Bottom sediments (3-cm-deep surface layer) 137 44  5 Cs exchangeable fractiona, % Bottom sediments (3-cm-deep surface layer) 137 3.6 Cs exchangeable fractiona, % Kd, L/kgb 1.4  105 (0.4–3.5)  105 b ex K d , L/kg 5.5  104 (0.2–1.5)  105 ex RIP (K ), meq/kg 1450 a

Suzuuchi

Funasawa

22.6  2.3

19.6  4.4

0.4

3.0

1.2  105 (0.6–2.3)  105 2.9  104 (1.4–5.2)  104 2250

1.6  105 (0.3–3.8)  105 3.3  104 (0.8–7.5)  104 2250

Wakiyama et al. (2019a) Arithmetic mean; in brackets—range

b

As shown by the post-Chernobyl studies, most soils and sediments are characterized by K FES c ðNH 4 =K Þ  5 , and hence Eq. 4.2 can be rewritten as follows (Konoplev 2020; Konoplev et al. 2002a; Wauters et al. 1996): K ex d ¼

RIPex ðK Þ
 ½K  þ 5 NH þ 4 þ

ð4:5Þ

Generally speaking, Eq. 4.5 could be expanded with a term accounting for stable cesium, considering that stable cesium (133Cs) is characterized by higher selectivity for FES than potassium and ammonium: K FES (Cs/K) 1000 (Wauters et al. 1996). c However, the stable Cs concentrations in ponds are more than five orders of magnitude lower than the concentrations of potassium and ammonium (Table 4.1), and therefore, the contribution of the naturally occurring stable Cs to the 137Cs desorption of FES can be ignored compared to potassium and ammonium. Table 4.5 presents data on the 137Cs exchangeability in the surface (top 3 cm) layer of sediment in the three ponds in comparison with similar soils’ data for the surrounding catchment (Wakiyama et al. 2019a). The latter soils were characterized by extremely low fractions of exchangeable 137Cs, including when compared with the Chernobyl values (Bobovnikova et al. 1977; Konoplev 2020; Konoplev et al. 1988), whereas the sediment surface layer, formed as a result of sedimentation, is characterized by a relatively high fraction of exchangeable 137Cs (up to 40% in Inkyozaka). Using the data of the particulate and dissolved 137Cs in the ponds, and the 137Cs exchangeable fraction in the surface layer of bottom sediments, we calculated the apparent distribution coefficient (Kd) and exchangeable distribution coefficient (K ex d ) in the Okuma ponds, and the radiocesium interception potential RIPex using Eq. 4.5 (Table 4.5). The obtained estimates of RIPex(K ), were quite high, ranging from 1450 meq/kg for Inkyozaka to 2250 meq/kg for Suzuuchi and Funasawa. This demonstrates the ability of sediment particles to selectively adsorb

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

a

b 1,0E+05

Kdex(137Cs), L/kg

1,0E+05

Kdex(137Cs), L/kg

49

8,0E+04 6,0E+04

4,0E+04 2,0E+04

8,0E+04 6,0E+04 4,0E+04 2,0E+04

R2=0.85

R2=0.84 0,0E+00

0,0E+00 0

10

20

30

40

0

50

10

Kdex(137Cs), L/kg

30

40

50

([K]+5[NH4])-1, L/meq

([K]+5[NH4])-1, L/meq

c

20

1,0E+05 8,0E+04 6,0E+04 4,0E+04 2,0E+04 R2=0.85 0,0E+00 0

10

20

30

40

50

([K]+5[NH4])-1, L/meq

Fig. 4.9 Correlation between K ex d and inverse effective concentration of the major Cs competitive 

1 cations ½K þ  þ 5 NH þ for the Okuma ponds: (a) Inkyozaka, (b) Suzuuchi, and (c) Funasawa 4 137

Cs, with the results being consistent with the values measured in laboratory experiments (Uematsu et al. 2015; Nakao et al. 2014; Yamaguchi et al. 2017) for the soils typical of the contaminated areas around Fukushima. The ability of the studied sediments to selectively adsorb radiocesium is even higher than of the sediments in Lake Constance in Europe (Konoplev et al. 2002a). For all three ponds studied, a linear trend was obtained for the dependences of K ex d (L/kg) on ([K+] + 5[NH4+])
1 (L/mEq), with the slopes conforming to RIPex(K ) values calculated about 2000 mEq/kg (Fig. 4.9). The apparent Kd(137Cs) values in the Okuma ponds (>105 L/kg) appeared to be about two orders of magnitude higher than those for Chernobyl-derived 137Cs in the closed lakes Svyatoe and Vorsee and close to the values in Lake Constance (Konoplev et al. 1998, 2002a). They are also similar to those in the rivers and dam reservoirs of the FDNPP contaminated area (Konoplev 2015; Yoshimura et al. 2015a). This can be attributed to two factors. The first is the high binding ability of the soils and sediments in the Fukushima-contaminated area (Nakao et al. 2014), and the second factor is the occurrence of insoluble glassy hot particles in the Fukushima fallout (Ikehara et al. 2018) and their higher content in the FDNPP near zone. Konoplev et al. (2020) have detected the slow reduction of annual average 137Cs apparent distribution coefficient Kd in three investigated ponds of Okuma town for

50

A. Konoplev et al.

Fig. 4.10 Time dependence of the annual mean 137Cs apparent distribution coefficients Kd for the suspended sediment–water system in the Okuma town ponds of the FDNPP exclusion zone: Inkyozaka, Suzuuchi, and Funasawa on the basis of data from Konoplev et al. (2021a) Table 4.6 Estimations of 137Cs leaching from CsMPs based on the apparent distribution coefficients time changes in the three Okuma town ponds for sediments–water system (Konoplev et al. 2021a) Pond kl estimate, year
1 T1/2 estimate, year

Inkyozaka 0.15 4.7

Suzuuchi 0.18 3.9

Funasawa 0.12 5.8

longer term (Fig. 4.10). The trend of decline in the apparent distribution coefficient Kd revealed for all three ponds (Fig. 4.10) can be due to 137Cs leaching from glassy hot particles which can make as high as 40% of the deposition inventory in the FDNPP near zone (Ikehara et al. 2018, 2020). The process of decomposition of glassy hot particles and 137Cs leaching from them is believed to be very slow, even slower than that of Chernobyl-derived fuel particles (Beresford et al. 2016; Konoplev 2020). It is much slower than the 137Cs fixation by micaceous clay minerals, the rate constant of which being kf ¼ 4.4–5.5 year
1 (Konoplev et al. 1992b). Therefore, 137Cs leaching from glassy hot particles is a limiting stage in the process of 137Cs transformation in the sediment–water system. This means that the rate constant of Kd’s decline should be corresponding to the rate constant of 137Cs leaching from glassy hot particles (Fig. 4.3). As can be seen from Table 4.6, the rate constants of 137Cs leaching from glassy hot particles kl, estimated using the above assumption, are not more than 0.12–0.18 yr.
1 and the time scale of the corresponding process should be not less than 5–8 years. Bearing in mind the assumptions made in obtaining those values, this is a lower bound estimate. In reality, the process can be even slower. The derived estimates agree with the

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

51

outcomes of laboratory experiments studying radiocesium leaching from CsMPs in various salt solutions (Okumura et al. 2019).

4.3

Radiocesium Vertical Migration in Soil

With time after deposition of radionuclides on the ground surface, the contamination migrates vertically down through the soil profile. The dynamic pattern of vertical distribution of radionuclides in soil is critical from the standpoint of external dose rate, availability of radionuclides for transfer to surface runoff and wind resuspension in the boundary atmospheric layer, and availability of radionuclides for root uptake by plants and penetration to groundwater (Konoplev et al. 2016a). To understand and estimate radionuclide wash-off by surface runoff, it is important to know the radionuclide concentration in the topsoil layer (Bulgakov et al. 2000; Konoplev et al. 2021b). Radionuclides migrate vertically in solution with infiltration water flow or attached to fine soil particles (Bulgakov et al. 1991a; Konoplev and Golubenkov 1991; Konoplev et al. 1992b; Prokhorov 1981; Van Genuchten and Wierenga 1986). Transport of radiocesium in solution by infiltration is slower than the water flow because of sorption–desorption and fixation on soil particles. Fine soil particles containing radiocesium (R1 + R3 + R4) can move by penetrating through pores, cracks, and cavities, as well as with infiltration flow (lessivage), and as a result of vital activity of plants and biota (bioturbation) (Bulgakov et al. 1991b; Bulgakov and Konoplev 2002). Nevertheless, the vertical migration of radionuclides in undisturbed soils can be described by the advection–dispersion equation using the effective values of dispersion coefficient and advective velocity (Prokhorov 1981; Van Genuchten and Wierenga 1986; Bulgakov et al. 2000; Smith et al. 1995). Simultaneous solution of respective equations written separately for specific radiocesium chemical forms in soil Ri (R1, R2, R3, and R4) with allowance for their transformation (Konoplev and Golubenkov 1991; Konoplev et al. 1992b) is the most accurate way of representing radiocesium migration by the advection–dispersion model:   X ∂Ri ∂ ∂Ri ∂R X ¼ kji R j
k ij Ri Di
vi i þ ∂t ∂x ∂x ∂x with the initial conditions Ri ¼ R0i δðx
0Þ and boundary conditions

ð4:6Þ

52

A. Konoplev et al.

Ri jx¼1 ¼ 0 where Di and vi are the effective dispersion coefficient and effective advective velocity for each chemical form i, respectively. It should be kept in mind that the properties of soil governing its sorption and fixation capacity can vary as a function of depth within the soil profile. Additionally, soil characteristics that alter radionuclide migration rates such as porosity, density, and hydraulic conductivity can vary with depth as well. In radioecological studies, simplified versions of the model are often used for one-time releases based on the approximation of analytical solution of Eq. 4.6 for the total radionuclide concentration R ¼ ∑Ri (Bossew and Kirchner 2004; Prokhorov 1981):
σ0 Rðx, t Þ ¼ pffiffiffiffiffiffiffiffiffiffiffiffi e πDeff t



ðx
vt Þ2 4Deff t þλt

ð4:7Þ

where σ0 is the initial radionuclide deposition on soil and λ is the rate constant of radionuclide decay. This approximation is valid for long term: t > > 2D/v2 (Bossew and Kirchner 2004). The studies of vertical migration of radiocesium in undisturbed soils of grassland and forests show that as a rule 137Cs transport due to dispersion prevails over advective transport (Bulgakov et al. 1991a; Ivanov et al. 1997; Konoplev et al. 2016b, 2020). Therefore, Eq. 4.7 can be further simplified as
σ0 Rðx, t Þ ¼ pffiffiffiffiffiffiffiffiffiffiffiffi e πDeff t



x2 4Deff t þλt

ð4:8Þ

Figure 4.11 shows as an example the vertical distribution of Fukushima-derived radiocesium in four soil cores of FDNPP exclusion zone in Okuma town collected 3 years after the accident. Eq. (4.8) is used to derive Deff for these vertical distributions which are presented in Table 4.7 (Konoplev et al. 2016a). The obtained effective dispersion coefficients in forest and grassland soils of the Fukushima near field exceeded the values for the Chernobyl near field (Bulgakov et al. 1991a; Konoplev et al. 1992b). Faster migration of radiocesium in the Fukushima area, as compared to the Chernobyl zone, was also reported in another work (Kato et al. 2012) for the surface layer of cultivated soil at earlier phase after the accident. There are several factors which can contribute to this. First, the mean annual precipitation in the Fukushima area is more than 2 times higher than that in the Chernobyl area (Laceby et al. 2016; Konoplev et al. 2016a). As a result, a more active infiltration flow can lead to higher migration rates of both mobile forms of radiocesium (R2) and immobile radiocesium (R1 + R3 + R4), which are entrained by infiltration flow when moving down the soil through pores and cracks (Bulgakov et al. 1991a; Konoplev et al. 2016b).

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

a

Activity concentration, Bq/kg 1,0E+01

1,0E+02

1,0E+03

1,0E+04

b

1,0E+05 1,0E+06

1,0E+02

1,0E+03

1,0E+04

1,0E+05

0-5

Depth, cm

4-7

Depth, cm

Activity concentration, Bq/kg 1,0E+01

0-4

53

7-10 10-13 13-18

5-10

10-15

18-23 15-20

23-29 Cs-134 Cs-137

c

Cs-134

Activity concentration, Bq/kg

d

Activity concentration, Bq/kg 1,0E+01

1,0E+00 1,0E+01 1,0E+02 1,0E+03 1,0E+04 1,0E+05 1,0E+06 0-3

Cs-137

1,0E+02

1,0E+03

1,0E+04

1,0E+05

0-2

3-5

Depth, cm

Depth, cm

2-5 5-8 8-13 13-18

5-10

10-20

18-24 20-28.5

24-30 Cs-134

Cs-134

Cs-137

134

Cs-137

137

Fig. 4.11 Radiocesium ( Cs and Cs) vertical distribution in soils on the catchments of the Okuma town water bodies: pond Suzuuchi (a); Sakashita Dam reservoir (b); pond Inkyozaka (c) and pond Kashiramori (d) in the close-in area of the FDNPP (from Konoplev et al. 2016a) Table 4.7 137Cs effective dispersion coefficients for undisturbed soils from the water bodies catchments in Okuma town in the vicinity of the FDNPP 3 years after the accident (from Konoplev et al. 2016a) Site Suzuuchi Inkyozaka Kashiramori Sakashita.

Land use Grassland Grassland Forest Forest

Soil type Alluvial sandy Alluvial meadow sandy Light brown loamy Loamy organic

137

Cs deposition, kBq/m2 5900 4250 865 672

Deff, cm2/year 2.5 2.2 9.3 5.0

Second, the bioturbation rate in the upper soil layer of the Fukushima zone may be higher due to plant growth and root systems, as well as due to soil fauna activity. Given the geoclimatic features of the Fukushima zone, the biological activity in the surface soil is higher as compared to the Chernobyl zone (Bulgakov and Konoplev 2002). Apart from that, temperature regime in the soil should be taken into account as well. For the Chernobyl zone, low temperatures in the surface soil layers occur when the soil is frozen. The temperature regime is known to have a significant impact on the pace of physical and chemical processes in soil. Importantly, all radiocesium vertical distributions demonstrated maximum of its activity concentrations in the topsoil layer. Even 30 years after the Chernobyl

54

A. Konoplev et al.

accident, this is valid for most soils in the Chernobyl exclusion zone (Konoplev et al. 2020). To evaluate the ambient dose equivalent rates which are strongly dependent on radionuclide vertical distribution in soil for the early phase after the accident, a purely empirical exponential model has been used in both post-Chernobyl (Bossew and Kirchner 2004) and post-Fukushima studies (Malins et al. 2016; Matsuda et al. 2015; Takahashi et al. 2015): Rðx, t Þ ¼ R0 ðt Þeð
βÞ x

ð4:9Þ

where β is the empirical parameter, called the relaxation length; R0(t) is the radiocesium activity concentration in the topsoil layer at x ¼ 0 for the time t. Since the bulk density changes essentially with depth, instead of the actual depth x (cm), the mass depth xm (g/cm2) is often used: Zx xm ¼

ρðxÞdx

ð4:10Þ

0

where ρ (g/cm3) is the soil bulk density. However, this empirical exponential model is unable to describe radiocesium long-term dynamics in soil and specifically in the topsoil layer which is critically important for consideration of water bodies’ secondary contamination by surface runoff.

4.4

Radiocesium Wash-off from Contaminated Catchments and River Transport

After the FDNPP accident, the river basins of the area were exposed to 137Cs deposition and became a long-term source of natural waters contamination due to its wash-off by surface runoff (Konoplev 2016; Konoplev et al. 2021b; Wakiyama et al. 2019b; Yoshimura et al. 2015b).

4.4.1

Quantitative Characteristics of Radionuclide Wash-off and their Parameterization

The most important quantitative characteristic of radionuclide wash-off used for predicting secondary contamination of water bodies is the wash-off coefficient, representing the fraction of radionuclide inventory in the catchment transported by surface runoff to water bodies (Bobovnikova et al. 1977; Borzilov et al. 1988;

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

55

Makhon’ko et al. 1977; Pisarev et al. 1972). In surface runoff, radionuclides can occur in dissolved and particulate forms. Since the physicochemical state of the radionuclide, to a great extent, determines its further behavior in surface waters, dissolved and particulate wash-off should be considered separately. The particulate wash-off coefficient (Wp) and dissolved wash-off coefficient (Wd) represent the fractions of radionuclide washed off from the catchment as suspended particles and in solution, respectively. They are determined for the period of interest or a single runoff event and are defined as follows (Borzilov et al. 1988; Garcia-Sanchez and Konoplev 2009; Konoplev 2016): RT Wp ¼

0

cp ðt Þmðt Þdt cp M R ¼ ; Wd ¼ σS s σ ðsÞds

RT 0

cd ðt ÞQðt Þdt cd V R ¼ σS s σ ðsÞds

ð4:11Þ

where cp(t) and cp are the instantaneous and weighted average radionuclide activity concentrations on suspended particles (Bq/g), respectively; m(t) and М are the erosion rate (g/s) and the sediment yield (g) for the period of interest or the single runoff event, respectively; cd(t) and cd are the instantaneous and weighted average radionuclide concentrations in solution (Bq/L), respectively; Q(t) is the runoff rate (L/s); V is the total runoff volume (L); σ(s) is the local radionuclide deposition for the watershed section ds (Bq/m2); σ is the radionuclide mean deposition over the entire catchment area (Bq/m2); S is the catchment area (m2); 0 and T are the starting time and duration of the runoff event or the period of interest(s), respectively; and t is the time (s). The values of the wash-off coefficients thus defined depend on the hydrological characteristics of a runoff event or period of interest such as the runoff volume from a catchment area (runoff layer) (Wd) and the mass of eroded material from the area (Wp). Usually, the average wash-off coefficient for 1 year of observations, that is, the fraction of radionuclide washed off the catchment over 1 year, is used (Bobovnikova et al. 1977). For research and prediction purposes, however, wash-off coefficients for events with different time scales are required (from a single shower or snowmelt event forming surface runoff to long-term multi-year observations). Therefore, for practical application, the wash-off coefficients have been normalized through the runoff characteristics essential for radionuclide wash-off. For dissolved wash-off, this is the runoff volume divided by the area of catchment, that is, runoff depth, while the particulate wash-off coefficient is proportional to the sediment yield from the catchment area (Borzilov et al. 1988; Konoplev et al. 1992b). Dividing Wd by the runoff depth and Wp by the mass of eroded material from the area, we get the normalized dissolved and particulate wash-off coefficients, which, according to Eq. 4.11, are equal to the ratio of the corresponding mean activity concentration and the average deposition of the radionuclide on the catchment:

56

A. Konoplev et al.

Nd ¼

W p W p S сp W d W d S cd ¼ ; Np ¼ ¼ ¼ ¼ V h σ q σ M

ð4:12Þ

where h is the runoff depth (m), q is the mass of eroded material from the catchment area (g/m2), Nd is the normalized dissolved wash-off coefficient (m
1) for the period of interest, and Np is the normalized particulate wash-off coefficient (m2/g) for the period of interest. We will refer to these parameters as Nd, dissolved wash-off ratio, and Np, particulate wash-off ratio. The proportionality of the dissolved wash-off coefficient Wd and runoff depth was noted in early studies of radionuclide wash-off resulting from the global nuclear weapon tests fallout (Rovinsky et al. 1976; Rovinsky and Sinitsyna 1979). To study the time dependence of wash-off characteristics, the use of instantaneous wash-off ratios is convenient (Bulgakov et al. 1991b; Konoplev et al. 2021b): nd ¼

cp ð t Þ cd ð t Þ ; np ¼ σ σ

ð4:13Þ

A major advantage of the wash-off ratios is that they can be used to predict the radionuclide wash-off from contaminated catchments and their concentrations in rivers and other water bodies (Borzilov et al. 1989, 1993). To estimate the fraction of the radionuclide washed off in solution, the dissolved wash-off ratio is multiplied by the expected runoff depth for a given runoff event or period of interest. The fraction of radionuclide washed off with sediments is estimated by multiplying the particulate wash-off ratio by the predicted sediment yield during the runoff event or period of interest (Borzilov et al. 1988; Konoplev et al. 1992b).

4.4.2

Semi-Empirical Modeling of Mid- to Long-Term Dynamics of Radionuclide Wash-off from Contaminated Catchments

To parameterize wash-off characteristics and their mid- to long-term dynamics, a semi-empirical model of radionuclide wash-off was proposed by Konoplev et al. (2020). The model is based on the presumption that the main source of suspended particles for surface runoff is the top layer of catchment soil. Multiple studies of wash-off of contaminants of different nature have shown that the depth of the soil layer from which the contaminant is effectively transferred to the surface runoff is several millimeters (up to 1 cm) (Ahuja et al. 1981; Borzilov et al. 1989; Bulgakov et al. 1991b, 2000; Donigian Jr et al. 1977; Knisel 1980). The radionuclide concentration in topsoil decreases over time due to its vertical migration to deeper soil layers. The analysis of the vertical distribution of radionuclides, such as 137Cs and 241 Am, strongly bound to soil particles, has indicated that as late as 30 years after the Chernobyl accident the highest concentrations of these radionuclides occurred in the

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

57

top layer of undisturbed meadow and forest soils (Konoplev et al. 2020). Therefore, the vertical migration of 137Cs in soil after a one-time fallout from the atmosphere, as is the case with a nuclear accident, with certain simplifications, can be approximated by a “quasi-diffusion” model and described by Eq. 4.8. In this case, the time dependence of radionuclide concentration in the topsoil layer (at x ¼ 0), and hence on suspended particles entrained by the surface runoff, is approximated by the following equation (Konoplev et al. 2020): σ сp ðtÞ ¼ pffiffiffiffiffiffiffiffiffiffiffiffi e
λt ρ πDeff t

ð4:14Þ

Here, it is assumed that radionuclide deposition and land-use types are uniform across the catchment area. The dissolved radionuclide concentration in the surface runoff or river is related to the particulate radionuclide concentration through the distribution coefficient Kd by Eq. 4.1. More than 30 years of 137Cs monitoring in the major rivers Pripyat and Dnieper of the Chernobyl zone has shown that after the early phase (1–2 years), the Kd(137Cs) value displayed some variation, but no statistically significant trend was observed (Fig. 4.8) (Konoplev et al. 2020). Then, with allowance for Eqs. 4.14 and 4.1, the time dependence of dissolved 137Cs concentration in a river can be approximated by the following equation (Konoplev et al. 2020): сd ð t Þ ¼

σ pffiffiffiffiffiffiffiffiffiffiffiffi e
λt ρKd πDeff t

ð4:15Þ

Substituting Eqs. (4.14) and (4.15) into Eq. 4.13, and assuming the 137Cs deposition in the catchment decreases with time only due to radioactive decay, that is, ignoring the minor losses due to wash-off, soil–plant transfer and wind resuspension with subsequent atmospheric transport, we get the following (Konoplev et al. 2021b):  pffiffiffiffiffiffiffiffiffiffiffi
1 n0p 1 np ðt Þ ¼ pffiffiffiffiffiffiffiffiffiffiffiffi ¼ pffi ; n0p ¼ ρ πDeff t ρ πDeff t nd ð t Þ ¼

pffiffiffiffiffiffiffiffiffiffiffi
1  n0 1 pffiffiffiffiffiffiffiffiffiffiffiffi ¼ pdffi ; n0d ¼ ρK d πDeff t ρK d πDeff t

ð4:16Þ ð4:17Þ

Equations (4.16) and (4.17) estimate the wash-off ratios after the early phase of a nuclear accident and predict the mid- and long-term dynamics of radionuclide washoff. It should be emphasized that the equations are not applicable for the first days, weeks, and months after the accident, because at t ! 0 eqs. (4.16) and (4.17) are meaningless. The proposed semi-empirical “diffusional” approach enables prediction of the radionuclide wash-off and secondary contamination of surface waters for the

58

A. Konoplev et al.

mid- and long-term phases after a nuclear accident, using only two key physicochemical parameters of radionuclide dispersion and distribution in the sediment– water system. The main parameter determining particulate radionuclide wash-off is the effective dispersion coefficient (Deff) for catchment topsoil, which depends on radionuclide chemical properties, sorption and fixation ability of catchment soil, and climatic conditions (e.g., mean annual rainfall, mean annual air temperature, etc.). Another critical parameter controlling dissolved radionuclide wash-off, besides Deff, is the distribution coefficient Kd in the sediment–water system. These parameters can be estimated from literature data regarding vertical migration of radionuclides in soil and their distribution in the sediment–water system (Bulgakov et al. 1991a, 2000; Ivanov et al. 1997; Konoplev 2015; Konoplev et al. 2016b). The values n0p and n0d can also be derived from monitoring data of the first few years after the accident and then used for mid- and long-term predictions (Konoplev et al. 2021b). An alternative approach to modeling the dynamics of radionuclide wash-off is P application of an empirical multi-exponential transfer function f ðtÞ ¼ wi e
ki t . This empirical model was developed and widely used in post-Chernobyl studies (Garcia-Sanchez 2008; Hilton 1997; Monte et al. 2004; Smith et al. 2005). Such an approach was followed by a number of researchers in post-Fukushima investigations of radiocesium dynamics in rivers (Funaki et al. 2020; Iwagami et al. 2017; Takahashi et al. 2020; Yoshimura et al. 2016). The semi-empirical diffusional model has important advantages as compared to the empirical multi-exponential model. For one thing, the diffusional model has only two parameters Deff and Kd which have a clear physical meaning. Equally as important, it integrates the processes of radionuclide migration in catchment soils with the processes of its wash-off and river transport. Notably, the diffusional model is valid for long-lived radionuclides strongly bound to soil particles, such as 137Cs, but cannot be applied to short-lived radioisotopes and radionuclides passing into solution easily, as the conditions are not met in such cases (Konoplev et al. 2021b). Another constraint is associated with a potential impact of remediation. It is assumed by the model that radionuclide concentration in the topsoil layer changes due to natural environmental processes. In case of remediation, the natural environment is disturbed and, strictly speaking, application of the model can be problematic. However, if the scale and efficacy of remediation for a specific catchment are known, its impact can be allowed through certain corrections.

4.4.3

Analysis of Fukushima-Derived Radiocesium Wash-off from Contaminated Catchments and its Mid- to Long-Term Dynamics

Characteristics of Fukushima-derived radiocesium wash-off at early phase after the FDNPP accident have been calculated and summarized for Fukushima Rivers

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

59

Table 4.8 Comparison of radiocesium wash-off ratios for catchments in the Fukushimacontaminated areas and Chernobyl area for initial periods (1–2 years) after the accidents (updated after Konoplev et al. 2016a). 137

River Fukushima Ukedo (Ukedo) Hiso

Period

Cs average deposition, kBq/m2

Nd, m
1

Np, m2/kg

Reference

12.2012

2500

1.3  10–4

2.2  10
2

07–11.2011

2000

5.0  10
4

7.5  10
2

Wariki

07–11.2011

2000

2.0  10
4

8.9  10
2

Niida (Haramachi) Odaka

12.2012

890

0.8  10
4

1.5  10
2

12.2012

730

1.0  10
4

1.1  10
2

Takase

12.2012

690

0.5  10
4

4.1  10
2

Mano

12.2012

490

2.9  10
4

6.4  10
2

Kuchibuto (Yamakiya) Hirose (Tsukidate) Abukuma (Fukushima) Abukuma (Iwanuma) Chernobyl Dnieper (Nedanchichi) Pripyat (Chernobyl)

12.2012

440

0.3  10
4

3.0  10
2

12.2012

230

0.6  10
4

6.7  10
2

01.2013

116

1.7  10
4

6.9  10
2

12.2012

88

1.6  10
4

3.8  10
2

Yoshimura et al. (2015a) Ueda et al. (2013) Ueda et al. (2013) Yoshimura et al. (2015a) Yoshimura et al. (2015a) Yoshimura et al. (2015a) Yoshimura et al. (2015a) Yoshimura et al. (2015a) Yoshimura et al. (2015a) Tsuji et al. (2014) Yoshimura et al. (2015a)

1988

147

14  10
4

4.5  10
2

1988

97

18  10
4

6.7  10
2

DB “RUNOFF” DB “RUNOFF”

(Konoplev 2016; Konoplev et al. 2016a). They were also compared with correspondent values for Chernobyl Rivers at early phase after the Chernobyl accident. Table 4.8 presents radiocesium dissolved and particulate wash-off ratios (Nl and Ns) calculated for rivers in contaminated areas at early phase after the Fukushima accident in 2011–2013 on the basis of available radiocesium monitoring data in rivers (Ueda et al. 2013; Tsuji et al. 2014; Yoshimura et al. 2015a) and for Chernobyl area in 1986–1988 according to DB “RUNOFF” (French-German Initiative [FGI] 2006; Konoplev et al. 2002a, b). The values of Ns are comparable for Fukushima and Chernobyl areas when compared at similar times post-accident. The same conclusion can be made comparing data on Ns obtained in runoff plot experiments in Fukushima (Yoshimura et al. 2015b; Wakiyama et al. 2019b) with data from Chernobyl (Borzilov et al. 1988; Garcia-Sanchez et al. 2005; Garcia-Sanchez and Konoplev 2009; Konoplev et al. 1992b) for early time after the accident. Ns is the

60

A. Konoplev et al.

Fig. 4.12 Comparison of magnitudes and time variations of the mean annual values of particulate 137 Cs wash-off ratios (Np) for the catchments of the Pripyat River at Chernobyl and the Dnieper River at Nedanchichi in the Chernobyl area (based on data from Konoplev et al. 2020), and for the Ukedo and Ohta Rivers in the Fukushima area (based on data from Nakanishi and Sakuma 2019; Takahashi et al. 2020; Yoshimura et al. 2015a), and comparison with the semi-empirical diffusional modeling at Deff ¼ 0.5 cm2/year and Deff ¼ 5 cm2/year (after Konoplev et al. 2021b)

ratio of the mean particulate concentration and mean inventory on the catchment area. In turn, the particulate concentration is close to its concentration in the upper erosion soil layer. For the first years after the accidents, the ratios were comparable for Chernobyl and Fukushima. For some Fukushima Rivers, Ns is slightly less than in Chernobyl, most likely, because of higher values Deff in Fukushima (Konoplev et al. 2021b). This is illustrated in Fig. 4.12 comparing magnitudes and time variations of the mean annual values of particulate 137Cs wash-off ratios (Np) for the catchments of the Pripyat River at Chernobyl and the Dnieper River at Nedanchichi in the Chernobyl area (based on data from Konoplev et al. 2020), and for the rivers Ukedo and Ohta in the Fukushima area (based on data from Nakanishi and Sakuma 2019; Takahashi et al. 2020; Yoshimura et al. 2015a), and comparing with the semi-empirical diffusional modeling at Deff ¼ 0.5 cm2/year and Deff ¼ 5 cm2/year. The comparison with the Ukedo and Ohta Rivers in the FDNPP zone shows that the Np values for these rivers are 2–3 times lower than those for the Pripyat River (Chernobyl) and similar to those for the Dnieper River (Nedanchichi). This is consistent with the higher Deff(137Cs) values in the Fukushima soils compared to the typical Deff(137Cs) values for the Chernobyl soils, which is associated with higher annual precipitation, higher mean annual air temperature, and higher biological activity in the Fukushima soils (Konoplev et al. 2016a,b). The Np(137Cs) value and its time dependence for the Ukedo and Ohta Rivers in the FDNPP zone are in

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

61

Fig. 4.13 Mean annual dissolved 137Cs wash-off ratios (Nd) from the catchments of the Ukedo and Ohta Rivers in the Fukushima area (based on data from Nakanishi and Sakuma 2019; Yoshimura et al. 2015a, b) in comparison with those for Pripyat River at Chernobyl and the Dnieper River at Rechitsa in the Chernobyl area (based on data from Konoplev et al. 2020), and comparison with the estimation by the semi-empirical diffusional model. The Y-axis is log-scaled (after Konoplev et al. 2021b)

agreement with the calculations using semi-empirical diffusional model at Deff(137Cs) of about 5 cm2/year. Figure 4.13 shows the magnitudes and their time changes of the mean annual dissolved 137Cs wash-off ratios Nd (m
1) for two river catchments (Ukedo and Ohta) in the Fukushima area calculated on the basis of published data (Nakanishi and Sakuma 2019; Yoshimura et al. 2015a, b) in comparison with two river catchments Pripyat (Chernobyl) and Dnieper (Rechitsa) in the Chernobyl area calculated on the basis of data (Konoplev et al. 2020). For the catchments of the Ukedo and Ohta Rivers in the Fukushima area, the Nd(137Cs) values are an order of magnitude lower than those for the Chernobyl catchments, which confirms previous findings (Konoplev 2016; Konoplev et al. 2016a) for the first years after the FDNPP accident (Table 4.8). This difference is mainly explained by the values of Kd(137Cs), at least an order of magnitude higher for most of the rivers in the Fukushima area (Konoplev 2015) (Fig. 4.8). The changes in Nd(137Cs) values over time for both catchments of the Ukedo and Ohta Rivers are described by the diffusional model for dissolved 137Cs wash-off using a catchment area average of Deff(137Cs) ¼ 5 cm2/year (Konoplev et al. 2016b) and mean Kd(137Cs) ¼ 2.5  105 L/kg (Konoplev 2015; Nakanishi and Sakuma 2019; Taniguchi et al. 2019).

62

4.5

A. Konoplev et al.

Conclusions

Fukushima-derived radiocesium is strongly bound by soil and sediment particles. Radiocesium apparent distribution coefficient Kd in Fukushima Rivers is considerably (at least an order of magnitude) higher than that in rivers of the Chernobyl area, which is most likely due to two reasons: high binding ability of soils and sediments in Fukushima-contaminated areas, and the presence of water-insoluble hot glassy microparticles (CsMP) in the Fukushima accidental fallout. Studies of the 137Cs behavior in ponds in the vicinity of FDNPP have shown that the concentrations of this radionuclide are higher in ponds than in rivers and dam reservoirs of the region. The highest levels of the dissolved 137Cs in the studied ponds were observed from June to October. Additional dissolution of 137Cs in the summer can be attributed to temperature dependence of 137Cs desorption from FES and its remobilization by ammonium. Despite radiocesium in closed and semi-closed ponds being relatively persistent, a decline in both particulate and dissolved 137Cs activity concentrations was revealed. The reduction rate of particulate 137Cs activity concentrations was much higher than that for dissolved 137Cs. Thus, the apparent distribution coefficient Kd(137Cs) in the sediment–water system decreased with the rate constant 0.12–0.18 yr.
1. Proceeding on the assumption that the decrease in Kd is associated with the decomposition of glassy Cs-rich microparticles, the timescale of 137Cs leaching from them in the ponds under study was estimated to be 5–8 years. The obtained estimates are consistent with the findings of recent laboratory experiments. Higher mean annual precipitation and air temperature promote faster vertical and lateral radiocesium migration in Fukushima as compared with Chernobyl. Wash-off is the principle long-term process responsible for radiocesium secondary contamination of surface waters on accidentally contaminated areas. For its characterization, the particulate and dissolved wash-off ratios are used. Both were found to decrease in the mid- and long-term as a result of radiocesium depletion in the topsoil layer due to its vertical migration in catchment soils. The processes of wash-off, river transport, and radionuclide vertical migration in catchment soils were considered in an integrated way by the semi-empirical diffusional model. This approach enables description of changes in the particulate and dissolved 137Cs wash-off ratios using only two physically meaningful parameters Deff—the 137Cs effective dispersion coefficient in the topsoil layer and Kd—the 137Cs apparent distribution coefficient. Particulate 137Cs wash-off ratios from the catchments of the Fukushima area display only minor differences compared to those in the Chernobyl area, being at the lower limit of the Chernobyl values. Somewhat lower values of Np(137Cs) in the Fukushima area are explained by higher values of the effective dispersion coefficient Deff(137Cs) in typical Fukushima soils. Dissolved 137Cs wash-off ratios for Fukushima catchments are at least an order of magnitude lower than those for Chernobyl, mainly due to an order of magnitude difference in the 137Cs distribution coefficients for the Fukushima and Chernobyl Rivers.

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

63

Acknowledgments This research was supported by the Japan Society for the Promotion of Science, Grant-in-Aid for Scientific Research (B) (KAKENHI 18H03389).

References Abe Y, Iizawa Y, Terada Y, Adachi K, Igarashi Y, Nakai I (2015) Detection of uranium and chemical state analysis of individual radioactive microparticles emitted from the Fukushima nuclear accident using multiple synchrotron radiation X-ray analyses. Anal Chem 88:8521– 8525 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 3:2554 Ahuja LR, Sharpley AN, Yamamoto M, Menzel RG (1981) The depth of rainfall-runoff-soil interaction as determined by 32P. Water Resource Res 17:969–974 Antsiferov SM, Kosyan RD (1986) Suspended material in the upper part of the continental shelf. Nauka, Moscow, p 224. (in Russian) Beresford N, Fesenko S, Konoplev A, Skuterud L, Smith JT, Voigt G (2016) Thirty years after the Chernobyl accident: what lessons have we learnt? J Environ Radioact 157:77–89 Bobovnikova TI, Avramenko AS, Makhon’ko KP, Dibtseva AA, Volokitin AA, Chumichev VB (1977) Strontium-90 concentration in surface waters on the territory of the USSR. Meteorologia i Gidrologia No 9:56–61. (In Russian) Borzilov VA, Konoplev AV, Revina SK, Bobovnikova TI, Lyutik PM, Shveikin YV, Shcherbak AV (1988) An experimental study of the washout of radionuclides fallen on soil in consequence of the Chernobyl failure. Soviet Meteorol Hydrol. 11:27–33 Borzilov VA, Sedunov YS, Novitskii MA, Vozzhennikov OI, Konoplev AV, Dragolyubova IV (1989) Physico-mathematical modeling of washout of long-lived radionuclides from drainage basins in the 30-km zone around the Chernobyl nuclear Power Station. Soviet Meteorol Hydrol 1:1–8 Borzilov VA, Novitsky MA, Konoplev AV, Vozzhennikov OI, Gerasimenko AC (1993) A model for prediction and assessment of surface water contamination in emergency situations and methodology of determining its parameters. Rad Prot Dosim 50(2–4):349–351 Bossew P, Kirchner G (2004) Modelling the vertical distribution of radionuclides in soil. Part 1: the convection-dispersion equation revisited. J Environ Radioact 73:127–150 Bulgakov AA, Konoplev AV (2002) Modelling of 137Cs vertical soil transfer by a tree root system. Radiation biology. Radioecology 42(5): 556–560 (In Russian) Bulgakov AA, Konoplev AV, Popov VE, Bobovnikova TI, Siverina AA, Shkuratova IG (1991a) Mechanisms of vertical migration of long-lived radionuclides in soils within 30 kilometers of the Chernobyl nuclear power station. Soviet Soil Sci 23:46–51 Bulgakov AA, Konoplev AV, Popov VE, Scherbak AV (1991b) Removal of long-lived radionuclides from soil by surface runoff near the Chernobyl nuclear power station. Soviet Soil Sci 23: 124–131 Bulgakov AA, Konoplev AV, Shkuratova IG (2000) Distribution of 137Cs in the topmost soil layer within a 30-km zone around the Chernobyl nuclear power plant. Pochvovedenie 9:1149–1152. (In Russian) Chaisan K, Smith JT, Bossew P, Kirchner G, Laptev GV (2013) Worldwide isotope ratios of the Fukushima release and early-phase external dose reconstruction. Sci Rep 3:2520 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 the Fukushima Daiichi nuclear power plant into the atmosphere. J Nucl Sci Technol 48:1129–1134

64

A. Konoplev et al.

Comans RNJ, Middelburg JJ, Zonderhuis J, Woittiez JRW, De Lange GJ, Das HA, Van der Weijden CH (1989) Mobilization of radiocaesium in pore water of lake sediments. Nature 339:367–369 Comans RNG, Hilton J, Voitsekhovich O, Laptev G, Popov V, Madruga MJ, Bulgakov A, Smith JT, Movchan N, Konoplev A (1998) A comparative study of radiocesium mobility measurements in soils and sediments from the catchment of a small upland oligotrophic lake (Devoke water, UK). Water Res 32(9):2846–2855 Cremers A, Elsen A, De Preter P, Maes A (1988) Quantitative analysis of radiocesium retention in soils. Nature 335:247–249 De Preter P (1990) Radiocaesium retention in aquatic, terrestrial and urban environment: a quantitative and unifying analysis. Ph.D. Thesis. K.U. Leuven, Belgium, 93 p Donigian AS, Jr, Beyerlein DC, Davis HH, Crawford NH (1977) Agricultural runoff management (ARM) model, version II: refinement and testing. EPA 600/3-77-098, environ. Res. Lab., US EPA, Athens, GA, U.S. government printing office, Washington, DC Evrard O, Chartin C, Onda Y, Lepage H, Cerdan O, Lefevre I, Ayrault S (2014) Renewed soil erosion and remobilization of radioactive sediment in Fukushima coastal rivers after the 2013 typhoons. Sci Rep 4:4574 Evrard O, Laceby JP, Lepage H, Onda Y, Cerdan O, Ayrault S (2015) Radiocesium transfer from hillslopes to the Pacific Ocean after the Fukushima nuclear power plant accident: a review. J Environ Radioact 148:92–110 French-German Initiative [FGI] (2006) The French-German Initiative for Chernobyl. Program 2: Study of the radioecological consequences. GRS/IRSN: 109 p Fukushima T, Arai H (2014) Radiocesium contamination of lake sediments and fish following the Fukushima nuclear accident and their partition coefficient. Inland Waters 4:204–214 Funaki H, Sakuma K, Nakanishi T, Yoshimura K, Katengeza EW (2020) Reservoir sediments as a long-term source of dissolved radiocesium in water system: a mass balance case study of an artificial reservoir in Fukushima, Japan. Sci Total Environ 743:140668 Garcia-Sanchez L (2008) Watershed wash-off of atmospherically deposited radionuclides: review of the fluxes and their evolution with time. J Environ Radioact 99:563–573 Garcia-Sanchez L, Konoplev AV (2009) Watershed wash-off of atmospherically deposited radionuclides: a review of normalized entrainment coefficients. J Environ Radioact 100:774–778 Garcia-Sanchez L, Konoplev A, Bulgakov A (2005) Radionuclide entrainment coefficients by wash-off derived from plot experiments near Chernobyl. Radioprotection 40(C1):S519–S524 Hilton J (1997) Aquatic radioecology post Chernobyl – a review of the past and look to the future. In: Desmet G, Blust RJ, Comans RNJ, Fernandez JA, Hilton J, De Bettencourt A (eds) Freshwater and Estuarine Radioecology. Elsevier. 47–74 Hirose K (2012) 2011 Fukushima Dai-ichi nuclear power plant accident: summary of regional radioactive deposition monitoring results. J Environ Radioact 111:13–17 Hirose K, Aoyama M, Sugimura Y (1990) Plutonium and cesium isotopes in river waters in Japan. J Radioanal Nucl Chem 141(1):191–202 Igarashi Y, Kogure T, Kurihara Y, Miura H, Okumura T, Satou Y, Takahashi Y, Yamaguchi N (2019) A review of Cs-bearing microparticles in the environment emitted by the Fukushima Dai-ichi nuclear power plant accident. J Environ Radioact 205-206:101–118 Ikehara R, Suetake M, Komiya T, Furuki G, Ochiai A, Yamasaki S, Bower WR, Law GTW, Ohnuki T, Grambow B, Ewing RC, Utsunomiya S (2018) Novel method of quantifying radioactive cesium-rich microparticles (CsMP) in the environment from the Fukushima Daiichi nuclear power plant. Environ Sci Technol 52:6390–6398 Ikehara R, Morooka K, Suetake M, Komiya T, Kurihara E, Takehara M, Takami R, Kino C, Horie K, Takehara M, Yamasaki S, Ohnuki T, Law GTW, Bower W, Grambow B, Ewing RC, Utsunomiya S (2020) Abundance and distribution of radioactive cesium-rich microparticles released from the Fukushima Daiichi nuclear power plant into the environment. Chemosphere 241:125019

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

65

International Atomic Energy Agency [IAEA] (2006) Applicability of Monitored Natural Attenuation at Radioactively Contaminated Sites. Technical Reports Series No. 445. Vienna: 105 p International Atomic Energy Agency [IAEA] (2010). Handbook of parameter values for the prediction of radionuclide transfer in terrestrial and freshwater environments. Technical reports series No. 472. Vienna: 194 p Ivanov YA, Lewyckyj N, Levchuk SE, Prister BS, Firsakova SK, Arkhipov NP, Kruglov SV, Alexakhin RM, Sandalls J, Askbrant S (1997) Migration of 137Cs and 90Sr from Chernobyl fallout in Ukrainian, Belarussian and Russian soils. J Environ Radioact 35:1–21 Iwagami S, Onda Y, Tsujimura M, Abe Y (2017) Contribution of radioactive 137Cs discharge by suspended sediment, coarse organic matter, and dissolved fraction from a headwater catchment in Fukushima after the Fukushima Dai-ichi nuclear power plant accident. J Environ Radioact 166:466–474 Kaneyasu N, Ohashi H, Suzuki F, Okuda T, Ikemori F (2012) Sulfate aerosol as a potential transport medium of radiocesium from the Fukushima nuclear accident. Environ Sci Technol 46:5720– 5726 Kato H, Onda Y, Teramage M (2012) Depth distribution of 137Cs, 134Cs and 131I in soil profile after Fukushima Dai-ichi nuclear power plant accident. J Environ Radioact 111:59–64 Knisel WG (ed) (1980) CREAMS: a field-scale model for chemicals runoff, and erosion from agricultural management systems. US DA Conservation Research Report No 26:643 p Konoplev AV (2015) Distribution of radiocesium of accidental origin between the suspended alluvium and solution in rivers: comparison of Fukushima and Chernobyl. Radiochemistry 57(5):552–556 Konoplev AV (2016) Comparative analysis of radioactive cesium wash-off from contaminated catchment areas after accidents at the Fukushima Dai-ichi and Chernobyl nuclear power plants. Geochem Int 54(6):522–528 Konoplev A (2020) Mobility and bioavailability of the Chernobyl-derived radionuclides in soilwater environment: review. In: Konoplev A, Kato K, Kalmykov S (eds) Behavior of radionuclides in the environment II: Chernobyl. Tokyo, Springer Nature, pp 157–193 Konoplev AV, Bobovnikova TI (1991). Comparative analysis of chemical forms of long-lived radionuclides and their migration and transformation in the environment following the Kyshtym and Chernobyl accidents. Proceedings of seminar on comparative assessment of the environmental impact of radionuclides released during three major nuclear accidents: Kyshtym, Windscale, Chernobyl. Luxembourg, 1-5 October 1990, 1, 371-396 Konoplev AV, Bulgakov AA (1996) Modelling the transformation processes of Chernobyl origin Cs-137 and Sr-90 speciation in soil and bottom sediments. In: environmental impact of radioactive releases. Proceedings of the international symposium Vienna, may 1995. IAEA, 311–324 Konoplev AV, Bulgakov AA (2000) 90Sr and 137Cs exchangeable distribution coefficient in soilwater systems. Atomic Energy 88(2):158–163 Konoplev AV, Golubenkov AV (1991) Modeling of the vertical radionuclide migration in soil (as a result of a nuclear accident). Meteorol Gidrol. 10:62–68. (In Russian) Konoplev AV, Konopleva IV (1999) Determination of radiocesium reversible selective sorption characteristics by soils and sediments. Geochem Int 37(2):207–214 Konoplev AV, Borzilov VA, Bobovnikova TI, Virchenko EP, Popov VE, Kutnyakov IV, Chumichev VB (1988) Distribution of radionuclides in the soil-water system fallen out in consequence of the Chernobyl failure. Russ Meteorol Hydrol 12:63–74 Konoplev AV, Kopylova LP, Bobovnikova TI, Bulgakov AA, Siverina AA (1992a) 90Sr and 137Cs distribution in the bottom sediments-water system of water bodies of the Chernobyl NPP exclusion zone. Soviet Meteorol. Hydrol. 1:35–42 Konoplev AV, Bulgakov AA, Popov VE, Bobovnikova TI (1992b) Behaviour of long-lived radionuclides in a soil-water system. Analyst 117:1041–1047 Konoplev AV, Bulgakov AA, Zhirnov VG, Bobovnikova TI, Kutnyakov IV, Siverina FA, Popov VE, Virchenko EP (1998) Study of the 137Cs and 90Sr behavior in lakes Svyatoe and Kozhanovskoe, Bryansk region Russian. Meteorol Hydrol 11:45–53

66

A. Konoplev et al.

Konoplev AV, Kaminski S, Klemt E, Konopleva I, Miller R, Zibold G (2002a) Comparative study of 137Cs partitioning between solid and liquid phases in lakes Constance, Lugano and Vorsee. J Environ Radioact 58:1–11 Konoplev AV, Deville-Cavelin G, Voitsekhovich O, Zhukova OM (2002b) Transfer of Chernobyl 137 Cs and 90Sr by surface run-off. Radioprotection 37(C1):315–318 Konoplev A, Golosov V, Laptev G, Nanba K, Onda Y, Takase T, Wakiyama Y, Yoshimura K (2016a) Behavior of accidentally released radiocesium in soil-water environment: looking at Fukushima from a Chernobyl perspective. J Environ Radioact 151:568–578 Konoplev AV, Golosov VN, Yoschenko VI, Nanba K, Onda Y, Takase T, Wakiyama Y (2016b) Vertical distribution of radiocesium in soils of the area affected by the Fukushima Dai-ichi nuclear power plant accident. Eurasian Soil Sci 49(5):570–580 Konoplev A, Golosov V, Wakiyama Y, Takase T, Yoschenko V, Yoshihara T, Parenyuk O, Cresswell A, Ivanov M, Carradine M, Nanba K, Onda Y (2018a) Natural attenuation of Fukushima-derived radiocesium in soils due to its vertical and lateral migration. J Environ Radioact 186:23–33 Konoplev A, Wakiyama Y, Wada T, Golosov V, Nanba K, Takase T (2018b) Radiocesium in ponds in the near zone of Fukushima Dai-ichi NPP. Water Resour 45(4):589–597 Konoplev A, Kanivets V, Laptev G, Voitsekhovich O, Zhukova O, Germenchuk M (2020) Longterm dynamics of the Chernobyl-derived radionuclides in rivers and lakes. In: Konoplev A, Kato K, Kalmykov S (eds) Behavior of radionuclides in the environment II: Chernobyl. Springer, Tokyo, pp 323–348 Konoplev A, Wakiyama Y, Wada T, Udy C, Kanivets V, Ivanov V, Komissarov M, Goto A, Nanba K (2021a) Radiocesium distribution and mid-term dynamics in the ponds of the Fukushima Dai-ichi nuclear power plant exclusion zone. Chemosphere 265:129058 Konoplev A, Kanivets V, Zhukova O, Germenchuk M, Derkach G (2021b) Mid- to long- term radiocesium wash-off from contaminated catchments at Chernobyl and Fukushima. Water Res 188:116514. https://doi.org/10.1016/j.watres.2020.116514 Konshin O (1992) Mathematical model of 137Cs migration in soil: analysis of observations following the Chernobyl accident. Health Phys 63(3):301–306 Laceby JP, Chartin C, Evrard O, Onda Y, Garcia-Sanchez L, Cerdan O (2016) Rainfall erosivity in catchments contaminated with fallout from the Fukushima Daiichi nuclear power plant accident. Hydrol Earth Syst Sci 20:2467–2482 Makhon’ko KP, Avramenko AS, Bobovnikova TI, Chumichev VB (1977) Strontium-90 and Cesium-137 wash-off coefficient from soil of river basin. Meteorologia i Gidrologia 10:62– 66. (In Russian) Malins A, Kurikami H, Nakama S, Saito T, Okumura M, Machida M, Kitamura A (2016) Evaluation of ambient dose equivalent rates influenced by vertical and horizontal distribution of radioactive cesium in soil in Fukushima. J Environ Radioact 151:38–49 Mamikhin SV, Golosov VN, Paramonova TA, Shamshurina EN, Ivanov MM (2016) Vertical distribution of 137Cs in alluvial soils of the Lokna River floodplain (Tula oblast) long after the Chernobyl accident and its simulation. Eurasian Soil Sci 49(12):1432–1442 Matsuda N, Mikami S, Shimoura S, Takahashi J, Nakano M, Shimada K, Uno K, Hagiwara S (2015) Depth profiles of radioactive cesium in soil using a scraper plate over a wide area surrounding the Fukushima Dai-ichi nuclear power plant, Japan. J Environ Radioact 139:427– 434 Matsunaga T, Amano H, Yanase N (1991) Discharge of dissolved and particulate 137Cs in the Kuji River, Japan. Appl Geochem 6:159–167 Ministry of Education, Culture, Sports, Science and Technology of Japan [MEXT] (2012) Results of the (i) Fifth Airborne Monitoring Survey and (ii) Airborne Monitoring Survey Outside 80 km from the Fukushima Dai-ichi NPP. Tokyo. http://radioactivity.nsr.go.jp/en/contents/6000/5790/ 24/203_092814e.pdf Miura H, Kurihara Y, Sakaguchi A, Tanaka K, Yamaguchi N, Higaki S, Takahashi Y (2018) Discovery of radiocesium-bearing microparticles in river water and their influence on the solidwater distribution coefficient (Kd) of radiocesium in the Kuchibuto River in Fukushima. Geochem J 52:1–10

4 Behavior of Fukushima-Derived Radiocesium in the Soil–Water Environment:. . .

67

Miura H, Kurihara Y, Yamamoto M, Sakaguchi A, Yamaguchi N, Sekizawa O, Nitta K, Higaki S, Tsumune D, Itai T, Takahashi Y (2020) Characterization of two types of cesium-bearing microparticles emitted from the Fukushima accident via multiple synchrotron radiation analyses. Sci Reports 10:11421 Monte L, Brittain J, Hakanson L, Smith JT, Van der Perk M (2004) Review and assessment of models for predicting the migration of radionuclides from catchments. J Environ Radioact 75: 83–103 Nagao S, Kanamori M, Ochiai S, Tomihara S, Fukushi K, Yamamoto M (2013) Export of 134Cs and 137 Cs in the Fukushima river systems at heavy rains by typhoon Roke in September 2011. Biogeosciences 10:6215–6223 Nakanishi T, Sakuma K (2019) Trend of 137Cs concentration in river water in the medium term and future following the Fukushima nuclear accident. Chemosphere 215:272–279 Nakao A, Ogasawara S, Sano O, Ito T, Yanai J (2014) Radiocesium sorption in relation to clay mineralogy of paddy soils in Fukushima. Japan Sci Total Environ 468–469:523–529 Naulier M, Eyrolle-Boyer F, Boyer P, Metivier J-M, Onda Y (2017) Particulate organic matter in rivers of Fukushima: an unexpected carrier phase for radiocesium. Sci Total Environ 579:1560– 1571 Niimura N, Kikuchi K, Tuyen ND, Komatsuzaki M, Motohashi Y (2015) Physical properties, structure and shape of radioactive Cs from the Fukushima Daiichi nuclear power plant accident derived from soil, bamboo and shiitake mushroom measurements. J Environ Radioact 139:234– 239 Okumura T, Yamaguchi N, Dohi T, Iijima K, Kogure T (2019) Dissolution behavior of radiocesium-bearing microparticles released from the Fukushima nuclear plant. Sci Reports 9: 3520 Pisarev VV, Koloskov IA, Kuznetsova VM, Tsybizov IS (1972) Strontium-90 washout from soil by surface waters. Pochvovedenie 3:66–75. (In Russian) Prokhorov VM (1981) Migration of radioactive contaminants in soil. Energoizdat, Moscow, p 98. (In Russian) Rovinsky FY, Sinitsyna ZL (1979) Prediction of river water quality spring flood. Meteorologia i Gidrologia 6:74–77. (In Russian) Rovinsky FY, Sinitsyna ZL, Cherhanov YP (1976) About strontium-90 migration from soil with surface runoff. Pochvovedenie 8:52–55. (In Russian) Saito K, Tanihata I, Fujiwara M, Saito T, Shimoura S, Otsuka T, Onda Y, Hoshi M, Ikeuchi Y, Takahashi F, Kinouchi N, Saegusa J, Takemiya H, Shibata T (2015) Detailed deposition density maps constructed by large scale soil sampling for gamma-ray emitting radioactive nuclides from the Fukushima Dai-ichi nuclear power plant accident. J Environ Radioact 139:308–319 Satou Y, Sueki K, Sasa K, Yoshikawa H, Nakama S, Minowa H, Abe Y, Naklai I, Ono T, Adachi K, Igarashi Y (2018) Analysis of two forms of radioactive particles emitted during the early stages of the Fukushima Dai-ichi nuclear Power Station accident. Geochem J 52:137–143 Smith JT, Hilton J, Comans RNJ (1995) Application of two simple models to the transport of 137Cs in an upland organic catchment. Sci Total Environ 168:57–61 Smith JT, Belova NV, Bulgakov AA, Comans RNJ, Konoplev AV, Kudelsky AV, Madruga MJ, Voitsekhovich OV, Zibold G (2005) The “AQUASCOPE” simplified model for predicting 89,90 Sr, and 134,137Cs in surface waters after a large-scale radioactive fallout. Health Phys 89(6):628–644 Takahashi J, Tamura K, Suda T, Matsumura R, Onda Y (2015) Vertical distribution and temporal changes of 137Cs in soil profiles under various land uses after the Fukushima Dai-ichi nuclear power plant accident. J Environ Radioact 139:351–361 Takahashi Y, Fan Q, Suga H, Tanaka K, Sakaguchi A, Takeichi Y, Ono K, Mase K, Kato K, Kanivets V (2017) Comparison of solid-water partitions of radiocesium in river water in Fukushima and Chernobyl areas. Sci Rep 7:12407 Takahashi Y, Sakaguchi A, Fan Q, Tanaka K, Miura H, Kurihara Y (2020) Difference in the solidwater distribution of radiocesium in rivers in Fukushima and Chernobyl. In: Kato K, Konoplev A, Kalmykov S (eds) Behavior of radionuclides in the environment I: function of particles in aquatic system. Springer Nature, Singapore, pp 115–150

68

A. Konoplev et al.

Taniguchi K, Onda Y, Smith HG, Blake W, Yoshimura K, Yamashiki Y, Kuramoto T, Saito K (2019) Transport and redistribution of radiocesium in Fukushima fallout through rivers. Environ Sci Technol 53:12339–12347 Tsuji H, Yasutaka T, Kawabe Y, Onishi T, Komai T (2014) Distribution of dissolved and particulate radiocesium concentrations along rivers and the relations between radiocesium concentration and deposition after the nuclear power plant accident in Fukushima. Water Res 60:15–27 Tsukada H, Nihira S, Watanabe T, Takeda S (2017) The 137Cs activity concentration of suspended and dissolved fractions in irrigation waters collected from the 80 km zone around TEPCO’s Fukushima Daichi nuclear Power Station. J Environ Radioact 178-179:354–359 Ueda S, Hasegawa H, Kakiuchi H, Akata N, Ohtsuka Y, Hisamatsu S (2013) Fluvial discharges of radiocesium from watersheds contaminated by the Fukushima Dai-ichi nuclear Powerer plant accident. Japan J Environ Radioact 118:96–104 Uematsu S, Smolders E, Sweek L, Wannijn J, Van Hees M, Vandenhove H (2015) Predicting radiocesium sorption characteristics with soil chemical properties for Japanese soils. Sci Total Environ 524-525:148–156 Valcke E, Cremers A (1994) Sorption-desorption dynamics of radiocesium in organic matter soils. Sci Total Environ 157:275–283 Van Genuchten MT, Wierenga PJ (1986) Solute dispersion coefficients and retardation factors. Methods of soil analysis. Part 1 physical and mineralogical methods, Madison, Wisconsin USA: 1025-1054 Wada T, Konoplev A, Wakiyama Y, Watanabe K, Furuta Y, Morishita D, Kawata G, Nanba K (2019) Strong contrast of cesium radioactivity between marine and freshwater fish in Fukushima. J Environ Radioact 204:132–142 Wakiyama Y, Konoplev A, Wada T, Takase T, Byrnes I, Carradine M, Nanba K (2017) Behavior of 137 Cs in ponds in the vicinity of the Fukushima Dai-ichi nuclear power plant. J Environ Radioact 178-179:367–376 Wakiyama Y, Konoplev A, Wada T, Takase T, Igarashi Y, Nanba K, Byrnes I (2019a) Temporal trends of 137Cs activity concentration in pond waters in the vicinity of Fukushima Dai-ichi nuclear power plant. Proc IAHS 381:101–106 Wakiyama Y, Onda Y, Yoshimura K, Igarashi Y, Kato H (2019b) Land use types control solid wash-off rate and entrainment coefficient of Fukushima-derived 137Cs, and their time dependence. J Environ Radioact 210:105990 Wauters J, Madruga MJ, Vidal M, Cremers A (1996) Solid phase speciation of radiocesium in bottom sediments. Sci Total Env 187:121–130 Yamaguchi N, Tsukada H, Kohyama K, Takata Y, Takeda A, Isono S, Taniyama I (2017) Radiocesium interception potential of agricultural soils in Northeast Japan. Soil Sci Plant Nutr 63(2):119–126 Yamashiki Y, Onda Y, Smith HG, Blake WH, Wakahara T, Igarashi Y, Matsuura Y, Yoshimura K (2014) Initial flux of sediment-associated radiocesium to the ocean from the largest river impacted by Fukushima daiichi nuclear power plant. Sci. Rep. 4:3714 Yoshikawa N, Obara H, Ogasa M, Miyazu S, Harada N, Nonaka M (2014) 137Cs in irrigation water and its effect on paddy fields in Japan after the Fukushima nuclear accident. Sci Total Environ 481:252–259 Yoshimura K, Onda Y, Sakaguchi A, Yamamoto M, Matsuura Y (2015a) An extensive study of the concentrations of particulate/dissolved radiocaesium derived from the Fukushima Dai-ichi nuclear power plant accident in various river systems and their relationship with catchment inventory. J Environ Radioact 139:370–378 Yoshimura K, Onda Y, Kato H (2015b) Evaluation of radiocesium wash-off by soil erosion from various land uses using USLE plots. J Environ Radioact 139:362–369 Yoshimura K, Onda Y, Wakahara T (2016) Time dependence of the 137Cs concentration in particles discharged from rice paddies to freshwater bodies after the Fukushima Daiichi NPP accident. Environ Sci Technol 50:4186–4193

Chapter 5

Dynamics of Radiocesium Solid–Liquid Distribution Coefficient in Soil and Water Environments Sadao Eguchi and Noriko Yamaguchi

Abstract This study aimed to elucidate the dynamic features of an apparent solid– liquid distribution coefficient (Kd(a)) of radiocesium that regulates radiocesium’s dissolved concentration and affects its transfer from soil to crops. According to our analysis of data from the literature and our use of a theoretical formulation that considers the competitive adsorption of Cs+ with K+ and NH4+ on a frayed edge site (FES), the solid-phase radiocesium interception potential (RIP) and liquid-phase water chemistry of dissolved K+ and NH4+ concentrations or electrical conductivity (EC) are key factors in controlling Kd(a) for radiocesium and stable cesium in different soil and water environments. The basic concept of the formulation was applicable independent of the radiocesium source (atmospheric nuclear tests and different nuclear power station accidents) and the time elapsed from its deposition although Kd(a) values for stable cesium were up to several times higher than those for radiocesium. The regional-scale Japanese government database indicated that the Kd (a) derived from the 1 M CH3COONH4-extractable radiocesium in soil tends to increase with soil RIP. The management of crops, fertilization, irrigation water volume and quality, etc., as well as the soil RIP in the target field, should be considered when examining and deciding the most effective strategy to reduce radiocesium transfer from soil to crops. Keywords Apparent solid–liquid distribution coefficient (Kd(a)) · Radiocesium interception potential (RIP) · Frayed edge site (FES) · Dissolved monovalent cation concentrations · Electrical conductivity (EC) · “Exchangeable” radiocesium or stable cesium concentration in soil (Ex-Cs)

S. Eguchi (*) · N. Yamaguchi Institute for Agro-environmental Sciences, National Agriculture and Food Research Organization, Ibaraki, Japan e-mail: [email protected]; [email protected] © The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2022 K. Nanba et al. (eds.), Behavior of Radionuclides in the Environment III, https://doi.org/10.1007/978-981-16-6799-2_5

69

70

5.1

S. Eguchi and N. Yamaguchi

Ideal and Apparent Solid–Liquid Distribution Coefficients (Kd)

Radiocesium in soil can be transported by water and absorbed by crops in its dissolved form. The concentration of dissolved radiocesium in irrigation water or soil–water in rice paddy fields is positively correlated with the radiocesium concentration in brown rice (Fujimura et al. 2013; Suzuki et al. 2015; Yoshikawa et al. 2019) regardless of radiocesium concentration in soil. Dissolved concentration of radiocesium in soil–water contact with solid phase depends on various physicochemical characteristics of soils and water. Thus, determining the dynamic dissolved radiocesium concentration in soil and water is key to understanding the mechanism of radiocesium transfer from soil to crops. The solid–liquid distribution coefficient (Kd, L/kg) in soil is a commonly used parameter when elucidating and/or predicting the fate and transport of radiocesium in soil–plant systems because it determines the activity concentration of dissolved radiocesium in soil–water. Kd is defined as the ratio of the mass activity density (Q, Bq/kg) of the specified solid phase (usually on a dry weight basis) to the volumetric activity density (C, Bq/L) of the specified liquid phase (International Atomic Energy Agency [IAEA] 2010). Kd ¼

Q C

ð5:1Þ

Ideally, the concept of Kd can be defined and applied under totally reversible equilibrium sorption conditions, independent of the dissolved concentrations of radiocesium and other ion species. However, such ideal conditions generally do not occur under either laboratory or field conditions because of the physical and chemical non-equilibrium associated with variable reaction times and dynamic feature of soil/water environment. The Kd that is observed under non-equilibrium or irreversible conditions has been termed an “apparent Kd (Kd(a))” in freshwater and marine systems (IAEA 2020) or “effective Kd” in soils (Eguchi 2017). Radiocesium in the soil can be distributed among various adsorption sites that include frayed edge sites (FES) (Cremers et al. 1988) and regular exchange sites (RES) (Sanchez et al. 2002). Moreover, it can occur in various non-adsorption sites, including collapsed interlayer sites (CIS) (Yin et al. 2017), iron oxide-hydroxide, or other mineral precipitates in which radiocesium can be occluded through processes such as co-precipitation, and living or dead biological cells in which radiocesium has been absorbed mainly through biological activities. Furthermore, the skeletal structure of a secondary mineral or parent rock in which stable cesium (133Cs) is originally incorporated can be partially replaced by radiocesium. Thus, the Kd(a) value may be regarded as the sum of different mechanism-based “partial” Kd values (Eguchi 2017):

5 Dynamics of Radiocesium Solid–Liquid Distribution Coefficient in Soil. . .

71

QFES þ QRES þ QCIS þ QPR þ QBI þ QST C

ð5:2Þ

K dðaÞ ¼

¼ K dFES þ K dRES þ K dCIS þ K dPR þ K dBI þ K dST

where subscripts FES, RES, CIS, PR, BI, and ST are used to indicate the partial Q or Kd values for FES, RES, CIS, precipitates, biological cells, and the structure of minerals or rocks, respectively. Among these partial Kd values, KdFES typically attracts the most attention and is investigated more in depth because of its high selectivity for cesium. Theoretically, KdFES may be formulated as (Eguchi 2017) K dFES ¼ RIP=

X K FES cðKiÞ C i

ð5:3Þ

i

where RIP is the radiocesium interception potential (mol/kg), K FES cðKiÞ is the selectivity coefficient of a monovalent cation i over potassium ion (K+) within the FES, and Ci is the dissolved concentration (mol/L) of i in the liquid phase. RIP is a quantitative index of the high affinity of Cs+ to FES and is experimentally determined by measuring liquid-phase K+ and radioactive Cs+ concentrations under competitive adsorption equilibrium against FES after masking negative charges other than FES by AgTU or Ca2+ (Cremers et al. 1988; Wauters et al. 1996). The cation species i include monovalent cations with low hydration energy, such as K+, + ammonium ion (NH4+), 133Cs+, etc. The K FES cðKiÞ value for K is 1, and those for NH4+ and Cs+ have been experimentally determined to be approximately 5 and 103, respectively (Cremers et al. 1988; Wauters et al. 1996).

5.2 5.2.1

Different Methods for Determining Apparent Kd (Kd(a)) Soil–Water Sampling Method

The Kd(a) in soil is rarely measured in the field because the radiocesium activity concentration in soil–water is generally low and obtaining this measurement requires a large soil–water sample for radiological analysis. Despite this challenge, Tegen and Dörr (1996) used stainless steel lysimeter plates to collect >1 L of soil–water in organic-rich soil contaminated with 137Cs emitted from the Chernobyl Nuclear Power Plant (CNPP). They reported seasonal changes in the dissolved 137Cs activity concentration exceeding one order of magnitude and calculated the average Kd(a) as 7  103  5  103 (mean  standard deviation) L/kg. In such a case, when water in contact with soil is directly sampled (i.e., soil–water sampling method), Kd(a) can be directly determined by applying Eq. (5.1) using the experimentally determined in situ C and Q values:

72

S. Eguchi and N. Yamaguchi

Q¼MCwg

ð5:4Þ

where M is the total activity concentration in the soil (Bq/kg), w is the gravimetric water content in soil (kg/kg), and g is the specific gravity of water (1 kg/L). The value of Q in Eq. (5.1) is sometimes assumed to be approximately equal to that of M because generally M > > Cwg. Instead of the field experiments conducted under actual environmental conditions, pot soil experiments are a useful way to determine Kd(a) by applying the soil– water sampling method. Yoshikawa et al. (2020) conducted a pot soil experiment for paddy rice under different fertilization treatments using three different soils contaminated with 137Cs emitted from the Fukushima Daiichi Nuclear Power Plant (FDNPP) of Tokyo Electric Power Company Holdings, Inc. (TEPCO). They collected 10 mL/d of soil–water for 106 days and used approximately 1 L of the composited water sample per pot to determine the time-averaged 137Cs activity concentrations and natural 133Cs concentrations in soil–water. As a result, they obtained Kd(a) values ranging from 3.3  102 to 4.7  104 L/kg for 137Cs and from 2.3  103 to 3.4  104 L/kg for 133Cs.

5.2.2

Laboratory Extraction Method

The “exchangeable” radiocesium activity concentration in soil (Ex-Cs, Bq/kg)— determined by extracting soil with a soil/liquid ratio of 1:v (v ¼ 10 or 20 in general) using 1 M CH3COONH4, deionized water, or other arbitrary fit-for-purpose solutions (i.e., laboratory extraction method)—may be used to identify the Kd(a) value under specified laboratory conditions: K extr dðaÞ ¼

M  Ex‐Cs Ex‐Cs=v

ð5:5Þ

where K extr dðaÞ is the solution extraction-based Kd(a). The parameter Ex-Cs using 1 M CH3COONH4 should consist mainly of the dissolved (Cwg) and adsorbed Cs (QRES and a part of QFES). With the exception of repeated extraction experiments, it typically shows the highest Ex-Cs value in field soils compared with those using the other solutions with neutral pH. Thus, K extr dðaÞ using 1 M CH3COONH4 can be regarded as the approximate minimum value of Kd(a) that can be observed in field soils under typical environmental conditions.

5 Dynamics of Radiocesium Solid–Liquid Distribution Coefficient in Soil. . .

5.2.3

73

Variation in Kd(a) between Different Methods

The laboratory extraction method can be easily applied to any soil sample. However, significant differences in Kd(a) are generally observed between the values determined using the soil–water sampling method and laboratory extraction method. Figure 5.1 shows the Kd(a) values plotted against RIP/(CK + 5CNH4) using the data obtained from the soil–water sampling method in the paddy rice pot experiment (Yoshikawa et al. 2020) and those calculated from the Ex-Cs data for the same soils managed under different crop and fertilization conditions (Yoshikawa et al. 2019). The K extr dðaÞ values obtained via the laboratory extraction method using 1 M CH3COONH4 are one to two orders of magnitude lower than the Kd(a) values obtained through the soil– water sampling method. The positive correlation between Kd(a) and RIP/(CK + 5CNH4) (Fig. 5.1) indicates that KdFES defined using Eq. (5.3) appears to play a key role in determining the variation of the soil Kd(a) value. This analysis suggests that soil Kd(a) varies considerably depending on which method is used to extract radiocesium, but there is a possibility that it can be estimated using the RIP/(CK + 5CNH4) value under a wide range of water chemistry conditions if the extraction solution pH is approximately neutral. Moreover, most Kd(a) values in Fig. 5.1 are more than one order of magnitude larger than those expected from the KdFES values (1:1 line), indicating that the quantitative importance of the other partial Kd values in Eq. (5.2), KdCIS in particular, should be much higher than KdFES. Figure 5.1 also shows that the significance of KdCIS (distance from the 1:1 line) tends to increase with the dissolved K+ and NH4+ concentrations. The CIS appears to be a reversible or semi-reversible site, at least in the differing water chemistry conditions indicated in Fig. 5.1. The regression line for natural 133Cs is nearly parallel to that for TEPCO’s FDNPP-derived 137Cs, and the difference between these two lines is within one 100000 y = 611.4x0.4673 R² = 0.8961

Kd(a) (L/kg)

10000

1:1 line Kd_137Cs L/kg

Kd_133Cs L/kg 1000

137 water Kd_137Cs Cs Soil L/kg

sampling L/kg method

133 Kd_133Cs Cs

100

y = 203.26x0.4301 R² = 0.6406

137 Kd-Ex_137Cs L/kg Cs Laboratory

extraction methodL/kg by 1 M CH3COONH4

133 Kd-Ex_133Cs Cs

10 0.01

0.1

1

10

100

1000

RIP/(CK + 5CNH4) (L/kg)

Fig. 5.1 Relationship between Kd(a) and RIP/(CK + 5CNH4)

10000

Water extraction-based Kd(a) (L/kg)

74

S. Eguchi and N. Yamaguchi

1000000

100000

1:1 line

10000

Water extraction-based Kd(a)

1000

Soil-Kd by water L/kg Quantified Soil-Kd (NDdetermined data) L/kg from Lower limit the detection limit

100 1

10

100

1000

10000

1 M CH3COONH4 extraction-based Kd(a) (L/kg) Fig. 5.2 Relationship between pure water extraction-based Kd(a) and 1 M CH3COONH4 extraction-based Kd(a) values. The black circles indicate the quantified Kd(a) values. The gray circles indicate the lower possible Kd(a) values calculated from the detection limit for the dissolved concentration of radiocesium when the dissolved concentration of radiocesium in the filtrate of the extract is too low to be determined. The gray vertical lines show the range of higher possible Kd (a) values

order of magnitude (Fig. 5.1). This indicates that the increase of Kd(a) for 137Cs due to the “aging” effect would not exceed more than 10 times that of the current Kd(a) determined 3 years after TEPCO’s FDNPP accident by Yoshikawa et al. (2019, 2020). The soil Kd(a) value determined via the laboratory extraction method using pure water can be regarded as almost the maximum value of Kd(a) in field soils. Figure 5.2 shows that the pure water extraction-based Kd(a) value tends to be more than one to two orders of magnitude higher than the 1 M CH3COONH4 extraction-based Kd(a) value (Yamaguchi et al. 2015; and unpublished; Yoshikawa et al. 2019). By measuring these maximum and minimum possible soil Kd(a) values, it would be possible to estimate the range of fluctuation in Kd(a) for the field soil under typical environmental conditions. These results indicate that the large variation in Kd(a) value between different determination methods can be primarily explained using the water chemistry and soil RIP based on the concept of Eq. (5.3). It should be noted that CIS is not irreversible and is generally much more important than FES as sorption sites for radiocesium and stable cesium; it also plays a significant role in regulating Kd(a) under different water chemistry and RIP conditions.

5 Dynamics of Radiocesium Solid–Liquid Distribution Coefficient in Soil. . .

75

Variability of Kd(a) in Soil and Water Environments

5.3 5.3.1

Relationship Between Kd(a) and Electrical Conductivity

Terrestrial freshwater systems, such as rivers, lakes, ponds, etc., provide irrigation water to rice paddy fields. These aquatic systems also include soil particles eroded from the soil profile in the catchment. Therefore, the spatiotemporal variability in Kd (a) values in the soil and suspended solids of freshwater ecosystems should be elucidated on the basis of an agreed-upon understanding of the Kd(a) variability. Figure 5.3 shows the clear negative correlation between Kd(a) and electrical conductivity (EC) values across various river and soil systems in Japan before and after the CNPP accident and after TEPCO’s FDNPP accident (Hirose et al. 1990; Matsunaga et al. 1991; Tsuji et al. 2014; Ochiai et al. 2015; Eguchi 2017; Yoshikawa et al. 2020). This indicates that the relatively higher Kd(a) values obtained for rivers after TEPCO’s FDNPP accident (Eguchi 2014, 2017) as compared to the freshwater ecosystem Kd(a) values after CNPP accident (International Atomic Energy Agency [IAEA] 2010) could, to an extent, be a result of relatively lower EC values (generally associated with lower K+ and NH4+ concentrations) in river water; it also shows that

107

C C C

106

C C

Kd(a) (L/kg)

(a) River water in 1985– 1986, before CNPP

C

105

(b) River water in 1986– 1988, after CNPP

C C

104

C C C

103 102 0.01

(c) River water in 2012– 2013, after FDNPP (d) Soil water in 2014, after FDNPP

0.1

1

10

Electrical conductivity (EC) (dS/m) Fig. 5.3 Relationship between river and soil Kd(a) and electrical conductivity (EC) in Japan before and after the Chernobyl Nuclear Power Plant (CNPP) accident and after TEPCO’s FDNPP accident. The data plots are grouped for (a) Tone, Ishikari, and Kitakami rivers in 1985–1986 before CNPP accident (Hirose et al. 1990); (b) Tone, Ishikari, Kuzuryu, and Kuji Rivers in 1986–1988 after CNPP accident (Hirose et al. 1990; Matsunaga et al. 1991); (c) Abukuma, Kuchibuto, Shakado, Ota, Niida, and other small rivers in 2012–2013 after FDNPP accident (Tsuji et al. 2014; Ochiai et al. 2015); and (d) soil–water samples taken from the three different soils in Fukushima (Yoshikawa et al. 2020). The EC values for (a) and (b) were estimated by Eguchi (2017) using the Water Information System of the Ministry of Land, Infrastructure, Transport, and Tourism (http://www1. river.go.jp/). Different shapes of the plots indicate the data obtained from different rivers/soils or different authors

76

S. Eguchi and N. Yamaguchi

this relationship is universally applicable regardless of the differences in radiocesium sources (atmospheric nuclear tests, CNPP accident, and TEPCO’s FDNPP accident) and in time elapsed from its deposition (Eguchi 2017). This suggests that the basic concept of Eq. (5.3) may be applicable to not only soil but also freshwater systems. Irrigation water EC can be measured in the field using EC meters, which are practically more efficient than measuring dissolved ion concentrations. The water EC value can be considered an important reference information for irrigation water managers when selecting the source of irrigation water with the lowest dissolved concentration of radiocesium in the agricultural catchment.

5.3.2

Relationship Between Kd(a) and RIP

Table 5.1 shows soil K extr dðaÞ values using 1 M CH3COONH4 for TEPCO’s FDNPPderived radiocesium grouped according to the RIP value. For this analysis, we have constructed a database on soil K extr dðaÞ and RIP in Japan using the national database for rice paddy fields in Fukushima from 2011 to 2017 gathered by the Ministry of Agriculture, Forestry and Fishery, as well as data from the available literature (Sakuma and Sato 2014; Kato et al. 2015; Kohyama et al. 2015; Kondo et al. 2015; Saito et al. 2015; Yamaguchi et al. 2015, 2017; and unpublished; Yoshikawa et al. 2019, 2020). The geometric mean and 95th percentile values of K extr dðaÞ tend to increase with the RIP value (Table 5.1) in accordance with the basic concept of Eq. (5.3). This indicates that the soil RIP is an important factor controlling K extr dðaÞ or Ex-Cs despite the large variation in agricultural practices such as fertilization and water management. Table 5.1 Soil K extr dðaÞ values (L/kg, using 1 M CH3COONH4) for TEPCO’s FDNPP-derived radiocesium grouped according to the RIP value RIP mmol/kg

Laboratory extraction-based K d(a) n

L/kg

GM RIP :F U L ,

 1 n u þ unR 0 2 L ,  1 n n u þ uR < 0 if 2 L if

! ! c c where U L, U R—predicted values of conservative variables on the left and right side of the boundary, and unL , unR —water flow velocity components normal to the boundary (positive means water flow from left to right).

236

M. Zheleznyak et al.

For triangular element, the ratio r for each component of vector flux is defined as below:

r¼

8 2 X > > 1 > >   ΔF i  dsi , > < ΔF  ds i¼1

2 > X > 1 > > ΔF j  ds j ,  > : ΔF  ds j¼1

if

 1 n uL þ unR 0 2

 1 n if u þ unR < 0 2 L

,

where i loops through the other two edges of the left cell and j loops through the other two edges of the right cell with positive fluxes directed outward of the cells. The positive flux ΔF is directed from left to right. Another, stronger version, is now used. ! 2 2 X X 1 1 r ¼ min    ΔF i  dsi , ΔF j  ds j : ΔF  ds i¼1 ΔF  ds j¼1 For the situation in the absence of waves, the next semi-implicit scheme for the calculation of bottom-stresses’ influence is used, where the upper subscript k stands for the values at previous time step. qkþ1  qkx qkþ1 x ¼ Ckb   x 2 Δt hk ¼

Ckb

qkþ1 y   2 hk

rffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi ffi  2   2 k k qx þ qy ,

qkþ1  qky y Δt

rffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi ffi  2   2 k k qx þ qy :

During the simulation, the time step is changed to satisfy the Courant condition. As the system of equations being solved is not of linear form, the Courant number must include all possible movements of liquid. The largest velocities are those of pffiffiffiffiffi gravity waves given by u ¼ gh. For Godunov-type scheme, the maximum stable Courant number is 0.4. COASTOX-UN code includes two finite-difference algorithms for numerical solution of sediment transport equations: explicit and implicit. The first one is used in conjunction with water flow equations’ numerical solution. The second one is used when only sediment transport is solved. Then, it allows using large time steps not restricted by Courant number. Both algorithms are simple upwind and have first order both in time and space. The numerical solution of the transport equation (Eq. 11.28) is split into two steps. At the first step, the integral of advection-dispersion part without erosion \deposition over a control volume V and time step Δt is taken.

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

237

t þΔt Z

 Z ∂ðhSÞ ∂ ∂ ∂S þ ðqi SÞ  hDij dVdt ¼ 0 ∂t ∂xi ∂xi ∂x j t

V

Again, using the divergence theorem one gets: t þΔt Z

t

0 @

Z V

∂ðhSÞ dV þ ∂t

I




!d



1

!

u  nx þ v  ny  hS  F  n dSAdt ¼ 0:

S

Using fully implicit temporal discretization, and approximating the integrals as summations over the control volume surface, we obtain the following conservative implicit scheme: V

t X ðhC Þtþτ tþτ P  ðnC ÞP Aγ F tþτ ¼ γ þ V ðγC þ f C ÞP , τ γ

ð11:31Þ

γ ¼ W, E, S, N where Aγ is the surface between volume elements VP and Vγ . The surface flux terms Fγ contain both diffusive and advective flux terms. The diffusive flux terms are central differenced, whereas the advective flux terms are upwind or donor cell differenced. F γ ¼ ½hDi Pγ

CP  Cγ  þ ½hui þ Pγ C γ þ ½hui Pγ C P , ΔPγ

γ ¼ W, E, S, N

where U+ ¼ max (0, U ) and U ¼ min (0, U ). The terms delimited with brackets indicate a suitable interface averaging (arithmetic, harmonic, geometric, or upwind). ΔPγ is the distance between the nodal points P and γ, and gPγ is the component of gravitational acceleration in the direction from P to γ. Due to approximation of advective flux terms by upwind or donor cell differences, the finite difference scheme (Eq. 11.31) is a monotonic scheme of first order, accurate in both space and time. The discretization of the water flow, transport, and radionuclides transport equations described above was obtained for nodes positioned within the interior of the computational domain. For nodes located adjacent to the domain boundary, the discretization of the governing equations differs to account for conditions at the boundary. Boundary conditions are enforced both at predictor and corrector steps of numerical solution of shallow water equations (SWE). The variables at boundaries are approximated either by using the prescribed field variables or with surface fluxes on the boundary surfaces, or the values from the adjacent cells, or by using the Riemann invariants which allow the internally generated waves to freely leave the model

238

M. Zheleznyak et al.

domain. Fourteen types of boundary conditions for SWE are implemented in COASTOX-UN code. Boundary conditions for sediment and radionuclide transport are specified either with field variables or with surface fluxes on the boundary surfaces. Boundary conditions of the former type are referred to as Dirichlet, whereas the latter type are referred to as Neumann. Also available is outflow-type boundary condition that acts the same as corresponding hydrodynamical boundary condition. Implementation of Dirichlet boundary conditions requires relatively few modifications to the discretized governing conservation equations. These modifications are replacing the following field variables at the boundary nodes by their values on the boundary surfaces. Neumann-type boundary conditions are accommodated by substituting the specified surface fluxes directly into the discretized form of the conservation equation. If for some boundary faces no hydro-dynamical boundary condition was specified, it is automatically assumed to be of Velocity type with zero value (i.e., the wall). For sediment part, the default is Neumann type with zero value, which also means the wall. The finite difference discretization of the surface water flow equations leads to a system of nonlinear equations: F ð xÞ ¼ 0

F : Rn ! Rn

ð11:32Þ

The nonlinear equations (Eq. 11.16) are solved iteratively, using a multivariable, residual-based Newton technique xsþ1 ¼ xs þ ps

s ¼ 0, 1, 2, . . . ,

ps ¼ J ðxs Þ1 F ðxs Þ

where s denotes an iteration number; and J(xs) is the Jacobian of F(x). Iteration continues until the criterion is satisfied.  sþ1   x  xs  i i   ε max xs  1in i

where ε is a user-provided error tolerance. The iteration routine starts, at a given time step, with an estimate of liquid pressures based on the previous time step. This estimate is used to update values of the governing equation residuals and to evaluate all of the partial derivatives that make up the Jacobian matrix. The resulting system of linear equations is then solved by using the conjugate gradient method with incomplete LU factorization. Solution of the system of equations yields changes to the liquid pressures. Iteration ends by updating the liquid pressures with the changes computed from the system of linear equations. If the criterion is satisfied, then the procedure is determined to have converged and a new time step begins. Otherwise, a new iteration starts. If the criterion is not satisfied within a specified number of iterations, the system is considered nonconvergent. Nonconvergent systems are handled by reducing the simulation step, resetting the liquid pressures to their previous time-step values, and reinitiating the time-step procedures.

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

11.4.1.3

239

Parallelization Algorithms for Ordinary and GPU Processors

Numerical schemes of the model are implemented with FORTRAN programming language. They are explicit with local computations. So for parallelization of model numerical scheme for shared memory multicore workstations and distributed memory clusters, the conventional domain decomposition approach with halo boundary structures and message-passing updating is used. To decompose an unstructured model grid, METIS graph partition library (METIS 2021) is used. So every processors computes numerical scheme in own subdomain. For calculation in boundary cells, every subdomain is extended with hallo structures—fictitious grid elements corresponding to the true boundary elements of other subdomains. Their values are updated with data calculated by other processors. The structure of halo elements is determined by the stencils of numerical schemes. 1. To calculate the flux across the face between cells, the values of the hydrodynamic variables specified in the centers of these cells are needed. 2. For correction of flux across any face, the flux values from the other two faces of the cells adjacent to that boundary (left or right, depending on the flux direction) are used. 3. To calculate the hydrodynamic variables in the cell, the values of face fluxes and the value of variable in cell from the previous stage are required. 4. The values of the variables in the grid node are obtained by averaging over all the cells surrounding the node. The information dependencies and halo structures are shown on Fig. 11.19. For data exchange, the message passing interface (MPI) technology is used (Walker and Dongarra 1996). MPI is parallel programming technology for Fig. 11.19 Computational grid at the subdomain boundary. The arrows indicate the information dependencies for the basic stencils of numerical schemes. By dark gray and light gray, the halo elements for the right subdomain are shown

240

M. Zheleznyak et al.

distributed memory clusters, but also may be used on shared memory systems such as workstations with multicore processors. Halo values updating is performed by calling of nonblocking functions MPI_ISEND(. . .), MPI_IRECV(. . .) with call competition function MPI_WAITALL(. . .). After the grid decomposition, each processor knows the neighboring processors, as well as the true grid elements, whose values will be sent, and the fictitious grid elements whose values will be updated by the received data. The structure of send and receive values is determined using MPI-derived types constructed using the functions MPI_TYPE_CREATE_INDEXED_BLOCK(. . .) and MPI_TYPE_COMMIT(. . .). A nonlocal operation performed at the beginning of each time step is calculation of a time step value from the CFL criteria. For its parallelization, each processor searches a minimum time step in its own subdomain, and then from the obtained values the minimum is selected using the global reduction operation MPI_ALLREDUCE(. . ., MPI_MIN,. . .). Data output with the prescribed time interval is performed by master processor. For this purpose, all other processors send to it the values from all elements of their subdomains. This procedure is implemented with the functions MPI_ISEND (. . .), MPI_IRECV (. . .), and MPI_WAITALL (. . .), and MPI-derived data types. All MPI function calls are bounded by the preprocessor directives. So user can optionally build parallel or serial model. To develop GPU-accelerated version of model OpenACC, directive-based programming interface is used (OpenACC 2021). Numerical schemes of the model are implemented in the form of loops for cells, nodes, and faces with independent iterations because of scheme explicitness and locality of computations. So, this iterations may be computed by GPU cores in parallel. OpenACC directives inserted in Fortran code specify for compiler the loops for which it is need to generate parallel code. Only for GPU computing, it is necessary to explicitly control the loading of CPU data to the GPU and the uploading of data from the GPU to the CPU for further processing.

11.4.1.4

COASTOX-UN Implementation to Simulate 137Cs Transport in the River Reservoirs of the Eastern Coastal Part of the Fukushima Prefecture

Behavior of 137Cs in the river reservoirs in the eastern coastal part of the Fukushima Prefecture (Figs. 11.1 and 11.3) was modeled within several modeling researches overviewed in Sect. 11.2.3. Such modeling studies should quantify the role of the reservoirs in the trapping of the contaminated sediments transported by the river flows from the contaminated mountain watersheds to the ocean through the densely populated coastal lowlands of the eastern part of the Fukushima Prefecture and through the areas that were under the Evacuation Order since 2011 in the vicinity of the FDNPP. In the Institute of Environmental Radioactivity of Fukushima

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

241

Fig. 11.20 Watershed and Takanokura Dam Reservoir, Mizunashi River: (a) border of the watershed of Takanokura Dam Reservoir (brown line) on the elevation map of the eastern part of the Niida River watershed; (b) distribution of the density of 137Cs fallout Bq/m2 on the watershed (MEXT 2011); (c) the satellite picture of the reservoir (Google Maps); and (d) depth distribution (m) in the reservoir over one of the simulated scenario of the currents

University, the reservoir modeling has started at the end of 2013 by the customization of the above described version of the COASTOX-UN model for the reservoirs at Mano Dam—Mano River, at Yokokawa Dam—Ota River, and at Takanokura Dam—Mizunashi River (Nanba et al. 2016). All these mountain reservoirs are rather deep—a maximum depth of up to 30–50 m. These are used as water storages for the agricultural, municipal, and industrial needs of the downstream populated lowlands regions. The watershed of Takanokura Dam Reservoir on Mizunashi River—the right tributary to Niida River—, includes high contaminated mountain part of the Niida watershed (Fig. 11.20); therefore, it was important to simulate behavior of 137Cs in this reservoir during the extreme flood events. The water inflow to the reservoir from Mizunashi River and two inflowing creeks is monitored by the operational staff of Takanokura Dam, but the data on the 137Cs concentrations in the inflowing water were absent. Therefore, we used the DHSVM-R code, for which hydrological model was verified versus the observed inflow (Fig. 11.21), to simulate the fluxes of suspended sediments and radionuclides to the reservoirs from the Mizunashi River and two inflowing creeks as the boundary conditions for the COASTOX-UN model. The chain of the sediment transport modules of DHSVM-R was validated versus the data on the long-term sedimentation rate in three cross-sections of the reservoir (Fig. 11.22). The modeling of the dynamics of the bottom contamination during the typhoon-generated flood of September, 2021 (Fig. 11.23), shows that sedimentation of the contaminated particles in the shallow upper part of the reservoir creates the “tail” in the contaminated bottom propagated to the deeper part of the reservoir.

Fig. 11.21 Measured and simulated by DHSVM-R discharge of Mizunashi River inflow to the Takanokura Dam Reservoir for the floods 20–26 September 2021 and 29 May–10 June 2011

242 M. Zheleznyak et al.

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

165

170

No. 13 1975

160

No.8 1975 simulated 1991

160

simulated 1991

155

243

150

160

No.1 1975 simulated 1991

150 140

150

140

145

130

140

120

130

0

40

80

120

120 110

0

50

100

150

200

250

0

100

200

300

Fig. 11.22 Bottom elevation of the Takanokura Dam Reservoir measured in 1975 and 1991 in the cross-sections shown in Fig. 11.20d and simulated changes in the bottom elevation due to sedimentation

Fig. 11.23 Density of 137Cs in the bottom deposition (Bq/m2) in Takanokura Dam due to the sedimentation of the inflowing suspended sediments simulated for fourth and sixth days of the high flood of September 2011 (Nanba et al. 2016)

The simulation for the longer period has demonstrated that during the initial phase of the next high floods, the deposited contaminated sediments in the shallow part of the reservoir can be resuspended and also transported to the deeper part with the contaminated sediments of a “new flood” for the next deposition. By such mechanism, the “tail” of the contaminated sediments could move continuously in the direction of the dam, however, depending on the bathymetry of the reservoir such movement practically stopped at the depth deeper than 20 m, where resuspension of the bottom sediment practically stopped. Such process that was modeled for each considered reservoirs demonstrates that deep reservoirs of a maximum depth of more than 25–20 m have deposited the major part of the particulate 137Cs that was transported by these rivers from the contaminated mountain watersheds.

11.5

Conclusions

The overview of the modeling of the Fukushima-derived radiocesium in watershedriver-reservoir system presented in this Chapter provides a possibility to make the following conclusions. The modeling is an important tool for the understanding of

244

M. Zheleznyak et al.

the processes driving radiocesium transport in watershed-river systems in the specific conditions of mountain landscape and typhoon-generated floods of the Japanese river basins impacted by the accidental atmospheric fallout from FDNPP. The developed models were used to assess the fluxes of 137Cs from the watersheds to the river channels and from the rivers to the ocean coastal waters to predict the increasing 137Cs concentration in river water during the flood, to quantify the balance of the radionuclides in the numerous river reservoirs in the Fukushima fallout region, and to simulate the influence of the decontamination activities in the watersheds on the diminishing 137Cs concentration in river water. After the accident, the field measurements quickly revealed the main difference between 137Cs transport in the radioactive contaminated European, Siberian, and United Stated lowland rivers and Japanese upland rivers—the much higher intensity of the hillslope erosion in Japan and following much higher concentration of suspended sediments in the river flow. As a result, the major input in 137Cs river flux during the flood provides particulate 137Cs. The models developed until 2011 describing the radionuclide transport inside the water bodies included the submodels of the particulate transport in different spatial resolution, for example, TODAM, FETRA, FLESCOT, RIVTOX, COASTOX, THREETOX, and others. However, the models of the watershed radionuclide transport till 2011 simulates mainly the washing-off dissolved radionuclides from the watersheds. Since 2011, the main research activities were directed on the customization of the above mentioned new models to the Fukushima river-reservoir systems, with specialized efforts for the development of the watershed models describing with a good accuracy the slope erosion and particulate radionuclide wash-off from the watersheds. The physicalbased distributed models of the watershed hydrology and erosion models developed from the end of 1990s and based on USLE, KINEROS, and other erosion formulas were used as a basis for new distributed models of particulate radionuclide transport from the watersheds. The introduction of a more comprehensive description of the “solid-water” exchange processes—consideration not only for fixed and diluted 137 Cs but also some intermediate forms of it with the relevant two-step kinetics for the parametrization of such processes—started to be a tendency for more adequate simulation of the 137Cs transport from watersheds. The wider use of the comprehensively distributed physically based models requires execution of parallel numerical methods on high-performance computers and also stimulates implementation of less-expensive GPU technologies that are illustrated in this Chapter by the description of the methodology of the parallel numerical solution of the equations of 2-D finite volume model COASTOX-UN for GPU. New distributed watershed model of radionuclide transport DHSVM-R described in the Chapter was used in the hillslope scale to simulate 137Cs wash-off from the experimental USLE plots. We can conclude after the comparison with the measured data that this tool can be used for the analyses of the role of the different watershed factors (steepness of the slope, soil types, and other) on the total flux of 137Cs from the watershed and the flux distribution between it soluble and particulate 137Cs. The implementation of DHSVM-R in river watershed scale to simulate radioactive contamination of Niida river system was directed to the quantification of the

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

245

influence of the decontamination activities at the western part of the watershed on the contamination of the Niida river water at the eastern lowland part of the basin. The results of the numerical modeling shows that due to the limitation of the decontamination works only by the territories that are not covered by the forest such decontamination covering only small part of the territory could not decrease significantly the concentration of 137Cs in lowland Niida River. Such a range of the changes in the concentrations cannot be identified by the monitoring measurements as the consequences of the decontamination taking into account natural hydrological variability of the river flow. The modeling study by the chain of DHSVM-R and COSTOX-UN models’ 137Cs transport in reservoirs confirms the important role of deep mountain reservoirs at the coastal Fukushima Rivers as “traps” for the contaminated sediments. These reservoirs settle down the major part of the inflowing contaminated sediments into the bottom depositions of the reservoirs preventing the significant contamination of the lowlands parts of the rivers. The introducing of the more comprehensive description of the “solid-water” exchange processes within last years: the consideration not only fixed on sediments and diluted forms of 137Cs but also modeling of some intermediate forms of it is a tendency for more accurate simulation of the 137Cs transport from watersheds. The task of further improvement of the models by introducing more detailed description of the radionuclide exchanges in “water-sediment” systems of mountain watersheds will require closer interactions of the modelers with the field and laboratory researchers of radioactivity of river systems. Acknowledgment This research was supported by the Japan Society for the Promotion of Science, Grant-in-Aid for Scientific Research (B) (KAKENHI 18H03389), and partly supported for the model development by the Japanese—Ukrainian project “Strengthening of the Environmental Radiation Control and Legislative Basis for the Environmental Remediation of Radioactively Contaminated Sites” (2017–2022)—, conducting within the frame of SATREPS (Science and Technology Research Partnership for Sustainable Development) funded by JICA and JST, and by the project No. 2020.01/0421 of the National Research Foundation of Ukraine.

References ALOS (2016) High-resolution land use and land cover map products. In: High-resolution land use and land cover map of Japan. https://www.eorc.jaxa.jp/ALOS/a/en/dataset/lulc/lulc_jpn_e.htm. Accessed 17 Jun 2021 Arnold JG, Moriasi DN, Gassman PW, Abbaspour KC, White MJ, Srinivasan R, Santhi C, Harmel RD, Van Griensven A, Van Liew MW, Kannan N (2012) SWAT: model use, calibration, and validation. Trans ASABE 55(4):1491–1508. https://doi.org/10.13031/2013.42256 Bagnold RA (1966) An approach of sediment transport model from general physics. In: US Geological Survey Prof. Paper 422-J Begnudelli L, Sanders BF (2006) Unstructured grid finite-volume algorithm for shallow-water flow and scalar transport with wetting and drying. J Hydraul Eng 132(4):371–384. https://doi.org/10. 1155/2014/209562

246

M. Zheleznyak et al.

Benavidez R, Jackson B, Maxwell D, Norton K (2018) A review of the (Revised) Universal Soil Loss Equation ((R) USLE): with a view to increasing its global applicability and improving soil loss estimates. Hydrol Earth Syst Sci 22(11):6059–6086. https://doi.org/10.5194/hess-226059-2018 Bencala KE (1983) Simulation of solute transport in a mountain pool-and-riffle stream with a kinetic mass transfer model for sorption. Water Resour Res 19(3):732–738. https://doi.org/10. 1029/WR019i003p00718 Bezhenar R, Heling R, Ievdin I, Iosjpe M, Maderich V, Willemsen S, de With G, Dvorzhak A (2016) Integration of marine food chain model POSEIDON in JRODOS and testing versus Fukushima data. Radioprotection 51(2):137–139. https://doi.org/10.1051/RADIOPRO/ 2016050 Bijker EW (1971) Longshore transport computations. J Waterw Harb Coast Eng Div 97 (4):687–703 CRIED (2021) Center for Research in Isotopes and Environmental Dynamics at University of Tsukuba, Database Web site: http://www.ied.tsukuba.ac.jp/en/database/, https://www.ied. tsukuba.ac.jp/~fukushimafallout/a.html Delmas M, Garcia-Sanchez L, Nicoulaud-Gouin V, Onda (2017) Improving transfer functions to describe radiocesium wash-off fluxes for the Niida River by a Bayesian approach. J Environ Radioact 167:100–109. https://doi.org/10.1016/j.jenvrad.2016.11.002 Delmas M, Garcia-Sanchez L, Onda Y (2019) Factors controlling the variability of 137Cs concentrations in 5 coastal rivers around Fukushima Daiichi power plant. J Environ Radioact 204:1–11. https://doi.org/10.1016/j.jenvrad.2019.03.013 DHSVM (2018) Distributed hydrology soil vegetation model. Source code. https://github.com/ pnnl/DHSVM-PNNL. Accessed 17 Jun 2021 Dibajnia M, Watanabe A (1993) Sheet flow under nonlinear waves and currents. In: Proc. 23rd Int. Conf. on Coastal Engineering. ASCE, New York, NY, pp 2015–2028 Engelund F, Hansen E (1967) A monograph on sediment transport in alluvial streams. Teknisk Forlag, Copenhagen Evrard O, Laceby JP, Nakao A (2019) Effectiveness of landscape decontamination following the Fukushima nuclear accident: a review. Soil 5(2):333–350. https://doi.org/10.5194/soil-5333-2019 Evrard O, Chartin C, Laceby JP, Onda Y, Wakiyama Y, Nakao A, Cerdan O, Lepage H, Jaegler H, Vandromme R, Lefèvre I, Bonté P (2021) Radionuclide contamination in flood sediment deposits in the coastal rivers draining the main radioactive pollution plume of Fukushima Prefecture, Japan (2011–2020). Earth Syst Sci Data 13(6):2555–2560. https://doi.org/10.5194/ essd-13-2555-2021 Falconer RA (1980) Modelling of planform influence on circulation in harbour. In: Proceedings of the Seventeenth Coastal Engineering Conference. American Society of Civil Engineers, ASCE Press, New York, pp 2726–2744 Funaki H, Sakuma K, Nakanishi T, Yoshimura K, Katengeza EW (2020) Reservoir sediments as a long-term source of dissolved radiocaesium in water system; a mass balance case study of an artificial reservoir in Fukushima, Japan. Sci Total Environ 743:140668. https://doi.org/10.1016/ j.scitotenv.2020.140668 Gallego E, Papush L, Ievdin I, García-Ramos A, Pato-Martínez R, Monte L (2016) Integration of long-term radionuclide transport models MOIRA-LAKE and MOIRA-RIVER into Hydrological Dispersion Module of JRODOS. Radioprotection 51:141–143. https://doi.org/10.1051/ RADIOPRO/2016051 Garcia-Sanchez L, Konoplev AV (2009) Watershed wash-off of atmospherically deposited radionuclides: a review of normalized entrainment coefficients. J Environ Radioact 100(9):774–778. https://doi.org/10.1016/j.jenvrad.2008.08.005 Goodrich DC, Burns IS, Unkrich CL, Semmens DJ, Guertin DP, Hernandez M, Yatheendradas S, Kennedy JR, Levick LR (2012) KINEROS2/AGWA: model use, calibration, and validation. Trans ASABE 55(4):1561–1574. https://doi.org/10.13031/2013.42264

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

247

Heling R, Zheleznyak M, Raskob W, Popov A, Borodin R, Gofman D, Lyashenko G, Marinets A, Pokhil A, Shepeleva T, Tkalich P (1997) Overview of the modelling of hydrological pathways in RODOS. Radiat Prot Dosim 73(1-4):67–70. https://doi.org/10.1093/oxfordjournals.rpd. a032165 IAEA (1985) Hydrological dispersion of radioactive material in relation to nuclear power plant siting. In: IAEA safety series N50-SG-S6. International Atomic Energy Agency, Vienna. 116 p IAEA (2006) Radiological conditions in the Dnieper River basin: assessment by an international expert team and recommendations for an action plan. In: Radiological assessment reports series STI/DOC/1230. International Atomic Energy Agency, Vienna. 185 p, International experts team: Balonov M, Levins D, Konoplev A, Ranguelova V, Voitsekhovich O, Zheleznyak M, Zhukova O Ikenoue T, Shimadera H, Kondo A (2020) Impact of soil erosion potential uncertainties on numerical simulations of the environmental fate of radiocesium in the Abukuma River basin. J Environ Radioact 25:106452. https://doi.org/10.1016/j.jenvrad.2020.106452 iRIC (2021) Software Nays2D solver manual. In: Shimizu et al (eds) iRIC project. https://i-ric.org/ en/solvers/nays2dh/. Accessed 17 Jun 2021 ISRIC (2021) World soil information. http://soilgrids1km.isric.org. Accessed 17 Jun 2021 Iwasaki T, Nabi M, Shimizu Y, Kimura I (2015) Computational modeling of 137Cs contaminant transfer associated with sediment transport in Abukuma River. J Environ Radioact 139:416–426. https://doi.org/10.1016/j.jenvrad.2014.05.012 Jia Y, Ni G, Kawahara Y, Suetsugi T (2001) Development of WEP model and its application to an urban watershed. Hydrol Process 15(11):2175–2194. https://doi.org/10.1002/HYP.275 JMA (2021) Past meteorological data search. Japan Meteorological Agency, Tokyo. www.data.jma. go.jp/obd/stats/etrn/upper/index.php. Accessed 17 Jun 2021 Kato H, Onda Y (2018) Determining the initial Fukushima reactor accident-derived cesium-137 fallout in forested areas of municipalities in Fukushima prefecture. J Forest Res 23(2):73–84. https://doi.org/10.1080/13416979.2018.1448566 Kinouchi T, Yoshimura K, Omata T (2015) Modeling radiocesium transport from a river catchment based on a physically-based distributed hydrological and sediment erosion model. J Environ Radioact 139:407–415. https://doi.org/10.1016/j.jenvrad.2014.07.022 Kitamura A, Yamaguchi M, Kurikami H, Yui M, Onishi Y (2014) Predicting sediment and cesium137 discharge from catchments in eastern Fukushima. Anthropocene 5:22–31. https://doi.org/ 10.1016/j.ancene.2014.07.001 Kitamura A, Kurikami H, Yamaguchi M, Oda Y, Saito T, Kato T, Niizato T, Iijima K, Sato H, Yui M, Machida M (2015) Mathematical modeling of radioactive contaminants in the Fukushima environment. Nucl Sci Eng 179(1):104–118. https://doi.org/10.13182/NSE13-89 Kitamura A, Kurikami H, Sakuma K, Malins A, Okumura M, Machida M, Mori K, Tada K, Tawara Y, Kobayashi T, Yoshida T, Tosaka H (2016) Redistribution and export of contaminated sediment within eastern Fukushima Prefecture due to typhoon flooding. Earth Surf Process Landf 41:1708–1726. https://doi.org/10.1002/esp.3944 Kivva S (2003) Data assimilation for 2-D advection-dispersion equations. In: Sloot PMA et al (eds) Computational science ICCS 2003. Lecture notes in computer science, vol 2658. Springer, New York, NY. https://doi.org/10.1007/3-540-44862-4_66 Kivva S, Zheleznyak M (2005) Two-dimensional modeling of rainfall runoff and sediment transport in small catchment areas. Int J Fluid Mech Res 32(6):703–717. https://doi.org/10.1615/ InterJFluidMechRes.v32.i6.50 Kivva S, Zheleznyak M, Pylypenko O, Yoschenko V (2020) Open water flow in a wet/dry multiplyconnected channel network: a robust numerical modeling algorithm. Pure Appl Geophys 177:3421–3458. https://doi.org/10.1007/s00024-020-02416-0 Kojiri T, Hamaguchi T, Ode M (2008) Assessment of global warming impacts on water resources and ecology of a river basin in Japan. J Hydro Environ Res 1:164–175. https://doi.org/10.1016/j. jher.2008.01.002

248

M. Zheleznyak et al.

Konoplev A (2020) Mobility and bioavailability of Chernobyl-derived radionuclides in soil-water environment: review. In: Konoplev A, Kato K, Kalmykov S (eds) Behavior of radionuclides in the environment II: Chernobyl. Springer Nature, Cham, pp 157–193. https://doi.org/10.1007/ 978-981-15-3568-0_3 Konoplev AV, Bulgakov AA, Hoffman FO, Kanyar B, Lyashenko G, Nair SK, Popov A, Raskob W, Thiessen KM, Watkins B, Zheleznyak M (1999) Validation of models of radionuclide wash-off from contaminated watersheds using Chernobyl data. J Environ Radioact 42 (2–3):131–141. https://doi.org/10.1016/S0265-931X(98)00050-2 Konoplev A, Golosov V, Laptev G, Nanba K, Onda Y, Takase T, Wakiyama Y, Yoshimura K (2016) Behavior of accidentally released radiocesium in soil–water environment: looking at Fukushima from a Chernobyl perspective. J Environ Radioact 151:568–578. https://doi.org/10. 1016/j.jenvrad.2015.06.019 Konoplev А, Kanivets V, Zhukova О, Germenchuk М, Derkach H (2021) Mid-to long-term radiocesium wash-off from contaminated catchments at Chernobyl and Fukushima. Water Res 188:116514. https://doi.org/10.1016/j.watres.2020.116514 Konoplev A, Wakiyama Y, Wada T, Igarashi Y, Kanivets V, Nanba K (2022) Behavior of Fukushima-derived radionuclides in soil-water environment: review. In: Nanba K, Konoplev A, Wada T (eds) Behavior of radionuclides in the environment III: Fukushima. Springer Nature, Singapore, pp 33–68 Kraus NC, Larson M (1991) NMLONG: numerical model for simulating the longshore current. In: Report 1, Model development and tests. Technical report DRP-91-1. U.S. Army Engineer Waterways Experiment Station, Vicksburg, MS Krone RB (1962) Flume studies of the transport of sediment in estuarial shoaling processes. Hydraulic Engineering Laboratory and Sanitary Engineering Research Laboratory, University of California at Berkeley, Berkeley, CA Kurikami H, Kitamura A, Yokuda ST, Onishi Y (2014) Sediment and 137Cs behaviors in the Ogaki Dam Reservoir during a heavy rainfall event. J Environ Radioact 137:10–17. https://doi.org/10. 1016/j.jenvrad.2014.06.013 Kurikami H, Funaki H, Malins A, Kitamura A, Onishi Y (2016) Numerical study of sediment and 137 Cs discharge out of reservoirs during various scale rainfall events. J Environ Radioact 164:73–83. https://doi.org/10.1016/j.jenvrad.2016.07.004 Le Provost C, Genco ML, Lyard F, Vincent P, Cenceill P (1994) Spectroscopy of the world ocean tides from a hydrodynamic finite element model. J Geophys Res 99(12):777–797 Liu X, Machida M, Kurikami H, Kitamura A (2019) Long-term simulations of radiocesium discharge in watershed with improved radioesium wash-off model: applying the model to Abukuma River basin of Fukushima. J Environ Radioact 203:135–146. https://doi.org/10. 1016/j.jenvrad.2019.03.001 Maderich V, Heling R, Bezhenar R, Brovchenko I, Jenner H, Koshebutskyy V, Kuschan A, Terletska K (2008) Development and application of 3D numerical model THREETOX to the prediction of cooling water transport and mixing in the inland and coastal waters. Hydrol Process Int J 22(7):1000–1013. https://doi.org/10.1002/hyp.6985 Maderich V, Brovchenko I, Dvorzhak A, Ievdin I, Koshebutsky V, Perianez R (2016) Integration of 3D model THREETOX in JRODOS, implementation studies and modelling of Fukushima scenarios. Radioprotection 51(2):133–135. https://doi.org/10.1051/radiopro/2016049 Margvelashvily N, Maderich V, Zheleznyak M (1999) Simulation of radionuclide fluxes from the Dnieper-Bug Estuary into the Black Sea. J Environ Radioact 43(2):157–171. https://doi.org/10. 1016/S0265-931X(98)00089-7 McCuen RH, Knight Z, Cutter AG (2006) Evaluation of the Nash–Sutcliffe efficiency index. J Hydrol Eng 11(6):597–602. https://doi.org/10.1061/(ASCE)1084-0699(2006)11:6(597) METIS (2021) Serial graph partitioning and fill-reducing matrix ordering. METIS. http://glaros.dtc. umn.edu/gkhome/views/metis. Accessed 17 Jun 2021

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

249

MEXT (2011) Results of the fourth Airborne monitoring Survey by MEXT. Ministry of Education, Culture, Sports, Science and Technology, Tokyo. http://radioactivity.nsr.go.jp/en/contents/ 4000/3179/24/1270_1216.pdf. Accessed 17 Jun 2021 Mizoguchi M (2019) Challenges of agricultural land remediation and renewal of agriculture in Iitate village by a collaboration between researchers and a non-profit organization. In: Nakanishi TM et al (eds) Agricultural implications of the Fukushima nuclear accident (III). Springer, Singapore, pp 177–190. https://doi.org/10.1007/978-981-13-3218-0_16 MOE (2019) Environmental remediation. In: Decontamination projects for radioactive contamination discharged by Tokyo Electric Power Company Fukushima Daiichi Nuclear Power Station Accident (Full Version). Ministry of Environment, Government of Japan, Tokyo. http://josen. env.go.jp/en/policy_document. Accessed 17 Jun 2021 Monte L (1996) Analysis of models assessing the radionuclide migration from catchments to water bodies. Health Phys 70:227–237. https://doi.org/10.1097/00004032-199602000-00011 Monte L, Perianez R, Kivva S, Laptev G, Angeli G, Barros H, Zheleznyak M (2006) Assessment of state-of-the-art models for predicting the remobilisation of radionuclides following the flooding of heavily contaminated areas: the case of Pripyat River floodplain. J Environ Radioact 88 (3):267–288. https://doi.org/10.1016/j.jenvrad.2006.02.006 Monte L, Boyer P, Brittain JE, Goutal N, Heling R, Kryshev A, Kryshev I, Laptev G, Luck M, Periañez R, Siclet F, Zheleznyak M (2008) Testing models for predicting the behaviour of radionuclides in aquatic systems. Appl Radiat Isot 66(11):1736–1740. https://doi.org/10.1016/j. apradiso.2007.09.020 Monte L, Brittain JE, Gallego E, Håkanson L, Hofman D, Jiménez A (2009) MOIRA-PLUS: a decision support system for the management of complex fresh water ecosystems contaminated by radionuclides and heavy metals. Comput Geosci 35(5):880–896. https://doi.org/10.1016/j. cageo.2008.03.008 Morel-Seytoux HJ, Khanji J (1974) Derivation of an equation of infiltration. Water Resour Res 10 (4):795–800 Mori K, Tada K, Uchiyama Y, Yamane M, Tosaka H (2011) Development of a coupled simulation technique of terrestrial fluid-flow and sediment transport. J Jpn Soc Civ Eng Ser B1 Hydraulic Eng 67(4):739–744. https://doi.org/10.2208/jscejhe.67.I_739 Mori K, Tada K, Tawara Y, Ohno K, Asami M, Kosaka K, Tosaka H (2015) Integrated watershed modeling for simulation of spatiotemporal redistribution of post-fallout radionuclides: application in radiocesium fate and transport processes derived from the Fukushima accidents. Environ Model Softw 72:126–146. https://doi.org/10.1016/j.envsoft.2015.06.012 Nagao S, Kanamori M, Ochiai S, Inoue M, Yamamoto M (2015) Migration behavior of 134Cs and 137Cs in the Niida River water in Fukushima prefecture, Japan during 2011–2012. J Radioanal Nucl Chem 303:1617–1621. https://doi.org/10.1007/s10967-014-3686-9 Nanba K, Zheleznyak M, Kivva S, Konoplev A, Maderich V, Koshebutsky V, Gallego E, Papush L, Mikhalskyi O (2016) Implementation of hydrological dispersion module of JRODOS for the assessment of 137Cs transport and fate in rivers, reservoirs and ponds of the Fukushima Prefecture. Radioprotection 51:145–148. https://doi.org/10.1051/radiopro/2016052 Nanba K, Moritaka S, Igarashi Y (2022) Dynamics of radiocesium in urban river in Fukushima-city. In: Nanba K, Konoplev A, Wada T (eds) Behavior of radionuclides in the environment III: Fukushima. Springer Nature, Singapore, pp 137–152 Nishimura H (1988) Computation of nearshore current. In: Horikawa K (ed) Nearshore dynamics and coastal processes. University of Tokyo Press, Tokyo, pp 271–291 Onishi Y (1981) Sediment and contaminant transport model. J Hydraul 107(9):1089–1107 Onishi Y (1987) A three-dimensional flow, energy, salinity, sediment and contaminant transport (FLESCOT) model for ocean disposal of low-level radioactive waste. In: Proceedings of the Workshop on Ocean Modeling Efforts at EPA, EPA-520/1-87-013. United States Environmental Protection Agency, Office Radiation Programs, Washington, DC, pp 37–49 Onishi Y, Serne RJ, Arnold EM, Cowan CE, Thompson FL (1981) Critical review: radionuclide transport, sediment transport, and water quality mathematical modeling; and radionuclide

250

M. Zheleznyak et al.

adsorption/desorption mechanisms. In: NUREG/CR1322, PNL-2901. Pacific Northwest Laboratory, Richland, WA Onishi Y, Kivva SL, Zheleznyak MJ, Voitsekhovich OV (2007) Aquatic assessment of the Chernobyl nuclear accident and its remediation. J Environ Eng 133(11):1015–1023. https:// doi.org/10.1061/(ASCE)0733-9372(2007)133:11(1015) OpenACC (2021) The OpenACC specification. https://www.openacc.org/specification. Accessed 17 Jun 2021 Partheniades E (1962) A study of erosion and deposition of cohesive soils in salt water. Thesis presented to the University of California at Berkeley, California Patin J, Onda Y, Yoda H (2013) Two year monitoring of soil and radiocesium redistribution at the hillslope scale after the 2011 Fukushima nuclear accident fallout. In: EGU General Assembly Conference Abstracts. EGU2013-12494 Periáñez R (2004) Testing the behaviour of different kinetic models for uptake/release of radionuclides between water and sediments when implemented in a marine dispersion model. J Environ Radioact 71(3):243–259. https://doi.org/10.1016/S0265-931X(03)00172-3 Periáñez R, Brovchenko I, Jung T, Kim KO, Maderich V (2018) The marine Kd and water/sediment interaction problem. J Environ Radioact 192:635–647. https://doi.org/10.1016/j.jenvrad.2018. 02.014 Phillips OM (1977) The dynamics of the upper ocean. Cambridge University Press, New York, NY. 336 p PNNL (2021) DHSVM: distributed hydrology soil vegetation model. Pacific Northwest National Laboratory, Richland, WA. https://www.pnnl.gov/projects/distributed-hydrology-soil-vegeta tion-model. Accessed 17 Jun 2021 Pratama MA, Yoneda M, Shimada Y, Matsui Y, Yamashiki Y (2015) Future projection of radiocesium flux to the ocean from the largest river impacted by Fukushima Daiichi Nuclear Power Plant. Sci Rep 5(1):1–9. https://doi.org/10.1038/srep08408 Pylypenko O, Zheleznyak M, Demchenko R, Kivva S, Sorokin M, Dykyi P (2020) Modelling of Tsunami inundation in 2011 at the sites of Three Nuclear Power Plants-Onagawa, Fukushima Daiichi and Fukushima Daini. In: EGU General Assembly Conference Abstracts, 13940. https:// doi.org/10.5194/egusphere-egu2020-13940 Raskob W, Trybushnyi D, Ievdin I, Zheleznyak M (2011) JRODOS: platform for improved long term countermeasures modelling and management. Radioprotection 46(6):731–736. https://doi. org/10.1051/radiopro/20116865s Rawls WJ, Brakensiek DL (1989) Estimation of soil water retention and hydraulic properties. In: Unsaturated flow in hydrologic modeling. Springer, Dordrecht, pp 275–300 Reid RO (1990) Water level changes-tides and storm surges. In: Herbich JB (ed) Handbook of coastal and ocean engineering. Gulf publishing, Houston, TX, pp 533–590 van Rijn RLC (1993) Principles of sediment transport in rivers, estuaries and coastal seas. Aqua Publications, Amsterdam Rios LM, Sahinidis NV (2013) Derivative-free optimization: a review of algorithms and comparison of software implementations. J Glob Optim 56(3):1247–1293. https://doi.org/10.1007/ s10898-012-9951-y Sakuma K, Kitamura A, Malins A, Kurikami H, Machida M, Mori K, Tada K, Kobayashi T, Tawara Y, Tosaka H (2017) Characteristics of radio-cesium transport and discharge between different basins near to the Fukushima Daiichi Nuclear Power Plant after heavy rainfall events. J Environ Radioact 169:137–150. https://doi.org/10.1016/j.jenvrad.2016.12.006 Sakuma K, Malins A, Funaki H, Kurikami H, Niizato T, Nakanishi T, Mori K, Tada K, Kobayashi T, Kitamura A, Hosomi M (2018) Evaluation of sediment and 137Cs redistribution in the Oginosawa River catchment near the Fukushima Daiichi Nuclear Power Plant using integrated watershed modeling. J Environ Radioact 182:44–51. https://doi.org/10.1016/j. jenvrad.2017.11.021 Sakuma K, Nakanishi T, Yoshimura K, Kurikami H, Nanba K, Zheleznyak M (2019a) A modeling approach to estimate the 137Cs discharge in rivers from immediately after the Fukushima

11

Modeling of Behavior of Fukushima-Derived Radionuclides in Freshwater Systems

251

accident until 2017. J Environ Radioact 208:106041. https://doi.org/10.1016/j.jenvrad.2019. 106041 Sakuma K, Tsuji H, Hayashi S, Funaki H, Malins A, Yoshimura K, Kurikami H, Kitamura A, Iijima K, Hosomi M (2019b) Applicability of Kd for modelling dissolved 137Cs concentrations in Fukushima river water: case study of the upstream Ota River. J Environ Radioact 210:105815. https://doi.org/10.1016/j.jenvrad.2018.09.013 Saxton KE, Rawls WJ (2006) Soil water characteristic estimates by texture and organic matter for hydrologic solutions. Soil Sci Soc Am J 70(5):1569–1578. https://doi.org/10.2136/sssaj2005. 0117 Schureman P (1958) Manual of harmonic analysis and prediction of tides, Special publication #98. Coast and Geodetic Survey, U.S. Department of Commerce, U.S. Government Printing Office, Washington, DC. 1958 Edition with corrections Shirazi MA, Boersma L (1984) A unifying quantitative analysis of soil texture. Soil Sci Soc Am J 48(1):142–147 Smith RE, Parlange JY (1978) A parameter-efficient hydrologic infiltration model. Water Resour Res 14:533–538 Smith RE, Goodrich DC, Quinton JN (1995a) Dynamic, distributed simulation of watershed erosion: the KINEROS2 and EUROSEM models. J Soil Water Conserv 50(5):517–520 Smith RE, Goodrich DC, Woolhiser DA, Unkrich CL (1995b) KINEROS – a kinematic runoff and erosion model. In: Singh VP (ed) Computer models of watershed hydrology. Water Resources Publication, Highland Ranch, CO, pp 697–732 Smith JT, Belova NV, Bulgakov AA, Comans RN, Konoplev AV, Kudelsky AV, Madruga MJ, Voitsekhovitch OV, Zibold G (2005) The “AQUASCOPE” simplified model for predicting 89, 90Sr, 131I, and 134,13 Cs in surface waters after a large-scale radioactive fallout. Health Phys 89(6):628–644. https://doi.org/10.1097/01.hp.0000176797.66673.b7 Takata H, Kumamoto Y (2022) Spatiotemporal variation of radiocesium in coastal and oceanic seawater. In: Nanba K, Konoplev A, Wada T (eds) Behavior of radionuclides in the environment III. Fukushima. Springer Nature, Singapore, pp 255–284 Taniguchi K, Onda Y, Smith HG, Blake W, Yoshimura K, Yamashiki Y, Kuramoto T, Saito K (2019) Transport and redistribution of radiocesium in Fukushima fallout through rivers. Environ Sci Technol 53(21):12339–12347. https://doi.org/10.1021/acs.est.9b02890 Tateda Y, Tsumune D, Misumi K, Aono T, Kanda J, Ishimaru T (2017) Biokinetics of radiocesium depuration in marine fish inhabiting the vicinity of the Fukushima Daiichi Nuclear Power Plant. J Environ Radioact 166:67–73. https://doi.org/10.1016/j.jenvrad.2016.02.028 Tsumune D, Tsubono T, Misumi K, Tateda Y, Toyoda Y, Onda Y, Aoyama M (2020) Impacts of direct release and river discharge on oceanic 137Cs derived from the Fukushima Daiichi Nuclear Power Plant accident. J Environ Radioact 214:106173. https://doi.org/10.1016/j. jenvrad.2020.106173 Van der Perk M, Jetten VG, Karssenberg D, He Q, Walling DE, Laptev GV, Voitsekhovitch OV, Svetlichnyi AA, Slavik O, Linnik VG, Korobova EM (2000) Assessment of spatial redistribution of Chernobyl-derived radiocaesium within catchments using GIS-embedded models. IAHS Publ 263:277–284 Velleux ML, England JF Jr, Julien PY (2008) TREX: spatially distributed model to assess watershed contaminant transport and fate. Sci Total Environ 404(1):113–128. https://doi.org/ 10.1016/j.scitotenv.2008.05.053 Wada T, Nemoto Y, Fujita T, Kawata G, Kamiyama K, Sohtome T, Narita K, Watanabe M, Enomoto M, Suzuki S, Amano Y, Sato M, Iwasaki T, Mizuno T, Nanba K (2022) Cesium radioactivity in marine and freshwater products and its relation to the restoration. In: Nanba K, Konoplev A, Wada T (eds) Behavior of radionuclides in the environment III. Fukushima. Springer Nature, Singapore, pp 313–353 Wahr JM (1981) Body tides on an elliptical, rotating, elastic and oceanless earth. Geophys J R Astron Soc 64:677–703 Wakiyama Y, Onda Y, Yoshimura K, Igarashi Y, Kato H (2019) Land use types control solid washoff rate and entrainment coefficient of Fukushima-derived 137Cs, and their time dependence. J Environ Radioact 210:105990. https://doi.org/10.1016/j.jenvrad.2019.105990

252

M. Zheleznyak et al.

Wakiyama Y, Konoplev A, Golosov V, Thoa N, Niida T, Tsukada Y, Takase T, Nanba K, Zheleznyak M (2022) Temporal variations in particulate and dissolved Cs-137 activity concentration in the Abukuma River during high-flow events. In: Behavior of radionuclides in the environment III. Fukushima. Springer Nature, Singapore, pp 153–175 Walker DW, Dongarra JJ (1996) MPI: a standard message passing interface. Supercomputer 12:56–68 Wei L, Kinouchi T, Yoshimura K, Velleux ML (2017) Modeling watershed-scale 137Cs transport in a forested catchment affected by the Fukushima Daiichi Nuclear Power Plant accident. J Environ Radioact 171:21–33. https://doi.org/10.1016/j.jenvrad.2017.01.017 Wigmosta MS, Vail LW, Lettenmaier DP (1994) A distributed hydrology-vegetation model for complex terrain. Water Resour Res 30(6):1665–1679. https://doi.org/10.1029/94WR00436 Yamada S, Kitamura A, Kurikami H, Yamaguchi M, Malins A, Machida M (2015) Sediment and 137Cs transport and accumulation in the Ogaki Dam of eastern Fukushima. Environ Res Lett 10 (1):014013. https://doi.org/10.1088/1748-9326/10/1/014013 Yamaguchi M, Kitamura A, Oda Y, Onishi Y (2014) Predicting the long-term 137Cs distribution in Fukushima after the Fukushima Daiichi Nuclear Power Plant accident: a parameter sensitivity analysis. J Environ Radioact 135:135–146. https://doi.org/10.1016/j.jenvrad.2014.04.011 Yoshimura K, Onda Y, Sakaguchi A, Yamamoto M, Matsuura Y (2015) An extensive study of the concentrations of particulate/dissolved radiocaesium derived from the Fukushima Daiichi Nuclear Power Plant accident in various river systems and their relationship with catchment inventory. J Environ Radioact 139:370–378. https://doi.org/10.1016/j.jenvrad.2014.08.021 Zheleznyak MJ, Demchenko RI, Khursin SL, Kuzmenko YI, Tkalich PV, Vitiuk NY (1992) Mathematical modeling of radionuclide dispersion in the Pripyat-Dnieper aquatic system after the Chernobyl accident. Sci Total Environ 112(1):89–114. https://doi.org/10.1016/0048-9697 (92)90241-j Zheleznyak MJ, Tkalich PV, Lyashenko GB, Marinets AV (1993) Aquatic dispersion models-first approaches to integration into the CEC Decision Support System based on post-Chernobyl experience. Radiat Prot Dosim 50(2–4):235–241. https://doi.org/10.1093/oxfordjournals.rpd. a082093 Zheleznyak M, Blaylock G, Garnier-Laplace J, Gontier G, Konoplev AV, Marinets A, Monte L, Onishi Y et al (1995) Modelling of radionuclide transfer in rivers and reservoirs: validation study within the IAEA/CEC VAMP programme. In: International Symposium on Environmental Impact of Radioactive releases, 8-12 May 1995, IAEA-SM-339/175. IAEA, Vienna, pp 85–86 Zheleznyak M, Shepeleva T, Sizonenko V, Mezhueva I (1997) Simulation of countermeasures to diminish radionuclide fluxes from the Chernobyl zone via aquatic pathways. Radiat Prot Dosim 73(1–4):181–186. https://doi.org/10.1093/oxfordjournals.rpd.a032129 Zheleznyak M, Potempski S, Bezhenar R, Boyko A, Ievdin I, Kadlubowski A, Trybushnyi D (2010) Hydrological dispersion module of JRODOS: development and pilot implementation – the Vistula river basin. Radioprotection 45(5):113–122. https://doi.org/10.1051/radiopro/ 2010044 Zheleznyak M, Kivva S, Ievdin I, Boyko O, Kolomiets P, Sorokin M, Mikhalskyi O, Gheorghiu D (2016) Hydrological dispersion module of JRODOS: renewed chain of the emergency response models of radionuclide dispersion through watersheds and rivers. Radioprotection 51:129–131. https://doi.org/10.1051/radiopro/2016048 Zheleznyak M, Dykyi P, Kivva S, Pylypenko O, Sorokin M, Aoyama M, Tsumune D (2018) Modelling of Cs-137 transport in the nearshore zone of Fukushima – Daiichi NPP under the combined action of waves, currents and fluxes of sediments. In: EGU General Assembly 2018 Geophysical Research Abstracts 20, EGU2018-19294. EGU, Munich Zia A, Banihashemi MA (2007) Simple efficient algorithm (SEA) for shallow water flows with shock wave on dry and irregular beds. Int J Numer Methods Fluids 56(11):2021–2043. https:// doi.org/10.1002/fld.1585

Part IV

Radionuclide Behavior in Coastal and Marine Environment

Chapter 12

Spatiotemporal Variation of Radiocesium in Coastal and Oceanic Seawater Hyoe Takata and Yuichiro Kumamoto

Abstract We review the marine radioactive monitoring, sources, and spatiotemporal change of dissolved radiocesium released from the Fukushima Dai-ichi Nuclear Power Plant (FNPP1) accident in 2011 during the past decade. The monitoring stations in the coastal and offshore areas increased according to the spread of radiocesium in the first year, Fiscal Year 2011 (FY2011), and have been fixed since FY2012. A problem arose from different detection limits from 1 to 1000 Bq m
3 with radiocesium data reported by different facilities that have different monitoring purposes. The probable range of the two major sources of the FNPP1-derived radiocesium, the atmospheric deposition and direct discharge, is estimated to be 9–12 PBq (petabequerel) and 3–5 PBq, respectively. Despite the decline of ongoing direct release from the FNPP1, the activity concentrations in the coastal area are still higher than the concentrations before the accident. This is mainly due to riverine transport of the FNPP1-derived radiocesium deposited on the land. The radiocesium deposited on and discharged into the coastal and offshore areas of the FNPP1 was transported eastward along the surface currents. It had reached the North American Continent by 2016 and then returned westward along with the subarctic gyre current. The radiocesium deposited just south of the Kuroshio Front was conveyed southward in subsurface layers along with the formation/ subduction of the Subtropical Mode Water (STMW) in the North Pacific Ocean. This resulted in the transport of the FNPP1-derived radiocesium to the western end of the subtropical area, the eastern East China Sea, and then the Japan Sea by 2013. Keywords Dissolved radiocesium · Marine radioactive monitoring · Atmospheric deposition · Direct discharge · Riverine input · Subarctic and subtropical gyres · Mode waters

H. Takata (*) Institute of Environmental Radioactivity, Fukushima University, Fukushima, Japan e-mail: [email protected] Y. Kumamoto Research Institute for Global Change, Japan Agency for Marine-Earth Science and Technology, Kanagawa, Japan © The Author(s), under exclusive license to Springer Nature Singapore Pte Ltd. 2022 K. Nanba et al. (eds.), Behavior of Radionuclides in the Environment III, https://doi.org/10.1007/978-981-16-6799-2_12

255

256

12.1

H. Takata and Y. Kumamoto

Introduction

After the giant tsunami attack following the Great East Japan Earthquake on March 11, 2011, the coastal and offshore areas of Fukushima and nearby prefectures in Japan were contaminated by radionuclides released from the damaged Fukushima Dai-ichi Nuclear Power Plant (FNPP1) operated by the Tokyo Electric Power Company (TEPCO). Just after the accident, a large release of a short half-life radionuclide, 131I (about 8 days half-life), was a public concern. Then, the concern shifted to radionuclides with a longer half-life, 134Cs (about 2 years), 137Cs (about 30 years), and 90Sr (about 30 years). A total release of radiocesium (134Cs and 137Cs) was more than 10 times larger than that of 90Sr (Buesseler et al. 2017). And most of the FNPP1-derived radiocesium in the marine environment is dissolved in seawater because of its solubility. As a result, radiocesium deposited on the seabed around Fukushima is estimated to be about 1  1014 Bq (bequerel) (Kusakabe et al. 2013; Otosaka and Kato 2014) that is less than 1% of the total release to the ocean (Sect. 12.3). Therefore, here we focus on the spatiotemporal variation of radiocesium dissolved in seawater during the past decade. Many studies on the spatiotemporal variation of the FNPP1-derived radiocesium in seawater have been reviewed by several researchers. Povinec et al. (2013a) and the United Nations Scientific Committee on the Effects of Atomic Radiation (UNSCEAR) (2014) reviewed the behavior of FNPP1-derived radionuclides in the marine environment by 2013. Nakata and Sugisaki (2015) compiled studies on contamination of the marine environment but they mainly focused on the impacts of the FNPP1-derived radionuclides on the fishery in Japan. Hirose (2016) and Kaeriyama (2017) widely discussed the spatiotemporal change in the FNPP1derived radiocesium in seawater in the coastal, offshore, and open ocean areas by 2016. Buesseler et al. (2017) reviewed the FNPP1-derived radionuclides in seawater, sinking particle, sediment, and biota comprehensively. However, their description of the FNPP1-derived radiocesium in seawater is concise. A book recently published (Nakajima et al. 2019) includes a chapter entitled “Ocean transport of radioactive materials.” This chapter, however, focused on the serious contamination in the early stage of the FNPP1 accident because it originated from a report published in 2014 in Japanese. As a result, the spatiotemporal change in the FNPP1-derived radiocesium in seawater after 2016 has not been reviewed yet. We first review the radioactive monitoring activity in the coastal and offshore areas of the FNPP1 and the procedures of seawater sampling and measurement methods for radiocesium in seawater (Sect. 12.2). Although there were confusion and update in the procedures during the radioactive monitoring, they have not been reviewed in the previous works. Because of the FNPP1 location in the North Pacific shoreline of Japan (37.4 N/141.0 E) and the predominant westerly wind, most of the radiocesium released from the plant was deposited on the North Pacific Ocean (Sect. 12.3). Just after the accident, TEPCO started the monitoring of radionuclides in the seawater within the 30-km radius from the FNPP1. Marine Ecology Research Institute (MERI) contracted by the Japanese Government started the marine radioactivity monitoring in the coastal and offshore areas of the FNPP1. Since then, several facilities have started monitoring in the marine environment. However, the

12

Spatiotemporal Variation of Radiocesium in Coastal and Oceanic Seawater

257

procedures of seawater sampling and measurement methods for radiocesium in seawater among the facilities including TEPCO and MERI are not unified because the purposes of their monitoring are different (radioactive impacts on human, fishery, and environment). Secondly, we propose probable ranges for sources of the FNPP1-derived radiocesium in the North Pacific Ocean, which are derived from 32 studies published by 2019 (Sect. 12.3). To understand the fate of radiocesium in the marine environment, it is essential to estimate how much of radiocesium was deposited on and discharged to the ocean. However, unlike radiocesium deposited on land, radiocesium in the ocean was immediately diluted by water mixing and transported by water currents. Thereby, it is difficult to estimate the total amount of radiocesium input to the ocean, which resulted in a large range of estimated values among previous studies. Finally, we compile the data of radiocesium activity concentration in seawater between 2011 and 2019 and review its spatiotemporal variation in the coastal and offshore areas (Sect. 12.4) and open ocean (Sect. 12.5) during the past decade. The radiocesium activity concentrations within 10 km of Fukushima and neighboring prefectures are still higher than those before the accident. Direct release of contaminated water from the plant, reentry from sediment through the submarine groundwater discharge, and desorption from riverine particle following to their inputs via rivers probably have maintained the higher concentrations in the nearshore areas (Sanial et al. 2017: Takata et al. 2020a). Those in the coastal area within 30-km distance from the FNPP1 peaked in early April 2011, decreased exponentially within a year, and then reached the pre-accident level before 2011. The transport from the coastal area to the open ocean was derived from (1) water mixing between the contaminated coastal seawater and the less-contaminated seawater in the open ocean, (2) eastward transport in the surface layer along with the Kuroshio-Oyashio Current system, and (3) subduction of surface water into subsurface layers in the midwinter (Buesseler et al. 2017). It should be mentioned that the FNPP1-derived 134Cs and 137Cs activities were equivalent (Buesseler et al. 2017). Before the FNPP1 accident, 137Cs was also deposited on the North Pacific Ocean mainly by the atmospheric testing of nuclear weapons in the 1950s and 1960s. This bomb-derived 137Cs in the North Pacific Ocean remained just before the FNPP1 accident and is still there. In contrast, virtually all the 134Cs released before the FNPP1 accident, which were derived from the testing of nuclear weapons and the Chernobyl accident, was negligible because of its small amount and radioactive decay. As a result, 134Cs could be a more suitable tracer than 137Cs for the tracing of the FNPP1-derived radiocesium. Also, the radiocesium activity concentration should be corrected for the radioactive decay to discuss its spatiotemporal change due to seawater mixing and transport. In Sects. 12.3 and 12.5, we review the data of 134Cs with the decay-correction to the date of the FNPP1 accident because we discuss the inventory and transport of the FNPP1derived radiocesium in the North Pacific Ocean where the bomb-derived 137Cs remains. On the other hand, in Sect. 12.4, we review the data of 137Cs without the decay-correction to the accident date to compare the activity concentrations before and after the accident in the past decade.

258

H. Takata and Y. Kumamoto 38'30’

a

b

45o

Latitude (N)

Latitude (N)

38'00’

40o

S1 S2 1 – A 1–B 1–1 1–2 1–3

37'30’

1–4 2–1 2–2 2–3 2–6 2–5 2–4 S4 S3

b c,d 35o

37'00’

30o 130o

135o

140o

145o

36'30’ 140'30’

150o

Longitude (E) 39o

c

Latitude (N)

38o

10

2

K1 K2K3

8

11

12

B1 B3

35o 140o

141o

11

15

14

15

17

18

20

21

19

20

21

7

22

23

25

26

27

25

26

27

142o

143o

144o

142o

143o

144o

B5

C1 C3 D1 D3 E5 E1 E3 F1 F3 G0G1 G3 G4 H1 H3 I0I1 I3

4

5

J1IB2J3 K1 IB4

L1L2L3L4 24

10

A1 A3

13

14

G0G1G2G3 G4 16 H1H2H3 I0I1 l2l3 9 6 19 J1 J2J3

36o

142'00’

MI4

C1C2C3 D1D2D3 E5 E1E2E3E4 3 F1F2F3

37o

141'30’

Longitude (E)

d

1 A1A2A3 a1 B1B2B3B4

141'00’

L1 L3 M1

145o 140o

141o

Longitude (E)

145o

Longitude (E)

Fig. 12.1 Temporal change of sampling stations conducted by MERI. (a) Map of Japanese Islands. (b) Sampling stations (red circles) from March 23 to May 7, 2011, and additional stations (blue circles) from March 28 to May 7, 2011. (c) Stations from May 9, 2011 to February 21, 2012. (d) Current stations from May 2012. The stars indicate the FNPP1

12.2

Radioactive Monitoring and Sample Pretreatment/Analysis

Because the FNPP1 locates in the North Pacific shoreline of Japan (Fig. 12.1), most of the radiocesium released from the plant was deposited on the North Pacific Ocean. Just after the accident, TEPCO started radioactive monitoring at stations within a 30-km radius of the FNPP1 and the monitoring area had expanded to the nearshore

12

Spatiotemporal Variation of Radiocesium in Coastal and Oceanic Seawater

259

areas of Fukushima, Miyagi, and Ibaraki Prefectures. According to a contract with the Ministry of Education, Culture, Sports, Science and Technology (MEXT) (Nuclear Regulatory Agency (NRA) after April 2013), MERI started the radioactive survey along the lines of 30-km radius from March 23, 2011, and then the survey was implemented by expanding the sampling area in the western North Pacific Ocean (Oikawa et al. 2013; Kusakabe and Takata 2020). In addition, the three prefectures have been monitoring regularly in the nearshore and coastal areas of their prefectural shorelines, and the Japan Coast Guard has also been conducting the monitoring in the open ocean (NRA 2020). Some facilities also collected seawater samples for radiocesium measurement for their scientific research, not for the monitoring activity (Buesseler et al. 2012; Kumamoto et al. 2019). The monitoring activities by TEPCO and NRA continue, although most of the research surveys by many facilities in Japan and overseas have finished. Herein, we review the monitoring of MERI contracted by MEXT. The first cruise commenced on March 23, 2011, and then three other cruises followed until May 7, 2011 at eight sampling stations located 30 km off the coast from the FNPP1 (Fig. 12.1b). In the middle of the survey period, eight more stations were added on both northmost and southmost sides of the first sampling stations, so that the stations formed a U-shape surrounding the FNPP1 (total stations: 16, Fig. 12.1b). After May 9, 2011, the monitoring was expanded to the western North Pacific Ocean and continued till the end of February 2012 (maximum 51 stations, Fig. 12.1c). The results from these monitoring activities by February 2012 have been reported by Oikawa et al. (2013). In May 2012, the monitoring stations were changed again and are being continued until now (42 stations, Fig. 12.1d). The sampling intervals varied from once every quarter to once a year. On the other hand, in the monitoring from the nearshore to the coastal area conducted by TEPCO, the sampling frequency is measured every day in the inside of the harbor of the FNPP1, and once/twice a month in the outside of the harbor and nearshore areas from Miyagi (northmost) to Ibaraki (southmost) prefectures. These differences in the sampling intervals resulted from that the activity concentration of the FNPP1-derived radiocesium decreased from the coastal area to the open ocean. Details of sampling and its intervals for the monitoring at present are available on the web pages of TEPCO (TEPCO 2020) and NRA (NRA 2020). During the MEXT monitoring from March 23 to May 7, 2011, 0.5–2 L of seawater samples was collected and the radiocesium in the seawater sample was measured using a germanium (Ge)-detector without any sample pretreatment because the purpose of the monitoring was the urgent survey. As a result, the detection limits of 134Cs and 137Cs were high, 1000–10,000 Bq m
3. After a rapid decrease in the activity concentration in the coastal and offshore areas, the detection limits have been lowered to about 1 Bq m
3 by the preconcentration of seawater using ammonium phosphomolybdate before the measurement with a Ge-detector. Fukushima and neighboring prefectures (Iwate, Miyagi, and Ibaraki Prefectures) and other facilities also conducted the marine radioactivity monitoring according to their procedures for their purposes (radioactive impacts on humans, fishery, and environment). As a result, the radiocesium activity concentrations reported by the different

260

H. Takata and Y. Kumamoto

facilities could not be compared directly because the pretreatment method and/or the detection limit were different. For instance, the detection limit of radiocesium measurement in the nearshore, offshore, and open ocean areas conducted by NRA was about 1 Bq m
3, while those conducted by other facilities varied from 1 to 1000 Bq m
3 (NRA 2020). Now, the activity concentration of 137Cs in the nearshore area is less than 1000 Bq m
3 (Takata et al. 2020a). Therefore, some of the monitoring data without the preconcentration in the nearshore area have been “ND (not detected: