Artificial Radionuclides in the Environment
2018
環境における人工放射能の研究（2018 年版）

Meteorological Research Institute
JMA
March 2020

気象庁

気象研究所

最終ページに図の詳しい説明があります。
See figure caption at the end page of this publication.

「環境における人工放射能の研究（2018 年版）」について

気象研究所では、1954 年以来、主に現在の原子力規制庁放射能調査研究費により大気及び海洋の環境放射能の研究を実施してきました。特に人工放射性核種の降下量を 60 年余りの長期にわたり、東京・つくばで精密測定してきました。この観測は、世界でも最長の定点観測です。このモニタリング観測の期間では、1950 年代から 1960 年代にかけては、旧ソ連や米国等による核実験が行われ、また、その後も 1986 年には旧ソ連のチェルノブイリ原子力発電所の事故が発生するなどがありましたが、この観測によって、それらの影響を含めた
バックグラウンドの人工放射性核種の長期的変動を明らかにし、国民の安全・安心に寄与してきました。2011 年には不幸にして東日本大震災に伴う福島第一原子力発電所事故が発生し、それによって放出された放射性核種の推移を調査するという課題も担うこととなりました。今後も長期的視点で、人工放射性核種の変動を監視する必要があります。
放射性物質の再飛散・輸送・沈着による変動のメカニズムを正しく理解し、将来の変動を予測するためには、長期観測や再飛散観測、サンプルの分析等に加えて、数値計算モデルを用いた解析が必要です。気象研究所では、気候変動や降水予測の研究のため、大気中の微粒子の動きを表現する数値モデルを開発してきましたが、そのような数値モデルを利用した、
バックグラウンド放射能の変動メカニズムを解析する研究も行ってきました。
この研究によるこれまでの観測や解析から、長期的変動に対しては、一旦地面などに沈着した人工放射性核種が、土壌粒子・生物由来の粒子等を通じて空気中へ再飛散する過程が重要である可能性がわかりました。
本論文集「環境における人工放射能の研究（2018 年版）」 は、その研究成果を､関係省庁の担当者の方々及び大学や試験研究機関の研究者の方々に広く周知するために、2016 年から 2018 年までの期間に出版された論文（主に英語論文）を、過去から現在までの成果と最近のトピックスに関するテーマ毎に分類し、各テーマの冒頭に簡単な日本語の解説を加えて､一冊にまとめたものです。

令和 2 年 3 月
気象研究所 研究総務官 高薮 出

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Contents
0. Preface
1. Temporal trends of radioactivity concentrations in the atmosphere and observational
approaches for their resuspension mechanisms
2. Physical and chemical characteristics of the radioactive materials in the atmosphere
3. Evaluations of resuspension , transportation and deposition processes of radioactive materials
using numerical models
4. Publications

2016-2018

5. Publication list 1954-2018
6. Publication list (in Japanese)

1954-2018

7. History of the studies at the Meteorological Research Institute
8. A caption of cover art

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目次
0. はじめに
1. 大気中の放射性物質の濃度変動、再浮遊観測
2. 大気中の放射性物質の物理・化学性状3. 放射性物質の再浮遊、輸送、沈着過程のモデルによる評価
4. 出版論文 2016-2018
5. 論文リスト（英文誌）1954-2018
6. 論文リスト（和文誌）1954-2018
7. 表：研究の歴史
8. 表紙説明-4-

0. はじめに気象研究所（2019 年 3 月現在は環境・応用気象研究部および予報研究部）では、1954 年以来、環境放射能の観測・測定法の開発、放射能汚染の実態の把握、大気や海洋における物質輸送解明のトレーサーとしての利用を目的として、60 年余にわたって環境放射能の研究を実施してきた。1957 年以降、各省庁がそれぞれの所掌で実施してきた環境放射能調査研究関連業
務は、まず旧科学技術庁、その後文部科学省が所管してきた。さらに原子力規制委員会と原子力規制庁の発足に伴って 2013 年からは、原子力規制庁が所管することとなった。このような長期にわたり多数の研究者が係わり、観測・研究を継続してきた結果、環境放射能について世界的にも他に類を見ない貴重な時系列データが内外に提供され、また様々な気象学・海洋学的発見をもたらしてきた。この間の研究成果は 200 編以上の原著論文や解説資料として内外の雑誌や成書で公表されている。
1954 年 3 月 1 日に米国によりビキニ環礁で行われた水爆実験により、危険水域外で操業していた第五福竜丸乗組員が放射性物質を含む降灰（いわゆる死の灰）による被ばくを受けた事件を契機にして、日本における環境放射能研究が本格的に始まった。当時の気象研究所地球化学研究室は環境の放射能の分析・研究において日本で有数の研究室であり、海洋及び大気中の放射能汚染の調査・研究に精力的に取り組んだ。その結果、当時予想されていなかった海洋の
放射能汚染、さらに大気を経由した日本への影響など放射能汚染の拡大の実態を明らかにすることができた。1958 年には、放射能調査研究費による特定研究課題の一つである「放射化学分析（落下塵・降水・海水中の放射性物質の研究）」を開始し、以降、地点の変動はあるが、札幌、仙台、東京、大阪、福岡の五つの管区気象台、秋田、稚内、釧路、石垣島の 4 地方気象
台、輪島、米子の 2 測候所の全国 11 気象官署及び観測船で採取した海水中の人工放射性核種（90Sr、137Cs、3H 及びプルトニウム）の分析を実施してきた。
東京・つくばにおける大気中の人工放射性核種の降下量は 1961 年から 1962 年に行われた大規模な大気圏核実験の翌年である 1963 年に最大値を観測した。その後、「部分的核実験禁止条約」の締結により米ソの大気圏核実験が中止された結果、降下量はおよそ１年の半減時間で減少した。この放射性核種の降下量の減少率は成層圏にまきあげられた物質の滞留時間を反映している。その後も、中国及びフランスにより大気圏核実験は続けられ、人工放射性核種の降下量は増減を繰り返した。1980 年の中国による最後の大気圏核実験の後、フォールアウト
は再度 1 年の半減期間で減少し、1985 年には 1957 年の観測開始以降最も低いレベルになった。しかし、1986 年の旧ソ連のチェルノブイリ原子力発電所事故により、大気中の人工放射性核種濃度（特に揮発性の高い 131I、134Cs、137Cs など）は日本でも 1963 年に近いレベルに達するほど著しく増加した。大部分の放射性核種は対流圏の滞留時間（25 日）で減少したが一部 137Cs は成層圏にも輸送されていることが分かった。1988 年以降は低いレベルで推移しているが、明瞭な減少の傾向は見られなかった。この原因は一度地上に降下した放射性核種の再飛散に由来すると考えている。さらに、再飛散がどこで起るかについて研究を進め、東アジアの砂漠域で発生する黄砂が有力な候補であることを明らかにした。黄砂の発生は大陸域の環境変化と関連しており、降下物中の人工放射性核種は、砂漠化や関連する植生変化など大陸域の環境変化の指標となりうることが分かった。
大気フォールアウトの研究とともに、海洋における放射性核種の挙動も長期的に調査研究を実施してきた。日本周辺海域ばかりでなく、太平洋の広域に亘って海水試料の採取を実施
し、-5-放射能汚染の実態を明らかにした。1960 年代後半から 1970 年代の調査で、海洋表面水中の放射能が北半球の中緯度で高いことを明らかにし、フォールアウトの緯度分布を反映している
ことが分かった。その後、海洋表面水中の放射性核種は海洋の物質循環に支配されていることを明らかにした。さらに、海水中の人工放射性核種の分析法の高度化を実現し少量の試料で分析可能にし、海洋の 137Cs 濃度の精密鉛直断面を描き、核実験由来の 137Cs の主な部分は北太平洋の亜熱帯中層に存在していることを明らかにした。フォールアウトによる人工放射性核種の海洋への主な降下以来、数 10 年以上経過し、その広がりは北太平洋から、インド洋など
他の海盆に及んでいる。これに関連して 2003－4 年に実施された BEAGLE2003 の航海で採取された海水について人工放射性核種の分析を行って、その広がりについて知見が得られた。
1993 年旧ソ連/ロシアによる放射性廃棄物の日本海等への海洋投棄の実態が明らかにされ、それに伴う日本海の放射能調査の実施に参加した。放射性廃棄物による影響は検出されなかったが、調査の結果を踏まえ、日本海における固有水の生成過程及び生成場所（ウラジオストッ
ク沖）についての知見を得ることができた。これらは、海洋放射能のデータベースとして公開されている。
1990 年以降の環境放射能汚染として、1997 年の旧動力炉核燃料開発事業団「アスファルト
固化処理施設」の火災爆発事故や 1999 年の JCO ウラン燃料工場の臨界事故があるが、いずれも環境中に放出された放射能は極めて低いレベルで、その影響は殆どなかった。しかしながら、
137Cs
2011 年 3 月に起きた東日本大震災にともなう東京電力福島第一原子力発電所事故により、の総量だけで 10 PBq（ペタベクレル）以上という過去日本では経験したことのない大量の人
工放射性核種が環境中に放出され、その 2 割程度が日本の国土に降下し、残りの 8 割は北太平洋に降下したと考えられている。これらの影響評価やその後の環境中での拡散状況の把握のためにも、環境放射能調査・研究は今後とも重要である。
2006 年 4 月より 2011 年 3 月まで、気象研究所では放射能調査研究費による特定研究課題
として「放射性降下物の長期変動と再浮遊に関する研究」及び「海洋環境における放射性核種の長期挙動に関する研究」の二課題で環境放射能研究に取り組んできた。また、2011 年 8月からは、「大気を通じた人工放射性核種の陸圏・水圏への沈着およびその後の移行過程の解明研究」の課題のもと、過去の知見を十分に活用しながら、同年 3 月 11 日に発生した東日本大震災にともなう福島第一原子力発電所の事故以降の新たな事態に対応し、放出された
大量の人工放射性核種の挙動を解明する研究を行った。目の前の大気に加えて、北太平洋におけるこれらの核種による汚染実態の把握やその動態の解明に取り組んだ。
過去 60 年余にわたり実施されてきたこうした研究成果を踏まえ、2015 年 4 月からは「人工放射性核種のバックグラウンド大気監視と数値解析に関する研究」を新たに開始した。本報告書では、この研究課題で得られた成果を含め、最新の成果やトピックスをテーマ毎に記述した。関係各位の今後の研究や業務に役立つ資料とすべく、編集作業を行った。我が国における環境放射能研究や大気科学研究に多少なりとも寄与できたとすれば、著者一同の望外の喜びである。

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1. 大気中の放射能物質の濃度変動、再浮遊観測
1）はじめに気象研究所では、大気圏内核実験が盛んに実施された 1950 年代後期から現在までの期間にわたり、大気圏での人工放射性核種の濃度水準の実態とその変動要因を明らかにすべく、
環境影響の大きい重要な核種について観測を継続してきた。90Sr（半減期 28.8 年）および137
Cs（半減期 30.2 年）の月間降下量（大気降下物）の長期観測結果は満 60 年を超え、世界でも稀にみる長期変動に関して重要な知見をもたらしている。また、2003 年 3 月からは大気中浮遊粒子の採取、及び放射能分析も開始し、2011 年 3 月の東京電力福島第一原子力発電所事故（福島第一原発事故）に関する環境影響に関して重要な知見を取得し、また大気モデル改善に貢献し、その結果をいち早く学会で報告するとともに、気象研究所のホームペー
ジにも掲載し一般に公開してきた。複数台有する Ge 半導体分析装置は他研究機関の持つ環
境試料の測定にも貢献し、福島第一原発事故の影響解明にも大きく貢献してきた。前報のIgarashi et al., (2015) では福島第一原発事故以降の三年間に着目した大気降下物および大気
中放射能について報告をしてきたが、事故後七年が経過し、その濃度は徐々に下がってきた
ため、より高精度な測定技術が求められてきた。本報ではより高精度に 134Cs を測定するための定量方法の開発結果について報告し、事故後数年経過した現在における放射性物質の再飛散（二次飛散）や環境中での動態プロセスに関する考察について解説する。

2）研究方法月間大気降下物の捕集は、茨城県つくば市長峰にある気象研究所の露場観測実験棟屋上に設置した大気降下物捕集用のプラスチック製水盤（面積 4m2）で 1980 年代以降実施してきた。福島第一原発事故以降（2011 年 4 月以降）は放射能水準の上昇を考慮し、気象研究所
屋上にて 1 m2 水盤２基での捕集を実施した。2015 年 8 月以降は放射能水準の低下を考慮して 1m2 水盤４基での大気降下物採取を開始した。大気中浮遊粒子は、上記露場観測実験棟前に設置したハイボリューム・エアサンプラー(HV-1000R、柴田科学)に石英繊維フィルター(QR-100、柴田科学)を装着して、空気流量 700LPM で一週間単位でのサンプリングを行った。
得られた大気降下物試料はポリ製保管容器で保存しつつ、ロータリーエバポレーターや蒸発皿等を使用して全量を蒸発濃縮・固化し、Ge 半導体検出器によりγ線放出核種（放射性Cs 等）を測定した。次いで試料の一部を分取して発煙硝酸処理・放射化学分離により 90Sr を精製し、最終的に炭酸 Sr として固定、低バックグラウンド 2πガスフロー検出器でβ放射能を測定した。福島第一原発事故後の試料については、事故により放出された 89Sr の影響があるため、生成固定した炭酸 Sr 線源のβ放射能を繰り返し測定することで、90Y の放射平衡達成と 89Sr 放射能の減衰の様子とを確認しつつ評価し、必要な場合は計算で 89Sr の影響を除去した。フィルターサンプルの場合には 8 等分の短冊にし、金型と油圧プレスで成型することで試料形状を一定形状にし、Ge 半導体検出器によりγ線放出核種（放射性 Cs 等）を測定した。
134Cs/137Cs 放射能比は発生源の特定につながる重要なパラメーターであるが、前述の通り

134Cs（半減期 2.06 年）の濃度は事故直後から見て大きく低下し、より高精度での測定が必要となってきたため、その解析手法について開発検討を行った。着目した開発点は 134Cs の補正方法である。134Cs はカスケードピークを持つことからサム効果による測定影響が生じる。同標準線源でピーク効率を求める際に、一般にはサム補正は不必要とされているが、本研究のフィルター試料成型では試料内での放射能分布に偏りがあることが考えられ、その際に誤差を生じると考えられる。そこで、サム効果補正が可能であるソフトウェア(P-SCAN、
スカラベ・コーポレーション)を使用してサム補正を行い、補正の有無でどの程度定量に差が出るかを比較検討した。

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3.1) 134Cs/137Cs 放射能比の分析結果
P-SCAN ではγ線スペクトルの光電ピーク面積と、それらの計数効率・全効率の間に成り立つ高次・多元連立方程式を直接解くという方法を採用し、壊変後にカスケードγ線を放出する核種については観測されるピーク数をパターン化して数式化している。ピークの測定に
は気象研究所の 4 台(#6、#7、#8、#10)のゲルマニウム半導体検出器を使用した。標準試料の測定は通常通り行い、次に福島第一原発事故時の放射能強度が高い環境試料を測定、サム効
果補正計算を行い、そうでない場合と比較した。
大気中浮遊粒子試料の測定結果を Table1 に示す。137Cs ではサム効果は見られないため、
P-SCAN の使用有無による有意な違いは生じなかった。一方で 134Cs では P-SCAN で得られた結果の方が数%低い結果となった。134Cs/137Cs 比については PSCAN を使用した場合の方がばらつきは非常に小さかった。このことから、134Cs/137Cs 比を測定する場合には P-SCANを使用した方がよいと考えられるが、絶対値の正しさについては今後の検討が必要である。
Table1. Comparison when using P-SCAN and normal calculation in 4 detectors (result of measuring thesame sample).

二種類のγ線エネルギーの 605keV/662keV 比はサムピークの比を表す。サムピークの発生
確率は検出器と試料の距離に関係し、距離が離れるほどサムピークは生じにくくなる。相対距離 0cm から 5cm ごとに距離を離してサム効果の比率を求めた結果を Fig. 1 に示す。また、
サムピークの代表値として 1400keV/605keV 比を求めた結果を Fig. 2 に示す。
605keV/662keV 比より、15cm 以上、検出器と試料の距離を保てば光電ピーク比が安定することが分かった。また、1400keV/605keV 比から 15cm 以上離せばサムピークの影響が無視できることが分かった。

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3.2)

モンゴル土壌中の 137Cs、およびプルトニウム、ウラン同位体の測定

モンゴル土壌中には過去に行われた核実験などにより放出された人工放射性同位体が含まれており、アジアダストのトレーサーとしてしばしば使われている。Hirose et al. (2017)の結果、モンゴル土壌中のプルトニウム同位体・ウラン同位体比を始めて明らかにした。ここで得られたプルトニウム同位体比は、大気降下物として測定される核実験により放出された
放射性核種（グローバル・フォールアウト）の同位体比と一致しており、モンゴル土壌中の人工放射性核種がグローバル・フォールアウトに起因するものだということが分かった。
137Cs とも関係性が見られ、つくば土壌でも同じように関係性が見られることが分かったが、その傾きは大きく異なった。その原因は放射性核種それぞれの物理的・化学的な移動過程の違いによるものと考えられる。

- 10 -

3.3)

福島第一原発事故に起因する放射性セシウムとバイオエアロゾル

Igarashi et al. (2019) および Kinase et al. (2018) では 2011 年 7 月～2015 年 3 月までに福島県の浪江町・川俣町にて採取した大気中粒子試料の 134Cs,137Cs を測定し、その季節変化を明らかにした(Fig. 3)。その結果、つくばで観測される大気中放射性セシウムの季節変化とは明らかに異なることが分かり、5–6 月、および 8–9 月に極大となる二山ピークを持つことが分かった。これは、過去における研究で明らかにされてきた再飛散プロセスと、福島高濃度
汚染地域における主要な再飛散プロセスが異なっていることを示唆する結果である。

Fig. 3. Activity concentrations of 134,137Cs (Bq m− 3) measured at the FD, FC, GD, and GC sites in
the Fukushima prefecture in July 2011 to March 2015. These results show high concentrations in latespring, summer, and early autumn and low concentrations in winter and early spring at all of thepresented observation locations.

- 11 -

気象研究所内の走査型電子顕微鏡(SEM)及び EDX によるフィルター試料の観察結果より、
大気中放射性セシウム濃度の上昇がみられる試料には、胞子や菌類などのバイオエアロゾルが多く見つかることが分かった（Fig. 4）。

The normalized particle counts of the bioaerosols using SEM (in relative values, theerror bar shows the Poisson standard deviation) and atmospheric concentrations of 137Cs activity
sampled at the GD site during summer (July and August 2014) and winter (December 2014 and January2015). The counts of microorganisms include those of spores (see the text).
上記結果を踏まえ、Igarashi et al. (2019)では SEM-EDX 分析に加え、DAPI 染色による蛍光
顕微鏡による観察(Fig. 5)、メタゲノム解析によるバイオエアロゾル種の解析を行った(Fig.
6)。その結果、大気中放射性セシウム濃度とバイオエアゾルの大気中濃度には相関がみられ(Fig. 7)、バイオエアロゾルの 1/3～半数が担子菌及び子嚢菌だということが分かった。

- 12 -

Fig. 6. Results of the metagenomic analysis showing the relative abundance of identified fungal
taxa in the samples collected at the forested and bare soil sites in Namie and contained in rain watercollected at the site during summer 2015. Sampling dates are expressed as mm/dd. The compositional
differences among sampling dates may reflect seasonal and weather differences.

- 13 -

3.4)

福島第一原発事故に起因する河川及び地下水中の水溶性放射性セシウム

Iwagami et al. (2017)では、福島県山木屋地区で河川水および地下水中の水溶性放射性セシ
ウムの長期的観測を実施した。その結果、チェルノブイリ事故と比較して河川・地下水の汚染は低く、その汚染低下のプロセスには二つあることが分かった。まず一つ目のプロセスは、
大量の雨による表層土ごと洗い流す過程であり、事故後初期の段階で周辺の水中放射性セシ
ウムを急激に低下させた。もう一つのプロセスは、土壌中への放射性セシウムの浸透と固定化による、水中放射性セシウム濃度低下の鈍化であり、事故後しばらくたってからの影響として現在も観測されている。この研究により、河川水及び二地点での地下水について環境中での半減期が明らかとなり、それぞれ 0.69–1.5 y、0.46–0.58y、0.50–3.3y となった（Fig. 8）。
河川水中の 137Cs の減少傾向はキャノピーに見られる放射性セシウムの脱離傾向と似ていることが分かった。

- 14 -

- 15 -

4）まとめと今後に向けて本研究所における、より高精度な測定に向けた 134Cs 測定方法開発、モンゴル土壌における人工放射能の観測、福島第一原発事故後の大気中人工放射能観測の結果と他地域には見られない新しい再飛散プロセスの提案、水中に溶存する放射性セシウムの環境中における動態の傾向について報告した。前報にあるように、事故後 7 年が経過した現時点では、福島第一原発事故直後と比較すると、つくばでの降下量は数桁も低下し、甚大な汚染というべき水準ではない。90Sr については、事故以前の降下量とほぼ同じ水準に戻りつつある。しかし、本報で報告した 134Cs 測定技術の改善により、今後も高精度での人工放射能観測が期待できる。
また、137Cs については依然として 3～4 桁より大きな値となっている。水準低下にはまだ時間がかかると考えられ、継続した観測と動態傾向の分析が肝要である。
事故直後に見られた一次放出に替わり、大気へのこれらの核種の供給を続けている過程は、
いわゆる再飛散（二次放出）であり、現在はこの再飛散プロセスにより大気中・降下物中の人工放射能濃度が支配されていると考える。再飛散は周辺の環境に由来し、周辺の汚染度がつくばに比し数桁高い地域においては、大気中濃度・降下量もそれに比例し数桁高い。従って、再飛散プロセスの解明は必須な科学的課題である。
従来、汚染した地表面からの表土ダスト、汚染植生からの再浮遊、または汚染したゴミの燃焼、野焼き等による放射性物質の大気への揮散などが想定されてきた。しかし、本報で報告したように、福島原発事故により高濃度に汚染された地域では、通常の汚染地域とは異なり森林生態系からの放出、すなわちバイオエアロゾルの寄与がとても大きいと分かってきた。
特にキノコに由来する胞子や担子菌・子嚢菌が放射性セシウムの担体として重要と考えられるが、まだ定量的には解明できていない。今後は、バイオエアロゾル観測の定量化と、継続的な連続観測を実施することで定量的な理解を得ていくことが肝要である。そのためには、
森林内での各高度や樹間密度ごとのバイオエアロゾル放出量（フラックス）を定量的に観測する必要があり、顕微鏡と目視による粒子同定だけでなく、リアルタイムで測定可能な手法の開発が望まれる。
上記に加えて、こうした長期の観測においては分析やデータの品質管理が必ず問題となる。
このような点についても常に留意しながら、観測の継続を図っていくことが肝要である。

- 16 -

参考文献
Hirose, K., Kikawada, Y., Igarashi, Y., Fujiwara, H., Jugder, D., Matsumoto, Y., Oi, T., and Nomura, M.
Plutonium, 137Cs and uranium isotopes in Mongolian surface soils, Journal of Environmental
Radioactivity, 166, 97̶̶-103, 2017. http://dx.doi.org/10.1016/j.jenvrad.2016.01.007
Igarashi, Y., Kita., K., Maki, T., Kinase, T., Hyashi, N., Hotaka, K., Adachi, K., Kajino, M., Ishizuka, M.,
Sekiyama, T., T., Zaizen, Y., Takenaka, C., Ninomiya, K., Okochi, H., and Sorimachi, A. Fungal
spore involvement in the resuspension of radiocaesium in summer, Scientific Reports, 9:1-10, 2019.
2019. DOI:10.1038/s41598-018-37698-x
Iwagami, S., Tsujimura, M., Onda, Y., Nishino, M., Konuma, R., Abe, Y., Hada, M., Pun, I., Sakaguchi, ,A., Kondo, H., Yamamoto, M., Miyata, Y., and Igarashi, Y. Temporal changes in dissolved 137Csconcentrations in groundwater and stream water in Fukushima after the Fukushima Dai-ichi Nuclear
Power Plant accident. Journal of Environmental Radioactivity, 166, 458-465, 2017.
http://dx.doi.org/10.1016/j.jenvrad.2015.03.025
木村徹,五十嵐康人,財前裕二,「ガンマ線スペクトル解析ソフト P-SCAN を用いた高精度な
134Cs/137Cs 放射能比の導出」Proceedings of the 18th Workshop Environmental Radioactivity.
KEK Proceedings 2017
Kinase, T., K, Kita, Igarashi, Y., Adachi, K., Ninomiya, K., Shinohara, A., Okochi, H., Ogata, H., Ishizuka,M., Toyoda, S., YamadaN, K., Yoshida, N., Zaizen, Y., Mikami, M., Demizu, H., and Onda, Y. The
seasonal variations of atmospheric 134,137Cs activity and possible host particles for their resuspension
in the contaminated areas of Tsushima and Yamakiya, Fukushima, Japan, Progress in Earth andPlanetary Science 5:12, 2018. https://doi.org/10.1186/s40645-018-0171-z- 17 -

- 18 -

2. 大気中の放射性物質の物理・化学性状
1）はじめに
大気中で観測される放射性物質は、ガスや粒子（エアロゾル）として大気中に漂い、やがて雨による除去や重力沈降によって地上や海水面に降下する。このような大気中での放射性物質の振る舞いはその放射能や核種に依存せず、物理的な粒径や密度、形態、また水溶性な
どの化学性質に依存する。そのため、放射性物質の大気拡散や環境動態、また発生メカニズ
ムを理解するには、放射性物質がどのような物理化学性状を持ち、またどのようなエアロゾ
ルを担体としているかを解明する必要がある。本章では、その放射性物質の物理化学性状の理解を目指した例として、東日本大震災で発生した福島第一原子力発電所事故によって直接放出された放射性物質に関する研究成果を報告する。
2011 年 3 月に発生した事故によって様々な放射性物質が大気・海洋に放出されたが、放出された放射性物質の重量は非常に微量であり、放射能測定以外の分析手法で環境試料から検出することは一般には困難とされてきた。しかしながら、電子顕微鏡や放射光分析などの微小領域化学分析を用いることで、放射性物質の物理化学特性を明らかにすることが可能であることが気象研究所の研究によって明らかになってきた。本研究においては、それらの手法を用いて 2011 年 3 月 14-15 日につくば市で採取されたフィルター試料中の放射性粒子の特定を行った先駆的な研究（Adachi et al., 2013）を発端として、多くの研究事例が報告されている。ここでは、それらの研究例のうち、前回の報告書から後に発表された小野ら（2017）
や Satou (2016,2018)の論文、またこれらの研究例をまとめた総説論文（足立、2017）を紹介する。
2）主な研究成果
東日本大震災で発生した福島第一原子力発電所事故によって直接放出された放射性物質
は、本放射能調査研究費研究でモニタリングを継続していた気象研究所におけるフィルター
試料から発見された（Adachi et al., 2013）。この研究では、走査型電子顕微鏡（SEM-EDS）
を用いて 2011 年 3 月 14－15 日にかけてつくばに飛来した放射性粒子を分析し、その物理
化学性状を初めて明らかにした。その結果、セシウム、鉄、亜鉛や多くの核燃料由来と考えられる元素を含み、水に溶けにくい２マイクロメートル前後の球状ガラス様粒子であることが明らかとなった。これらの詳細な物理特性は、事故時における粒子の拡散モデルの精度向上に重要であることを、モデル比較実験で明らかにした。その後、Abe et al. (2014）やYamaguchi et al. (2016)によって、粒子のさらに詳細な微量成分や内部構造が明らかとなった。
これらの先駆的な研究を受けて、本研究期間ではエアロゾル試料やその大気中微粒子が沈着した結果生じた表層土壌汚染地域からの試料を採取して、そこにふくまれる放射性粒子を単離して、電子顕微鏡や放射光分析を用いてその個別粒子の物理化学性状を明らかにし、また少なくとも 2 種類の異なった物理化学性質を持つ放射性粒子が存在することを明らかにした。
Satou et al (2016)では、事故があった原子力発電所から 20ｋｍほど離れた地点の土壌から、６㎛ほどの放射性粒子を新たに開発し手法で単離し、それらがエアロゾル試料で見つかった粒子と似た、ケイ素、鉄、亜鉛、セシウムなどを含む化学組成をしていることが分かった。
これらの粒子は大気中のエアロゾル試料に見つかった粒子よりも大きく、また放射能強度も強い特徴を持ち、近隣地域にはより大きな粒子が重力沈降で沈着した可能性を示唆した。
小野ら（2017）の論文では、さらに土壌試料中に比放射能（単位質量あたりの放射能）はエアロゾルから分離された粒子より低いものの、大きさが 100μm ほどの放射性粒子を発見
し、放射光分析でその詳細な化学成分を明らかにした。これらの大きな粒子は、球状粒子が複数溶融したような形態をしており、ケイ素を主成分とする特徴を示し、また 20 元素以上の様々な微量成分が検出された。これらの粒子は、Adachi et al (2013)で見つかった粒子と異なる物理化学性質を持っており、異なった放出プロセスがあったことを示唆する。Satou et- 19 -

al (2018)では、それらの粒子のセシウム 134、137 の比が異なることを見つけて、その同位体比の違った粒子のうち比放射能が高く、数 μ 程度で比較的球状に近い粒子を A タイプとし、比放射能は低いものの大きな粒子を B タイプとして分類した。
足立（2017）の論文ではこれらの成果をまとめて、今後の研究の発展に資するデータのまとめを行った。その一例の表を下記に示す。

Table 1 Summary of insoluble radioactive particles from nine studies
References

Sampling
place

Sample type

Adachi et al(2013)
Abe et al(2014)
Yamaguchi
et al. (2016)
Satou et al.,(2016)
Kogure et al.
(2016)
Furuki et al(2017)
Yamaguchi
et al. (2017)
Ono et al.
(2017)
Higaki et al.
(2017)

IbarakiAerosol filter

IbarakiAerosol filter

Fukushima
Fukushima
Fukushima
Fukushima
Fukushima
Fukushima
Fukushima

Diameter(μm) ofspherical
particles

Radioactivity (Bq) of137
Cs

Particlenumber

3.3-0.73

1.4-2.8

1.1-1.5

3

Non-woven fabriccloth & Cedar
Soil

2.2-2.5

3.1-5.0

2

5-67

4

Non-woven fabriccloth & Wheat
Soil & Gravel

0.8-2.7

1.9-3.3

3

3.4

0.9-11.33

1.4-3.6

0.5-1.94

2.0-2.6

3.0-4.0

2.0-3.3

Diameter(μm) of nonsphericalparticles

4.5-6.4

Aerosol filter &Plant tissue
Soil

2.6-3.6

50-25033-109

7

Mask

2.1

6.6-121-70

4足立（2017）エアロゾル研究 32 巻 4 号より。
3）まとめ
本放射能調査費によって行われた研究成果により、福島第一原子力発電所事故において放出された放射性物質の大気中の物理化学特性の一部が明らかとなった。これらの研究をもとに、今後事故から放出された放射性物質の発生源メカニズムの解明や環境動態の予測への貢献が期待される。

- 20 -

参考文献
Y. Satou, K. Sueki, K. Sasa, H. Yoshikawa, S. Nakama, H. Minowa, Y. Abe, I. Nakai, T. Ono, K. Adachi
and Y. Igarashi, Analysis of two forms of radioactive particles emitted during the early stages of the
Fukushima Dai-ichi Nuclear Power Station accident, Geochemical Journal, Vol. 52, 2018、
足立光司、原子力発電所事故で放出された不溶性放射性粒子の形態と組成、エアロゾル研究 324, 255-260, 2017
小野貴大, 飯澤勇信, 阿部善也, 中井 泉, 寺田靖子, 佐藤志彦, 末木啓介, 足立光司, 五十嵐康人、福島第一原子力発電所事故により 1 号機から放出された放射性粒子の 放射光マイク
ロビーム X 線分析を用いる化学性状の解明、分析化学 66（4） 2017
Y. Satou, K. Sueki, K. Sasa, K. Adachi, Y. Igarashi, First successful isolation of radioactive particles fromsoil near the Fukushima Daiichi Nuclear Power Plant, Anthropocene, Volume 14, Pages 71-76, 2016
Abe, Y., Iizawa, Y., Terada, Y., Adachi, K., Igarashi, Y., and Nakai, I., Detection of uranium andchemical state analysis of individual radioactive microparticles emitted from the Fukushima nuclear
accident using multiple synchrotron radiation X-ray analyses, Analytical Chemistry, 86, 8521-8525,
2014.
Adachi, K., Kajino, M., Zaizen, Y., and Igarashi, Y., Emission of spherical cesium-bearing particles froman early stage of the Fukushima nuclear accident, Scientific Reports, 3, 2554, 2013.
Yamaguchi, N., Mitome, M., Kotone, A. H., Asano, M., Adachi, K., and Kogure, T., Internal structure ofcesium-bearing radioactive microparticles released from Fukushima nuclear power plant. Scientific
Reports, 6, 20548, 2016.

- 21 -

- 22 -

3.放射性物質の再浮遊、輸送、沈着モデルによる評価
1） 福島原子力事故後の放射性セシウム大気濃度の長期評価：裸地と森林生態系からの再飛散
福島原子力事故による汚染土壌や森林からの 137Cs 再飛散の長期的影響を、汚染地域（福島県浪江町）の土壌からの再飛散実験結果，および汚染地域内（浪江）・外（茨城県つくば市）の大気濃度測定，および数値シミュレーションを併用して実施した（Kajino et al., 2016）。
評価期間は 2013 年の 1 年間である。推定された再飛散率は約 10-6/day であり、福島県におけるガンマ線線量率の減少率（10-4-10-3/day）よりも明らかに低かった．すなわち，再飛散は
陸上の放射能の減少にはほとんど寄与していなかったことが分かる。再飛散メカニズムは、
特に森林生態系に関しては，未知のままである。今後は，メカニズム解明の実験研究に基づいた影響評価を行っていく必要がある。
Fig. 1 に計算された 2013 年の 1 年間の再飛散量と、再沈着量を示す。領域・時間積算再飛散量は、初期の沈着量 2.68 PBq（NRA, 2012）の 0.048%に相当した。

Fig. 1. Annual total amounts of (a) the re-suspension and (b) the re-deposition of 137Cs. The total
activities are presented as numbers.
2） 福島原発事故時の地上プリューム移流予測に及ぼす地上風データ同化の影響
2011 年 3 月の福島原発事故で排出された放射性セシウムを大気トレーサーとして用い、対流圏下層のプルーム移流の予測可能性と AMeDAS 地上風データ同化の影響を調べた（Sekiyama et al., 2017）。気象庁非静力学モデルと局所アンサンブルカルマンフィルタ（JMANHM-LETKF）データ同化システムを用いて、実験を行った。40 の観測地点における、
放射性セシウムのプルーム到着時間について調べたところ、標準実験（標準的な観測データ
のみ同化）のプルーム到着時間の平均誤差は 82.0 分であった。それに対して、AMeDAS 実験（標準的な観測データと AMeDAS 地上風観測の両方を同化する）の平均誤差は 72.8 分であり、これは標準実験に比べて 9.2 分（11％）高精度であった。本研究は、プルーム予測モデル
が緊急時対応のために十分な精度を有し、予測結果が地上風データ同化によって著しく改善
され得ることを示すものである。
Fig. 2 に、埼玉県久喜市において観測された大気濃度（SPM テープ：浮遊粒子状物質観測用フィルタ上の 137Cs 濃度、Tsuruta et al., 2014; Oura et al., 2015）と、データ同化前後の計算結果の比較を示す。データ同化を行っていないものを STD Exp.（実線）、地上風データ同化を行ったものを AMeDAS Exp.（破線）としている。プリュームの到達時間に明らかな改善が見られている。

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Fig. 2 An example of time series of the radioactive cesium (137Cs) concentrations derived from the SPMtape sampling observations and two model experiments (STD and AMeDAS) from 12:00 UTC on March
14 to 12:00 UTC on March 15, 2011. The advection of the plumes was examined using the data from Kuki
City, Saitama Prefecture.
3） 福島原子力事故後の大気モデリング研究から学んだ教訓：アンサンブルシミュレーション、
データ同化、素過程モデリング、逆解析
福島第一原子力発電所事故に伴う放射性セシウムの大気拡散と沈着に関するモデリング研究について、フランス放射線防護・原子力安全研究所と気象庁気象研究所の共同研究に着目しながら、レビューを行った（Kajino et al., 2018）。これまでのモデル研究により、湿性沈着過程が重要なプロセスであることが判明しているが、その影響の大きさは不明のままである。また、モデルによって計算結果が大きく異なることが判明しているが、その原因は定かではない。本共同研究では、大気素過程、特に湿性沈着過程の理解のために、アンサンブ
ルシミュレーション、データ同化、素過程モデリング、逆解析などのさまざまな数値技術を使用して、シミュレーションの各要素の不確実性を定量化することを目指した。本論文では、これらの共同研究の成果を紹介するとともに、今後の大気モデル研究の将来の方向性についても議論した。

4） 福島第一原子力発電所から放出された 137Cs プルームの質量フラックス解析
Sekiyama and Iwasaki (2018) では、137Cs プルームの質量フラックス解析を行い、2011 年 3月に福島第一原子力発電所から放出された 137Cs のプルームについて質量フラックスの流れを定量的に示すことができた。この 137Cs のプルームは 3km の水平分解能を持つオイラー型の大気移流拡散モデルによって計算した。鉛直積分した質量フラックスは 137Cs 濃度の地上観測値に基づく移流経路推定とほぼ一致していたが、地上とプルーム層との間に風速シアーがある場合には不一致が生じた。この不一致は、地表面の水平質量フラックスと鉛直積分した質量フラックスの二つを組み合わせることによって理解できる（Fig. 3 参照）。質量密度と風速の積である質量フラックスの概念は気象力学においては一般的であるが、大気汚染の研究においてはほとんど使用されてこなかった。しかし質量フラックス解析は、数値モデルが計算する４次元（x,y,z,t）の出力値を２次元（x,y）プロットに縮小することによって、大気微量成分の主要な流れの位置、方向、および沈着の振る舞いを明瞭にすることができるた
め定量的な可視化ツールとして最適である。本研究では、大気中の放射性物質（この場合は137Cs）の移流拡散を調べることによって、大気汚染の研究においても質量フラックス解析が非常に有用な手法であることを示すことができた。

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Fig. 3 Time-integrals from 15:00 to 21:00 JST on 15 March 2011 of (a) the column-integrated mass fluxof plume P3 and (b) the horizontal mass flux of plume P3 at the ground surface. Color shading indicatesthe magnitude of the mass flux. Small arrowheads indicate the direction of the mass flux. A large grayarrow illustrates the depiction in Nakajima et al. (2017) for plume P3 at 18:00 JST on March 15, shown in
Fig. 2. (c) Latitude-altitude cross-sections at 140.5E for 137Cs concentrations (gray shading), meridional
winds (arrows), and potential temperature (h; contours) at 17:00 JST on 15 March 2011. The cross-section
is located at the green line in panel (b).

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2018
Hirose, K., Kikawada, Y., Igarashi, Y., Fujiwara, H., Jugder, D., Matsumoto, Y., Oi, T., andNomura, M. Plutonium, 137Cs and uranium isotopes in Mongolian surface soils, Journal ofEnvironmental
Radioactivity,
166,
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http://dx.doi.org/10.1016/j.jenvrad.2016.01.007
Igarashi, Y., Kita., K., Maki, T., Kinase, T., Hyashi, N., Hotaka, K., Adachi, K., Kajino, M.,
Ishizuka, M., Sekiyama, T., T., Zaizen, Y., Takenaka, C., Ninomiya, K., Okochi, H., andSorimachi, A. Fungal spore involvement in the resuspension of radiocaesium in summer,
Scientific Reports, 9:1-10, 2019. 2019. DOI:10.1038/s41598-018-37698-x
Iwagami, S., Tsujimura, M., Onda, Y., Nishino, M., Konuma, R., Abe, Y., Hada, M., Pun, I.,
Sakaguchi, , A., Kondo, H., Yamamoto, M., Miyata, Y., and Igarashi, Y. Temporal changes indissolved 137Cs concentrations in groundwater and stream water in Fukushima after the
Fukushima Dai-ichi Nuclear Power Plant accident. Journal of Environmental Radioactivity, 166,458-465, 2017. http://dx.doi.org/10.1016/j.jenvrad.2015.03.025
Kajino, M.., Sekiyama, T. T., Mathieu, A,. Korsakissok, I., Périllat, R., Quélo, D., Quérel, A.,
Saunier, O., Adachi, K., Girard, S., Maki, T., Yumimoto, K., Didier, D., Masson, O., andIgarashi, Y., Lessons learned from atmospheric modeling studies after the Fukushima nuclear
accident: Ensemble simulations, data assimilation, elemental process modeling, and inverse
modeling, Geochemical Journal, Vol. 52, 2018.
Kinase, T., Kita, K., Igarashi, I., Adachi, K., Ninomiya, K., Shinohara, A., Okochi, H., Ogata,
H., Ishizuka, M., Toyoda, S., Yamada, K., Yoshida, N., Zaizen, Y., Mikami, M,. Demizu, H.,
and Onda, Y., The seasonal variations of atmospheric 134,137Cs activity and possible hostparticles for their resuspension in the contaminated areas of Tsushima and Yamakiya,
Fukushima, Japan, Progress in Earth and Planetary Science, 5:12, 2018.
Satou, Y., Sueki, K., Sasa, K., Yoshikawa, H., Nakama, S., Minowa, H., Abe, Y., Nakai, I., Ono,
T., Adachi, K., and Igarashi, Y., Analysis of two forms of radioactive particles emitted during
the early stages of the Fukushima Dai-ichi Nuclear Power Station accident, Geochemical Journal,
Vol. 52, 137-143, 2018. doi:10.2343/geochemj.2.0514.
Sekiyama, T. T. and T. Iwasaki, Mass flux analysis of 137Cs plumes emitted from the Fukushima
Daiichi Nuclear Power Plant, Tellus B, 70(1), 1–11, doi:10.1080/16000889.2018.1507390, 2018.
2017
Iwagami, S., Tsujimura, M., Onda, Y., Nishino, M., Konuma, R., Abe, Y., Hada, M., Pun, I.,
Sakaguchi, A., Kondo, H., Yamamoto, M., Miyata, Y., and Igarashi, Y., Temporal changes indissolved 137Cs concentrations in groundwater and stream water in Fukushima after the
Fukushima Dai-ichi Nuclear Power Plant accident, Journal of Environmental Radioactivity,166,
458-465, 2017.
Hirose, K., Kikawada, Y., Igarashi, Y., Fujiwara, H., Jugder, D., Matsumoto, Y., Oi, T., Nomura,M., Plutonium, 137Cs and uranium isotopes in Mongolian surface soils, Journal of Environmental
Radioactivity, 166, 97-103, 2017.
Sekiyama, T. T., Kajino, M., and Kunii, K., The Impact of Surface Wind Data Assimilation on
the Predictability of Near-Surface Plume Advection in the Case of the Fukushima Nuclear
Accident, Journal of the Meteorological Society of Japan, Vol. 95, 447-454, 2017. doi:
10.2151/jmsj.2017-025.
2016
Kajino, M., Ishizuka, M., Igarashi, Y., Kita, K., Yoshikawa, C., and Inatsu, M., Long-term
assessment of airborne radiocerium after the Fukushima nuclear accident: re-suspension frombare soil and forest ecosystems, Atmos. Chem. Phys, 16, 13149-13172, 2016.
Satou,Y., Sueki, K., Sasa, K., Adachi, K., and Igarashi, Y., First successful isolation ofradioactive particles from soil near the Fukushima Daiichi Nuclear Power Plant, Anthropocene,
14, 71-76, 2016.
Yamaguchi, N., Mitome, M., Kotone, A-H., Asano, M., Adachi, K., and Kogure, T., Internal
structure of cesium-bearing radioactive microparticles released from Fukushima nuclear power
plant. Scientific Reports. 6, 20548, 2016.

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2.

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2015
Igarashi, Y., Kajino, M., Zaizen, Y., Adachi, K., and Mikami, M., Atmospheric radioactivity
over Tsukuba, Japan: a summary of three years of observations after the FDNPP accident,
Progress in Earth and Planetary Science 2:44, 2015. doi: 10.1186/s40645-015-0066-1.
Kajino, M., G-3. NHM-Chem: Sensitivity of Cs deposition to the size and hygroscopicity of Csbearing aerosols, Technical Reports of the Meteorological Research Institute, 76, 132–142,
2015.
Sekiyama, T. T., Kunii, M., Kajino, M., and Shinbori, T., Horizontal Resolution Dependenceof Atmospheric Simulations of the Fukushima Nuclear Accident Using 15-km, 3-km, and 500m Grid Models, J. Meteorol. Soc. Jpn., 93B, 49–64, 2015.

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2014
Abe, Y., Iizawa, Y., Terada, Y., Adachi, K., Igarashi, Y., and Nakai, I., Detection of uranium
and chemical state analysis of individual radioactive microparticles emitted from the Fukushima
nuclear accident using multiple synchrotron radiation X-ray analyses, Analytical Chemistry, 86,
8521-8525, 2014.
2013
Aoyama, M., Uematsu, M., Tsumune, D., and Hamajima, Y., Surface pathway of radioactive
plume of TEPCO Fukushima NPP1 released 134Cs and 137Cs, Biogeosciences, 10, 30673078, 2013. doi: 10.5194/bg-10-3067-2013.
Tsumune, D., Tsubono, T., Aoyama, M., Uematsu, M., Misumi, K., Maeda, Y., Yoshida, Y.,
and Hayami, H., One-year, regional-scale simulation of 137Cs radioactivity in the ocean
following the Fukushima Daiichi Nuclear Power Plant accident, Biogeosciences, 10, 5601-5617,
2013.
Povinec, P.P., Aoyama, M., Biddulph, D., Breier, R., Buesseler, K., Chang, C., Golser, R., Hou,X., Ješkovský, M., and Jull, A., Cesium, iodine and tritium in NW Pacific waters-a comparisonof the Fukushima impact with global fallout. Biogeosciences, 10, 5481-5496. 2013.
Kumamoto, Y., Murata, A., Kawano, T., and Aoyama, M., Fukushima-derived radiocesium inthe northwestern Pacific Ocean in February 2012, Applied Radiation and Isotopes, 81, 335-339,
2013. doi: 10.1016/j.apradiso.2013.03.085.
Toyama, C., Muramatsu, Y., Igarashi, Y., and Aoyama, M., Atmospheric fallout of 129I in Japan
before the Fukushima accident: Regional and global contributions, 1963-2005, Environmental
Science & Technology, 47(15), 8383-8390, 2013. doi: 10.1021/es401596z.
Adachi, K., Kajino, M., Zaizen, Y., and Igarashi, Y., Emission of spherical cesium-bearing
particles from an early stage of the Fukushima nuclear accident. Scientific Reports, 3, Article
number 2554, 2013. doi:10.1038/srep02554.
2012
Toyama, C., Muramatsu, Y., Uchida, Y., Igarashi, Y., Aoyama, M., and Matsuzaki, H.,
Variations of 129I in the atmospheric fallout of Tokyo, Japan 1963-2003, Journal ofEnvironmental Radioactivity, 113, 116-122, 2012. doi: 10.1016/j.jenvrad.2012.04.014.
Inomata, Y., Aoyama, M., Tsumune, D., Motoi, T., and Nakano, H., Optimum interpolation
analysis of basin-scale 137Cs transport in surface seawater in the North Pacific Ocean, Journalof Environmental Monitoring, 14, 3146-3155, 2012. doi: 10.1039/C2EM30707C.
Aoyama, M., Tsumune, D., Uematsu, M., Kondo, F., and Hamajima, Y., Temporal variation of
134Cs and 137Cs activities in surface water at stations along the coastline near the Fukushima Daiichi Nuclear Power Plant accident site, Japan. Geochem. J, 46, 321-325, 2012.
Aoyama, M., Tsumune, D., and Hamajima, Y., Distribution of 137Cs and 134Cs in the North
Pacific Ocean: impacts of the TEPCO Fukushima-Daiichi NPP accident, Journal ofRadioanalytical and Nuclear Chemistry, 1-5, 2012. doi: 10.1007/s10967-012-2033-2.
Honda, M. C., Aono, T., Aoyama, M., Hamajima, Y., Kawakami, H., Kitamura, M., Masumoto,Y., Miyazawa, Y., Takigawa, M., and Saino, T., Dispersion of artificial caesium-134 and-137
in the western North Pacific one month after the Fukushima accident, Geochem. J, 46, e1-e9,
2012.
Povinec, P. P., Hirose, K., and Aoyama, M., Radiostrontium in the Western North Pacific:
characteristics, behavior, and the Fukushima impact, Environmental Science & Technology, 46,10356-10363, 2012. doi: 10.1021/es301997c.
Tsumune, D., Tsubono, T., Aoyama, M., and Hirose, K., Distribution of oceanic 137Cs from the
Fukushima Dai-ichi Nuclear Power Plant simulated numerically by a regional ocean model,
Journal of Environmental Radioactivity, 111, 100-108, 2012.

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2011
Aoyama, M., Fukasawa, M., Hirose, K., Hamajima, Y., Kawano, T., Povinec, P.P., and SanchezCabeza, J. A., Cross Equator transport of 137Cs from North Pacific Ocean to South Pacific Ocean,
BEAGLE2003
cruises,
Progressin
Oceanography,
89,
7-16,
2011.
doi:10.1016/j.pocean.2010.12.003.
Buesseler, K., Aoyama, M., and Fukasawa, M., Impacts of the Fukushima nuclear power plants
on marine radioactivity. Environmental Science & Technology, 45(23), 9931-9935, 2011. doi:
10.1021/es202816c.
Povinec, P. P., Aoyama, M., Fukasawa, M., Hirose, K., Komura, K., Sanchez-Cabeza, J.A.,
Gastaud, J., Jeskovsky, M., Levy, I., and Sykora, I., 137Cs water profiles in the South Indian
Ocean - an evidence for accumulation of pollutants in the subtropical gyre. Progress in
Oceanography, 89 , 17-30, 2011. doi:10.1016/j.pocean.2010.12.004.
Gastaud, J., Povinec, P.P., Aoyama, M., Hirose, K., Sanchez-Cabeza, J.A., Levy, I., Roos, P.,
Eriksson, M., Bosc, E. and Rezzoug, S., Transport and scavenging of Pu in surface waters of theSouthern Hemisphere Oceans, Progress in Oceanography, 89, 92-100, 2011.
doi:10.1016/j.pocean.2010.12.009.
Hirose, K., Kim, C.S., Yim, S.A., Aoyama, M., Fukasawa, M., Komura, K., Povinec, P.P., andSanchez-Cabeza, J.A., Vertical profiles of plutonium in the central South Pacific, Progress in
Oceanography, 89, 101-107, 2011. doi:10.1016/j.pocean.2010.12.010.
Tsumune, D., Aoyama, M., Hirose, K., Bryan, F., Lindsay, K. and Danabasoglu, G., Transportof 137Cs to the Southern Hemisphere in an Ocean General Circulation Model, Progress in
Oceanography, 89, 38-48, 2011. doi:10.1016/j.pocean.2010.12.006.
Sanchez-Cabeza, J.A., Levy, I., Gastaud, J., Eriksson, M., Osvath, I., Aoyama, M., Povinec, P.P.
and Komura, K., Transport of North Pacific 137Cs labeled waters to the south-eastern Atlantic
Ocean. Progress in Oceanography, 89, 31-37, 2011. doi:10.1016/j.pocean.2010.12.005.
Levy, I., Povinec, P.P., Aoyama, M., Hirose, K., Sanchez-Cabeza, J.A., Comanducci, J-F.,
Gastaud, J., Eriksson, M., Hamajima, Y., Kim, C.S., Komura, K., Osvath, I., Roos, P. and Yim,
S.A., Marine anthropogenic radiotracers in the Southern Hemisphere: new sampling andanalyticalstrategies,
Progressin
Oceanography,
89,
120-133,
2011.
doi:10.1016/j.pocean.2010.12.012.
2010
Inomata,Y., Igarashi, Y., Yoshioka, K., Tanaka, T. Y., and M, Chiba., Temporal variation of
222Rn at the summit of Mt. Fuji associated with the Asian continental outflow, Atmospheric
Environment, 44,3856-3865, 2010.
Nakano, H., Motoi, T., Hirose, K., and Aoyama, M., Analysis of 137Cs concentration in the
Pacific using a Lagrangian approach. Journal of Geophysical Research, 115, C06015, 2010.
doi:10.1029/2009JC005640.
Hirose, K., Igarashi, Y., Aoyama, M., and Inomata, Y., Depositional behaviors of plutonium
and thorium isotopes at Tsukuba and Mt. Haruna in Japan indicate the sources of atmosphericdust. Journal of Environmental Radioactivity, 101,106-112, 2010.
Zhang, Y., Zeng, J., Yamada, M., Wu, F., Igarashi, Y., and Hirose, K., Characterization of
Pu concentration and its isotopic composition in a reference fallout material. Science of the
Total Environment, 408(5), 1139-1144, 2010.
2009
Aoyama, M., Hamajima, Y., Fukasawa, M., Kawano, T., and Watanabe, S., Ultra low level deep
water 137Cs activity in the South Pacific Ocean, Journal of Radioanalytical and Nuclear
Chemistry, 2009. doi:10.1007/s10967-009-0253-x.
Hirose, K., Aoyama, M., and Povinec, P.P., 239, 240Pu /137Cs ratios in the water column of theNorth Pacific: a proxy of biogeochemical processes, Journal of Environmental Radioactivity,
100, 258-262, 2009.
Igarashi, Y., Inomata, Y., Aoyama, M., Hirose, K., Takahashi, H., Shinoda, Y., Sugimoto, N.,
Shimizu, A., and Chiba, M., Possible change in Asian dust source suggested by atmosphericanthropogenic radionuclides during the 2000s, Atmospheric Environment, 43, 2971-2980, 2009.
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Inomata, Y., Aoyama, M., and Hirose, K., Analysis of 50-y record of surface 137Csconcentrations in the global ocean using the HAM-global database, Journal of Environmental
Monitoring, 11(1), 116-125, 2009. doi: 10.1039/b811421h.
Hirose, K., Igarashi, Y., Aoyama, M., and Inomata, Y., Depositional behaviors of plutonium andthorium isotopes at Tsukuba and Mt.Haruna in Japan indicate the sources of atmospheric dust,
106-112,2009.
JournalofEnvironmental
Radioactivity,
101(2),
doi:10.1016/j.jenvrad.2009.09.003.
Igarashi Y., Kikawada, Y., Oda, K., Yamauchi, R., Nomura, M., Honda, T., Oi, T., and Hirose,K., Anomalous Uranium Isotope Ratio in Atmospheric Deposits in Japan, Journal of Nuclear
Science and Technology, 46(12), 1094-1098. 2009.

2008
Aoyama, M., Hirose, K., Nemoto, K., Takatsuki, Y., and Tsumune, D., Water masses labeledwith global fallout 137Cs formed by subduction in the North Pacific, Geophysical Research
Letters, 35, L01604, 2008. doi:10.1029/2007GL031964.
Hirose, K., Igarashi, Y., and Aoyama, M., Analysis of the 50-year records of the atmosphericdeposition of long-loved radionuclides in Japan, Applied Radiation and Isotopes, 66, 1675-1678,
2008.
Hirose, K., Aoyama, M., Igarashi, Y., and Komura, K., Improvement of 137Cs analysis in small
volume seawater samples using the Ogoya underground facility, Journal of Radioanalytical andNuclear Chemistry, 276,3,795-798, 2008.
Aoyama, M., Hirose, K., Nemoto, K., Takatsuki, Y., and Tsumune, D., Water masses labeledwith global fallout 137Cs formed by subduction in the North Pacific, Geophysical Research
Letters, 35, L01604, 2008. doi:10.1029/2007GL031964.
2007
Hirose, K., Aoyama, M., and Kim, C.S., Plutonium in Seawater of the Pacific Ocean, Journal ofRadioanalytical and Nuclear Chemistry, 274 No. 3, 635-638, 2007.
Hirose, K., Igarashi, Y., and Aoyama, M., 50 years records of atmospheric deposition of longlived anthropogenic radionuclides in Japan, International Conference on Environmental
Radioactivity: From Measurements and Assessments to Regulation, IAEA-CN-145, 95-96,
2007.
Hirose, K., Igarashi, Y., and Aoyama, M., Recent trends of plutonium fallout observed in Japan:
Comparison with natural lithogenic radionuclides, thorium isotopes, Journal of Radioanalyticaland Nuclear Chemistry, 273 No.1, 115-118, 2007.
Hirose, K., Aoyama, M., Fukasawa, M., Kim, C. S., Komura, K., Povinec, P. P., and SanchezCabeza, J. A., Plutonium and 137Cs in surface water of the South Pacific Ocean, Science of the
Total Environment, 381, 243-255, 2007.
2006
Yamamoto, M., Sakaguchi, A., Sasaki, K., Hirose, K., Igarashi, Y., and Kim, C.K., Seasonaland spatial variation of atmospheric 210Pb and 7Be deposition: features of the Japan Sea side ofJapan, Journal of Environmental Radioactivity, 86, 110-131, 2006.
Hirose, K., Aoyama, M., Kim, C.S., Kim, C.K., and Povinec, P.P., Plutonium isotopes inseawater of the North Pacific: effects of close-in fallout, INTERNATIONAL CONFERENCE
ON ISOTOPES AND ENVIRONMENTAL STUDIES. Eds. P. Povinec, J.A. Sanchez-Cabeza.
Vol. 8 of Radioactivity in the Environment, Elsevier, London, 2006.
Aoyama, M., Hirose, K., and Igarashi, Y., Re-construction and updating our understanding on
the global weapons tests 137Cs fallout, Journal of Environmental Monitoring, 8, 431-438, 2006.
Aoyama, M., Fukasawa, M., Hirose, K., Mantoura, R.F.C., Povinec, P.P., Kim, C.S., andKomura, K., Southern Hemisphere Ocean Tracer Study (SHOTS): An overview and preliminary
results, INTERNATIONAL CONFERENCE ON ISOTOPES AND ENVIRONMENTALSTUDIES. Eds. P. Povinec, J.A. Sanchez-Cabeza, Vol. 8 of Radioactivity in the Environment,
Elsevier, London, 2006.

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7.

Inoue, H.Y., Matsueda, H., Igarashi, Y., Sawa, Y., Wada, A., Nemoto, K., Sartorius, H., andSchlosser, C., Seasonal and long-term variations in atmospheric CO2 and 85Kr in Tsukuba,
Central Japan, Journal of the Meteorological Society of Japan, 84/6, 959-968, 2006.

2005
Hirose, K., Aoyama, M., Igarashi, Y., and Komura, K., Extremely low backgroundmeasurements of 137Cs in seawater samples using an underground facility (Ogoya), Journal ofRadioanalytical and Nuclear Chemistry, 263, 349-353, 2005.
Povinec, P.P., Aarkrog, A., Buesseler, K.O., Delfanti, R., Hirose, K., Hong, G.H., Ito, T.,
Livingston, H.D., Nies, H., Noshkin, V.E., Shima, S., and Togawa, O., 90Sr, 137Cs and 239,
240
Pu concentration surface water time series in the Pacific and Indian Oceans- WOMARS
results, Journal of Environmental Radioactivity, 81, 63-87, 2005.
Igarashi, Y., Aoyama, M., Hirose, K., Povinec, P.P., and Yabuki, S., What anthropogenic
radionuclides (90Sr and 137Cs) in atmospheric deposition, surface soils and Aeoliandusts suggest
for dust transport over JAPAN, Water, Air, and Soil Pollution: Focus, 5, 51-69, 2005.
2004
Hirose, K., Kim, C.K., Kim, C.S., Chang, B.W., Igarashi, Y., and Aoyama, M., Wet and dry
depositions of plutonium in Daejeon, Korea, The Science of the Total Environment, 332, 243252, 2004.
Povinec, P.P., Hirose, K., Honda, T., Ito, T., MarianScott, E., and Togawa, O., Spatial
distribution of 3H, 90Sr, 137Cs and 239, 240Pu in surface waters of the Pacific and Indian Oceans GLOMARD database, Journal of Environmental Radioactivity, 76, 113-137, 2004.
Aoyama, M., and Hirose, K., Artificial Radionuclides database in the Pacific Ocean: Ham
database, TheScientificWorldJOURNAL, 4, 200-215, 2004.
Hirota, M., Nemoto, K., Wada, A., Igarashi, Y., Aoyama, M., Matsueda, H., Hirose, K.,
Sartorius, H., Schlosser, C., Schmid, S., Weiss, W., and Fujii, K., Spatial and Temporal
Variations of Atmospheric 85Kr Observed During 1995-2001 in Japan: Estimation of
Atmospheric 85Kr inventory in the Northern Hemisphere, Journal of Radiation Research,
45, 405-413, 2004.
Hirose, K., Chemical Speciation of Thorium in Marine Biogenic Particulate matter, The
Scientific World JOURNAL 4, 67-76, 2004.
2003
Hirose, K., and Aoyama, M., Analysis of 137Cs and 239, 240Pu concentrations in surface waters of
the Pacific Ocean, Deep-Sea Research II, 50, 2675-2700, 2003.
Hirose, K., Aoyama, and Povinec, P.P., Concentration of particulate plutonium in surface anddeep-water, samples collected during the IAEA’97 expedition, Deep-Sea Research II, 50, 26392647, 2003.
Povinec, P.P., Livingston, H.D., Shima, S., Aoyama, M., Gastaud, J., Goroncy, I., Hirose, K.,
Huynh-Ngoc, L., Ikeuchi, Y., Ito, T., Rosa, J.L., Kwong, L.L.W., Lee, S.H., Moriya, H., Mulsow,
S., Oregioni, B., Pettersson, H., and Togawa, O., IAEA’97 expedition to the NW Pacific OceanResults of oceanographic and radionuclide investigations of the water column, Deep-Sea
Research II, 50,2607-2637, 2003.
Ito, T., Povinec, P.P., Togawa, O., and Hirose, K., Temporal and Spatial variation ofanthropogenic radionuclides in seawater of the Japan Sea, Deep-Sea Research II, 50, 2701-2711,
2003.
Tsumune, D., Aoyama, M., and Hirose, K., Behavior of 137Cs and concentrations in the North
Pacific in an ocean general circulation model, Journal of Geophysical Research, 108, C8, 3262,
2003.
Hirose, K., and Aoyama, M., Present Background Levels of Surface 137Cs and 239, 240Pu
Concentrations in the Pacific, Journal of Environmental Radioactivity, 69(1-2), 53-60, 2003.
Tsumune, D., M. Aoyama, and K. Hirose, Numerical simulation of 137Cs and 239, 240Pu
concentrations by an ocean general circulation model, Journal of Environmental Radioactivity,69(1-2), 61-84, 2003.
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Aoyama, M., and Hirose, K., Temporal variation of 137Cs water column inventory in the North
Pacific since the 1960s, Journal of Environmental Radioactivity, 69(1-2),107-117, 2003.
9. Hirose, K., Igarashi, Y., Aoyama, M., Kim, C.K., Kim, C.S., Chang, B.W., Recent trends ofplutonium fallout observed in Japan: plutonium as a proxy for desertification, J. Environ. Monit,
5, 302-307, 2003.
10. Igarashi, Y., Aoyama, M., Hirose, K., Miyao, T., Nemoto, K., Tomita, M., and Fujikawa, T.,
Resuspension: Decadal Monitoring Time Series of the Anthropogenic Radioactivity Depositionin Japan, Journal of Radiation Research, 44, 319-328, 2003.
11. Kim, C.K., Kim, C.S., Chang, B.U., Choi, S.W., Chung, C.S., Hong, G.H., Hirose, K., andIgarashi, Y., Plutonium isotopes in seas around the Korean Peninsula, The Science of the Total
Environment, 318, 197-209, 2003.
12. Kim, C.K., Kim, C.S., Chang, B.U., Choi, S.W., Chung, C.S., Hong, G.H., Hirose, K., andPettersson, H.B.L., 240Pu /239Pu atom ratios in the bottom sediments of the NW Pacific
Ocean, Journal of Radioanalytical and Nuclear Chemistry, 258(2), 265-268, 2003.
13. Lee, H.N., Tanaka, T., Chiba, M., and Igarashi, Y., Long range transport of Asian dust from dust
storms and its impact on Japan, Water, Air, and Soil Pollution, 3, 231-243, 2003
14. Hirose, K., Implication of POC/234Th ratios in oceanic particulate matter: an approach to particle
aggregation, Papers of Meteorology and Geophysics, 53, 4, 109-118, 2003.
1.
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5.
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1.

2002
Hirose, K., Miyao, T., Aoyama, M., and Igarashi, Y., Plutonium isotopes in the Sea of Japan,
Journal of Radioanalytical and Nuclear Chemistry, 252, 293-299, 2002.
Hirose, K., and Aoyama, M., Chemical speciation of plutonium in seawater, Analytical andBioanalytical Chemistry, 372, 418-420, 2002.
Hong, G.H., Kim, Y.I., Lee, S.H., Cooper, L.W., Choe, S.M., Tkalin, A.V., Lee, T., Kim, S.H.,
Chung, C.S., and Hirose, K., 239+240Pu and 137Cs concentrations for zooplankton and nekton in
the Northwest Pacific and Antarctic Oceans 1993-1996, Marine Pollution Bulletin, 44, 660-665,
2002.
2001
Igarashi, Y., Aoyama, M., Nemoto, K., Hirose, K., Miyao, T., Fushimi, K., Suzuki, M., Yasui,
S., Asai, Y., Aoki, I., Fujii, K., Yamamoto, S., Sartorius, H., and Weiss, W., 85Kr measurement
system for continuous monitoring at Meteorological Research Institute, Japan, Journal ofEnvironmental Monitoring, 3, 688-696, 2001.
Igarashi, Y., Aoyama, M., Hirose, K., Miyao, T., and Yabuki, S., Is it Possible to use 90Sr and137Cs as tracers for the aeolian transport?, Water. Air. And Soil Pollution, 130, 349-354, 2001
Hirose, K., Igarashi, Y., Aoyama, M., and Miyao, T., Long-term trends of plutonium fallout
observed in Japan, In Plutonium in the Environment, 251-266, 2001.
Aoyama, M., Hirose, K., Miyao, T., Igarashi, Y., and Povinec, P.P., 137Cs activity in surface
water in the western North Pacific, J. Radioanal. Nucl. Chem., 248, 3, 789-793, 2001.
Aoyama, M., Hirose, K., Miyao, T., Igarashi, Y., and Povinec, P.P., Temporal variation of 137Cs
inventory in the western North Pacific, J Radioanal Nucl Chem, 248, 3, 785-787, 2001
Hirose, K., Aoyama, M., Miyao, T., and Igarashi, Y., Plutonium in seawaters on the western
North Pacific, J Radioanal Nucl Chem, 248, 3, 771-776, 2001.
Tsumune. D., Aoyama, M., Hirose, K., Maruyama, K., and Nakashiki, N., Caluculation of
Artificial Radionuclides in the Ocean by an Ocean General Circulation Model, J Radioanal Nucl
Chem, 248, 3, 777-783, 2001.
2000
Shiraishi, K., Kimura, S., Yonehara, H., Takada, J., Ishikawa, M., Igarashi, Y., Aoyama, M.,
Komura, K., and Nakazima, T., Survey of external dose around the JCO facility using sugar
samples and ESR method, Adv. ESR Appl, 16, 9-14, 2000.

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Komura, K, Yamamoto, M., Muroyama, T., Murata, Y., Nakanishi, T., Hoshi, M., Takada, J.,
Ishikawa, M., Kitagawa, K., Suga, S., Endo, A., Tozaki, N., Mitsugashira, T., Hara, M.,
Hashimoto, T., Takano, M., Yanagawa, Y., Tsuboi, T., Ichimasa, M., Ichimasa, Y., Imura, H.,
Sasajima, E., Seki, R., Saito, Y., Kondo, M., Kojima, S., Muramatsu, Y., Yoshida, S., Shibata,S., Yonehara, H., Watanabe, Y., Kimura, S., Shiraishi, K., Bannai, T., Sahoo, S.K., Igarashi, Y.,
Aoyama, M., Hirose, K., Uehiro, M., Doi, T., and Matsuzawa, T., The JCO criticality accident
at Tokai-mura, Japan: an overview of the sampling campaign and preliminary results, J. Environ.
Radioactivity, 50, 3-14, 2000.
Igarashi, Y., Miyao, T., Aoyama, M., Hirose, K., Sartorius, H., and Weiss, W., Radioactive noble
gases in the surface air monitored at MRI, Tsukuba, before and after the JCO accident, J.
Environ. Radioactivity, 50, 107-118, 2000.
Miyao, T., Hirose, K., Aoyama, M., and Igarashi, Y., Trace of the recent deep water formationin the Japan Sea deduced from historical 137Cs data, Geophys. Res. Lett., 27, 22, 3731-3734,
2000.
Aoyama, M., Hirose, K., Miyao, T., and Igarashi, Y., Low level 137Cs measurements in deep
seawater samples, Appl. Radiat. Isot, 53, 159-162, 2000.
Igarashi, Y., Sarutorius, H., Miyao, T., Weiss, W., Fushimi, K., Aoyama, M., Hirose, K., andInoue, H.Y., 85Kr and 133Xe monitoring at MRI, Tsukuba and its importance, J. Environ.
Radioactivity, 48, 191-202, 2000.

1999
Hirose, K., Amano, H., Baxter, M.S., Chaykovskaya, E., Chumichev, V.B., Hong, G.H., Isogai,
K., Kim, C.K., Kim, S.H., Miyao, T., Morimoto. T, Nikitin, A., Oda, K., Pettersson, H.B.L.,
Povinec, P.P., Seto, Y., Tkalin, A., Togawa, O., and Veletova, N.K., Anthropogenic
radionuclides in seawater in the East Sea/Japan Sea: Results of the first-stage Japanese-KoreanRussian expedition, J. Environ. Radioactivity, 43, 1-13, 1999.
Igarashi, Y., Aoyama, M., Miyao, T., Hirose, K., Komura, K., and Yamamoto, M., Airconcentration of radiocaesium in Tsukuba, Japan following the release from the Tokai waste
treatment plant: comparisons of observations with predictions, Applied Radiation and Isotopes,
50, 1063-1073, 1999.
Igarashi, Y., Aoyama, M., Miyao, T., Hirose, K., and Tomita, M., Anomalous 90Sr deposition
during the fall, 1995 at MRI, Tsukuba, Japan, J. Radioanal. Nucl. Chem, 239(3), 539-542, 1999.
Ikeuchi, Y., Amano, H., Aoyama, M., Berezhnov, V.I., Chaykovskaya, E., Chumichev, V.B.,
Chung, C.S., Gastaud, J., Hirose, K., Hong, G.H., Kim, C.K., Kim, S.H., Miyao, T., Morimoto,T., Nikitin, A., Oda, K., Pettersson, H.B.L., Povinec, P.P., Tkalin, A., Togawa, O., and Veletova,
N.K., Anthropogenic radionuclides in seawater of the Far Eastern Seas, Sci. Total Environ.,
237/238, 203-212, 1999.
Pettersson, H.B.L., Amano, H., Berezhnov, V.I., Chaykovskaya, E., Chumichev, V.B., Chung,C.S., Gastaud, J., Hirose, K., Hong, G.H., Kim, C. K., Lee, S.H., Morimoto, T., Nikitin, A., Oda,
K., Povinec, P.P., Suzuki, E., Tkalin, A., Togawa, O., Veletova, N.K., Volkov, Y., and Yoshida,K., Anthropogenic radionuclides in sediments in the NW Pacific Ocean and its marginalseas:results of the 1994-1995 Japanese-Korean-Russian expeditions, Sci. Total Environ,
237/238, 213-224, 1999.
1998
Igarashi, Y., Hirose, K., and Otsuji-Hatori, M., Beryllium-7 Deposition and Its Relation to
Sulfate Deposition, J. Atmos. Chem, 29, 217-231, 1998.
Miyao, T., Hirose, K., Aoyama, M., and Igarashi, Y., Temporal variation of 137Cs and 239, 240Pu
in the sea of Japan, J. Environ. Radioactivity, 40, 239-250, 1998.
1997
Hirose, K., Complexation-scavenging of plutonium in the ocean, Radioprotection - Colloq, 32,
C2-225 - C2-230, 1997.

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1996
Igarashi, Y., Otsuji-Hattori, M., and Hirose, K., Recent deposition of 90Sr and 137Cs observed in
Tsukuba, J. Environ. Radioactivity, 31, 157-169, 1996.
Otsuji-Hatori. M., Igarashi, Y., and Hirose, K., Preparation of a Reference Fall out Material for
Activity Measurements, J. Environ Radioactivity, 31, 2, 143-155, 1966.
1995
Aoyama, M., and Hirose, K., The temporal and spatial variation of 137Cs concentration in thewestern North Pacific and its marginal seas during the period from 1979 to 1988, J. Environ.
Radioactivity, 29, 57-74, 1995.
Hirose, K., Geochemical studies on the Chernobyl radioactivity in environmental samples, J.
Radioanal. Nucl. Chem., Articles, 197, 315-335, 1995.
1994
Hirose, K., Takatani, S., and Aoyama, M., Deposition of 90Sr and plutonium isotopes derived
from the Chernobyl accident in Japan, J. Radioanal. Nucl. Chem., Articles, 182, 349-358, 1994.
1993
Hirose, K., and Sugimura, Y., Chemical speciation of particulate 238U, 239, 240Pu and Th isotopes
in seawater, Sci. Total Environ., 130/131, 517-524, 1993.
Hirose, K., Takatani, S., and Aoyama, M., Wet deposition of radionuclides derived from the
Chernobyl accident, J. Atmos. Chem., 17, 61-71, 1993.
Shiraishi, K., Igarashi, Y., Yamamoto, M., and Nakajima, T., Concentrations of thorium anduranium in freshwater samples collected in the former USSR, J. Radioanal. Nucl. Chem.,
185,157-165, 1993.
1992
Aoyama, M., Hirose, K., and Takatani, S., Particle size dependent dry deposition velocity of theChernobyl radioactivity, In Precipitation Scavenging and Atmosphere - Surface
Exchange.volume 3, The 22e Summers Volume: Application and Appraisals. Coordinated by
S.E. Schwartz and W.G.N. Slinn. Hemisphere Publishing Corporation, Washington, 1581-1593,
1992.
Hirose, K., Sugimura, Y., and Aoyama, M., Plutonium and 137Cs in the western North Pacific:
estimation of residence time of plutonium in surface water, Appl. Radiat. Isot, 43, 349-359,
1992.
1991
Aoyama, M., Hirose, K., and Sugimura, Y., The temporal variation of stratospheric fallout
derived from the Chernobyl accident, J. Environ. Radioactivity, 13, 103-115, 1991.
Hirose, K., and Sugimura, Y., Chemical speciation of particulate uranium in seawater, J.
Radioanal. Nucl. Chem., Articles, 149, 83-96, 1991.
1990
Hirose, K., and Sugimura, Y., Plutonium isotopes in the surface air in Japan: effect of Chernobyl
accident, J. Radioanal. Nucl. Chem., Articles, 138, 127-138, 1990.
Hirose, K., Aoyama, M., and Sugimura, Y., Plutonium and cesium isotopes in river waters inJapan, J. Radioanal. Nucl. Chem., Articles, 141, 191-202, 1990.
Hirose, K., Aoyama, M., and Sugimura, Y., Short and long term effects of Chernobyl
radioactivity on deposition and air concentrations in Japan, IAEA-SM 306/129, 141-149, 1990.
1988
Aoyama, M., Evidence of stratospheric fallout of caesium isotopes from Chernobyl accident,
Geophys. Res. Lett., 15, 327-330, 1988.
Miyake, Y., Saruhashi, K., Sugimura, Y., Kanazawa, T., and Hirose, K., Contents of 137Cs,
plutonium and americium isotopes in the southern ocean waters, Pap. Met. Geophys, 39, 95113, 1988.
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1987
Aoyama, M., Hirose, K., and Sugimura, Y., Deposition of gamma-emitting nuclides in Japanafter the reactor-IV accident at Chernobyl, J. Radioanal. Nucl. Chem., Articles, 116, 291-306,
1987.
Hirose, K., Aoyama, M., Katsuragi, Y., and Sugimura, Y., Annual deposition of Sr-90, Cs-137
and Pu-239, 240 from the 1961-1980 nuclear explosions: a simple model, J. Meteor. Soc. Japan,
65, 259-277, 1987.
1986
Aoyama, M., Hirose, K., Suzuki, Y., and Sugimura, Y., High level radioactive nuclides in Japanin May, Nature, 321, 819-820, 1986.
Hirose, K., Sugimura, Y., and Katsuragi, Y., 90Sr and 239+240Pu in the surface air in Japan: theirconcentrations and size distributions, Pap. Met. Geophys, 37, 255-269, 1986.
Katsuragi, Y., and Aoyama, M., Seasonal variation of Sr-90 fallout in Japan through the end of
1983, Pap. Met. Geophys., 37, 15-36, 1986.
Suzuki, Y., Inoue, H., Katsuragi, Y., and Sugimura, Y., The distribution of 85Krin the air over
the North and South Pacific Ocean Mem., Natl. Inst. Polar Res., Spec. Issue, 40, 462-466, 1986.
1985
Hirose, K., and Sugimura, Y., A new method of plutonium speciation in seawater, J. Radioanal.
Nucl. Chem, Articles, 92, 363-369, 1985.
1984
Hirose, K., and Sugimura, Y., Plutonium in the surface air in Japan, Health Phys., 46, 12811285, 1984.
Inoue, H., Katsuragi, Y., and Shigehara, K., Tritiated water vapor in the surface air at Tokyo,
Pap. Met. Geophys, 35, 11-20, 1984.
1983
Katsuragi, Y., A study of 90Sr fallout in Japan, Pap. Met. Geophys., 33, 277-291, 1983.
Katsuragi, Y., Kawamura, K., and Inoue, H., Tritium fallout in Tokyo, Pap. Met. Geophys, 34,
21-30, 1983.
1982
Inoue, H., and Katsuragi, Y., A study of tritium fallout in Japan, Pap. Met. Geophys, 32, 21-28,
1982.
Katsuragi, Y., Hirose, K., and Sugimura, Y., A study of plutonium fallout in Japan, Pap. Met.
Geophys, 33, 85-93, 1982.
1981
Hirose, K., and Sugimura, Y., Concentration of Uranium and the activity ratio of 234U /238U insurface air-effect of atmospheric burn-up of cosmos-954, Pap. Met. Geophys., 32, 317-322,
1981.
Hirose, K., and Sugimura, Y., Plutonium content of river water in Japan, Pap. Met. Geophys, 32,
301-305, 1981.
Sugimura, Y., and M. Mayeda, An improved method of analysis of uranium in sea water using
chelating resin, Pap. Met. Geophys, 32, 167-171, 1981.
1980
Sugimura, Y., The uranium content and the activity ratio 234U /238U in sea water in the Pacificocean, Isotope Marine Chemistry, 211-246, 1980.
1979
Sugiura, Y., Saruhashi, K., and Miyake, Y., Evaluation on the disposal of radioactive wastes
into the North Pacific. - the effect of steady flow and up - welling, Proc. The 3rd NEA Seminaron Marine Radioecology, 1979.
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1978
Miyake, Y., Nuclear weapons and radioactive pollution of the earth's environment, In, Proc.
Internal Symp. On the damage and after-effects of the atomic bombing of Hiroshima andNagasaki July 21-Aug., 9, 1977, Tokyo, 164-188, 1978.
1976
Miyake, Y., and Saruhashi, K., A critical study on the IAEA definition of high level radioactive
waste unsuitable for dumping at sea, Pap. Met. Geophys. 27, 75-80, 1976.
Miyake, Y., and Saruhashi, K., Disposal of radioactive waste into deep seas, J. Radiat. Res., 17,
42-43, 1976.
Miyake, Y., and Sugimura, Y., The plutonium content in the Pacific Ocean waters, Proc. of
Symp. on Transuranium Nuclides in the Environment, San Francisco, 17-21 Nov. 1975,
USERDA and IAEA, IAEA-SM-199/22, 91-105,1976.
Sugimura, Y., Symposium II plutonium in environment, II-1.Distribution and behavior ofplutonium in the global environments, J. Radiat. Res., 17, 4-5, 1976.
Sugiura, Y., Saruhashi, K., and Miyake, Y., Evaluation on the disposal of radioactive wastes
into the North Pacific, Pap. Met. Geophys, 27, 81-87, 1976.
1975
Miyake, Y., Shimada, T., Kawamura, K., Sugimura, Y., Shigehara, K., and Saruhashi, K.,
Distribution of tritium in the Pacific ocean, Rec. Oceanogr. Works Japan, 13, 17, 1975.
Miyake, Y., Katsuragi, Y., and Sugimura, Y., Plutonium fallout in Tokyo, Pap. Met. Geophys,
26, 1-8, 1975.
Miyake, Y., Sugimura, Y., and Hirao, Y., Uranium, thorium and potassium contents in granitic
and basaltic rocks in Japan, In, The Natural Radiation Environment II, II, Proc. 2nd Intnl. Symp.
On Natural Radiation Environment, Aug. 7-11, 1972, Houston, Texas, USA, ed. J. A. S. Adams,
W. M. Lowder, T. F. Gesell, 535-558,1975.
Saruhashi, K., I-1 The global radioactive contamination due to nuclear weapon testings, J. Radiat.
Res., 16, 47, 1975.
Saruhashi, K., Katsuragi, Y., Kanazawa, T., Sugimura, Y., and Miyake, Y., 90Sr and 137Cs in the
Pacific waters, Rec. Oceanogr. Works Japan, 13, 1-15, 1975.
1973
Miyake, Y., and Saruhashi, K., The radio-carbon model of the ocean, Pap. Met. Geophys, 24,
263-271, 1973.
Miyake, Y., Saruhashi, K., and Sugimura, Y., The isotopic ratio 234U /238U in sea water and its
bearings on the isotopic ratio in river waters, Oceanogr. Works Japan, 12, 23-25, 1973.
Miyake, Y., Sugimura, Y., and Saruhashi, K., Content of plutonium in river water in Japan, Pap.
Met. Geophys, 24, 75-78, 1973.
Miyake, Y., Sugimura, Y., and Yasujima, T., Thorium isotope content in river water in Japan,
Pap. Met. Geophys, 24, 67-73, 1973.
1972
Miyake, Y., Sugimura, Y., and Uchida, T., A new method of spectrophotometric determinationof uranium in sea water and uranium content with 234U /238U ratio in the Pacific water, Rec.
Oceanogr. Works Japan, 11, 53-63, 1972.
1970
Miyake, Y., Katsuragi, Y., and Sugimura, Y., A study on plutonium fallout, J. Geophys. Res.,75, 2329-2330, 1970.
Miyake, Y., Sugimura, Y., and Mayeda, M., The uranium content and the activity ratio 234U
/238U in marine organisms and sea water in the western North Pacific, J. Oceanogr. Soc. Japan,
26, 123-129, 1970.

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1969
Miyake, Y., Radioactive contamination of the ocean, Bull. Jpn. Soc. Fisheries Oceanogr., Nov.,
1969, 1969.
1968
Miyake, Y., and Sugimura, Y., Plutonium content in the western North Pacific waters, Pap. Met.
Geophys, 19, 481-485, 1968.
Miyake, Y., Katsuragi, Y., and Sugimura, Y., Deposition of plutonium in Tokyo through the endof 1966, Pap. Met. Geophys, 19, 267-276, 1968.
1967
Miyake, Y., and Kanazawa, T., Atmospheric ozone and radioactive fallout, Pap. Met. Geophys.,
18, 311-326, 1967.
Miyake, Y., Radioactive contamination of the ocean, J. Radiat. Res., 8, 1, 1967.
Miyake, Y., Sea, radioactivity in, International Dictionary of Geophysics, 1-7, 1967.
1966
Miyake, Y., and Saruhashi, K., On the radio-carbon age of the ocean waters, Pap. Met. Geophys.,
17, 218-223, 1966.
Miyake, Y., and Sugimura, Y., Ratio 234U /238U and the uranium concentration in seawater inthe western North Pacific, J. Geophys. Res., 71, 3083-3087, 1966.
1965
Kuroda, P. K., Miyake, Y., and Nemoto, J., Strontium isotopes Global circulation after the
Chinese nuclear explosion of 14 May 1965, Science, 150, 1289-1290, 1965.
Miyake, Y., Saruhashi, K., Katsuragi K., and Kanazawa, T., Radioactivity of dust and rain - theratio of Cs-137 and Sr-90 in the radioactive fallout, Proc. Internl. Conf. On Cloud Physics,Tokyo,
IAMAP, 395-399, 1965.
Park, K., George, M.J., Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Strontium90 and caesium-137 in Columbia river plume, July 1964, Nature, 208, 1084-1085, 1965.
1964
Miyake, Y., and Sugimura, Y., Uranium and radium in the western North Pacific waters, Studies
on Oceanography, Prof. K. Hidaka Sixtieth Birthday Commemorative Volume, Univ. of Tokyo
Press, Tokyo, 274-278, 1964.
Miyake, Y., and Ohtsuka, Y., Beryllium-7 in rain water, Pap. Met. Geophys, 15, 89-92, 1964.
Miyake, Y., A sequential procedure for the radiochemical analysis of marine material, Annex tosafety series 11."Method of surveying and monitoring marine radioactivity, IAEA, Vienna, 26,1964.
Miyake, Y., Atomic weapons and the pollution of the sea, The East, 3, 64-68, 1964.
Miyake, Y., Saruhashi, K., Katsuragi, Y., Kanazawa, T., and Sugimura, Y., Uranium, radium,
thorium, Ionium, strontium 90 and cesium 137 in coastal waters of Japan, In, Recent Researches
in the Fields of Hydrosphere, Atmosphere and Nuclear Geochemistry, published by Editorial
Committee of Sugawara Festival Volume, Maruzen Co. Ltd., 127-141, 1964.
Miyake, Y., Sugimura, Y., and Tsubota, H., Content of uranium, radium, and thorium in riverwaters in Japan, In, The Natural Radiation Environment, ed. By J. A. S. Adams and W. M.
Lowder, Univ. of Chicago Press, Chicago, Rice Univ. Semicentennial Series, 219-225, 1964.
Sugimura, Y., Torii, T., and Murata, S., Uranium distribution in drake passage waters, Nature,
204, 464-465, 1964.
1963
Folsom, T.R., and Saruhashi, K., A comparison of analytical techniques used for determinationof fallout cesium in sea water for oceanographic purpose, J. Radiat. Res., 4, 39-53, 1963.
Miyake, Y., and Tsubota, H., Estimation of the direct contribution of meteoric water to riverwaters by means of fall-out radio-cesium and radiostrontium, "Radioisotopes in Hydrology"
Proc. IAEA Symp., Tokyo. IAEA, Vienna, 425-431, 1963.
- 40 -

3.
4.
5.
6.

1.
2.
3.
4.

1.
2.

1.
2.
3.

1.
2.

1.
2.
3.
4.

1.
2.

Miyake, Y., Artificial radioactivity in the Pacific ocean, IUGG Monograph, 20 Radioactive
traces in oceanography, 1963.
Miyake, Y., Saruhashi, K., Katsuragi, K., and Kanazawa, T., Deposition of Cs-137 and Sr-90 in
Tokyo through the end of July 1963, Pap. Met. Geophys, 13, 180-181, 1963.
Miyake, Y., Saruhashi, K., Katsuragi, Y., Kanazawa, T., and Tsunogai, S., Deposition of 90Sr
and 137Cs in Tokyo through the end of July 1963, Pap. Met. Geophys, 14, 58-65, 1963.
Miyake, Y., Sugiura, Y., and Tsubota, H., II. Contents of uranium, radium, and thorium in riverwaters in Japan, The Natural Radiation Environment, RICE Univ. semicentennial Pub, 1963.

1962
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Seasonal variation of radioactive
fallout, J. Geophys. Res., 67, 189-193, 1962.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Penetration of 90Sr and 137Cs indeep layers of the Pacific and vertical diffusion rate of deep water, J. Radiat. Res., 3, 141-147,
1962.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., The peak in radioactive fallout inthe temperate zone of the northern hemisphere, J. Radiat. Res., 3,148-152, 1962.
Sugimura, Y., and Sugimura, T., Uranium in recent Japanese sediments, Nature, 194, 568-569,
1962.
1961
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Cesium 137 and strontium 90 insea water, J. Radiat. Res., 2 25-28, 1961.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Cesium 137 and strontium 90 insea water, Pap. Met. Geophys., 12, 85-88, 1961.
1960
Miyake, Y., and Saruhashi, K., Vertical and horizontal mixing rates of radioactive material in
the ocean, Disposal of Radioactive wastes IAEA Vienna, 167-173, 1960.
Miyake, Y., and Katsuragi, Y., Strontium 90 in western North Pacific surface waters, Pap. Met.
Geophys., 11, 188-190, 1960.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Radioactive fallout in Japan andits bearings on meteorological conditions, Pap. Met. Geophys., 11, 151-158, 1960.
1959
Miyake, Y., Saruhashi, K., and Katsuragi, Y., The Sr-90 fallout and the air motion, Pap. Met.
Geophys., 9, 172-176, 1959.
Miyake, Y., Special committee on oceanic research, Working Group on Radioactivity in the
Ocean, 1959.
1958
Miyake, Y., and Saruhashi, K., Distribution of man-made radio-activity in the North Pacificthrough summer 1955, J. Mar. Res., 17, 383-389, 1958.
Miyake, Y., and Sugiura, Y., The method of measurement of radioactivity in sea water, Pap.
Met. Geophys., 9, 48-50, 1958.
Miyake, Y., Hazards to human health of radioactive dust, Radio Japan, 2, 3-5, 1958.
Miyake, Y., The distribution of artificial radioactivity in the equatorial region in the Pacific in
the summer of 1956, The Proc. 9th Pacific Science Congress, 16, 227, 1958.
1957
Miyake, Y., and Saruhashi, K., The world-wide strontium 90 deposition during the period from1951 to the fall of 1955, Pap. Met. Geophys., 8, 241-243, 1957.
Miyake, Y., The biological effects of nuclear tests - warning by Japanese scientists, Radio Japan,
1, 3-5, 1957.

- 41 -

3.
4.
5.

1.
2.
3.

1.
2.
3.
4.
5.
6.
7.
8.

1.
2.

Miyake, Y., XI. Methods for the measurement of radioactivity in sea water, Annuals ofInternational Geophysical Year, 5, 1957.
Miyake, Y., XII. Radioactivity as a tracer of air motions in the atmosphere, Symp. On
radioactivity, IGY, Utrecht, 360, 1957.
Miyake, Y., Sugiura, Y., Saruhashi, K., and Kanazawa, T., The estimation of the amount of Sr90 deposition and the external infinite gamma does in Japan due to man-made radioactivity, Pap.
Met. Geophys., 8, 222-231, 1957.

1956
Miyake, Y., On the distribution of radioactivity in the North Pacific ocean in 1954-1955, Internl.
Conf. For Peaceful Uses of Atomic Energy in Geneva, August 1955, 13, pp.381-384, 1956.
Miyake, Y., Radioactivity in rain water and air, Proc. Internl. Conf. For Peaceful Uses of Atomic
Energy in Geneva, August 1955, 13, 345-349, 1956.
Sugiura, Y., and Kanazawa, T., On the radioactive fallout collected in Tokyo on 26th November,
1955, Pap. Met. Geophys., 7, 128-135, 1956.
1955
Miyake, Y., and Sugiura, Y., Radiochemical analysis of radio-nuclides in sea water collected
near BIKINI atoll, Pap. Met. Geophys., 6, 90-92, 1955.
Miyake, Y., Effects of atomic explosions on the atmosphere and sea, The research in the effects
and influences of the nuclear bomb text explosions, 1-6, 1955.
Miyake, Y., Kigoshi, K., and Saruhashi, K., Radiochemical analysis of fission products
contained in the soil collected at Tokyo, May, 1954, Pap. Met. Geophys., 6, 1, 93-94, 1955.
Miyake, Y., The artificial radioactivity in rain water observed in Japan, 1954-1955, Research in
the Effects and Influences of the Nuclear Bomb Test Explosions Science, 151-159, 1955.
Miyake, Y., The artificial radioactivity in rain water observed in Japan, from autumn 1954 tospring 1955, Pap. Met. Geophys., 6, 26-32, 1955.
Miyake, Y., Sugiura, Y., and Kameda, K., On the artificial radioactivity in the sea near Japan,
Pap. Met. Geophys., 6, 90-92, 1955.
Miyake, Y., Sugiura, Y., and Kameda, K., On the distribution of radioactivity in the sea around
Bikini atoll in June, 1954, Pap. Met. Geophys., 5, 253-262, 1955.
Miyake, Y., Sugiura, Y., and Kameda, K., Research in the effects and influence of the nuclear
bomb test explosions, Soc. For Promotion of Science, 415-417, 1955.
1954
Miyake, Y., Kigoshi, K., Sugiura, Y., and Saruhashi, K., A study on the productivity in coastal
waters by means of the radio-carbon, Pap. Met. Geophys., 5, 89-94, 1954.
Miyake, Y., The artificial radioactivity in rain water observed in Japan from May to August,
1954, Pap. Met. Geophys, 5, 173-177, 1954.

- 42 -

6. Publication list (in Japanese) 1954-2018
論文リスト（和文誌）1954-2018- 43 -

1.
2.
3.

4.
5.

2017
足立光司, 福島原子力発電所事故で放出された不溶性放射性粒子の形態と組成, エアロゾル
研究, 32-4,255‒260, 2017.
五十嵐康人, 放射性物質の大気沈着・拡散および陸域からの再浮遊について, Proceedings ofthe 17th Workshop Environmental Radioactivity. KEK Proceedings, 2017-6, 50-55, 2017.
小野貴大, 飯澤勇信, 阿部善也, 中井 泉, 寺田康子, 佐藤志彦, 末木啓介, 足立光司, 五十嵐康人, 福島第一原子力発電所事故により１号機から放出された放射性粒子の放射光マイク
ロビーム X 線分析を用いる化学性状の解明, 分析化学 66（4）, 2017.
梶野瑞王, 大気エアロゾルの環境動態シミュレーション,安全工学, 57-6, 433-441, 2018
木村徹, 五十嵐康人, 財前祐二, ガンマ線スペクトル解析ソフト P-SCAN を用いた高精度な Cs134/Cs137 放射能比の導出, Proceedings of the 17th Workshop Environmental Radioactivity.
KEK Proceedings, 2017-6, 262-266, 2017.

1.

2016
木村徹, 五十嵐康人, 財前祐二, 2016: IAEA-TEL-2015-03 Proficiency Test 試料の測定,
Proceedings of the 17th Workshop Environmental Radioactivity. KEK Proceedings, 2016-8, 228-231.

1.
2.

2015
足立光司, 電子顕微鏡がとらえた放射性粒子, 地球化学, 49,185‒193, 2015.
木村徹, 五十嵐康人,「大気降下物試料の放射能分析の品質管理」，Proceedings of the 16th
Workshop Environmental Radioactivity. KEK Proceedings 2015-4 , 23-27, 2015.

1.

2014
青山道夫, 東電福島原発事故後の海洋での放射能汚染の推移 （With English Abstract）.
YAKUGAKU ZASSHI, Journal of the Pharmaceutical Society of Japan. 2014.

1.

2013
青山道夫, 極低レベル放射能測定：少量表層海水および極深層海水試料への応用.放射線,39(1), 17-20, 2013.

1.
2.
3.
4.

1.
2.

1.
2.

2012
青山道夫, 福島第一原子力発電所から放出された放射性セシウム同位体の北太平洋における総量と分布, 日本原子力学会誌, 54, 780-783, 2012.
青山道夫, 五十嵐康人, 1957 年から始まる気象研究所における環境放射能研究の歴史, 放射線, 38(3), 123-128, 2012.
青山道夫, 海洋に放出された放射性物質の長期地球規模での挙動, 検証！福島第一原発事故-放射性物質の実際と科学者たちの活動の記録（化学 4 月号別冊）. 4, 33-38, 2012.
青山道夫, 五十嵐康人, 広瀬勝己, 月間降下物測定 660 ヶ月が教えること-90Sr, 137Cs および Pu 降下量 1957 年 4 月～2012 年 3 月-. 科学, 82(4), 0442-0457, 2012.
2011
青山道夫, 福島原子力発電所事故による海洋汚染. 理大科学フォーラム, 11(通 巻 329 号),
42-45, 2011.
津旨大輔, 坪野考樹, 青山道夫, 廣瀬勝巳, 福島第一原子力発電所から漏洩した 137Cs の海
洋拡散シミュレーション(With English Abstract). 電力中央研究所報告書, V11002, PP.18, 2011.
2009
五十嵐康人, 放射性同位体分析, In 岩坂泰信, 西川雅高, 山田丸, 洪天祥（編）, 黄砂, 古今書院, pp81-87, 東京, 2009.
Igarashi, Y., Anthropogenic Radioactivity in Aerosol -A Review Focusing on Studies during the
2000s-, 保健物理, 44(3), 2009.
- 44 -

3.

1.
2.

1.
2.

1.
2.

猪股弥生, 五十嵐康人, 千葉長, 青山道夫, 廣瀬勝己, 再現期間を用いた降水時における高ガンマ線線量率の異常値判定. Radioisotopes, 58, 37-42, 2009.

2007
三浦誓也, 貝森優希, 工藤英嗣, 野呂幸男, 時枝隆之, 廣瀬勝己, 青森県における大気中クリプトン-85 濃度調査, 青森県原子力センター所報, 1, 43-44, 2007.
廣瀬勝己, 青山道夫, Kim、S.A. Kim, 太平洋の海水中のプルトニウムの挙動について, 日本
放射化学会誌 別冊, 8 supplement, 103, 2007.
2006
青山道夫, 大気と海洋深層における核実験起源放射性降下物の超低レベル放射能測定,
ADIOISOTOPES, 55, 429-438, 2006.
小村和久, 青山道夫, 秋山正和, 氏平祐輔, 大田裕二, 大塚良仁, 大西勝基, 櫻井敬久, 篠原
伸夫, Ge 検出器の使用状況に関するアンケート報告, RADIOISOTOPES, 55, 699-707,2006.
2004
廣瀬勝己, 化学トレーサーを利用した海洋変動の研究, 測候時報第 71 巻特別号, 71, S81S86, 2004.
五十嵐康人, 廣瀬勝己, 放射性降下物標準試料の調製を通じた 90Sr, 137Cs,プルトニウムなどの人工放射性核種分析の相互比較と参照値の決定, 放射化分析, No. 17, 39-42, 2004.
2002

1.

五十嵐康人, 90Sr

1.

1986
鈴木款, 井上久幸, 杉村行勇, 大気中の放射性クリプトンの分布, 地球化学, 20, 106-108,
1986.

1.
2.
3.

1.
2.
1.

2002.

と 137Cs を用いたダスト輸送過程と再飛散,

エアロゾル研究, 12, 17(4), 252,1978
葛城幸雄, 人工放射性物質, 天気, 25, 第 11 号, 20-23, 1978.
猿橋勝子, 海洋中の放射性物質を指標とした物質の移動・拡散, 水汚染の機構と解析－環境
科学持論－, 日本地球化学編, 東京, 産業図書, 271-287, 1978.
三宅泰雄, 猿橋勝子, 杉村行勇, 葛城幸雄, 金沢照子, 核兵器と地球環境の放射能汚染, 被爆の実相と被爆者の実情 －1977 年 NGO 被爆問題シンポジウム報告書, ISDA－ JNPC 編集,
朝日イブニングニュース社発行, 182-212, 1978.
1976
葛城幸雄, 日本における放射性降下物, 天気, 23, No. 7, 333-345, 1976.
杉村行雄, 環境のなかのプルトニウム, 自然, 3 月号, 1976.
1975
三宅泰雄, 島田利夫, 川村清, 重原好次, 地上大気のトリチウム含量と気象との関係, 日本
地球化学会 “環境問題特別号”, 131-135, 1975.

1.

1974
三宅泰雄, 葛城幸雄, 最近の 90Sr 降下物の起源, 天気, 21, 639-644, 1974.

1.

1971
葛城幸雄, 人工放射能による大気汚染, 気象研究ノート, 107, 258-282, 1971

- 45 -

1.
1.
2.

1970
三宅泰雄, 葛城幸雄, 最近の日本における放射性塵の降下, 天気, 17, 593-598, 1970.
1965
葛城幸雄, 日本における Cs-137 および Sr-90 降下について(I), 天気, 12, 323-328, 1965.
葛城幸雄, 日本における Cs-137 および Sr-90 降下について(II) "成層圏における人工放射性物質の滞留時間の推定", 天気, 12, 377-384, 1965.

1.

1964
三宅泰雄, 葛城幸雄, 金沢照子, 放射性降下物の現状と将来, 科学, 34, 142-148, 1964.

1.

1963
三宅泰雄, 放射性降下物の気象学の側面, 天気, 9, No. 1, 1962.

1962
1. 三宅泰雄, 放射性降下物の気象学的側面, 天気, 9, 1-6, 1962.
2. 三宅泰雄, 猿橋勝子, 放射能による水の汚染, 水利科学, 23, 1-14, 1962.
1.
1.

1.
2.
3.
4.
1.
2.
3.

1958
三宅泰雄, 猿橋勝子, 放射化学と海洋, 科学, 28, 510-513, 1958.
1956
三宅泰雄, 杉浦吉雄, 葛城幸雄, 1955 年 4 月旭川地方に降った放射性の落下塵,気象集誌, 第2 号, 34, 226-230, 1956.
1955
三宅泰雄, 放射能雨の性格, 測候時報, 22, 196-202, 1955.
三宅泰雄, 日本に降った人工放射性雨・雪[II]9 月～12 月, 天文と気象, 21, 1-8, 1955.
三宅泰雄, 杉浦吉雄, 亀田和久, 日本近海の海水放射能について, 昭和 29 年度海洋資源開発
調査報告（黒潮班）, 81-82, 1955.
三宅泰雄, 放射能雨の性格, 測候時報, 22, No. 7, 196-202, 1955.
1954
三宅泰雄, 日本に降った人工放射性雨(1954 年 5-7 月), 天文と気象, 20, 1-8, 1954.
三宅泰雄, 杉浦吉雄,亀田和久, ビキニ海域における人工放射能の分布とその海洋学的考察"
ビキニ灰とそれによる傷害 第三", 科学, 24, 601-605, 1954.
三宅泰雄, 有住直介, 磯野謙治, 田島英三, ビキニの灰とそれによる傷害（つづき）日本に降った放射能雨, 科学, 24, No. 8, 405-409, 1954.

- 46 -

年代・雑誌名不明
1.
2.
3.
4.

三宅泰雄, 放射能とその作用, 遺伝,（1975？）
三宅泰雄, 第 II 編 環境の放射能汚染と化学分析,（1970？）
三宅泰雄, 猿橋勝子, 杉村行勇, 海洋における放射性核種, 海洋化学講座, 第 6 巻, 海洋無機
化学, 第 4 章, 109-170,（1969？）
三宅泰雄, 田島英三, 日米放射能会議, 8-18,（1954？）- 47 -

7. History of the studies at the Meteorological Research Institute

表：研究の歴史- 48 -

表紙の図説明2018 年 3 月までの 90Sr および 137Cs 月間降下量
福島第一原発事故後の大気の放射能汚染の実態把握を継続しています。2011 年3 月の事故以降に確定できた90Sr および137Cs 月間降下量を対数軸で誤差と併せてプロットしています。
134Cs は事故直後には 137Cs とほぼ等量降下していますから、放射性セシウム全体ではこのプロットのほぼ倍量となります。ただし、90Sr、137Cs はどちらもおよそ 30 年の半減期で減衰しますが、134Cs はおよそ 2 年で半減します。図示した誤差は計測の統計誤差で、１シグマです。測定誤差は本来全ての測定値につき表示すべきですが、グラフが見づらくなるため、従来はあえて表示してきませんでした。また、数十年前のデータについては誤差データが伝えられていません。
福島第一原発事故以前に採取され、分析や前処理途上だった試料については、事故の汚染によって実験室環境および測定室環境、測定機器のバックグラウンドなどが大幅に上昇したため、蒸発濃縮工程においても試料の汚染（コンタミネーションと言います）の問題が発生します。そのため、観測値を求めること自体が困難になっていましたが、試料を汚染の水準の低い関西にて分析しこの問題の回避に努めました。しかし、細心の注意と努力にもかかわらず、2010 年の後半および 2011 年初のデータは欠測となってしまいました。
徐々に実験環境の除染や測定機器の入れ替えなどを実施して事故以前のきれいな実験環境を追求し、データを求めるように努めています。また、IAEA が実施した分析相互比較に参画するなどし、その精度の維持管理に努めていますので、現状のデータの信頼性は担保できていると考えています。

- 49 -

8. A caption of the cover art
Monthly depositions of 90Sr and 137Cs before March 2018.
We have been monitoring the deposition amounts for the purpose of understanding the actual conditionof radioactive pollution after the accident of Fukushima Daiichi Nuclear Power Plant (FDNPP). The
monthly deposition amounts of both 90Sr and 137Cs, which were successfully fixed values after the accident
in March 2011, are shown in logarithmic axis with error widths. Since the deposition amounts of 134Cs were
nearly equal to those of 137Cs, the total amounts of radioactive cesium were approximately two times of thevalues in this figure. We would like to note that the half-lives of 90Sr and 137Cs are approximately 30 years,
however, 134Cs decays to half in about 2 years. The error widths are one sigma in statistical error ofmeasurements. Although the error widths of all measurement values were better to be shown, error values
used to be not shown on purpose to avoid busy figure in the former versions. In addition, error data before
several decades are not handed down.
For the analysis of samples collected before the accident of FDNPP and not measured then, significant
increase of background values were severe problem, because environments including laboratories andinstruments were polluted, and it was also problem that samples could have been polluted in the operationsof concentration or other preprocesses (contamination problem). We overcame the difficulty by carrying theanalysis out in western Japan (Kansai area) where the pollution level was low, however, the data during thelatter part of 2010 and the former part of 2011 were lacked in spite of our efforts.
We are making efforts to recover clean working environment as before the accident, by cleaning of theenvironments, exchanging instruments, etc. for the purpose of acquire the correct data. In addition, we
maintain the precision of the data, for example, by participating inter-comparison programs performed by
IAEA, so that, the credibility of the data are assured (please refer to the section: quality control of radioactive
analysis of atmospheric deposition samples ).

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最終ページに図の詳しい説明があります。
See figure caption at the end page of this publication.

「環境における人工放射能の研究（2018 年版）」について

気象研究所では、1954 年以来、主に現在の原子力規制庁放射能調査研究費により大気及び海洋の環境放射能の研究を実施してきました。特に人工放射性核種の降下量を 60 年余りの長期にわたり、東京・つくばで精密測定してきました。この観測は、世界でも最長の定点観測です。このモニタリング観測の期間では、1950 年代から 1960 年代にかけては、旧ソ連や米国等による核実験が行われ、また、その後も 1986 年には旧ソ連のチェルノブイリ原子力発電所の事故が発生するなどがありましたが、この観測によって、それらの影響を含めた
バックグラウンドの人工放射性核種の長期的変動を明らかにし、国民の安全・安心に寄与してきました。2011 年には不幸にして東日本大震災に伴う福島第一原子力発電所事故が発生し、それによって放出された放射性核種の推移を調査するという課題も担うこととなりました。今後も長期的視点で、人工放射性核種の変動を監視する必要があります。
放射性物質の再飛散・輸送・沈着による変動のメカニズムを正しく理解し、将来の変動を予測するためには、長期観測や再飛散観測、サンプルの分析等に加えて、数値計算モデルを用いた解析が必要です。気象研究所では、気候変動や降水予測の研究のため、大気中の微粒子の動きを表現する数値モデルを開発してきましたが、そのような数値モデルを利用した、
バックグラウンド放射能の変動メカニズムを解析する研究も行ってきました。
この研究によるこれまでの観測や解析から、長期的変動に対しては、一旦地面などに沈着した人工放射性核種が、土壌粒子・生物由来の粒子等を通じて空気中へ再飛散する過程が重要である可能性がわかりました。
本論文集「環境における人工放射能の研究（2018 年版）」 は、その研究成果を､関係省庁の担当者の方々及び大学や試験研究機関の研究者の方々に広く周知するために、2016 年から 2018 年までの期間に出版された論文（主に英語論文）を、過去から現在までの成果と最近のトピックスに関するテーマ毎に分類し、各テーマの冒頭に簡単な日本語の解説を加えて､一冊にまとめたものです。

令和 2 年 3 月
気象研究所 研究総務官 高薮 出

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Contents
0. Preface
1. Temporal trends of radioactivity concentrations in the atmosphere and observational
approaches for their resuspension mechanisms
2. Physical and chemical characteristics of the radioactive materials in the atmosphere
3. Evaluations of resuspension , transportation and deposition processes of radioactive materials
using numerical models
4. Publications

2016-2018

5. Publication list 1954-2018
6. Publication list (in Japanese)

1954-2018

7. History of the studies at the Meteorological Research Institute
8. A caption of cover art

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目次
0. はじめに
1. 大気中の放射性物質の濃度変動、再浮遊観測
2. 大気中の放射性物質の物理・化学性状3. 放射性物質の再浮遊、輸送、沈着過程のモデルによる評価
4. 出版論文 2016-2018
5. 論文リスト（英文誌）1954-2018
6. 論文リスト（和文誌）1954-2018
7. 表：研究の歴史
8. 表紙説明-4-

0. はじめに気象研究所（2019 年 3 月現在は環境・応用気象研究部および予報研究部）では、1954 年以来、環境放射能の観測・測定法の開発、放射能汚染の実態の把握、大気や海洋における物質輸送解明のトレーサーとしての利用を目的として、60 年余にわたって環境放射能の研究を実施してきた。1957 年以降、各省庁がそれぞれの所掌で実施してきた環境放射能調査研究関連業
務は、まず旧科学技術庁、その後文部科学省が所管してきた。さらに原子力規制委員会と原子力規制庁の発足に伴って 2013 年からは、原子力規制庁が所管することとなった。このような長期にわたり多数の研究者が係わり、観測・研究を継続してきた結果、環境放射能について世界的にも他に類を見ない貴重な時系列データが内外に提供され、また様々な気象学・海洋学的発見をもたらしてきた。この間の研究成果は 200 編以上の原著論文や解説資料として内外の雑誌や成書で公表されている。
1954 年 3 月 1 日に米国によりビキニ環礁で行われた水爆実験により、危険水域外で操業していた第五福竜丸乗組員が放射性物質を含む降灰（いわゆる死の灰）による被ばくを受けた事件を契機にして、日本における環境放射能研究が本格的に始まった。当時の気象研究所地球化学研究室は環境の放射能の分析・研究において日本で有数の研究室であり、海洋及び大気中の放射能汚染の調査・研究に精力的に取り組んだ。その結果、当時予想されていなかった海洋の
放射能汚染、さらに大気を経由した日本への影響など放射能汚染の拡大の実態を明らかにすることができた。1958 年には、放射能調査研究費による特定研究課題の一つである「放射化学分析（落下塵・降水・海水中の放射性物質の研究）」を開始し、以降、地点の変動はあるが、札幌、仙台、東京、大阪、福岡の五つの管区気象台、秋田、稚内、釧路、石垣島の 4 地方気象
台、輪島、米子の 2 測候所の全国 11 気象官署及び観測船で採取した海水中の人工放射性核種（90Sr、137Cs、3H 及びプルトニウム）の分析を実施してきた。
東京・つくばにおける大気中の人工放射性核種の降下量は 1961 年から 1962 年に行われた大規模な大気圏核実験の翌年である 1963 年に最大値を観測した。その後、「部分的核実験禁止条約」の締結により米ソの大気圏核実験が中止された結果、降下量はおよそ１年の半減時間で減少した。この放射性核種の降下量の減少率は成層圏にまきあげられた物質の滞留時間を反映している。その後も、中国及びフランスにより大気圏核実験は続けられ、人工放射性核種の降下量は増減を繰り返した。1980 年の中国による最後の大気圏核実験の後、フォールアウト
は再度 1 年の半減期間で減少し、1985 年には 1957 年の観測開始以降最も低いレベルになった。しかし、1986 年の旧ソ連のチェルノブイリ原子力発電所事故により、大気中の人工放射性核種濃度（特に揮発性の高い 131I、134Cs、137Cs など）は日本でも 1963 年に近いレベルに達するほど著しく増加した。大部分の放射性核種は対流圏の滞留時間（25 日）で減少したが一部 137Cs は成層圏にも輸送されていることが分かった。1988 年以降は低いレベルで推移しているが、明瞭な減少の傾向は見られなかった。この原因は一度地上に降下した放射性核種の再飛散に由来すると考えている。さらに、再飛散がどこで起るかについて研究を進め、東アジアの砂漠域で発生する黄砂が有力な候補であることを明らかにした。黄砂の発生は大陸域の環境変化と関連しており、降下物中の人工放射性核種は、砂漠化や関連する植生変化など大陸域の環境変化の指標となりうることが分かった。
大気フォールアウトの研究とともに、海洋における放射性核種の挙動も長期的に調査研究を実施してきた。日本周辺海域ばかりでなく、太平洋の広域に亘って海水試料の採取を実施
し、-5-放射能汚染の実態を明らかにした。1960 年代後半から 1970 年代の調査で、海洋表面水中の放射能が北半球の中緯度で高いことを明らかにし、フォールアウトの緯度分布を反映している
ことが分かった。その後、海洋表面水中の放射性核種は海洋の物質循環に支配されていることを明らかにした。さらに、海水中の人工放射性核種の分析法の高度化を実現し少量の試料で分析可能にし、海洋の 137Cs 濃度の精密鉛直断面を描き、核実験由来の 137Cs の主な部分は北太平洋の亜熱帯中層に存在していることを明らかにした。フォールアウトによる人工放射性核種の海洋への主な降下以来、数 10 年以上経過し、その広がりは北太平洋から、インド洋など
他の海盆に及んでいる。これに関連して 2003－4 年に実施された BEAGLE2003 の航海で採取された海水について人工放射性核種の分析を行って、その広がりについて知見が得られた。
1993 年旧ソ連/ロシアによる放射性廃棄物の日本海等への海洋投棄の実態が明らかにされ、それに伴う日本海の放射能調査の実施に参加した。放射性廃棄物による影響は検出されなかったが、調査の結果を踏まえ、日本海における固有水の生成過程及び生成場所（ウラジオストッ
ク沖）についての知見を得ることができた。これらは、海洋放射能のデータベースとして公開されている。
1990 年以降の環境放射能汚染として、1997 年の旧動力炉核燃料開発事業団「アスファルト
固化処理施設」の火災爆発事故や 1999 年の JCO ウラン燃料工場の臨界事故があるが、いずれも環境中に放出された放射能は極めて低いレベルで、その影響は殆どなかった。しかしながら、
137Cs
2011 年 3 月に起きた東日本大震災にともなう東京電力福島第一原子力発電所事故により、の総量だけで 10 PBq（ペタベクレル）以上という過去日本では経験したことのない大量の人
工放射性核種が環境中に放出され、その 2 割程度が日本の国土に降下し、残りの 8 割は北太平洋に降下したと考えられている。これらの影響評価やその後の環境中での拡散状況の把握のためにも、環境放射能調査・研究は今後とも重要である。
2006 年 4 月より 2011 年 3 月まで、気象研究所では放射能調査研究費による特定研究課題
として「放射性降下物の長期変動と再浮遊に関する研究」及び「海洋環境における放射性核種の長期挙動に関する研究」の二課題で環境放射能研究に取り組んできた。また、2011 年 8月からは、「大気を通じた人工放射性核種の陸圏・水圏への沈着およびその後の移行過程の解明研究」の課題のもと、過去の知見を十分に活用しながら、同年 3 月 11 日に発生した東日本大震災にともなう福島第一原子力発電所の事故以降の新たな事態に対応し、放出された
大量の人工放射性核種の挙動を解明する研究を行った。目の前の大気に加えて、北太平洋におけるこれらの核種による汚染実態の把握やその動態の解明に取り組んだ。
過去 60 年余にわたり実施されてきたこうした研究成果を踏まえ、2015 年 4 月からは「人工放射性核種のバックグラウンド大気監視と数値解析に関する研究」を新たに開始した。本報告書では、この研究課題で得られた成果を含め、最新の成果やトピックスをテーマ毎に記述した。関係各位の今後の研究や業務に役立つ資料とすべく、編集作業を行った。我が国における環境放射能研究や大気科学研究に多少なりとも寄与できたとすれば、著者一同の望外の喜びである。

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1. 大気中の放射能物質の濃度変動、再浮遊観測
1）はじめに気象研究所では、大気圏内核実験が盛んに実施された 1950 年代後期から現在までの期間にわたり、大気圏での人工放射性核種の濃度水準の実態とその変動要因を明らかにすべく、
環境影響の大きい重要な核種について観測を継続してきた。90Sr（半減期 28.8 年）および137
Cs（半減期 30.2 年）の月間降下量（大気降下物）の長期観測結果は満 60 年を超え、世界でも稀にみる長期変動に関して重要な知見をもたらしている。また、2003 年 3 月からは大気中浮遊粒子の採取、及び放射能分析も開始し、2011 年 3 月の東京電力福島第一原子力発電所事故（福島第一原発事故）に関する環境影響に関して重要な知見を取得し、また大気モデル改善に貢献し、その結果をいち早く学会で報告するとともに、気象研究所のホームペー
ジにも掲載し一般に公開してきた。複数台有する Ge 半導体分析装置は他研究機関の持つ環
境試料の測定にも貢献し、福島第一原発事故の影響解明にも大きく貢献してきた。前報のIgarashi et al., (2015) では福島第一原発事故以降の三年間に着目した大気降下物および大気
中放射能について報告をしてきたが、事故後七年が経過し、その濃度は徐々に下がってきた
ため、より高精度な測定技術が求められてきた。本報ではより高精度に 134Cs を測定するための定量方法の開発結果について報告し、事故後数年経過した現在における放射性物質の再飛散（二次飛散）や環境中での動態プロセスに関する考察について解説する。

2）研究方法月間大気降下物の捕集は、茨城県つくば市長峰にある気象研究所の露場観測実験棟屋上に設置した大気降下物捕集用のプラスチック製水盤（面積 4m2）で 1980 年代以降実施してきた。福島第一原発事故以降（2011 年 4 月以降）は放射能水準の上昇を考慮し、気象研究所
屋上にて 1 m2 水盤２基での捕集を実施した。2015 年 8 月以降は放射能水準の低下を考慮して 1m2 水盤４基での大気降下物採取を開始した。大気中浮遊粒子は、上記露場観測実験棟前に設置したハイボリューム・エアサンプラー(HV-1000R、柴田科学)に石英繊維フィルター(QR-100、柴田科学)を装着して、空気流量 700LPM で一週間単位でのサンプリングを行った。
得られた大気降下物試料はポリ製保管容器で保存しつつ、ロータリーエバポレーターや蒸発皿等を使用して全量を蒸発濃縮・固化し、Ge 半導体検出器によりγ線放出核種（放射性Cs 等）を測定した。次いで試料の一部を分取して発煙硝酸処理・放射化学分離により 90Sr を精製し、最終的に炭酸 Sr として固定、低バックグラウンド 2πガスフロー検出器でβ放射能を測定した。福島第一原発事故後の試料については、事故により放出された 89Sr の影響があるため、生成固定した炭酸 Sr 線源のβ放射能を繰り返し測定することで、90Y の放射平衡達成と 89Sr 放射能の減衰の様子とを確認しつつ評価し、必要な場合は計算で 89Sr の影響を除去した。フィルターサンプルの場合には 8 等分の短冊にし、金型と油圧プレスで成型することで試料形状を一定形状にし、Ge 半導体検出器によりγ線放出核種（放射性 Cs 等）を測定した。
134Cs/137Cs 放射能比は発生源の特定につながる重要なパラメーターであるが、前述の通り

134Cs（半減期 2.06 年）の濃度は事故直後から見て大きく低下し、より高精度での測定が必要となってきたため、その解析手法について開発検討を行った。着目した開発点は 134Cs の補正方法である。134Cs はカスケードピークを持つことからサム効果による測定影響が生じる。同標準線源でピーク効率を求める際に、一般にはサム補正は不必要とされているが、本研究のフィルター試料成型では試料内での放射能分布に偏りがあることが考えられ、その際に誤差を生じると考えられる。そこで、サム効果補正が可能であるソフトウェア(P-SCAN、
スカラベ・コーポレーション)を使用してサム補正を行い、補正の有無でどの程度定量に差が出るかを比較検討した。

-8-3）結果・議論
3.1) 134Cs/137Cs 放射能比の分析結果
P-SCAN ではγ線スペクトルの光電ピーク面積と、それらの計数効率・全効率の間に成り立つ高次・多元連立方程式を直接解くという方法を採用し、壊変後にカスケードγ線を放出する核種については観測されるピーク数をパターン化して数式化している。ピークの測定に
は気象研究所の 4 台(#6、#7、#8、#10)のゲルマニウム半導体検出器を使用した。標準試料の測定は通常通り行い、次に福島第一原発事故時の放射能強度が高い環境試料を測定、サム効
果補正計算を行い、そうでない場合と比較した。
大気中浮遊粒子試料の測定結果を Table1 に示す。137Cs ではサム効果は見られないため、
P-SCAN の使用有無による有意な違いは生じなかった。一方で 134Cs では P-SCAN で得られた結果の方が数%低い結果となった。134Cs/137Cs 比については PSCAN を使用した場合の方がばらつきは非常に小さかった。このことから、134Cs/137Cs 比を測定する場合には P-SCANを使用した方がよいと考えられるが、絶対値の正しさについては今後の検討が必要である。
Table1. Comparison when using P-SCAN and normal calculation in 4 detectors (result of measuring thesame sample).

二種類のγ線エネルギーの 605keV/662keV 比はサムピークの比を表す。サムピークの発生
確率は検出器と試料の距離に関係し、距離が離れるほどサムピークは生じにくくなる。相対距離 0cm から 5cm ごとに距離を離してサム効果の比率を求めた結果を Fig. 1 に示す。また、
サムピークの代表値として 1400keV/605keV 比を求めた結果を Fig. 2 に示す。
605keV/662keV 比より、15cm 以上、検出器と試料の距離を保てば光電ピーク比が安定することが分かった。また、1400keV/605keV 比から 15cm 以上離せばサムピークの影響が無視できることが分かった。

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Fig.

1. Relationship between distance to sample and 605/662keV photo-peak count ratio.

Fig. 2. Relationship between distance to sample and 1400/605 keV photo-peak count ratio.
3.2)

モンゴル土壌中の 137Cs、およびプルトニウム、ウラン同位体の測定

モンゴル土壌中には過去に行われた核実験などにより放出された人工放射性同位体が含まれており、アジアダストのトレーサーとしてしばしば使われている。Hirose et al. (2017)の結果、モンゴル土壌中のプルトニウム同位体・ウラン同位体比を始めて明らかにした。ここで得られたプルトニウム同位体比は、大気降下物として測定される核実験により放出された
放射性核種（グローバル・フォールアウト）の同位体比と一致しており、モンゴル土壌中の人工放射性核種がグローバル・フォールアウトに起因するものだということが分かった。
137Cs とも関係性が見られ、つくば土壌でも同じように関係性が見られることが分かったが、その傾きは大きく異なった。その原因は放射性核種それぞれの物理的・化学的な移動過程の違いによるものと考えられる。

- 10 -

3.3)

福島第一原発事故に起因する放射性セシウムとバイオエアロゾル

Igarashi et al. (2019) および Kinase et al. (2018) では 2011 年 7 月～2015 年 3 月までに福島県の浪江町・川俣町にて採取した大気中粒子試料の 134Cs,137Cs を測定し、その季節変化を明らかにした(Fig. 3)。その結果、つくばで観測される大気中放射性セシウムの季節変化とは明らかに異なることが分かり、5–6 月、および 8–9 月に極大となる二山ピークを持つことが分かった。これは、過去における研究で明らかにされてきた再飛散プロセスと、福島高濃度
汚染地域における主要な再飛散プロセスが異なっていることを示唆する結果である。

Fig. 3. Activity concentrations of 134,137Cs (Bq m− 3) measured at the FD, FC, GD, and GC sites in
the Fukushima prefecture in July 2011 to March 2015. These results show high concentrations in latespring, summer, and early autumn and low concentrations in winter and early spring at all of thepresented observation locations.

- 11 -

気象研究所内の走査型電子顕微鏡(SEM)及び EDX によるフィルター試料の観察結果より、
大気中放射性セシウム濃度の上昇がみられる試料には、胞子や菌類などのバイオエアロゾルが多く見つかることが分かった（Fig. 4）。

Fig. 4. The normalized particle counts of the bioaerosols using SEM (in relative values, theerror bar shows the Poisson standard deviation) and atmospheric concentrations of 137Cs activity
sampled at the GD site during summer (July and August 2014) and winter (December 2014 and January2015). The counts of microorganisms include those of spores (see the text).
上記結果を踏まえ、Igarashi et al. (2019)では SEM-EDX 分析に加え、DAPI 染色による蛍光
顕微鏡による観察(Fig. 5)、メタゲノム解析によるバイオエアロゾル種の解析を行った(Fig.
6)。その結果、大気中放射性セシウム濃度とバイオエアゾルの大気中濃度には相関がみられ(Fig. 7)、バイオエアロゾルの 1/3～半数が担子菌及び子嚢菌だということが分かった。

- 12 -

Fig. 5. Fluorescent micrographs of DAPI-stained particles (indicated by red arrows) in thebioaerosol samples collected at Namie site in August and September 2015 (a). The aggregated particles
observed as yellow particles (b,c), yellow and blue particles (d), blue particles (e), black (indicated by
the white arrow) and white (indicated by the red arrow) particles (f), white particles (g) and spores
form particles that are likely ascospores (h,i). The bars indicate a length of 10 μm. In the photo, whiteand yellow particles may not exhibit the colours seen by the naked eye on the microscopic screen.

Fig. 6. Results of the metagenomic analysis showing the relative abundance of identified fungal
taxa in the samples collected at the forested and bare soil sites in Namie and contained in rain watercollected at the site during summer 2015. Sampling dates are expressed as mm/dd. The compositional
differences among sampling dates may reflect seasonal and weather differences.

- 13 -

Fig. 7. The relationship between the concentration of coloured fungal spores (countable without
DAPI staining by optical microscopy) and the 137Cs activity in air at the Namie site in summer 2016.
The sampling duration was 24 hours of daytime or nighttime (circled data; e.g., daytime data from
Aug. 24–25 indicates that sampling was performed from 6:00 to 18:00 on August 24 and 25, a total of
24 hours). Despite the large scatter, the spore number and 137Cs concentration exhibited a positive
correlation (rank correlation; significant at 8% based on a t-test). The slope of the fitted curve (1.66 ×
10−8 Bq/grain) corresponds to the lower range of estimated values (see Supplementary Tables S3–S5).
3.4)

福島第一原発事故に起因する河川及び地下水中の水溶性放射性セシウム

Iwagami et al. (2017)では、福島県山木屋地区で河川水および地下水中の水溶性放射性セシ
ウムの長期的観測を実施した。その結果、チェルノブイリ事故と比較して河川・地下水の汚染は低く、その汚染低下のプロセスには二つあることが分かった。まず一つ目のプロセスは、
大量の雨による表層土ごと洗い流す過程であり、事故後初期の段階で周辺の水中放射性セシ
ウムを急激に低下させた。もう一つのプロセスは、土壌中への放射性セシウムの浸透と固定化による、水中放射性セシウム濃度低下の鈍化であり、事故後しばらくたってからの影響として現在も観測されている。この研究により、河川水及び二地点での地下水について環境中での半減期が明らかとなり、それぞれ 0.69–1.5 y、0.46–0.58y、0.50–3.3y となった（Fig. 8）。
河川水中の 137Cs の減少傾向はキャノピーに見られる放射性セシウムの脱離傾向と似ていることが分かった。

- 14 -

Fig. 8. Time series of dissolved 137Cs concentrations in stream water in normal flow condition andgroundwater at (a) Koutaishi-yama, (b) Iboishi-yama, and only stream water at (c) Ishidaira-yama withexponential regression lines. Error bar shows a standard deviation from counting statistics.

- 15 -

4）まとめと今後に向けて本研究所における、より高精度な測定に向けた 134Cs 測定方法開発、モンゴル土壌における人工放射能の観測、福島第一原発事故後の大気中人工放射能観測の結果と他地域には見られない新しい再飛散プロセスの提案、水中に溶存する放射性セシウムの環境中における動態の傾向について報告した。前報にあるように、事故後 7 年が経過した現時点では、福島第一原発事故直後と比較すると、つくばでの降下量は数桁も低下し、甚大な汚染というべき水準ではない。90Sr については、事故以前の降下量とほぼ同じ水準に戻りつつある。しかし、本報で報告した 134Cs 測定技術の改善により、今後も高精度での人工放射能観測が期待できる。
また、137Cs については依然として 3～4 桁より大きな値となっている。水準低下にはまだ時間がかかると考えられ、継続した観測と動態傾向の分析が肝要である。
事故直後に見られた一次放出に替わり、大気へのこれらの核種の供給を続けている過程は、
いわゆる再飛散（二次放出）であり、現在はこの再飛散プロセスにより大気中・降下物中の人工放射能濃度が支配されていると考える。再飛散は周辺の環境に由来し、周辺の汚染度がつくばに比し数桁高い地域においては、大気中濃度・降下量もそれに比例し数桁高い。従って、再飛散プロセスの解明は必須な科学的課題である。
従来、汚染した地表面からの表土ダスト、汚染植生からの再浮遊、または汚染したゴミの燃焼、野焼き等による放射性物質の大気への揮散などが想定されてきた。しかし、本報で報告したように、福島原発事故により高濃度に汚染された地域では、通常の汚染地域とは異なり森林生態系からの放出、すなわちバイオエアロゾルの寄与がとても大きいと分かってきた。
特にキノコに由来する胞子や担子菌・子嚢菌が放射性セシウムの担体として重要と考えられるが、まだ定量的には解明できていない。今後は、バイオエアロゾル観測の定量化と、継続的な連続観測を実施することで定量的な理解を得ていくことが肝要である。そのためには、
森林内での各高度や樹間密度ごとのバイオエアロゾル放出量（フラックス）を定量的に観測する必要があり、顕微鏡と目視による粒子同定だけでなく、リアルタイムで測定可能な手法の開発が望まれる。
上記に加えて、こうした長期の観測においては分析やデータの品質管理が必ず問題となる。
このような点についても常に留意しながら、観測の継続を図っていくことが肝要である。

- 16 -

参考文献
Hirose, K., Kikawada, Y., Igarashi, Y., Fujiwara, H., Jugder, D., Matsumoto, Y., Oi, T., and Nomura, M.
Plutonium, 137Cs and uranium isotopes in Mongolian surface soils, Journal of Environmental
Radioactivity, 166, 97̶̶-103, 2017. http://dx.doi.org/10.1016/j.jenvrad.2016.01.007
Igarashi, Y., Kita., K., Maki, T., Kinase, T., Hyashi, N., Hotaka, K., Adachi, K., Kajino, M., Ishizuka, M.,
Sekiyama, T., T., Zaizen, Y., Takenaka, C., Ninomiya, K., Okochi, H., and Sorimachi, A. Fungal
spore involvement in the resuspension of radiocaesium in summer, Scientific Reports, 9:1-10, 2019.
2019. DOI:10.1038/s41598-018-37698-x
Iwagami, S., Tsujimura, M., Onda, Y., Nishino, M., Konuma, R., Abe, Y., Hada, M., Pun, I., Sakaguchi, ,A., Kondo, H., Yamamoto, M., Miyata, Y., and Igarashi, Y. Temporal changes in dissolved 137Csconcentrations in groundwater and stream water in Fukushima after the Fukushima Dai-ichi Nuclear
Power Plant accident. Journal of Environmental Radioactivity, 166, 458-465, 2017.
http://dx.doi.org/10.1016/j.jenvrad.2015.03.025
木村徹,五十嵐康人,財前裕二,「ガンマ線スペクトル解析ソフト P-SCAN を用いた高精度な
134Cs/137Cs 放射能比の導出」Proceedings of the 18th Workshop Environmental Radioactivity.
KEK Proceedings 2017
Kinase, T., K, Kita, Igarashi, Y., Adachi, K., Ninomiya, K., Shinohara, A., Okochi, H., Ogata, H., Ishizuka,M., Toyoda, S., YamadaN, K., Yoshida, N., Zaizen, Y., Mikami, M., Demizu, H., and Onda, Y. The
seasonal variations of atmospheric 134,137Cs activity and possible host particles for their resuspension
in the contaminated areas of Tsushima and Yamakiya, Fukushima, Japan, Progress in Earth andPlanetary Science 5:12, 2018. https://doi.org/10.1186/s40645-018-0171-z- 17 -

- 18 -

2. 大気中の放射性物質の物理・化学性状
1）はじめに
大気中で観測される放射性物質は、ガスや粒子（エアロゾル）として大気中に漂い、やがて雨による除去や重力沈降によって地上や海水面に降下する。このような大気中での放射性物質の振る舞いはその放射能や核種に依存せず、物理的な粒径や密度、形態、また水溶性な
どの化学性質に依存する。そのため、放射性物質の大気拡散や環境動態、また発生メカニズ
ムを理解するには、放射性物質がどのような物理化学性状を持ち、またどのようなエアロゾ
ルを担体としているかを解明する必要がある。本章では、その放射性物質の物理化学性状の理解を目指した例として、東日本大震災で発生した福島第一原子力発電所事故によって直接放出された放射性物質に関する研究成果を報告する。
2011 年 3 月に発生した事故によって様々な放射性物質が大気・海洋に放出されたが、放出された放射性物質の重量は非常に微量であり、放射能測定以外の分析手法で環境試料から検出することは一般には困難とされてきた。しかしながら、電子顕微鏡や放射光分析などの微小領域化学分析を用いることで、放射性物質の物理化学特性を明らかにすることが可能であることが気象研究所の研究によって明らかになってきた。本研究においては、それらの手法を用いて 2011 年 3 月 14-15 日につくば市で採取されたフィルター試料中の放射性粒子の特定を行った先駆的な研究（Adachi et al., 2013）を発端として、多くの研究事例が報告されている。ここでは、それらの研究例のうち、前回の報告書から後に発表された小野ら（2017）
や Satou (2016,2018)の論文、またこれらの研究例をまとめた総説論文（足立、2017）を紹介する。
2）主な研究成果
東日本大震災で発生した福島第一原子力発電所事故によって直接放出された放射性物質
は、本放射能調査研究費研究でモニタリングを継続していた気象研究所におけるフィルター
試料から発見された（Adachi et al., 2013）。この研究では、走査型電子顕微鏡（SEM-EDS）
を用いて 2011 年 3 月 14－15 日にかけてつくばに飛来した放射性粒子を分析し、その物理
化学性状を初めて明らかにした。その結果、セシウム、鉄、亜鉛や多くの核燃料由来と考えられる元素を含み、水に溶けにくい２マイクロメートル前後の球状ガラス様粒子であることが明らかとなった。これらの詳細な物理特性は、事故時における粒子の拡散モデルの精度向上に重要であることを、モデル比較実験で明らかにした。その後、Abe et al. (2014）やYamaguchi et al. (2016)によって、粒子のさらに詳細な微量成分や内部構造が明らかとなった。
これらの先駆的な研究を受けて、本研究期間ではエアロゾル試料やその大気中微粒子が沈着した結果生じた表層土壌汚染地域からの試料を採取して、そこにふくまれる放射性粒子を単離して、電子顕微鏡や放射光分析を用いてその個別粒子の物理化学性状を明らかにし、また少なくとも 2 種類の異なった物理化学性質を持つ放射性粒子が存在することを明らかにした。
Satou et al (2016)では、事故があった原子力発電所から 20ｋｍほど離れた地点の土壌から、６㎛ほどの放射性粒子を新たに開発し手法で単離し、それらがエアロゾル試料で見つかった粒子と似た、ケイ素、鉄、亜鉛、セシウムなどを含む化学組成をしていることが分かった。
これらの粒子は大気中のエアロゾル試料に見つかった粒子よりも大きく、また放射能強度も強い特徴を持ち、近隣地域にはより大きな粒子が重力沈降で沈着した可能性を示唆した。
小野ら（2017）の論文では、さらに土壌試料中に比放射能（単位質量あたりの放射能）はエアロゾルから分離された粒子より低いものの、大きさが 100μm ほどの放射性粒子を発見
し、放射光分析でその詳細な化学成分を明らかにした。これらの大きな粒子は、球状粒子が複数溶融したような形態をしており、ケイ素を主成分とする特徴を示し、また 20 元素以上の様々な微量成分が検出された。これらの粒子は、Adachi et al (2013)で見つかった粒子と異なる物理化学性質を持っており、異なった放出プロセスがあったことを示唆する。Satou et- 19 -

al (2018)では、それらの粒子のセシウム 134、137 の比が異なることを見つけて、その同位体比の違った粒子のうち比放射能が高く、数 μ 程度で比較的球状に近い粒子を A タイプとし、比放射能は低いものの大きな粒子を B タイプとして分類した。
足立（2017）の論文ではこれらの成果をまとめて、今後の研究の発展に資するデータのまとめを行った。その一例の表を下記に示す。

Table 1 Summary of insoluble radioactive particles from nine studies
References

Sampling
place

Sample type

Adachi et al(2013)
Abe et al(2014)
Yamaguchi
et al. (2016)
Satou et al.,(2016)
Kogure et al.
(2016)
Furuki et al(2017)
Yamaguchi
et al. (2017)
Ono et al.
(2017)
Higaki et al.
(2017)

IbarakiAerosol filter

IbarakiAerosol filter

Fukushima
Fukushima
Fukushima
Fukushima
Fukushima
Fukushima
Fukushima

Diameter(μm) ofspherical
particles

Radioactivity (Bq) of137
Cs

Particlenumber

3.3-0.73

1.4-2.8

1.1-1.5

3

Non-woven fabriccloth & Cedar
Soil

2.2-2.5

3.1-5.0

2

5-67

4

Non-woven fabriccloth & Wheat
Soil & Gravel

0.8-2.7

1.9-3.3

3

3.4

0.9-11.33

1.4-3.6

0.5-1.94

2.0-2.6

3.0-4.0

2.0-3.3

Diameter(μm) of nonsphericalparticles

4.5-6.4

Aerosol filter &Plant tissue
Soil

2.6-3.6

50-25033-109

7

Mask

2.1

6.6-121-70

4足立（2017）エアロゾル研究 32 巻 4 号より。
3）まとめ
本放射能調査費によって行われた研究成果により、福島第一原子力発電所事故において放出された放射性物質の大気中の物理化学特性の一部が明らかとなった。これらの研究をもとに、今後事故から放出された放射性物質の発生源メカニズムの解明や環境動態の予測への貢献が期待される。

- 20 -

参考文献
Y. Satou, K. Sueki, K. Sasa, H. Yoshikawa, S. Nakama, H. Minowa, Y. Abe, I. Nakai, T. Ono, K. Adachi
and Y. Igarashi, Analysis of two forms of radioactive particles emitted during the early stages of the
Fukushima Dai-ichi Nuclear Power Station accident, Geochemical Journal, Vol. 52, 2018、
足立光司、原子力発電所事故で放出された不溶性放射性粒子の形態と組成、エアロゾル研究 324, 255-260, 2017
小野貴大, 飯澤勇信, 阿部善也, 中井 泉, 寺田靖子, 佐藤志彦, 末木啓介, 足立光司, 五十嵐康人、福島第一原子力発電所事故により 1 号機から放出された放射性粒子の 放射光マイク
ロビーム X 線分析を用いる化学性状の解明、分析化学 66（4） 2017
Y. Satou, K. Sueki, K. Sasa, K. Adachi, Y. Igarashi, First successful isolation of radioactive particles fromsoil near the Fukushima Daiichi Nuclear Power Plant, Anthropocene, Volume 14, Pages 71-76, 2016
Abe, Y., Iizawa, Y., Terada, Y., Adachi, K., Igarashi, Y., and Nakai, I., Detection of uranium andchemical state analysis of individual radioactive microparticles emitted from the Fukushima nuclear
accident using multiple synchrotron radiation X-ray analyses, Analytical Chemistry, 86, 8521-8525,
2014.
Adachi, K., Kajino, M., Zaizen, Y., and Igarashi, Y., Emission of spherical cesium-bearing particles froman early stage of the Fukushima nuclear accident, Scientific Reports, 3, 2554, 2013.
Yamaguchi, N., Mitome, M., Kotone, A. H., Asano, M., Adachi, K., and Kogure, T., Internal structure ofcesium-bearing radioactive microparticles released from Fukushima nuclear power plant. Scientific
Reports, 6, 20548, 2016.

- 21 -

- 22 -

3.放射性物質の再浮遊、輸送、沈着モデルによる評価
1） 福島原子力事故後の放射性セシウム大気濃度の長期評価：裸地と森林生態系からの再飛散
福島原子力事故による汚染土壌や森林からの 137Cs 再飛散の長期的影響を、汚染地域（福島県浪江町）の土壌からの再飛散実験結果，および汚染地域内（浪江）・外（茨城県つくば市）の大気濃度測定，および数値シミュレーションを併用して実施した（Kajino et al., 2016）。
評価期間は 2013 年の 1 年間である。推定された再飛散率は約 10-6/day であり、福島県におけるガンマ線線量率の減少率（10-4-10-3/day）よりも明らかに低かった．すなわち，再飛散は
陸上の放射能の減少にはほとんど寄与していなかったことが分かる。再飛散メカニズムは、
特に森林生態系に関しては，未知のままである。今後は，メカニズム解明の実験研究に基づいた影響評価を行っていく必要がある。
Fig. 1 に計算された 2013 年の 1 年間の再飛散量と、再沈着量を示す。領域・時間積算再飛散量は、初期の沈着量 2.68 PBq（NRA, 2012）の 0.048%に相当した。

Fig. 1. Annual total amounts of (a) the re-suspension and (b) the re-deposition of 137Cs. The total
activities are presented as numbers.
2） 福島原発事故時の地上プリューム移流予測に及ぼす地上風データ同化の影響
2011 年 3 月の福島原発事故で排出された放射性セシウムを大気トレーサーとして用い、対流圏下層のプルーム移流の予測可能性と AMeDAS 地上風データ同化の影響を調べた（Sekiyama et al., 2017）。気象庁非静力学モデルと局所アンサンブルカルマンフィルタ（JMANHM-LETKF）データ同化システムを用いて、実験を行った。40 の観測地点における、
放射性セシウムのプルーム到着時間について調べたところ、標準実験（標準的な観測データ
のみ同化）のプルーム到着時間の平均誤差は 82.0 分であった。それに対して、AMeDAS 実験（標準的な観測データと AMeDAS 地上風観測の両方を同化する）の平均誤差は 72.8 分であり、これは標準実験に比べて 9.2 分（11％）高精度であった。本研究は、プルーム予測モデル
が緊急時対応のために十分な精度を有し、予測結果が地上風データ同化によって著しく改善
され得ることを示すものである。
Fig. 2 に、埼玉県久喜市において観測された大気濃度（SPM テープ：浮遊粒子状物質観測用フィルタ上の 137Cs 濃度、Tsuruta et al., 2014; Oura et al., 2015）と、データ同化前後の計算結果の比較を示す。データ同化を行っていないものを STD Exp.（実線）、地上風データ同化を行ったものを AMeDAS Exp.（破線）としている。プリュームの到達時間に明らかな改善が見られている。

- 23 -

Fig. 2 An example of time series of the radioactive cesium (137Cs) concentrations derived from the SPMtape sampling observations and two model experiments (STD and AMeDAS) from 12:00 UTC on March
14 to 12:00 UTC on March 15, 2011. The advection of the plumes was examined using the data from Kuki
City, Saitama Prefecture.
3） 福島原子力事故後の大気モデリング研究から学んだ教訓：アンサンブルシミュレーション、
データ同化、素過程モデリング、逆解析
福島第一原子力発電所事故に伴う放射性セシウムの大気拡散と沈着に関するモデリング研究について、フランス放射線防護・原子力安全研究所と気象庁気象研究所の共同研究に着目しながら、レビューを行った（Kajino et al., 2018）。これまでのモデル研究により、湿性沈着過程が重要なプロセスであることが判明しているが、その影響の大きさは不明のままである。また、モデルによって計算結果が大きく異なることが判明しているが、その原因は定かではない。本共同研究では、大気素過程、特に湿性沈着過程の理解のために、アンサンブ
ルシミュレーション、データ同化、素過程モデリング、逆解析などのさまざまな数値技術を使用して、シミュレーションの各要素の不確実性を定量化することを目指した。本論文では、これらの共同研究の成果を紹介するとともに、今後の大気モデル研究の将来の方向性についても議論した。

4） 福島第一原子力発電所から放出された 137Cs プルームの質量フラックス解析
Sekiyama and Iwasaki (2018) では、137Cs プルームの質量フラックス解析を行い、2011 年 3月に福島第一原子力発電所から放出された 137Cs のプルームについて質量フラックスの流れを定量的に示すことができた。この 137Cs のプルームは 3km の水平分解能を持つオイラー型の大気移流拡散モデルによって計算した。鉛直積分した質量フラックスは 137Cs 濃度の地上観測値に基づく移流経路推定とほぼ一致していたが、地上とプルーム層との間に風速シアーがある場合には不一致が生じた。この不一致は、地表面の水平質量フラックスと鉛直積分した質量フラックスの二つを組み合わせることによって理解できる（Fig. 3 参照）。質量密度と風速の積である質量フラックスの概念は気象力学においては一般的であるが、大気汚染の研究においてはほとんど使用されてこなかった。しかし質量フラックス解析は、数値モデルが計算する４次元（x,y,z,t）の出力値を２次元（x,y）プロットに縮小することによって、大気微量成分の主要な流れの位置、方向、および沈着の振る舞いを明瞭にすることができるた
め定量的な可視化ツールとして最適である。本研究では、大気中の放射性物質（この場合は137Cs）の移流拡散を調べることによって、大気汚染の研究においても質量フラックス解析が非常に有用な手法であることを示すことができた。

- 24 -

Fig. 3 Time-integrals from 15:00 to 21:00 JST on 15 March 2011 of (a) the column-integrated mass fluxof plume P3 and (b) the horizontal mass flux of plume P3 at the ground surface. Color shading indicatesthe magnitude of the mass flux. Small arrowheads indicate the direction of the mass flux. A large grayarrow illustrates the depiction in Nakajima et al. (2017) for plume P3 at 18:00 JST on March 15, shown in
Fig. 2. (c) Latitude-altitude cross-sections at 140.5E for 137Cs concentrations (gray shading), meridional
winds (arrows), and potential temperature (h; contours) at 17:00 JST on 15 March 2011. The cross-section
is located at the green line in panel (b).

- 25 -

参考文献
Kajino, M., M. Ishizuka, Y. Igarashi, K. Kita, C. Yoshikawa, M. Inatsu. Long-term assessment of
airborne radiocesium after the Fukushima nuclear accident: Re-suspension from bare soil and forest
ecosystems, Atmos. Chem. Phys., 16, 13149-13172, doi:10.5194/acp-16-13149-2016, 2016.
Kajino, M., T. T. Sekiyama, A. Mathieu, I. Korsakissok, R. Périllat, D. Quélo, A. Quérel, O. Saunier, K.
Adachi, S. Girard, T. Maki, K. Yumimoto, D. Didier, O. Masson, Y. Igarashi, Lessons learned fromatmospheric modeling studies after the Fukushima nuclear accident: Ensemble simulations, data
assimilation, elemental process modeling, and inverse modeling, Geochemical Journal 52(2), 85-101,
doi:10.2343/geochemj.2.0503, 2018.
Nakajima. T., S. Misawa, Y. Morino, H. Tsuruta, D. Goto, and co-authors. Model depiction of theatmospheric flows of radioactive cesium emitted from the Fukushima Daiichi Nuclear Power Station
accident. Prog. Earth Planet. Sci. 4:2, DOI:10.1186/s40645-017-0117-x, 2017.
NRA (Nuclear Regulation Authority), Airborne Monitoring Results in each prefecture, available at
http://radioactivity.nsr.go.jp/en/list/203/list-1.html, 2012. (last access: 21 January 2019)
Oura, Y., M. Ebihara, H. Tsuruta, T. Nakajima, T. Ohara, M. Ishimoto, H. Sawahata, Y. Katsumura, W.
Nitta. A database of hourly atmospheric concentration of radiocesium (134Cs and 137Cs) in
suspended particulate matter collected in March 2011 at 99 air pollution monitoring stations in EasternJapan. J. Nucl. Radiochem. Sci., 15(2), 15-26, 2015.
Sekiyama, T. T., M. Kajino, M. Kunii. The impact of surface wind data assimilation on the predictabilityof near-surface plume advection in the case of the Fukushima Nuclear Accident, J. Meteor. Soc. Japan,
95(6), 447-454, doi:10.2151/jmsj.2017-025, 2017.
Sekiyama, T. T. and T. Iwasaki: Mass flux analysis of 137Cs plumes emitted from the Fukushima Daiichi
Nuclear Power Plant, Tellus B, 70(1), 1–11, doi:10.1080/16000889.2018.1507390, 2018.
Tsuruta, H., Y. Oura, M. Ebihara, T. Ohara, T. Nakajima. First retrieval of hourly atmospheric radionuclides
just after the Fukushima accident by analyzing filter-tapes of operational air pollution monitoring
stations. Sci. Rep., 4, 6717, doi:10.1038/srep06717, 2014.

- 26 -

4. Publications 2016-2018
出版論文 2016-2018- 27 -

Journal of Environmental Radioactivity 166 (2017) 97e103

Contents lists available at ScienceDirect

Journal of Environmental Radioactivity
journal homepage: www.elsevier.com/locate/jenvrad

Plutonium, 137Cs and uranium isotopes in Mongolian surface soils
K. Hirose a, *, Y. Kikawada a, Y. Igarashi b, H. Fujiwara c, D. Jugder d, Y. Matsumoto a, T. Oi a,
M. Nomura eaDepartment of Material and Life Sciences, Faculty of Science and Technology, Sophia University, 7-1 Kioi-cho, Ciyoda-Ku, Tokyo 102-8554, Japan
Atmospheric Environment and Applied Meteorology Research Department, Meteorological Research Institute, Tsukuba, Ibaraki 305-0052 Japanc
Soil Environment Division, National Institute for Agro-Environmental Sciences, Tsukuba, Ibaraki 305-8601, Japand
Institute of Meteorology and Hydrology, Ulaanbaatar 46, MongoliaeResearch Laboratory for Nuclear Reactors, Tokyo Institute of Technology, Ookayama 2-12-1, Meguro-Ku, Tokyo 152-8550, Japanba r t i c l e i n f o

a b s t r a c t

Article history:
Received 29 October 2015
Received in revised form
12 January 2016
Accepted 14 January 2016
Available online 30 January 2016

Plutonium (238Pu and 239,240Pu), 137Cs and plutonium activity ratios (238Pu/239,240Pu) as did uranium
isotope ratio (235U/238U) were measured in surface soil samples collected in southeast Mongolia. The
239,240
Pu and 137Cs concentrations in Mongolian surface soils (<53 mm of particle size) ranged from
0.42 ± 0.03 to 3.53 ± 0.09 mBq g1 and from 11.6 ± 0.7 to 102 ± 1 mBq g1, respectively. The
238
Pu/239,240Pu activity ratios in the surface soils (0.013e0.06) coincided with that of global fallout. The
235 238
U/ U atom ratios in the surface soil show the natural one. There was a good correlation between the239,240
Pu and 137Cs concentrations in the surface soils. We introduce the migration depth to have better
understanding of migration behaviors of anthropogenic radionuclides in surface soil. We found a difference of the migration behavior between 239,240Pu and 137Cs from 137Cs/239,240Pu e 137Cs plots for the
Mongolian and Tsukuba surface soils; plutonium in surface soil is migrated easier than 137Cs.
© 2016 Elsevier Ltd. All rights reserved.

Keywords:
239,240
Pu
238
Pu/239,240Pu activity ratio
137
Cs/239,240Pu activity ratio
235
U/238U isotope ratio
Mongolian surface soil
Migration

1. Introduction
Increasing human activities and resulting environmental effects
such as climate change alter global environment, ex, desertiﬁcation.
In order to assess environmental effects of human activities,
therefore, it is necessary to better understand what kinds of environmental change occur globally. Radionuclides and their isotope
ratios have been used as indicators of the environmental changesuch as desertiﬁcation and erosion (Hirose, 2011). Anthropogenic
radionuclides, especially plutonium and ﬁssion products, in environmental samples, which were dominantly derived from global
fallout due to the atmospheric nuclear tests during the period of the
1950s to the early 1960s (UNSCEAR, 2000), are a useful tool to solve
long-term environmental changes. After 1990, plutonium andﬁssion products (137Cs, 90Sr and others) in aerosol and deposition
samples before the Fukushima Daiichi Nuclear Power Plant accident in 2011 (Povinec et al., 2013; Hirose and Povinec, 2015), whose
major sources are dust blown by severe storms in the deserts andarid regions, are recognized as a proxy of environmental change* Corresponding author.
E-mail address: hirose45037@mail2.accsnet.ne.jp (K. Hirose).
http://dx.doi.org/10.1016/j.jenvrad.2016.01.007
0265-931X/© 2016 Elsevier Ltd. All rights reserved.

due to human activities/climate change and others (Fujiwara et al.,2007; Hirose, 2011; Hirose et al., 2003, 2004; Igarashi et al., 2005,
2009, 2011).
Plutonium has been released into the environment due to theatmospheric nuclear-weapons tests, burn-up of nuclear-battery
satellite, nuclear reactor accidents, and others (Harley, 1980).
Plutonium, which is one of chemically reactive elements, is cycled
in the globe according to physical and biogeochemical processes.
Effectively to use radionuclides as a tracer, it is important toelucidate geographical distributions of radionuclides in continental land surface. In the past 50 years, the Pu distribution inland surface across many countries has been determined andpublished (Hardy et al., 1973). However, data on the geographicaldistribution of Pu in soil of East Asian continental regions
including China and Mongolia is scarce. Zhu et al. (2002) reported
the 239,240Pu concentrations in Chinese soils estimated from the137
Cs concentrations and the global fallout 239,240Pu/137Cs ratio(0.02). Zheng et al. (2009) studied concentrations and characterization of Pu isotopes in loess collected from Gansu Province,
Northwest China, which corresponds to the downwind regions of
Lop Nor Chinese nuclear test site. Dong et al. (2010) determined
239,240
Pu concentrations and 240Pu/239Pu atom ratios in soil

98

K. Hirose et al. / Journal of Environmental Radioactivity 166 (2017) 97e103

columns in central China. Recently Bu et al. (2015) reported
comprehensive research of Pu isotopes in loess collected from
Gansu Province to elucidate effects of local fallout due to theChinese nuclear testing. Within our knowledge, however, nosystematic study has been published in the literature on the distribution of 239,240Pu in Mongolian surface soil. It is important tohave better understanding of plutonium isotopes and uranium
isotopes in Mongolian surface soil because Mongolia is a potential
source region of the Asian dust (Igarashi et al., 2011).
The atomic abundance ratio between the two primordial isotopes of uranium, 235U and 238U, shows no signiﬁcant variation innature except in fossil natural reactors (Cowan, 1976). However,uranium isotope ratios in environmental samples have altered as aresult of anthropogenic activities such as burn up of nuclear satellites (Krey et al., 1979) and use of depleted uranium (Carvalho andOliveira, 2010). As did plutonium isotopes, uranium isotopes(234U/238U activity ratios and 235U/238U isotope ratios) are a tracerof anthropogenic activities and natural processes. In fact, anomalous uranium isotope ratios (235U/238U) have been recognized infallout samples (Kikawada et al., 2009, 2012, 2015) and archived
samples (Warneke et al., 2002). In the East Asian Continent,
including atmospheric nuclear test sites, it may be likely that surface soils have been contaminated by ﬁssile materials. It is, therefore, is important to have information on uranium isotope ratios(235U/238U) in surface soils in Mongolia, where is downstream of
the Chinese nuclear test site and also source region of the Asiandust.
In this report, we describe the ﬁrst sight of plutonium concentrations and uranium isotopes in surface ﬁne soils of the southeast
Mongolia, where the desertiﬁcation is a signiﬁcant recent problem,
which suggests that the Asian dust source region may have been
expanding after the 2000s (e.g. Igarashi et al., 2011; Kido et al.,2012). We also discuss about mobility of plutonium in surface soils.
2. Sampling and measurements
Undisturbed surface soil sampling campaign was conducted insoutheastern Mongolia in fall 2007 (Igarashi et al., 2011), in whichmost of the sampling sites are located in the desert-steppe zone;
annual precipitation amounts at the sampling sites of Mongolia
ranges from about 100 to 300 mm yr1 (Igarashi et al., 2011).
Annual precipitation amounts at Tsukuba, Japan as a reference
station were in the range from 1069 to 1607 mm yr1 in the 2000s.
Sampling locations are shown in Fig. 1 and Table S1. At each site, atspots as far from vehicular trafﬁc and ruts as possible, an approximate 1-kg of undisturbed surface soil was collected to a depth of
5 cm by using a core sampler. After sample preparation treatments
including an autoclave treatment, removal of stones, pebbles,
grasses and roots by hand and homogenization, a fraction thatpassed through a 53-mm mesh (nominal), which almost corresponds to silts and/or clay, was subjected to analysis of 137Cs andplutonium because this fraction was considered as a candidate fordust suspended and transported by strong wind (Igarashi et al.,2011). After being packed into a uniform plastic container, eachsample was subjected to gamma-ray spectrometry using a Ge
semiconductor detector for 137Cs measurements. After the 137Cs
measurements, 24 soil samples were used for plutonium analysis.
The soil samples were decomposed and dissolved by strong acids.
After the radiochemical separation and puriﬁcation (Otsuji-Hatori
et al., 1996), plutonium isotopes were measured by a-spectrometry.
Ten Mongolian soil samples were also subjected to the uranium
isotope measurement. For three soil samples, we divided into threeparticle size fractions (100e500 mm, 45e100 mm and <45 mm) byusing sieve. For the measurement, the samples were ﬁrst subjectedto the acid (4 M HNO3) leaching treatments (acid soluble fraction).

Residue (acid insoluble fraction) was decomposed by HF, HClO4 andconc. HNO3 and ﬁnally dissolved in 4 M HNO3 solution. After
chemical isolation by Eichrom UTEVA resin, the uranium concentrations and the 235U/238U ratios in the acid soluble and acidinsoluble fractions were measured by ICP-QMS separately. The
obtained 235U/238U ratios in the samples were corrected by that inseawater as a conventional standard, whose isotope ratio is regarded as the natural ratio. The details of these analytical proceduresare described in Kikawada et al. (2009).

3. Results and discussion
3.1. Levels of plutonium and 137Cs in soil
Concentrations of 239,240Pu and 137Cs in Mongolian surface soils(24 samples) are summarized in Table 1. The 239,240Pu concentrations in surface soil in the southeast Mongolia ranged from
0.42 ± 0.03 to 3.53 ± 0.09 mBq g1 with an average of 1.59 mBq g1,
and from 11.6 ± 0.7 to 102 ± 1 mBq g1 with an average of
46.9 mBq g1, respectively. The results reveal that there are large
spatial variations of the 239,240Pu and 137Cs concentrations in the
Mongolian surface soils. It is likely that precipitation amounts areclosely related to the 239,240Pu and 137Cs concentrations in surface
soil because plutonium and 137Cs in surface soil were deposited by
wet and dry depositions and they tightly adsorb on soil surface
lsson et al., 2013). Igarashi et al. (2011) examined the relation(Paship between the 137Cs concentrations in the Mongolian surface
soils and annual precipitation amounts, in which they found acorrelation between the 137Cs concentrations in the surface soil andannual precipitation amounts, suggesting that the level of the 137Csconcentrations in surface soil is mainly controlled by wet deposition. These ﬁndings suggest that the spatial distribution of theMongolian surface 239,240Pu is governed by wet deposition, as did
137
Cs.
We compared with the 239,240Pu concentrations in Japaneseand Chinese surface soils. Although direct comparison between
Mongolian and Chinese surface soil 239,240Pu is difﬁcult becauseof different sample preparations, the 239,240Pu concentrations in
the Mongolian surface soils seems to be generally higher thanreported ones for the Chinese soils; their values for Gansu (Zheng
et al., 2009) and central China (Hubei) (Dong et al., 2010) were0.02e0.16 and 0.3e0.5 mBq g1, respectively. On the other hand,the 239,240Pu concentrations between the Mongolian and Tsukuba
surface soils can be compared each other because soil fractions
for Tsukuba was passed through a 53-mm mesh, as did theMongolian soils. The level of 239,240Pu in the Mongolian surface
soils was about one order of magnitude greater than that in
Tsukuba,
Japan,
ranging
from
0.019
±
0.005
to
0.41 ± 0.03 mBq g1 as an average of 0.18 mBq g1. It is noteworthy that the soil samples collected at Tsukuba contained onesin cultivated areas. The 239,240Pu concentrations in surface soil
collected in central-eastern Japan around the 1970s were determined, where includes the Kanto area locating at Tsukuba (Yanget al., 2015). We examine frequency distributions of the 239,240Pu
concentrations in surface soils of Mongolia and central-easternJapan. The frequency distributions of the soil 239,240Pu (Fig. 2)
reveal that the 239,240Pu concentrations in surface soils ofMongolia are systematically greater than that in central-easternJapan. Especially, Yang et al. (2015) reported that the low
239,240
Pu concentrations in surface soils occurred in the Kanto
area. The similar results were observed for 137Cs and 90Sr in
Mongolian surface soils; median levels of activity concentrationsof 137Cs and 90Sr in the Mongolian surface soil were 2 and 8 times
higher than those in Tsukuba, respectively (Igarashi et al., 2011).

K. Hirose et al. / Journal of Environmental Radioactivity 166 (2017) 97e10399

Fig. 1. Sampling location of surface soil in Mongolia. Numerals correspond to sampling sites shown in Table 1.

Table 1
Concentrations of239,240Pu,238Pu and137Cs in surface soil in Mongolia.

1a
2
3
4
5
6
7
8
25
32
33
34
35
36
48
49
51
53
54
56
57
59
61
63

Pu mBq g1

LatitudeLongitude

239,240

47.49547.65847.68347.70247.91647.89848.01247.76
47.24846.95346.88846.99346.96946.75746.86646.85546.77846.71146.51346.38
46.19
46.05346.03445.746110.92111.08111.39111.65112.38112.66114.58114.88115.11115.24114.93114.64114.32113.81113.08112.87112.76112.54112.31112.12111.86111.72111.45111.090.71 ± 0.04
3.53 ± 0.09
2.43 ± 0.06
1.81 ± 0.07
2.48 ± 0.07
2.42 ± 0.07
1.85 ± 0.06
1.33 ± 0.05
0.74 ± 0.04
1.71 ± 0.06
2.05 ± 0.06
2.54 ± 0.07
2.74 ± 0.07
0.99 ± 0.05
1.21 ± 0.05
1.56 ± 0.06
1.44 ± 0.06
0.80 ± 0.04
0.46 ± 0.03
0.60 ± 0.04
1.92 ± 0.07
0.42 ± 0.03
1.24 ± 0.07
1.07 ± 0.06

238

Pu mBq g1

238

Pu/239,240Pu mBq g10.078 ± 0.013
0.107 ± 0.014
0.043 ± 0.01
0.114 ± 0.015
0.031 ± 0.008
0.111 ± 0.014
0.022 ± 0.007

0.022
0.044
0.024
0.046
0.013
0.06
0.017

0.062 ± 0.01
0.076 ± 0.011
0.085 ± 0.013
0.021 ± 0.008
0.03 ± 0.008

0.03
0.03
0.031
0.021
0.025

0.035 ± 0.008
0.024 ± 0.007

0.024
0.03

0.038 ± 0.01

0.02

Cs mBq g1

Cs/239,240Pu mBq g1

137

137

20.1 ± 0.5
102 ± 174.3 ± 0.9
62.4 ± 0.8
75.2 ± 0.9
78.0 ± 0.8
63.3 ± 1.1
42.4 ± 1
22.3 ± 0.4
57.5 ± 1
65.9 ± 1.1
74.3 ± 1.2
77.2 ± 1.2
26.2 ± 0.8
26.8 ± 0.8
42.3 ± 0.5
37.5 ± 0.9
25.9 ± 0.8
11.6 ± 0.7
14.7 ± 0.8
49.7 ± 1
13.1 ± 0.4
32.3 ± 0.8
30.1 ± 0.9

28.3
28.9
30.6
34.5
30.3
32.2
34.2
31.9
30.1
33.7
32.1
29.3
28.2
26.5
22.1
27.1
26
32.2
25.4
24.6
25.9
31.1
26.1
28.1

Radioactive decay of 137Cs and 238Pu were corrected at sampling time (Oct. 2007).
a
Number of sampling location as shown in Fig. 1.

3.2. 238Pu and 238Pu/239,240Pu activity ratio in Mongolian soils
The 238Pu/239,240Pu activity ratio is a kind of ﬁngerprint to
identify sources of plutonium in the environmental samples. The
238
Pu concentrations and 238Pu/239,240Pu activity ratio in the
Mongolian soils are shown in Table 1. The 238Pu/239,240Pu activity
ratios in the Mongolian surface soil were in the range of 0.013e0.06
with an average of 0.029. The similar 238Pu/239,240Pu activity ratios
were observed in the Sahara dust (Pham et al., 2005; Mason et al.,2010). The 238Pu/239,240Pu activity ratios coincide with that in global
fallout (0.03e0.05) from the 1961e62 large-scale atmosphericnuclear tests (UNSCEAR, 1982). These ﬁnding suggests that plutonium in the Mongolian surface soils is primarily derived fromglobal fallout.
3.3. Uranium isotopes in Mongolian soil
Uranium concentrations in Mongolian soil samples, summarized in Table S2, are in the range of 0.8e2.5 mg g1. The uranium
concentrations in Mongolian soils are the same order of magnitude
as that in igneous and sedimentary rocks (0.45e3.7 mg g1) (Faure,1991) and Chinese loess (2.72e4.08 mg g1) (Ding et al., 2001; Tayler

100

K. Hirose et al. / Journal of Environmental Radioactivity 166 (2017) 97e103

climate change (Igarashi et al., 2011).

3.5. Anthropogenic radionuclides as an indicator of erosion
processes

Fig. 2. Frequency distributions of 239,240Pu concentrations in surface soils of Mongolia
and central-eastern Japan. Red colored column: Mongolian soils, blue dot column:
central-eastern Japan soils (cited from Yang et al., 2015) (For interpretation of thereferences to color in this ﬁgure legend, the reader is referred to the web version of thisarticle).

et al., 1983). The percentages of acid soluble uranium to total uranium (Table S3) ranged from 8.3 to 50%. We examined 235U/238U
isotope ratios in both acid soluble and acid insoluble fractions of theMongolian soil samples. The 235U/238U isotope ratios in both theacid soluble and acid insoluble fractions of the Mongolian soil
samples were in ranges of 0.996 ± 0.003e1.004 ± 0.003 and0.997 ± 0.006e1.002 ± 0.005 as seawater base, respectively, whichcoincide with the natural ratio within analytical uncertainty,
although the leaching proportion of uranium varied sample bysample widely. We examined 235U/238U isotope ratios in three soil
particle size fractions (100e500 mm, 45e100 mm and <45 mm) of
the Mongolian soil samples. The 235U/238U isotope ratios in threeparticle fractions of the Mongolian soils (Table S3) were0.998 ± 0.006e1.005 ± 0.012, 0.999 ± 0.015e1.005 ± 0.008, and
0.997 ± 0.011e1.001 ± 0.005 for particle ranges of 100e500 mm,
45e100 mm and <45 mm, respectively. These ﬁndings suggest thatirrespective of chemical forms and particle size, there is no fractionation of the uranium isotope ratio in the Mongolian soils.

In order to have better understanding of behavior of plutonium
in the Mongolian surface soil, it is valuable to compare between
ﬁssion products and plutonium in the surface soils because both
sources are derived from the nuclear weapons tests. The differenceof migration behaviors between 90Sr and 137Cs in the terrestrial
environment has been recognized (Forsberg and Strandmark, 2001;
Igarashi et al., 2011). We examined the relationship between
plutonium and 137Cs in the Mongolian surface soils, in which
plutonium exists as insoluble oxides and/or organic complexes
 et al., 2002) whereas 137Cs is tightly incorporated into
(Druteikiene
clay minerals (Cremers et al., 1988; Fukushi et al., 2014; Hirose et al.,2015; Kogure et al., 2012). The plot of plutonium to 137Cs in the
Mongolian soils as did Tsukuba (Japan) soils is shown in Fig. 3, in
which radioactive decay of 137Cs was corrected at each soil sampling time. The 239,240Pu concentrations in the Mongolian andTsukuba surface soils well correlated with that of 137Cs in the corresponding soils, in which correlation factors for Mongolia andTsukuba are 0.982 and 0.968, respectively. Correlation slopes for
the Mongolian and Tsukuba surface soils were 0.032 and 0.022,
respectively, in which 239,240Pu/137Cs ratios of global fallout at thesoil sampling times of Mongolia (Oct. 2007) and Tsukuba (Jan.
2000) were 0.034 ± 0.003 and 0.029 ± 0.003, respectively. The
correlation slope for the Tsukuba surface soil is smaller than that of
the Mongolian surface soil, which suggests that migration behaviors of anthropogenic radionuclides in surface soil differ between
Mongolia and Tsukuba.
To elucidate the migration behaviors of plutonium in surface
soil, we introduce a simple model, in which migration depth ofanthropogenic radionuclide in surface soil as a new concept isdeduced. The concentrations of anthropogenic radionuclides insurface soil (Crad, soil) is calculated from cumulative deposition ofanthropogenic radionuclides (Drad, soil), where subscript rad denotes individual anthropogenic radionuclide, and corresponding
migration depth (dm) as following equation.

3.4. Plutonium as an indicator of Asian dust
Atmospheric Pu deposition observed in Japan changed from the
1990s to the 2000s; typically, higher 239,240Pu deposition occurredin spring of the 2000s with higher massic concentrations of239,240
Pu than those in the 1990s (Hirose et al., 2008). As its causes,
dust events in the East Asian continent desert and arid regions
increased in the 2000s; especially, the dust storm activities
increased in the Mongolian Plateau and adjacent source regions(Kurosaki and Mikami, 2003; Lim and Chun, 2006; Kim, 2008). The
high 239,240Pu concentrations in the Mongolian soils, which cansupport the high 239,240Pu massic concentrations in deposition at
the Meteorological Research Institute (MRI), Tsukuba, suggest thatthe source region of Asian dust changed not only from the Tarim
Basin but also from the Mongolian Plateau and adjacent regions.
These ﬁndings suggest that long-term change of the deposition ofanthropogenic radionuclides reﬂects environmental changes in the
East Asian continent, corresponding to desertiﬁcation due to human activities such as over-clopping and over-grazing and/or

Fig. 3. Relationships between 239,240Pu and 137Cs concentrations in surface soils.
Closed circle: Mongolia, open circle: Tsukuba.

K. Hirose et al. / Journal of Environmental Radioactivity 166 (2017) 97e103

Crad;soil ¼ Drad;soil ðdm  rÞ

(1)

where r denotes density of soil. Now we assume that the migration
depth of anthropogenic radionuclide in soil column is expressed asa sum of radionuclide-dependent and radionuclide-independent
terms, which correspond to chemical processes such as sorptiondesorption with soil particles and physical processes such as mixing and bio-perturbation; dm ¼ dm,c,rad þ dm,p. The latter term (dm,p)is common for both radionuclides. When this relationship applies
to 137Cs and 239,240Pu in surface soil, we have a following equation.

CCs137;soil CPu;soil ¼ DCs137;soil DPu;soil þ dm;c;Pu  dm;c;Cs137 r DPu;soil CCs137;soil
137

(2)

239,240

This equation means that theCs/
Pu activity ratios insurface soil are linearly related to the 137Cs concentrations in surface soil. We examined the relationship between the 137Cs/239,240Pu
activity ratios and 137Cs concentrations in the surface soils in
Mongolia and Tsukuba, in which radioactive decay of 137Cs wascorrected at each soil sampling time. The result is shown in Fig. 4.
We have linear relationships between the 137Cs/239,240Pu ratios andthe 137Cs concentrations in the Mongolian and Tsukuba surface
soils, whose correlation factors are 0.428 and 0.878, respectively.
The correlation intercepts, which correspond to the ratio of thedepositions of 137Cs and 239,240Pu; DCs-137,soil/DPu,soil, are 26.6 ± 1.3
and 29.5 ± 3.6 for the Mongolian and Tsukuba surface soils,
respectively, which are in good agreement with 137Cs/239,240Pu ratios of global fallout at soil sampling times (29 ± 3 in 2007 and
34 ± 3 in 2000) (Hodge et al., 1996). The positive correlations
suggest that the migration depths of plutonium in land surface inMongolia and Tsukuba are deeper than those of 137Cs, which suggests that the plutonium in soil column is easier migrated than137
Cs due to the chemical erosion processes. The correlation slopes
for the Mongolian and Tsukuba surface soils are 0.055 ± 0.025 and0.108 ± 0.026, respectively. Taken into account cumulative 239,240Pu
deposition at the MRI, Tsukuba (44 Bq m2) (Hirose et al., 2001), therelative migration depth of 239,240Pu (¼(dm,c,Pu e dm,c,Cs-137)) in
Tsukuba soils is calculated to be 0.30 cm, assuming the soil densityof 1.6 g cm3 (Feng et al., 2004). The relative migration depth ofplutonium in the Mongolian surface soil is smaller than that at

101

Tsukuba, if the 239,240Pu inventory in Mongolia is similar to that at
Tsukuba. In this case, the result may mean that plutonium in the
Tsukuba surface soil is easier migrated than that in the Mongolian
surface soil. The 239,240Pu inventories in the southeast Mongolia arecalculated from the 137Cs inventories (Fujiwara: personal information) and 137Cs/239,240Pu ratio in the Mongolian surface soil. The
calculated 239,240Pu inventories in the southeast Mongolia, ranging
from 7.6 to 75 Bq m2, are the same order of magnitude as that in
China (10e546 Bq m2) (Bu et al., 2014, 2015; Dong et al., 2010; Xu
et al., 2015; Zheng et al., 2009). The large spatial variability of the239,240
Pu inventories in the southeast Mongolia causes that of theannual precipitation amount; the 137Cs concentrations in the
Mongolian surface soils correlated with the annual precipitation
amount in the corresponding location (Igarashi et al., 2011).
Although there is a large variability of the 239,240Pu inventories inthe southeast Mongolia, we found the linear relationship between
the 137Cs/239,240Pu activity ratios and the 137Cs concentrations in the
Mongolia surface soils, suggesting that the relative migration depthof 239,240Pu in surface soil may be linearly related with the 239,240Pu
inventories. Assuming that the slope of the 137Cs/239,240Pu e 137Cs
plot for the Mongolian soils is constant, the relative migration
depth of 239,240Pu in the Mongolian soils is calculated to be in arange of 0.032e0.32 cm, which is shorter than that at Tsukuba. Thisﬁnding suggests that plutonium in soil column of deserts is
migrated chemically less than in higher precipitation area such as
Tsukuba, because chemical movement of plutonium in soil columns
occurs via pore water between soil particles. Concept of the relative
migration depth is an important tool to understand the migration
behavior of radionuclides in surface soil because it is related todifference of biogeochemical conditions and soil chemistry (inorganic complexation including hydrolysis and organic complexation) in addition to difference of precipitation amounts between
the Mongolian and Tsukuba surface soils; higher precipitation
amounts in Tsukuba may lead to increase of chemical and biogeochemical activities. These ﬁndings reveal that the 137Cs/239,240Pu e
137
Cs plots are effective to characterize the migration properties ofradionuclides including chemical erosion processes and related soil
chemical properties.
Igarashi et al. (2011) revealed that there is a good positive correlation between 137Cs/90Sr activity ratios and 137Cs concentrationsin the Mongolian and Tsukuba surface soils, which indicates that90
Sr in the both surface soils is easily migrated comparing with137
Cs. Taken into account cumulative 90Sr deposition at Tsukuba(1000 Bq m2 in 2000) (Hirose et al., 2001) and the correlation
slope of the 137Cs/90Sr e 137Cs plot for Tsukuba (Igarashi et al., 2011),
the relative migration depth of 90Sr (¼(dm,c,Sr-90 e dm,c,Cs137)) in
Tsukuba soils is calculated to be 34 cm. The correlation slope for90
Sr in the Tsukuba surface soils is about 10-times deeper than thatin the Mongolian soils, whereas the correlation slope for plutonium
in the Tsukuba surface soils is about twice deeper than that in the
Mongolian soils. These ﬁndings suggest that 90Sr in the surface soils
is easier migrated than plutonium in surface soil.
4. Conclusion

Fig. 4. Relationships between 137Cs/239,240Pu activity ratios and 137Cs concentrations insurface soils. Closed circle: Mongolia, open circle: Tsukuba.

This is the ﬁrst report that provides information on plutonium
isotopes and uranium isotope ratio (235U/238U) in surface soils ofMongolia. The 239,240Pu concentrations in surface ﬁne soil (<53 mm
of particle size) of Mongolia are generally higher than that in Tsukuba. The 238Pu/239,240Pu activity ratios in Mongolian surface soil
coincide with that of global fallout, which means that most of theplutonium in the Mongolian soils originated from global fallout.
The 239,240Pu concentrations in the Mongolian soils well correlate with the 137Cs concentrations in the corresponding soils as dosoil samples at Tsukuba. The correlation slope of the 239,240Pu e

102

K. Hirose et al. / Journal of Environmental Radioactivity 166 (2017) 97e103

137
Cs plots for the Mongolian soils differs from that for the Tsukuba
soils. We introduce migration depth of anthropogenic radionuclides in surface soil, consisting of radionuclide-independent andradionuclide-dependent terms, which correspond to physical andchemical processes, respectively. In this case, a linear relationshipbetween 137Cs/239,240Pu activity ratios and the 137Cs concentrations
in surface soil is expected. This relationship is conﬁrmed for the239,240
Pu, 137Cs and 90Sr concentrations in the Mongolian and Tsukuba soils (Igarashi et al., 2011). The analysis reveals that the
migration tendency of three radionuclides in surface soil is afollowing order: 90Sr>>239,240Pu > 137Cs.
The 235U/238U ratios in the Mongolian soil samples, irrespectiveof chemical forms and particle size of soils, coincide with the natural ratio within analytical uncertainty.

Acknowledgments
The authors thank the Mongolian Institute of Meteorology andHydrology for supporting our survey. All samples were exported
and imported following all regulations and with the approval ofboth Mongolian and Japanese authorities. Sample preparation andradiochemical analyses were carried out by K. Inukai, Y. Togashi, M.
Tomita, and T. Mizo, KANSO Co. Ltd. Part of the expense for conducting the Mongolian survey and the sample analysis was coveredby the fund for Radioactivity Survey in Japan from the Ministry of
Education, Culture, Sports, Science and Technology.
Appendix A. Supplementary data
Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.jenvrad.2016.01.007.
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Journal of Environmental Radioactivity 166 (2017) 458e465

Contents lists available at ScienceDirect

Journal of Environmental Radioactivity
journal homepage: www.elsevier.com/locate/jenvrad

Temporal changes in dissolved 137Cs concentrations in groundwaterand stream water in Fukushima after the Fukushima Dai-ichi Nuclear
Power Plant accident
Sho Iwagami a, *, Maki Tsujimura b, Yuichi Onda a, Masataka Nishino c, Ryohei Konuma d,
Yutaka Abe b, Manami Hada b, Ishwar Pun b, Aya Sakaguchi e, Hiroaki Kondo e,
Masayoshi Yamamoto f, Yoshiki Miyata f, Yasuhito Igarashi gaCenter for Research in Isotopes and Environmental Dynamics, University of Tsukuba, Tsukuba 305-8577, Japan
Department of Sustainable Environmental Studies, Graduate School of Life and Environmental Sciences, University of Tsukuba, Tsukuba 305-8572, Japanc
Department of Geoenvironmental Sciences, Graduate School of Life and Environmental Sciences, University of Tsukuba, 305-8572, Japand
College of Geoscience, School of Life and Environmental Sciences, University of Tsukuba, 305-8572, Japane
Department of Earth and Planetary Systems Science, Graduate School of Science, Hiroshima University, 739-8526, Japanf
Low Level Radioactivity Laboratory, Institute of Nature and Environmental Technology, Kanazawa University, 923-1224, Japang
Atmospheric Environment and Applied Meteorology Research Department, Meteorological Research Institute, 305-0052, Japanba r t i c l e i n f o

a b s t r a c t

Article history:
Received 16 May 2014
Received in revised form
11 February 2015
Accepted 8 March 2015
Available online 11 May 2015

The concentration of dissolved 137Cs in groundwater and stream water in the headwater catchments in
Yamakiya district, located ~35 km north west of Fukushima Dai-ichi Nuclear Power Plant (FDNPP), wasmonitored from June 2011 to July 2013, after the earthquake and tsunami disaster. Groundwater andstream water were sampled at intervals of approximately 2 months at each site. Intensive sampling was
also conducted during rainstorm events. Compared with previous data from the Chernobyl NPP accident,
the concentration of dissolved 137Cs in stream water was low. In the Iboishi-yama catchment, a trend wasobserved for the concentration of dissolved 137Cs in stream water to decline, which could be divided into
two phases by October 2011 (a fast ﬂush of activity as a result of rapid washoff and a slow decline as aresult of soil ﬁxation and redistribution processes). The highest 137Cs concentration recorded at Iboishiyama was 1.2 Bq/L on August 6, 2011, which then declined to 0.021e0.049 Bq/L during 2013 (in stream
water under normal water-ﬂow conditions). During the rainfall events, the concentration of dissolved
137
Cs in stream water increased temporarily. The concentration of dissolved 137Cs in groundwater at a
depth of 30 m at Iboishi-yama displayed a decreasing trend from 2011 to 2013, with a range from
0.039 Bq/L to 0.0025 Bq/L. The effective half-lives of stream water in the initial fast ﬂush and secondary
phases were 0.10e0.21 and 0.69e1.5 y, respectively in the three catchments. The effective half-life ofgroundwater was 0.46e0.58 y at Koutaishi-yama and 0.50e3.3 y at Iboishi-yama. The trend for theconcentration of dissolved 137Cs to decline in groundwater and stream water was similar throughout
2012e2013, and the concentrations recorded in deeper groundwater were closer to those in stream
water. The declining trend of dissolved 137Cs concentrations in stream water was similar to that of theloss of canopy 137Cs by throughfall, as shown in other reports of forest sites in the Yamakiya district.
© 2016 Published by Elsevier Ltd.

Keywords:
Dissolved 137Cs
Fukushima Dai-ichi Nuclear Power Plant
Stream water
Groundwater
Rainstorm event

1. Introduction
The Fukushima Dai-ichi Nuclear Power Plant (FDNPP) accident,
following the earthquake and tsunami disaster on March 11, 2011,* Corresponding author.
E-mail address: sho.iwagami@gmail.com (S. Iwagami).
http://dx.doi.org/10.1016/j.jenvrad.2015.03.025
0265-931X/© 2016 Published by Elsevier Ltd.

resulted in a substantial release of radionuclides to the environment. The 137Cs atmospheric deposition rate observed at theMeteorological Research Institute, Tsukuba, Japan, showed that theatmospheric deposition rate of 137Cs was less than 0.1 Bq/m2/
month during the 2000s, whereas it increased to approximately
130 Bq/m2/month after the Chernobyl reactor accident in April
1986 (Aoyama et al., 1991; Igarashi, 2009). In contrast, the present
research areas have received far heavier local fallout of

S. Iwagami et al. / Journal of Environmental Radioactivity 166 (2017) 458e465

0.3e1.0 MBq/m2 of radiocesium (134Cs and 137Cs) from the FDNPP
accident (MEXT, 2011). Most of the area in which the radionuclides
were deposited was forested (Hashimoto et al., 2012). Studies offorested areas in Fukushima have shown that radionuclides areremaining in the forest. For example, Kato et al. (in this issue)
showed the transfer of canopy-intercepted radiocesium to theforest ﬂoor by measuring throughfall, stemﬂow, and litterfall.
A radionuclide proﬁle of the soil has been reported in Kato et al.
(2012) and Takahashi et al. (2014). Takahashi et al. (2014) showeda vertical distribution of 137Cs in soil proﬁles under eight different
land uses including forests.
Generally, radiocesium that reaches the ground surface will be
rapidly absorbed into the surface soil (Walling, 1998). It has beenreported that 137Cs will readily absorb into “frayed edge sites” onclay minerals (Cremers et al., 1988). Tanaka et al. (2014) studied theparticle-size-dependent distribution of radiocesium in ﬂuvial suspended sediments in Fukushima after the FDNPP accident. They
reported that smaller particles had higher radiocesium concentrations, reﬂecting their larger speciﬁc surface areas. The contributionof clay-size fractions, which had the highest 137Cs concentration,
was small because of their low frequency. The downward migrationof 137Cs in the soil is reported to be slow. The 137Cs derived from the
Chernobyl accident is reported to migrate downward by less than
1 cm/y, and therefore more than 80% of 137Cs remains in the uppern et al., 1999). Plant roots
10 cm of the soil (Arapis et al., 1997; Rosealso extract dissolved 137Cs from the soil (Bunzl et al., 1999; Prorok
et al., 2006). Kliashtorin et al. (1994) reported that the washout ofradiocesium with soil water inﬁltration from the forest litter
accounted for only a small portion of the total radiocesium in thearea within 30 km of the Chernobyl NPP. The rate of transport ofradiocesium from the soil to groundwater and stream water may
depend on site-speciﬁc conditions as well as the physical andchemical properties of the deposited radiocesium.
We are focusing on the process whereby the radionuclides
deposited to the forested headwater catchment migrate to thegroundwater and stream water, from inﬁltration to discharge.
Cesium-137 contamination of groundwater intended for domestic
use and that of river water intended for agricultural use is of great
concern for the public. A trend for the concentration of dissolved
137
Cs to decline in river water after the Chernobyl accident wasn and Grip,
shown in a two-component exponential model (Nyle1997; Smith et al., 2000, 2002) and a three-component expon andnential model (Monte, 1997; Smith et al., 2004, 2005). Nyle
Grip (1997) reported that the concentration of 137Cs in stream
water in northern Sweden declined from approximately
20e0.1 Bq/L within 8 y after the Chernobyl NPP accident. The IAEA
(2006) reported that the observed 137Cs concentration in shallow
groundwater declined from approximately 8 Bq/L to less than
0.1 Bq/L from 1989 to 1999 in the Red Forest area near the Chernobyl accident site.
Several studies have reported on the transport of dissolved
radiocesium in river water in Fukushima (Sakaguchi et al., 2012;
Yasutaka et al., 2012; Ikeuchi, 2013). Yasutaka et al. (2012)
focused on Abukuma River, the riverine system of which broadly
covers a large part of the northwest of the contaminated FDNPP
area. Sakaguchi et al. (2012) reported that Kuchibuto River, which isone of the branches of the Abukuma River and is located in the
headwater area, showed higher 137Cs values (0.46 Bq/kg) than the
Abukuma River (0.25 Bq/kg). However, there is little information
about the headwater catchment. Because a large part of the area
affected by FDNPP fallout is forested, the forested area could be thesource of dissolved 137Cs in the local river systems. Therefore,
detailed studies of the forested headwater catchment are required.
Additionally, there is no report on the dissolved 137Cs in the
groundwater at the headwaters of the forested catchment. As the459transport of radionuclides through subsurface ﬂow is of great
concern around the Chernobyl NPP (Levchuk et al., 2012), it is also
important to evaluate the transport of radiocesium throughgroundwater in the Fukushima area.
There are reports of ﬂuctuations in radiocesium concentrationsin the river water during rainfall events. For example, Nagao et al.
(2013) measured the concentration of total (dissolved and suspended) radioactivity of 137Cs in river water during heavy rainfall
associated with a typhoon. The total 137Cs radioactivity in the riverwater increased during the high-water-level period as a result of
n and Grip
increased suspended contributions. In addition, Nyle
(1997) suggested that the 137Cs discharge increase was caused by
the occurrence of surface runoff. These reports assessed both dissolved and suspended radiocesium, and highlighted that radiocesium migration mainly occurs by suspended sediments (Ueda
et al., 2013).
The objective of this study was to investigate the trend anddistribution of dissolved 137Cs concentrations in groundwater andstream water in the headwater region of the Yamakiya district inFukushima, to clarify the process of dissolved 137Cs transport
through the water ﬂow path from a hydrological perspective. The
concentrations of dissolved 137Cs are also important data that canbe used in simulation models for the prediction of future radiocesium distributions.
2. Study area and methods
2.1. Study area
Hydrometric observations and dissolved radiocesium
concentrations were recorded in three small catchments in
Yamakiya District, Nihonmatsu City, Fukushima Prefecture,
Japan: Koutaishi-yama (KOU, 0.54 km2, 94% forested: 72% conifer,
22% broad-leaf), Iboishi-yama (IBO, 0.17 km2, 71% forested: 44%
conifer, 27% broad-leaf) and Ishidaira-yama (ISH, 0.075 km2, 92%
forested: 13% conifer, 79% broad-leaf). The catchments arelocated ~35 km northwest of FDNPP and received heavy local
fallout (0.30e1.0 MBq/m2 of 137Cs) after the FDNPP accident(MEXT, 2011). The average 137Cs deposition density in the KOU,
IBO and ISH catchment areas was 916, 544, and 298 kBq/m2
respectively (MEXT, 2011). The catchments are located at the
headwater of the Abukuma River, which is the major river inFukushima Prefecture and is an important industrial and domestic water resource. The study area and the location of thethree observation sites are shown in Fig. 1. Contour lines of 137Cs
deposition are also shown, which were obtained by aircraft
monitoring by the Ministry of Education, Culture, Sports, Scienceand Technology (MEXT, 2011).
2.2. Stream water sampling and observation
Sampling of stream water was conducted from June 6, 2011 toJuly 21, 2013. Stream water was sampled directly at gauging stations during each ﬁeld survey. The water level in the streams wasrecorded using a water level recorder (TruTrack WT-HR500, Intech
Instruments Ltd., Christchurch, New Zealand) installed on a Parshall ﬂume (Takeuchi Ironworks, Japan). Stream water discharge
was calculated based on the relationship between water level andmanually measured discharge. Rainfall was measured in 0.2-mm
increments by a tipping-bucket rain gauge (Davis #7852M, CA,
USA) close to the stream water observation point at IBO. During a
rainfall event, stream water was sampled using an automated water
sampler (Sigma SD900, HACH, Loveland, CO, USA) at intervals of afew hours.

460

S. Iwagami et al. / Journal of Environmental Radioactivity 166 (2017) 458e465

Fig. 1. Study area and locations of three observation sites with the contour line of 137Cs deposition obtained by aircraft monitoring (Ministry of Education, Culture, Sports, Scienceand Technology http://ramap.jmc.or.jp/map/).

Groundwater was sampled from boreholes, which were instrumented on August 31, 2011 at IBO and KOU. The boreholes at IBO were
170 m upslope from the stream water sampling point, while those at
KOU were 55 m upslope from the stream water sampling point.
The structure (depth and range of the screen for water intake) of
the borehole and the average water level at IBO and KOU during theobservation period are shown in Fig. 2. The diameter of the borehole
was 0.065 m. The upper part of the borehole was sealed with cement
and the ﬁnal part of the screen was made from silica sand. The depthsof the boreholes at IBO were 5, 15, and 30 m (the screen depth of eachborehole was 2e5, 5e15, and 20e30 m, respectively). The depths ofthe boreholes in KOU were 3, 6, and 20 m (the screen depth of eachborehole was 2e3, 3e6, and 12e20 m, respectively).
The groundwater was sampled using a portable water pump
(DIK-665A, Daiki Rika Kogyo Co., Ltd., Japan). During groundwater
sampling, care was taken not to contaminate the groundwater inthe borehole with material from the surrounding environment.
Fresh groundwater samples were collected after pumping out thepooled water in the borehole. Fresh groundwater was distinguishedbased on the change of water temperature and the amount of thepumped out ponded water. The groundwater level was monitoredcontinuously using a water level recorder (HOBO CO-U20-001-01,
Tempcon instrumentation Ltd., UK), which was suspended by awire rope inside each borehole.

Tsukuba (a 40% coaxial type Ge detector, GC4019, Canberra Industries Inc., Meriden, CT, U.S.A.) and Kanazawa University (a 40%
closed-end coaxial type Ge detector, GC4019, Canberra Industries,
Inc., Meriden, CT, U.S.A). The spectrometers were calibrated with (i)a standard prepared by the Low Level Radioactivity Laboratory ofKanazawa University using NBL (New Brunswick Laboratory)
Reference Material No.42-1 (4.04% U), (ii) a g-ray certiﬁed referencematerial (JSAC 0471) including 134Cs and 137Cs, from the JSAC (theJapan Society for Analytical Chemistry), and (iii) analytical grade(KCl). Early phase samples taken just after the FDNPP accident wereanalyzed at the MRI. Most of the samples were analyzed at
Kanazawa University to detect low-level dissolved 137Cs at lowbackground facility (Low Level Radioactivity Laboratory, Kanazawa University: Komura and Hamajima, 2004).
Samples with low-concentrations of 137Cs were concentratedusing the ammonium molybdophosphate (AMP)/Cs compound
method (Tanaka et al., 2006) at Hiroshima University. A portion of
the AMP powder was dissolved with 1.25% tetramethylammoniumhydroxide (TAMAPURE-AA) to determine the concentration of 133Cs
as a yield tracer. To measure 133Cs in the solubilized AMP samples,an ICP mass spectrometer (Agilent 7700: Agilent Technologies,
Santa Clara, CA, USA) was used. In this case, rhenium (10 ng/ml) was
employed as an on-line internal standard. The chemical yields forall of the processes were in the range 95e98%, and the concentration of 137Cs was recalculated with chemical yield. The physicaldecay of 137Cs was corrected to the sampling date.

2.4. Analysis of dissolved 137Cs

3. Results and discussion

All water samples were ﬁltered through a 0.45-mm pore-size
membrane (HA type, Millipore, Billerica, MA, USA). The 137Cs concentration was determined using g-ray spectrometry with agermanium semiconductor detector at the Meteorological ResearchInstitute (GEM-90205-P, Seiko EG&G Co., LTD, Japan), University of

3.1. The concentration of dissolved 137Cs in stream water

2.3. Groundwater sampling and observation

A time series of the concentration of dissolved 137Cs in stream
water at the three sites (KOU, IBO, and ISH) is shown in Fig. 3. The
dissolved 137Cs concentration of stream water in normal ﬂow

S. Iwagami et al. / Journal of Environmental Radioactivity 166 (2017) 458e465

461

condition is shown in Table S1 in supplementary materials. The
concentration of dissolved 137Cs in stream water was high in mid2011 (the highest concentration at each observation site was
1.1 Bq/L at KOU on June 25, 2011, 1.2 Bq/L at IBO on August 6, 2011,
and 0.56 Bq/L at ISH on August 17, 2011) and a rapidly declining
trend continued until the end of 2011. During 2012e2013, theconcentration of dissolved 137Cs in stream water decreased slightly
at each site. During normal water-ﬂow conditions, the majority ofdissolved 137Cs concentrations in the stream water ranged from0.02 to 0.08 Bq/L after 2012, whereas some measurements from
KOU and ISH showed concentrations of less than 0.01 Bq/L. On theother hand, the concentration of dissolved 137Cs in stream water
was comparatively high (up to one order of magnitude higher)
during rainfall events. The concentration of dissolved 137Cs in
stream water at KOU was lower than that at the other two catchments despite the higher initial 137Cs deposition density.
Kato et al. (in this issue) and Loffredo et al. (2014) reported thatthe 137Cs ﬂux from canopy to forest ﬂoor by throughfall, stemﬂow
and litterfall is higher in coniferous forests compared to broad-leaf
forests. However, in this study, the forest composition may not have
had an effect on the dissolved 137Cs concentrations in the stream
water. At the KOU site, 6% of the catchment is grassland that wasused as pasture before the NPP accident, but the stream water

Fig. 2. The structure and average water level at each borehole at Iboishi-yama andKoutaishi-yama.

Fig. 3. Time series of dissolved 137Cs concentrations in stream water at three sites(Koutaishi-yama, Iboishi-yama, and Ishidaira-yama). Error bar shows a standard deviation from counting statistics. Three arrows show rainfall events.

Fig. 4. Time series of dissolved 137Cs concentrations in stream water at Iboishi-yama
during (a) Event 1209, (b) Event 1210 and (c) Event 1307 with the hydrograph. Errorbar shows a standard deviation from counting statistics.

462

S. Iwagami et al. / Journal of Environmental Radioactivity 166 (2017) 458e465

passes through the forested area. In addition, at the IBO site, 21% of
the catchment is grassland that exists close to the spring point,
while at 8% of the catchment is grassland at ISH. Surface runoff will
rarely occur in a forested area except in an untreated artiﬁcial forest
with a hard bare soil surface and low inﬁltration capacity (Gomiet al., 2008). The USLE plot was constructed in Yamakiya town
(also at the KOU, IBO, and ISH sites) for monitoring soil erosion and137
Cs discharge from various land uses related to soil erosion(Yoshimura et al., 2014). The water discharge from the USLE plot
has been monitored and only occurred during rainfall events. Basedon the amount of discharge from the USLE plot, 3% of the surface
runoff from the grassland area in the catchment is assumed to be
stream water discharge from the IBO site, which includes thelargest area of grassland. Usually the discharge from an USLE plot
will not occur or negligibly small. Therefore, almost all of thestream water was assumed to originate from the spring, and theland use in this case is not considered to be a determining factor ofthe dissolved 137Cs concentration in stream water.

that the site is a discharge area of the groundwater ﬂow system. In
contrast, the hydraulic gradient in the borehole at IBO was downward, which suggests that IBO is a recharge area for the groundwater ﬂow system.

3.2. The concentration of dissolved 137Cs in stream water during
rainfall events
A time series of the concentration of dissolved 137Cs in stream
water at IBO during rainfall events in the periods (a) September 30- October 1, 2012 (Event 1209, total precipitation 47 mm), (b)October 17e19, 2012 (Event 1210, total precipitation 30 mm), and(c) July 13e19, 2013 (Event 1307, total precipitation 234 mm) is
shown with a hydrograph in Fig. 4. The dissolved 137Cs concentration of stream water in IBO during rainfall event is shown in
Table S2 in supplementary materials. The ﬂuctuation of dissolved
137
Cs concentration in stream water was 0.02e0.5 Bq/L during
Event 1209, 0.02e0.1 Bq/L during Event 1210, and 0.03e0.5 Bq/L
during Event 1307 respectively. During Event 1209, the highest
concentration of dissolved 137Cs in stream water was observed
before the start of rainfall, whereas during Event 1210 and Event
1307 an increase in the concentration of dissolved 137Cs in stream
water corresponded with the rise in discharge. During Event 1210,
the peak value of the concentration of dissolved 137Cs in stream
water almost corresponded to the discharge peak.
During Event 1209, an increase in the concentration of dissolved
137
Cs in stream water occurred before the rise in the water level.
The increase in the concentration of dissolved 137Cs could beelution from organic matter or suspended sediment. To understandthe timing of elution, intense sampling of organic matter or suspended sediment over a short time period is required. It is important to characterize where, when, and how dissolved 137Cs willenter the water ﬂow throughout the rainfall-runoff process. Event
1209 was a typhoon event, which was characterized by strong
winds and a spike of very intense rainfall, whereas Event 1307 wasa seasonal rainfall event, which was characterized by a weakerrainfall intensity and a lack of strong wind. It is possible that thisdifference in the characteristics of the rainfall events affected thetiming of the high concentration of dissolved 137Cs in stream water.
Since groundwater and soil water could contribute to stream waterby piston ﬂow, to investigate where, when, and how dissolved 137Cs
will be attached to the stream water ﬂow, it would be desirable tocombine inorganic ion concentrations and stable isotope ratio data.
3.3. Concentrations of dissolved 137Cs in groundwater
In Fig. 2, it can be seen that the groundwater level observed in
the deeper borehole was higher than the groundwater level in theshallower borehole throughout the observation period from midJuly to mid-August at KOU. Therefore, an upward hydraulicgradient was dominant around the borehole at KOU, suggestingFig. 5. Time series of dissolved 137Cs concentrations in stream water in normal ﬂow
condition and groundwater at (a) Koutaishi-yama, (b) Iboishi-yama, and only stream
water at (c) Ishidaira-yama with exponential regression lines. Error bar shows astandard deviation from counting statistics.

S. Iwagami et al. / Journal of Environmental Radioactivity 166 (2017) 458e465

where C (Bq/L) is the 137Cs concentration, a, b, g (Bq/L) and k1, k2, k3(y1) are empirically determined (radionuclide-speciﬁc) constants,
and t is time after the deposition. The k values may be expressed as
effective half-lives (Teff). The three exponential terms represent: (1)the fast ﬂush of activity as a result of rapid washoff (~1 y after
deposition), (2) a slow decline as a result of soil ﬁxation andredistribution processes (1e8 y), and (3) the very long-term equilibrium situation, respectively (Smith et al., 2005). Kato et al.
(in this issue) mentioned that the two-component exponential
model could well estimate 137Cs loss from the canopies bythroughfall, stemﬂow, and litterfall compared to the single exponential or offset exponential model, though all models could not
follow seasonal changes in the 137Cs loss. Loffredo et al. (2014)
showed the declining trend of fallout densities of 137Cs inthroughfall and stemﬂow, which indicated the inﬂection point of
the double exponential ﬁtting line to be around October 2011. The
samples for those measurements (Kato et al., in this issue; Loffredo
et al., 2014) were taken in forest sites within the study area of thisreport. In Fig. 5, the trend for the concentration of dissolved 137Cs todecline in stream water was divided into two phases from October
2011 (two-component exponential model), which is because of theclear change in the declining trend and because it coincided withthe end of the typhoon season in 2011. The concentration of dissolved 137Cs in groundwater is shown by a single exponential line
due to the lack of data for the fast ﬂush in 2011. Also there might be
no fast ﬂush phase in the trend for the concentration of dissolved
137
Cs in groundwater, however it cannot be conﬁrmed. The trend
for the concentration of dissolved 137Cs to decline in stream water
during the second phase was similar to that observed for groundwater. The Teff values for stream water in the initial fast ﬂush andsecondary phase were 0.10e0.21 and 0.69e1.5 y, respectively in thethree catchments. The Teff values for groundwater was 0.46e0.58 y
at KOU and 0.50e3.3 y at IBO. Parameters describing the dissolved
137
Cs concentration in stream water and groundwater in twocomponent exponential model and Teff value obtained from theparameters for three catchments are shown in Table 1. In contrast,
the Teff values for stream water in the initial and secondary phases
in Europe after the Chernobyl accident were approximately 1e3
and 7e8 y, respectively (Smith et al., 2005).
Smith et al. (2004) reported that the concentrations of radionuclides in water sources were strongly correlated with the area
covered by inland water. For Chernobyl, they showed that inland
water covered up to 20% of the land area studied, whereas in thearea considered in this study, the land covered by inland water was
extremely small, because the study area is at a headwater region.
This might be a reason for the short Teff of the stream water in thisstudy. The declining trend of dissolved 137Cs concentrations inTime series of the concentrations of dissolved 137Cs in stream
water in normal water ﬂow at all sites and those in groundwater atthree different depths at KOU and IBO, are shown with an exponential regression line in Fig. 5. The dissolved 137Cs concentration ofgroundwater in KOU is shown in Table S3 and dissolved 137Csconcentration of groundwater in IBO is shown in Table S4 in supplementary materials. The declining trend of groundwater was lessclear compared to stream water, but exponential regression lines
are shown to assess the effective half-live value. At both KOU and
IBO, the concentration of dissolved 137Cs in groundwater decreased
throughout the observation period. The concentration of dissolved
137
Cs in groundwater at KOU (depth of 20 m) was 0.013 Bq/L on
October 8, 2011, but decreased to 0.0011 Bq/L by January 18, 2013.
The concentration of dissolved 137Cs in groundwater at IBO (depthof 30 m) was 0.039 Bq/L on October 8, 2011, but decreased to
0.0025 Bq/L by January 18, 2013. The concentration of dissolved
137
Cs in groundwater was one order of magnitude lower than in thestream water. At both sites, the concentration of dissolved 137Cs indeeper groundwater was closer to that of stream water; the concentrations of dissolved 137Cs in groundwater at IBO, for example,
on June 22, 2012 at depths of 5, 15, and 30 m were 0.0039, 0.0018,
and 0.011 Bq/L, respectively. The deeper groundwater could
potentially contribute to the stream water, because upward or
horizontal groundwater ﬂow was dominant, as indicated by thehigher average groundwater level in the deeper boreholes (Fig. 2).
Contamination during sampling (from something on the groundthat entered the borehole) is a major concern for measurements ofdissolved 137Cs in groundwater. However, even if contamination ispresent, Cs can easily be adsorbed to the soil surrounding the insidewall of the borehole. The groundwater was ﬂuctuating dynamicallyand the concentration of dissolved 137Cs displayed a similar
declining trend to that of stream water. Therefore, in this study,
contamination was considered unlikely to be a serious problem
when measuring the concentration of dissolved 137Cs ingroundwater.
3.4. The trend for the concentration of dissolved 137Cs to decline ingroundwater and stream water
A trend for the concentration of dissolved 137Cs to decline inriver water after the Chernobyl accident was shown in a twon and Grip, 1997; Smith
component exponential model (Nyleet al., 2000, 2002) and in a three-component exponential model(Monte, 1997; Smith et al., 2004).

CðtÞ ¼ aek1t þ bek2t þ gek3t

463

(1)

Table 1
Parameters describing the dissolved 137Cs concentration in stream water and groundwater in two-component exponential model and Teff value obtained from theparameters for three catchments.
Catchment sample

Fast ﬂush

Secondary phase
Teff (y)b (Bq/L)k2 (y1)

Teff (y)

14
e
e
e

0.10
e
e
e

0.11
0.015
0.0073
0.023

1.0
1.4
1.2
1.5

0.69
0.50
0.58
0.46

7.8
e
e
e

11
e
e
e

0.15
e
e
e

0.24
0.012
0.0027
0.023

1.0
1.4
0.21
1.1

0.69
0.50
3.3
0.63

1.1

4.9

0.21

0.094

0.46

1.5

a (Bq/L)

k1 (y

KOU
Stream water
Groundwater at 3 m depth borehole
Groundwater at 6 m depth borehole
Groundwater at 20 m depth borehole

20
e
e
e

IBO
Stream water
Groundwater at 5 m depth borehole
Groundwater at 15 m depth borehole
Groundwater at 30 m depth borehole
ISH
Stream water

1

)

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S. Iwagami et al. / Journal of Environmental Radioactivity 166 (2017) 458e465

stream water is similar to that of the loss of canopy 137Cs bythroughfall, as shown by Kato et al. (in this issue) at a forest site in
the Yamakiya district. Therefore, there may be a relationship between dissolved 137Cs concentration in stream water and 137Cs
activity in throughfall at the forest site. The concentrations of dissolved 137Cs in this study were low compared to those recorded inEurope after the Chernobyl accident. Monitoring of radionuclides in
the aqueous phase has been performed for approximately 20 y after
the Chernobyl NPP accident (IAEA, 2006). It will also be necessary
to continuously monitor radionuclide concentrations in the subsurface and surface water in the area affected by the FDNPP
accident.
4. Conclusions
The concentration of dissolved 137Cs in groundwater and stream
water in Yamakiya district, approximately 35 km North West of
Fukushima Dai-ichi Nuclear Power Plant, was monitored from June
2011 (August 2011 for groundwater) to July 2013. A trend for theconcentration of dissolved 137Cs to decline in stream water wasobserved and could be divided into two phases, consisting of a fast
ﬂush and a slow decline. The highest concentration of dissolved
137
Cs in stream water was 1.2 Bq/L at Iboishi-yama on August 6,
2011, which then declined to 0.021e0.049 Bq/L during 2013 (instream water under normal ﬂow conditions). The effective halflives of stream water in the initial fast ﬂush and secondary phases
were 0.10e0.21 and 0.69e1.5 y, respectively in the three catchments. The effective half-life of groundwater was 0.46e0.58 y at
Koutaishi-yama and 0.50e3.3 y at Iboishi-yama. During rainfall
events, the concentration of dissolved 137Cs in stream water
temporarily increased. The concentration of dissolved 137Cs ingroundwater displayed a decreasing trend throughout 2011e2013,
decreasing from 0.039 Bq/L to 0.0025 Bq/L at a depth of 30 m at
Iboishi-yama. The trend for the concentration of dissolved 137Cs ingroundwater and stream water to decline was similar during
2012e2013. To gain a better understanding of the behavior andpotential impacts of dissolved 137Cs, it is necessary to combine
these results with other water quality data, such as the concentration of inorganic ions or stable isotopes, and to undertake further
analyses, such as hydrograph separation.
Acknowledgments
This work was conducted as a commissioned study from the
JAEA, as a part of the MEXT-funded FY2012 project entitled, ‘The
establishment of the grasp method to determine long-term effects
caused by radioactive materials from the Fukushima Dai-ichi Nuclear Power Plant accident’.
Appendix A. Supplementary data
Supplementary data related to this article can be found at http://dx.doi.org/10.1016/j.jenvrad.2015.03.025.
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ᨺᑕᛶ≀㉁䛾኱Ẽỿ╔䞉ᣑᩓ䛚䜘䜃㝣ᇦ䛛䜙䛾෌ᾋ㐟䛻䛴䛔䛶㻌㻌஬༑ᔒᗣே㻌 㻌 Ẽ㇟◊✲ᡤ䞉⎔ቃ䞉ᛂ⏝Ẽ㇟◊✲㒊ࡣࡌࡵ࡟㸸㻌 ᮏ✏䛷䛿ᖹᡂ  ᖺ䛛䜙㛤ጞ䛧䛯᪂Ꮫ⾡㡿ᇦ◊✲䛂⚟ᓥཎⓎ஦ᨾ䛻䜘䜚ᨺฟ䛥䜜䛯ᨺᑕᛶ᰾✀䛾⎔ቃᣲື䛻㛵䛩䜛Ꮫ㝿ⓗ◊✲䠄,6(75䠗㡿ᇦ௦⾲䠖 ᜠ⏣⿱୍⟃Ἴ኱ᩍᤵ䠅䛃䛾 $ ィ⏬◊✲⌜䛂ᨺᑕᛶ≀㉁䛾኱Ẽỿ╔䞉ᣑᩓ㐣⛬䛚䜘䜃㝣㠃┦஫స⏝䛾⌮ゎ䛃䠄$ ⌜䠅䛻䛚䛡䜛◊✲䛾඲యᴫ␎䜢⤂௓䛩䜛䠊ྠ⌜䛷䛿䠈䠍䠅⚟ᓥ஦ᨾ䛻䜘䜚኱Ẽ୰䛻ᨺฟ䛥䜜䛯ᨺᑕᛶ᰾✀䛜䛹䛾䜘䛖䛺≀⌮䞉໬Ꮫᛶ≧䛷኱Ẽ୰䜢ᾋ㐟䛧䛯䛛䠈䛚䜘䜃䛭䛾ᡂᅉ䜢⌮ゎ䛩䜛䛣䛸䠄୍ḟᨺฟ䠅䠈䠎䠅ᗈᇦⓗ䛺ᨺᑕᛶ≀㉁⃰ᗘ䛾᫬㛫䞉✵㛫ⓗኚືㄪᰝ䜔ỿ╔≀䛾 ᐃ䜢㏻䛨䠈኱Ẽ㍺㏦䞉ᣑᩓ䝰䝕䝹䛾⢭ᗘྥୖ䛸䠈䛭䜜䜢⏝䛔䛯஦ᨾᨺฟ㔞䛾᥎ᐃ䛻ᐤ୚䛩䜛䛣䛸䠈䠏䠅㝣ᇦ䠄ᅵተ䞉⏕ែ⣔䜢ྵ䜐䠅䜈䛾ỿ╔䛸኱Ẽ୰䜈䛾ᨺᑕᛶ᰾✀䛾෌㣕ᩓ㔞䠈䛭䛾≀⌮䞉໬Ꮫᙧ≧䛚䜘䜃ᨺฟ※䠄஧ḟᨺฟ䠅䜢⌮ゎ䛧䠈኱Ẽ䇷㝣㠃┦஫స⏝䜢ᐃ㔞໬䛩䜛䛣䛸䜢┠ⓗ䛸䛧䛯䠊䛣䜜䜙䜢㏻䛨䛶䠈㡿ᇦ◊✲඲య䛜┠ᶆ䛸䛩䜛ᨺᑕᛶ≀㉁䛾኱Ẽ䇷㝣ᇦ䇷Ỉᅪ䛾⛣⾜䝰䝕䝹䛾☜❧䜔䠈ྛᆅ䛷䛾࿧྾䛻䜘䜛ෆ㒊⿕䜀䛟㔞䛾᥎ᐃ䛾ᇶ♏䝕䞊䝍䛻䜒ᐤ୚䛩䜛䛣䛸䜢┠ᣦ䛩䛸䛸䜒䛻䠈ேయ䛾ෆ㒊⿕䜀䛟䛻ಀ䜟䜛኱Ẽᨺᑕ⬟䛾᝟ሗ䜢♫఍䛻ᥦ౪䛧䠈ᅜẸ䛾Ᏻ඲䞉Ᏻᚰ䛾☜ಖ䛻㈉⊩䛩䜛䛣䛸䜢┠ᣦ䛧䛯䠊◊✲ࡢ⫼ᬒ㸸 ཎᏊຊⓎ㟁ᡤࡢ஦ᨾ࡟ࡼࡿ⎔ቃࡸ೺ᗣᙳ㡪ࢆ⪃࠼ࡿ࠺࠼࡛኱Ẽ⛉ᏛࡢほⅬ࠿ࡽ㔜せ࡞ࡇ࡜ࡣ㸪ᨺฟ㸪㍺㏦࣭ᣑᩓ㸪ỿ╔ࡢ㔞ⓗ࡞㛵ಀࡢṇ☜࡞ᢕᥱ࡛࠶ࡿ㸬ࡇࡢ┠ⓗ㐩ᡂࡢࡓࡵ࡟ࡣ㸪ほ ࡜ࣔࢹࣝࡑࢀࡒࢀ୍㎶ಽ࡛ࡣ࡞ࡃ㸪୧⪅ࢆ᭷ᶵⓗ࡟㐃ᦠࡉࡏ஫࠸ࡢ▷ᡤࢆ⿵࠸㸪㛗ᡤࢆᙉ໬ࡍࡿࡇ࡜ࡀ኱ษ࡟࡞ࡿ㸬ほ࡛ᚓࡽࢀࡿࡢࡣ㸪᫬⣔ิ࠿㸰ḟඖ㸦ሙྜ࡟ࡼࡗ࡚ࡣ㸱ḟඖ㸧ࡢศᕸ࡛࠶ࡗ࡚㸪ࡑࢀ௨ୖࡣᴟᗘ࡟ᅔ㞴࡜࡞ࡿ㸬௚᪉㸪ࣔࢹ࡛ࣝࡣື⏬ࢆᐜ᫆࡟ᚓࡽࢀࡿࡀ㸪ᮍ▱ࡢࣉࣟࢭࢫࡀ࠶ࡗࡓሙྜ㸪⪃៖୙⬟࡛࠶ࡿ㸬ᚑࡗ࡚㸪ᨺฟ㸪㍺㏦࣭ᣑᩓ㸪ỿ╔࡟࠾࠸࡚୺せ࡞ࣉࣟࢭࢫࢆゎ᫂ࡋ㸪᪂つ࡟ࢫ࣮࣒࢟ᑟධࡍࡿ➼ࡋ࡚ࣔࢹࣝ
ࡢ⢭ᗘྥୖࡶᅗࡾࡘࡘ㸪෌⌧ィ⟬ࢆ㐍ࡵࡿࡇ࡜ࡀ⫢せ࡛࠶ࡿ㸬ࡇࢀࡽࡣ௚ࡢேⅭ㉳※ởᰁၥ㢟඲⯡࡟ࡘ࠸࡚ࡶᡂࡾ❧ࡘࡀ㸪ᨺᑕᛶ≀㉁࡟ࡼࡿ⎔ቃởᰁ࡛ࡶඹ㏻ࡍࡿ>@>@>@㸬ࡉࡽ࡟㸪ᨺᑕ⬟㔞࡛⾲ࡍ࡜⚟ᓥ➨୍ཎⓎ஦ᨾ࡟ࡼࡿᨺᑕᛶ≀㉁ࡢᨺฟ㔞ࡣ 3%T ࡜ᴟࡵ࡚኱ࡔࡀ㸪≀㉁㔜㔞࡟᥮⟬ࡍࡿ࡜ࢃࡎ࠿ᩘ NJ ࡟ࡋ࠿࡞ࡽࡎ㸪᭱኱ᨺฟ㔞࡜࡞ࡿ &V ࡟ࡋ࡚ࡶ㸪ࡏ࠸ࡐ࠸ NJ㸦ホ౯࡟ࡼࡾᨺฟ㔞ࡣ␗࡞ࡿࡀ㸪᭱኱ホ౯್࡛ࡶ ᩘ NJ㸧࡛࠶ࡿ㸬௚᪉㸪௦⾲ⓗ࡞኱Ẽ࢚࢔ࣟࢰࣝᵓᡂ≀㉁࡛࠶ࡿ◲㓟ࡢ๓㥑య 62 ࡢᡃࡀᅜ࡛ࡢ᪥᤼ฟ㔞ࡣ㸪NW ⛬ᗘ࡛>@㸪㔜㔞࡛㸴᱆௨ୖࡢᕪࡀ࠶ࡿ㸬ࡇࡢࡼ࠺࡟≀㉁㔞࡜ࡋ࡚ࡣࢃࡎ࠿࡞ᨺᑕᛶ᰾✀ࡣ㸪ᕼ࢞ࢫ㸪ࣚ࢘⣲ࢆ㝖ࡁ༢య࡛ᾋ㐟ࡍࡿࡶࡢࡣᑡ࡞ࡃ㸪ᨺฟ࠾ࡼࡧ㍺㏦㐣⛬࡟࠾࠸࡚ࡉࡲࡊ
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⅔䠅䠈ᨺฟ䜲䞁䝧䞁䝖䝸䠄㔞ⓗホ౯䠅䠈&V 䛾Ỉ⁐ᛶ䇷㠀Ỉ⁐ᛶ䛾๭ྜ➼䛻㛵䛧ゎ᫂䛜ồ䜑䜙䜜䜛䠊䛥䜙䛻䠈⎔ቃ୰䛷䛾ᑑ࿨䠄䛔䛴⭉㣗䞉ᾐ㣗䛻䜘䜚 &V 䛜⁐Ꮡែ໬䛩䜛䛾䛛䠅䠈⭉㣗䞉ᾐ㣗䛿ᚎ䚻䛻㐍䜐䛾䛛䠈䛭䜜䛸䜒䛒䜛᫬ᮇ䛻୍Ẽ䛻⏕䛪䜛䛾䛛䠈⤖ᯝᨺฟ䛥䜜䜛䛰䜝䛖Ỉ⁐ᛶ &V 䛾ᙳ㡪ホ౯䛺䛹䠈௒ᚋ䛾୍ᒙ䛾◊✲ᒎ㛤䛜ᮃ䜎䜜䜛䠊஦ᨾ䛾㝿䛻⅔ෆ
䛷Ⓨ⏕䛧䛯஦㇟䛾ゎ᫂䠈䜘䜚ຠ⋡ⓗ䛺ᗫ⅔సᴗ䞉㝖ᰁసᴗ䠈㝖ᰁᗫᲠ≀䛾ῶᐜ䞉ᅽ⦰䜢䛿䛨䜑䛸䛧)LJ 6XPPDU\RISUREOHPVUHPDLQLQJUHODWHGWRHQYLURQPHQWDOEHKDYLRURILQVROXEOH&VSDUWLFOHV䛶䠈ᖐ㑏ᆅᇦఫẸ䛾Ᏻ඲䞉Ᏻᚰ䛾୍ᒙ䛾ᢸಖ䛺䛹䛾㐩ᡂ䛾䛯䜑䠈⥅⥆ⓗ䛺◊✲ྲྀ䜚⤌䜏䛜ᚲせ䛷䛒䜛䠊䠎䠅஦ᨾ䛻䜘䜚ᨺฟ䛥䜜䛯ᨺᑕᛶ᰾✀䛿䠈〄ᆅ㠃䜔㎰ᆅ䠈᳃ᯘ䛺䛹䛾ᵝ䚻䛺ᆅ⾲䛻ỿ╔䛧䠈ᆅ⾲㠃ᅵተ䛻྾╔䛧䛯䜒䛾䛜஧ḟᨺฟ䛸䛧䛶ᙉ㢼䛻䜘䜚෌㣕ᩓ䛩䜛䛣䛸䛜༴᝹䛥䜜䛯䠊䛭䛣䛷䠈䜋䜌〄ᆅ䛾⚟ᓥ┴ᾉỤ⏫䛾Ꮫᰯ䜾䝷䜴䞁䝗䛻䛚䛔䛶䠈ᨺᑕᛶ≀㉁䛾෌㣕ᩓ䛾ほ 䜢⾜䛳䛯䠊䜾䝷䜴䞁䝗䛻 +9 䝃䞁䝥䝷䞊䜢タ⨨䛧䠈䜎䛯䠈䜹䝇䜿䞊䝗䜲䞁䝟䜽䝍䜢᥋⥆䛧䛶䠈⢏ᚄู䛾䝃䞁䝥䝸䞁䜾䜢⾜䛔䠈䛭䜜䛮䜜ᤕᤊ䛥䜜䛯ᅵ⢏Ꮚ䛾ᨺᑕ⬟⃰ᗘ䜢ศᯒ䛧䛯䠊䛣䛾䜶䜰䝻䝌䝹⢏ᚄูಶᩘ⃰ᗘ䛸 &V ⃰ᗘ䛾㞟୰ほ䠈ᦶ᧿㏿ᗘ䠄㢼ᛂຊ䠅➼Ẽ㇟ほ 䛾䝕䞊䝍ẚ㍑䛻䜘䜚䠈ᨺᑕᛶ≀㉁䛾෌㣕ᩓ⡆␎໬䝇䜻䞊䝮䠄' ᨺᑕ⬟㣕ᩓ䝰䝕䝹䠅䜢㛤Ⓨ䛧䠈ᅵ⢏Ꮚ䛻䜘䜛෌㣕ᩓ䝰䝕䝹䜢ᵓ
⠏䛧䛯>@䠊ᮏ䝰䝕䝹䛷䛿䠈ᅵተ䛾ởᰁ㔞䛛䜙ᅵ⢏Ꮚ  ಶ䛒䛯䜚䛾ᨺᑕ⬟䜢ィ⟬䛧䠈䝎䝇䝖䝣䝷䝑䜽䝇䜢஌䛨䜛䛣䛸䛷䠈㣕ᩓ䛧䛯ᅵ⢏Ꮚ䛻䜘䜛኱Ẽᨺᑕ⬟䝣䝷䝑䜽䝇䜢⟬ᐃ䛩䜛䠊䛭䛾㝿䠈㣕ᩓ䛧䛯ᅵ⢏Ꮚ䛾⢏ᚄศᕸ䛿䠈6DQG䠈/RDP\VRLO䠈6LOW\FOD\ORDP䠈&OD\ 䛺䛹䛾ᆅ⾲㠃ᅵተ䛾⢏ᚄ䝍䜲䝥䛻౫Ꮡ䛩䜛䝰䝕䝹䛸䛧䛯䠊䛭䛧䛶䠈㣕ᩓ䛧䛯䝎䝇䝖⢏Ꮚ䛾ΰྜᒙ㧗䛥䠈䝎䝇䝖䛾Ⓨ⏕㠃✚䠈䝎䝇䝖䜲䝧䞁䝖䛾⥅⥆᫬㛫䜢௬ᐃ䛧䠈⢏ᚄู䛾኱Ẽᨺᑕ⬟⃰ᗘ䜢᥎ᐃ䛧䛯䠊䛭䛾⤖ᯝ䠈⡆᫆ⓗ䛺䝰䝕䝹䛰䛜䠈 ᖺ෤Ꮨ䛸
)LJ 6LPXODWLRQRIUHVXVSHQVLRQE\DHURVROWUDQVSRUWPRGHODQG
FRPSDULVRQZLWKWKHREVHUYDWLRQ- 51 -

 ᖺ᫓Ꮨ䛻䛴䛝䠈⢏ᚄู䛾኱Ẽᨺᑕ⬟⃰ᗘ䜢᥎ᐃ䛩䜛䛣䛸䛜䛷䛝䛯䠊䛧䛛䛧䠈ᵝ䚻䛺௬ᐃ䛜䛒䜚୙☜ᐇᛶ䜢᭷䛩䜛䛯䜑䠈ᅵተ෌㣕ᩓ䝰䝕䝹䜢㡿ᇦ䜶䜰䝻䝌䝹㍺㏦䝰䝕䝹䜈⤌䜏㎸䜏䠈䛭䛾᳨ド䜢㐍䜑䜛䛸ྠ᫬䛻෌⌧ィ⟬䜢⾜䛳䛯>@䠊⏝䛔䛯䝰䝕䝹䛿䠈䝷䜾䝷䞁䝆䝳ᆺ⛣ὶᣑᩓ䝰䝕䝹䛷䛒䜚䠈Ẽ㇟ᗇ䝯䝋ゎᯒ䝕䞊䝍䠄⣙ NP ᱁Ꮚ䠈 ᫬㛫㛫㝸䠅䜢⏝䛔䛶䠈⛣ὶ䞉஘ὶᣑᩓ䠈஝ᛶ䞉‵ᛶỿ╔䠄UDLQRXWZDVKRXW䠅䠈ᨺᑕᛶቯኚ䜢ィ⟬䛩䜛䠊 ᖺ  ᖺ㛫䛾෌㣕ᩓ䛻䛴䛔䛶䠈ᩘ್䝰䝕䝹䛸㑊㞴༊ᇦෆ䠄ᾉỤ⏫䠅䛚䜘䜃༊ᇦእ䠄䛴䛟䜀ᕷ䠅䛷䛾኱Ẽほ䜢⏝䛔䛶䛭䛾㡿ᇦ཰ᨭ䛾ᐃ㔞ⓗホ౯䜢⾜䛳䛯䠊䛭䛾⤖ᯝ䠈ᮾ᪥ᮏ䛾ィ⟬㡿ᇦ䛛䜙䛿ᖺ㛫෌ᾋ㐟㔞䛿 7%䡍䛷䠈䛭䛾䛖䛱䛾 7%T 䛜ྠ㡿ᇦ䛻෌ỿ╔䛩䜛䛸ィ⟬䛥䜜䛯䠊䛣䜜䛿䠈䛭䜜䛮䜜䠈ྠ㡿ᇦ䛾ᆅ⾲㠃ởᰁ䛻ᑐ䛧䛶䠈䠂௨ୗ䛚䜘䜃 ௨ୗ䛸⪃䛘䜙䜜䠈෌ᾋ㐟䛿ᆅ⾲㠃ởᰁ䛾ῶᑡ䛻ᐤ୚䛧䛺䛔䛣䛸䛜䜟䛛䛳䛯䠊䛺䛚䠈ᮏィ⟬䛷䛿䠈ᅵተ䠈㝣ୖ⏕ែ⣔䛛䜙䛾஧䛴䛾 &V ෌㣕ᩓ䛸⚟ᓥ➨୍ཎⓎ䛛䜙䛾ప䝺䝧䝹䛷⥅⥆䛩䜛୍ḟᨺฟ䜢⪃៖䛧䛶䛔䜛䠊⤖ᯝ䛸䛧䛶䠈ᾉỤ⏫䛻ẚ䜉䛴䛟䜀ᕷ䛷䛿䠈኱Ẽ୰䛾 &V ⃰ᗘ䛜  ᱆ప䛔್䛸䛺䛳䛯䠊䛣䜜䛿ほ ⤖ᯝ䛾䜸䞊䝎䞊䜢෌⌧䛧䠈ᗈᇦⓗ䛺 &V 䛾⛣ὶᣑᩓ䜢Ⰻዲ䛻⾲⌧䛷䛝䛯䠄
)LJ ཧ↷䠅䠊䜎䛯䠈Ꮨ⠇ⓗ䛺ኚ໬䛷䛿䠈ᾉỤ⏫䛷䛿ኟᏘ䛻㧗䛟䠈෤Ꮨ䛻ప䛟䠈䛴䛟䜀ᕷ䛷䛿෤Ꮨ䛻㧗䛔≉ᚩ䛜ᚓ䜙䜜䛶䛚䜚䠈ୖグ䛻♧䛧䛯ᅵተ䛛䜙䛾෌㣕ᩓ㔞䜢  ಸ䛩䜛䛣䛸䛷෤Ꮨ䛾 &V ⃰ᗘ䜢෌⌧䛷䛝䛯䠊䛧䛛䛧䠈ᾉỤ⏫䛾ኟᏘ䛾 &V ⃰ᗘୖ᪼䛿෌⌧䛷䛝䛪䠈෌㣕ᩓಀᩘ  ẖ᫬䛾᳃ᯘ⏕ែ⣔䛛䜙䛾 &V ᨺฟ䜢௬ᐃ䛩䜛䛣䛸䛻䜘䛳䛶䛿䛨䜑䛶䠈ᾉỤ⏫䛷䛾ኟᏘ䛾 &V ୖ᪼䜢䜋䜌ᐃ㔞ⓗ䛻ㄝ᫂䛩䜛䛣䛸䛜䛷䛝䛯䠊ᚋ⪅䛿䠈᳃ᯘ⏕ែ⣔䛛䜙䛾෌ᾋ㐟䛾ṇయゎ᫂䛜ᴟ䜑䛶㔜せ䛺䛣䛸䜢♧၀䛩䜛䠊䛣䛾䜋䛛䠈඲⌫䜶䜰䝻䝌䝹㍺㏦䝰䝕䝹䛻䜘䜛෌⌧ィ⟬䛸䛭䛾ホ౯>@䠈඲⌫䜶䜰䝻䝌䝹㍺㏦䝰䝕䝹䛸㏫ゎᯒ䛻䜘䜛 &V ⥲ᨺฟ㔞᥎ᐃ>@䠈㡿ᇦ㍺㏦䝰䝕䝹䛾㧗✵㛫ศゎ⬟໬䜈䛾ᛂ⟅◊✲>@䛺䛹ᨺᑕᛶ≀㉁䜢ྵ䜐໬Ꮫ㍺㏦䝰䝕䝹䛾㧗⢭ᗘ໬䛻䛴䛺䛜䜛᩾᪂䛺◊✲䜢㐍䜑䛯䠊䠏䠅Ⲉᇛ┴䛴䛟䜀ᕷ䛷䛾㝆ୗ㔞䜔኱Ẽ⃰ᗘ䛾ኚື䛛䜙䠈⚟ᓥ➨୍ཎⓎ஦ᨾ⏤᮶䛾ᨺᑕᛶ≀㉁䛾෌ᾋ㐟䛻䛴䛔䛶ᐃ㔞ⓗ䛺㆟ㄽ䜢⾜䛳䛯>@䠊ḟ䛔䛷䠈⚟ᓥ┴ෆ䛾ởᰁᆅᇦ䛷ᨺᑕᛶ䝉䝅䜴䝮䛾኱Ẽ䜈䛾෌㣕ᩓ䜢㛗ᮇほ䛧䛯⤖ᯝ䠈ϸ䠅㒔ᕷ㒊䛷䛾ほ⤖ᯝ䛸␗䛺䜚䠈඾ᆺⓗ䛺㔛ᒣ䛷䛒䜛ほ Ⅼ䛷䛿ኟᏘ䛻ᨺᑕᛶ &V 䛾኱Ẽ୰⃰ᗘ䛜ୖ᪼䛧䠈Ϲ䠅䛣䜜䜢ᢸ䛖⢏Ꮚ䛿䠈ぢ䛯┠䜔ගᏛ㢧ᚤ㙾ീ䛛䜙䝎䝇䝖䛸ᛮ䜟䜜䛯䛜䠈ពእ䛻䜒䛭䛾኱㒊ศ䛜⏕≀⏤᮶䛷䛒䜛䛣䛸䜢ぢฟ䛧䛯>@䠄)LJ ཧ↷䠗 䠊㟁Ꮚ㢧ᚤ㙾䛷䝞䜲䜸䜶䜰䝻䝌䝹➼䜢ィᩘ䛧䠈&V ⃰ᗘ䛸ẚ㍑䛧䛯
౛䜢 )LJ 䛻♧䛩>@䠊䝞䜲䜸䜶䜰䝻䝌䝹䛾ಶᩘ⃰ᗘ䛿 &V ྠᵝ䛻ኟᏘ䛻㧗䛟෤Ꮨ䛻ప䛔䠊䜎䛯ⰼ⢊䛾ಶᩘ⃰ᗘ䜘䜚䜒䝞䜽䝔䝸䜰䜔⬊Ꮚ>@䛷䛿୧⪅䛾༊ู䛜㐩ᡂ䛷䛝䛶䛔䛺䛔䠅䛾ಶᩘ⃰ᗘ䛾ኚື䛾᪉䛜ኚືᖜ䛜኱䛝䛛䛳䛯䠊䛥䜙䛻㟁Ꮚ㢧ᚤ㙾䛷䛾Ⅳ⣲ ; ⥺䛾ᙉᗘኚື䛿 &V ⃰ᗘኚື䛸䜘䛟┦㛵䛧䠈䝞䜲䜸䜶䜰䝻䝌䝹䛜ᢸయ䛸䛺䛳䛶 &V 䛾෌ᾋ㐟䛜⏕䛨䛶䛔䜛䛸᥎ 䛥䜜䛯䠊ᚑ᮶▱ぢ䛛䜙䛩䜛䛸䠈┿⳦㢮
䛜 &V 䜢⃰⦰䛩䜛䛯䜑䠈⬊Ꮚ䜈䜒 &V 䛜⛣⾜䛧䛶䠈䛭䛾⬊Ꮚ䛜ᨺฟ䛥䜜䜛䛣䛸䛷䠈&V 䛾኱Ẽ䜈䛾෌ᾋ㐟䜢ᢸ䛖䛸⪃䛘䜙䜜䠈㛵㐃䛩䜛▱ぢ䛜㞟䜑䜙䜜䛴䛴䛒
䜛>@䠊෌ᾋ㐟䝥䝻䝉䝇䛾୙᫂Ⅼ䛜ゎᾘ䛥䜜䜛䛣䛸䛿䠈ఫẸ䛾Ᏻ඲䞉Ᏻᚰ䛾☜ಖ䛻㛵䜟䜚䛯䛔䜈䜣㔜せ䛷䛒)LJ2SWLFDODQGHOHFWURQPLFURVFRSLFLPDJHVRI+9ILOWHUVDPSOHFROOHFWHGLQWKHFRQWDPLQDWHGDUHD ,JDUDVKLHWDOLQ
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IAEA-TEL-2015-03 Proficiency Test 試料の測定○木村徹＊１，五十嵐康人＊２，財前祐二＊２＊１㈱アトックス＊２

気象研究所

１．はじめにIAEA Proficiency Test（PT）は、環境試料に含まれる人工・自然放射性核種の測定技能を確認することを目的とした、国際原子力機関（IAEA）の相互比較評価の一つに位置づけられている。ところで、アトッ
クスでは数年前から気象研究所で実施されている大気降下物中の 90Sr と 137Cs の観測の一部を請負い、環境・応用研究部にて放射能分析を実施してきた。この研究グループが参画している新学術領域研究「福島原発事故により放出された放射性核種の環境動態に関する学際的研究」において、分析の品質向上の目的
で、代表者の恩田裕一筑波大教授が率先して PT 参加を募った。適正な分析品質を確保する観点から積極的に PT に参加し測定技能を確認することは、重要な事である。今回は PT でどのように試料を分析・測定したか、またその精度について結果をまとめ報告する。
２．分析試料
技術水準の確認に用いた分析試料は IAEA から供された水・玄米・土壌試料(Fig.1)である。IAEA から供された試料のうち水試料の 1 つは既知量の放射性核種 152Eu（下限数量以下）が添加されたものであった。
このほか、気象研究所から供された標準試料(Fig.2)の分析も併せて行っている。この標準試料は 1963 年～1979 年にかけて日本全国１４地点で採取された降下物試料を、気象研究所において混合・調製したものである。なおこの試料は、調製後に他の分析機関においても分析が行われ、参照値が与えられている[1,2]。

Fig.1 IAEA-PT sample

Fig.2 Reference fallout materials

３．分析・測定方法
γ線放出核種は、水試料については 500ml ポリ瓶に分取、玄米・土壌試料については外径 6cm のプラス
チック容器(丸型 V 式容器 V-7)に詰め、高さを測定後、Ge 半導体検出器(Fig.3)で測定した。
水試料は、193.1Bq/kg の 152Eu が添加された試料 No.3 約 500g を元にエネルギー効率曲線(Fig.4)を作成し、検出した核種の定量を実施した。また、134Cs については、サム効果補正のために試料と検出器との距離を0～142mm まで 5 段階で測定を行い、137Cs の計数率を基準とし、距離 0mm での 134Cs の計数率とサム効果が無視できる距離での計数率との差からサム効果の補正値を算出した(Table-1)。玄米・土壌試料は、高さを変えた気象研降下物標準試料から、高さ効率曲線(Fig.5)を作成し、多核種混合線源から得られたエネルギ
ー効率曲線(Fig.6)で補正を行った。

90

Sr については水試料は 10g、40g、土壌試料は 10g、15g を分取しテフロン容器内に入れ、濃硝酸と過酸

化水素を加えて 200℃に加熱し有機物を分解後、不溶残渣をろ過で除去する。担体として Sr 50mg 相当を加えたろ液に水酸化ナトリウムを加えアルカリにして、炭酸塩沈殿法、つづいてシュウ酸塩沈殿法、発煙硝酸法によるカルシウム除去、クロム酸塩沈殿法によるバリウム・ラジウム除去、水酸化鉄沈殿法によるイットリウム除去により化学分離し、最終的に炭酸ストロンチウムとして回収しろ紙ばさみに固定した。二週間以上放置し、90Sr と 90Ｙとが放射平衡に達した後に、低バックグラウンドガスレス検出器(CANBERRA 社
製 iMatic)(Fig.7)で測定した[2]。
Sr の回収率については、以下の計算式による。

Y=

𝑊𝑊2
∗ 100
𝑊𝑊１

Y : Sr yield(%)、W1 ： The quantity of added strontium as a carrier (mg)、
W2 ： The quantity of recovered strontium (mg)

0.100

Efficiency

y = 0.3158 x-0.5355
0.010

0.001
100

Fig.3 A Ge semiconductor detector

1000
Energy(keV)

10000

Fig.4 Water sample energy efficiency curve

Table-1 Relative count rate and coincidence summing effect revision
Distance(mm)
0
35.5
71
106.5
142
71-142mm average

Relative count rate(605/662keV)

0.80 ±0.11
0.91 ±0.19
0.98 ±0.17
0.89 ±0.12
0.95 ±0.04
0.94 ±0.11

・・・a

(b/a)
Coi nci dence s ummi ng
ef f ect rev i s i on・・・b

1 .1 8 ±0 .2 1

30.00
1.000

20.00

y = 77.2448 x-1.0936
1/Eff

15.00

多項式 (1/Eff)

y = -0.005437 x2 + 0.579471 x + 15.409279
R² = 0.983788

10.00

Efficiency

1/Efficiency

25.00

0.100

0.010

5.00
0.00

0

5

10
15
Height(mm)20

Fig.5 Height efficiency curve

25

0.001
100

1000
Energy(keV)

10000

Fig.6 Rice/soil sample energy efficiency curve

Fig.7 A gas-less α/β detector
４．評価方法
分析値は IAEA が示す以下の手順で評価した。
正確度の評価については、以下の式による。

𝐵𝐵𝐵𝐵𝐵𝐵𝐵𝐵𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟 =𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟 − 𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡
× 100%𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡

Biasrelative : Relative bias、Valuereported : Analysis value、Valuetarget : Target value
測定工程や分析法を考慮し、測定対象ごとに定められた最高受容相対バイアス(MARB:Maximum Acceptable
Relative Bias)と上式の Biasrelative を比較し、Biasrelative≦MARB の場合、正確度は”受容可能”となる。
精度の評価については、以下の式による。

22𝑢𝑢𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡
𝑢𝑢𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟
P = ��
� +�
� × 100
𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡
𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑉𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟𝑟
2

P : Precision、Valuetarget : Target value、Valuereported : Analysis value、
utarget : Target uncertainty、ureported : Analisys uncertainty

P≦MARB 及び、Biasrelative≦k*P の場合、精度は“受容可能”となる。
k は包含係数で、95%の信頼度で 2.56 となる。
最終評価について

A:受容可能 正確度と精度の両方が受容可能であるとき
N:受容不可W:警告

正確度が受容不可であるとき

正確度は受容可能だが、精度が不可のとき

Table-2 Comparison of target value and analysis value
Sample code

1

1

1

2

2

4

4

4

5

5

AnalyteSr-90

Cs-134

Cs-137

Na-22

Zn-65

K-40

Cs-134

Cs-137

Sr-90

Cs-137

Target value[Bq/kg]

29.6
±0.8

30
±0.9

30.1
±0.9

65.1
±1.5

177.3
±5.7

72.1
±3.5

12
±0.4

38.6
±0.9

36.2
±2.7

715
±30

MARB

15%

15%

15%

15%

15%

25%

15%

15%

20%

15%

Analysis value[Bq/kg]

30.2
±1.3

35
±6.4

34
±1.7

55.2
±1.8

180
±2.4

83
±4.0

11
±2.5

35
±1.7

37.8
±1.9

640
±2.6

Relative bias

2.03%

16.67%

12.96%-15.51%

1.52%

15.12%-8.33%-9.33%

4.42%-10.49%Accuracy

A

N

A

N

A

A

A

A

A

A

P

5.08

18.53

5.83

4.00

3.48

6.84

22.97

5.39

8.99

4.22

Precision

A

N

A

A

A

A

N

A

A

A

Final score

A

N

A

N

A

A

W

A

A

A

・Sample code 1・2 is water sample,4 is brown rice sample,5 is soil sample
・IAEA evaluation criteria, A：Accepted, N：Not Accepted, W:Warning
５．結果と考察
測定結果について、各分析値と IAEA 発表の目標値との比較と評価を表－１に示す。測定結果は概ね良好
なばらつきの範囲内であったが、134Cs は独自に実施したサム効果補正の誤差影響で不確かさが大きくなった。90Sr については、分析の都合上、報告期限に間に合わなかった為、IAEA には未報告であるが、分析値は目標値とよく一致している。よって、分析・測定が全体として適正に行われていると判断している。結果として、アトックスが分析を請負って以降の気象研究所における環境放射能試料の分析・測定データの信頼性については、良好に保たれていると判断する。今後も適正な分析品質を確保する観点から積極的に
PT 等に参加していきたい。
「参考文献」[1] 五十嵐康人,広瀬勝己,「放射性降下物標準試料の調製を通じた 90Sr、137Cs、プルトニウムなどの人工放射性核種分析の相互比較と参照値の決定」、放射化分析研究会,17,39-42 (2004)[2] M. Otsuji-Hatori, Y. Igarashi & K. Hirose, Preparation of Reference Fallout Material for Activity
Measurements, J. Environ. Radioactivity, 31, 143-155 (1996)

The measurement of
the IAEA-TEL-2015-03 Proficiency Test sample○Toru Kimura *1, Yasuhito Igarashi *2, Yuji Zaizen *2
1 ATOX Co.,Ltd. , 2 Meteorological Research Institute
IAEA Proficiency Test (PT) was intended to confirm the measurement skills of artificial and naturalradionuclides contained in the environmental samples, are located in one of the mutual comparativeevaluation of the International Atomic Energy Agency (IAEA). By the way, a part of the observationof 90Sr and 137Cs in atmospheric deposition have been carried out in the Meteorological ResearchInstitute, was undertaken by ATOX Co. from a few years ago.How I analyze and measure the sample
in the PT, also reported results are summarized for its accuracy. The analysis sample which I used
for the confirmation of technical standards is water, unpolished rice, soil sample offered by IAEA.
The gamma ray-releasing nuclide analysis was measured with a Ge detector. For 90Sr analysis, after
decomposition of organic substances, Sr fraction was radiochemically-separated and finally purified
strontium carbonate was measured with a low background gas-less counter.Comparing the targeted valueof the IAEA announcement with the analytical ones, analysis results were generally within thefavorable range. However,the 134Cs analytical values uncertainty was grows big under error influenceof the coincidence summing effect correction value that I carried out originally.
The 90Sr, it is unreported in IAEA,but analytical values were in good agreement with the targetedvalues.Therefore, it is concluded that the measurement of the environmental radioactivity sample
in the Meteorological Research Institute has been done properly.

ガンマ線スペクトル解析ソフト P-SCAN を用いた高精度な
Cs-134/Cs-137 放射能比の導出○木村徹＊１，五十嵐康人＊２，財前祐二＊２＊１＊２㈱アトックス
気象研究所
１．はじめに気象研究所では、大気降下物中の 90Sr、134Cs 及び 137Cs の長期モニタリングを実施している。このう
ち 134Cs は、複数のγ線をカスケードで放出するため、ゲルマニウム半導体検出器での測定においては、
サム効果の補正が必要であるが、通常、134Cs の標準線源（面線源）でピーク効率を求める場合にはサム
効果補正は要しない 1)と言われている。大気中の Cs 測定用の試料は、ハイボリュームサンプラーで大気
エアロゾルを石英フィルター上に捕集した後、フィルターを短冊状に 8 分割し放射能が均等になるように金型に入れて油圧プレスで圧縮固化し、標準面線源と同一形状に加工している。しかし、この過程で、放射能分布は必ずしも完全に均一ではなくなり、標準線源の放射性核種の分布とのずれが僅かだが生じて、試料毎にサム効果が微妙に異なる可能性が考えられる。134Cs/137Cs 放射能比（134/137 比）の高精度の測定は、発生源の特定に繋がることもあり、重要視している。そこで、試料毎（スペクトル毎）にサ
ム効果補正を行うことが可能なソフトウェアである P-SCAN（㈱スカラベ・コーポレーション製）の活用を考えた。この研究では、このソフトウェアを用いて、一番端的な事例であるフィルター試料に関し
134/137 比を高精度で測定することを試み、サム効果補正を行わない定量方法（スペクトル解析ソフト
として ITECH Instruments 製 InterWinner を使用）との比較を行った。
２．分析試料
校正用の標準試料として、多核種混合面線源（Eckert & Ziegler 製 134Cs 118Bq、137Cs 213Bq （2015/12/1
における保証値））を用いた。これは、石英繊維フィルター（SIBATA 製 QR-100）を油圧プレスで圧力を加えて円形（直径約 51mm、厚さ約 2mm）に圧縮固化し、その表面に均等に RI 溶液を添加して容器に封入したものである。
分析試料は東日本大震災直後の 2011 年 4 月に気象研究所で採取した大気フィルター試料（134Cs はま
だ減衰しきっていない）を用いた。この試料はハイボリュームサンプラー（SIBATA 製 HV-1000F）で大
気中浮遊塵を採取した石英繊維フィルターを標準試料と同様に円形に圧縮固化し、標準面線源と同形状の容器に封入したものである。
３．P-SCAN による定量原理
ガンマ線スペクトルの光電ピーク面積と、それらの計数効率・全効率の間に成り立つ高次・多元連立方程式を直接解くという方法を採用している。134Cs の様な壊変後にカスケード関係にあるγ線を放出す
る核種においては、観測されるピークの数をパターン化して数式化する。このようにして得られた複数の連立方程式を解き、最尤適合関数を求めることで、ピーク計数効率関数を決定している 2-5)。試料測定により“InterWinner”から得られたチャンネルごとの計数値をこのソフトウェアに入力すると、ピークを再解析し、選択した核種の効率や放出比も掛け合わされたピーク毎の崩壊数計算値として出力される。サム効果が発生する対象核種については補正計算が実行されて、サム効果補正後の崩壊数計算値が出力される。ここで得られた計算値を測定時間、減衰補正係数、さらに試料量で除算することで放射能濃度を得ることができる。
４．測定方法気象研究所の 4 台(#6、#7、#8、#10)のゲルマニウム半導体検出器を使用した(Fig.1)。まず、標準試料で校正し、大気中浮遊塵試料を測定して“InterWinner”で計数値を得た。得られた計数値に対して以下の放射能計算式でデータを算出した。また、“P-SCAN”でピークを再解析しサム効果補正計算処理
を行った。
n=

𝑎𝑆 𝑎𝑏
−
𝑡𝑠 𝑡𝑏

,

A=

𝑛･𝑑
𝑃･𝜀･𝑉

n : Net count rate(cps)、as : Sample peak area(count)、ab : Background peak area(count)、ts : Sample measurementtime(s)、tb : Background measurement time(s)、A : Radioactivity concentration(Bq/m3)、d : Attenuation correction、
P : Gamma ray emission rate、ε : Counting efficiency(cps/Bq)、V : Air suction volume(m3)サム効果の比率を確認するため、134Cs が放出するγ線のうち放出率が 97.6%と最も高い 605keV の計
数値と、相対計数値の基準としてサム効果が発生しない 137Cs の 662keV の計数値の比を、#8 検出器で、標準面線源との距離を 0cm から 30cm まで 5cm 間隔で離して測定した。検出器から距離を離すことによ
り、同時にγ線が入射する確率が減るためにサム効果は起こりにくくなり、605keV/662keV の光電ピー
クの比が一定値に近づく。0cm のカウント比を 1 として、一定値に達したときの距離のカウント比からサム効果の比率が分かる。
さらに、134Cs のサムピークの代表値として 605keV と 796keV（放出率 85.4%）が同時に入射した際に検出される 1400keV と 605keV とのカウント比の関係についても算出した。

Table-1 Comparison when using P-SCAN and normal calculation in 4 detectors(results of measuring the same sample)
Detector

Sampling
periodCs-134
NormalUsing P-SCAN
134+137(Bq)calculation
mBq/m3Err
mBq/m3
Err

#6
#7
#8

2011/04/11～04/19#10

Cs-137
NormalUsing P-SCAN
calculation
mBq/m3Err
mBq/m3
Err

Normalcalculation
134/137 ratio

134/137 ratio

Using P-SCAN

60.1

0.4

59.0

0.4

55.2

0.1

55.3

0.1

1.09

1.07

62.0

0.3

57.8

0.3

55.3

0.1

54.5

0.1

1.12

1.06

61.7

0.3

58.6

0.3

55.3

0.1

54.6

0.1

1.12

1.07

60.6

0.4

57.6

0.4

55.4

0.1

54.3

0.1

1.09

1.06

σ

0.014

0.006

898

Table-2 Detector distance and count ratio
Ge#8
Count ratio
Ratio to 0 cm
cm
605/662 1400/605 605/662 1400/605
0
0.318
0.086
1.00
1.00
5
0.420
0.018
1.32
0.21
10
0.438
0.007
1.38
0.08
15
0.443
0.001
1.39
0.01
20
0.455
0.000
1.43
0.00
25
0.443
0.001
1.39
0.01
30
0.452
0.000
1.42
-0.01

Fig.3 Relationship between distance to sample

Fig.4 Relationship between distance to sample

and 605/662 keV photo-peak count ratio

and 1400/605 keV photo-peak count ratio

５．結果と考察
大気中浮遊塵試料の測定結果を Table-1 に示す。137Cs については、両手法による測定結果はよく一致したが、134Cs については P-SCAN で得たデータの方が数％低めの値であった。また、134/137 比については、P-SCAN データの方が検出器間のばらつきが非常に小さかった。このことから、134/137 比の測定については P-SCAN を用いた方が精度が良いと考えられるが、絶対値としてどちらがより正確な値であるかの判断はできない。今後、濃度の異なる他の面線源の測定を行い再現性の確認等データの検証を行う必要がある。また、通常の計算法と P-SCAN を用いた場合との差は 10%以内であったので、環境放射能試料の測定において測定値の再計算等が必要とされるような重大な影響は無い。

距離と光電ピーク比の関係を Table-2 および Fig.3 に、また、距離と 1400/605 光電ピーク比の関係
を Fig.4 に示す。15cm 以上離せば 605/662 光電ピーク比は安定し、1400/605 光電ピーク比の関係からもサムピークの影響は無視できるほど小さくなる（サム効果が弱まり 0 に近づくことが分かる）ことが確認された。体積線源を用いる大気降下物試料については今回の検討には含まれていないため、今後さらに、補正方法の改善や P-SCAN の適用を検討していく。
参考文献
1)

日本アイソトープ協会理工学部会、NaI(Tl)シンチレーションスペクトロメータによるγ線スペク

トロメトリーガイダンス(2015)2)

日本アイソトープ協会理工学部会次世代スペクトル解析専門委員会、ガンマ線スペクトル解析におけるピーク探査と面積計算実態と今後のスペクトル解析のあるべき姿、RADIOISOTOPES、65、267-285(2016)
3)

山田崇裕、石津秀剛、高野直人、加算同時計数効果を用いた Ge 検出器の計数効率関数決定法、第51 回アイソトープ・放射線研究発表会要旨集、50 (2014)4)

仁木工芸株式会社、P-SCAN カタログ (2014)5)文部科学省、放射能測定シリーズ 7 ゲルマニウム半導体検出器によるガンマ線スペクトロメトリー (1992)

Aiming at higher precision measurement of Cs-134 and Cs-137
radioactivity ratios by using P-SCAN gamma ray data analysis
software○ Toru Kimura * 1, Yasuhito Igarashi * 2, Yuji Zaizen * 2
1 ATOX Co.,Ltd. , 2 Meteorological Research InstituteAt the Meteorological Research Institute (MRI), long-term monitoring of 90Sr, 134Cs, and 137Cs in
atmospheric fallout is carried out. For the sample for measuring Cs in the atmosphere, after collecting
atmospheric aerosol on a quartz filter with a high volume sampler, it is compressed and solidified andprocessed into the same shape as the standard surface source. However, in this process, the radioactivity
distributes not necessarily completely uniform, slight deviation from the distribution of the radionuclideof the standard radiation source occurs, and the sum effect potentially varies slightly for each sample.
Therefore, we considered utilization of P-SCAN (manufactured by SCARABS CORPORATION), which issoftware capable of correcting sum effect for each specimen (spectrum). In this research, we tried tomeasure 134/137 radioactivity ratios for filter samples with high precision using this software
compared with a quantitative method

without sum effect correction.

andFirst, we calibrated with the standard source , then we measured airborne dust sample and got the countvalue by "InterWinner". Data was calculated for the obtained count value by the radioactivity calculation
formula. Moreover, the peak was re-analyzed with "P-SCAN" and the sum effect correction calculation
process was performed.
For 137Cs, the results were in good agreement, but for 134Cs, P-SCAN led lower values. As for theconcentration ratio of 134Cs/137Cs, P-SCAN gave a much smaller variation for all detectors. From thisresult, it can be considered that the precision of P-SCAN is better, but it cannot be judged which one isthe accurate measurement result by this measurement alone. In the future, it is necessary to measureother surface radiation sources with different concentrations and verify data as confirmation of
reproducibility. Also, since the difference between the usual calculation method and P-SCAN was within
10%, there is no need to recalculate/correct in the measurement of environmental radioactivity samples.

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12
https://doi.org/10.1186/s40645-018-0171-z

RESEARCH ARTICLE

Progress in Earth andPlanetary Science

Open Access

The seasonal variations of atmospheric
134,137
Cs activity and possible host particles
for their resuspension in the contaminatedareas of Tsushima and Yamakiya,
Fukushima, Japan
Takeshi Kinase1,2*, Kazuyuki Kita3, Yasuhito Igarashi4, Kouji Adachi4, Kazuhiko Ninomiya5, Atsushi Shinohara5,
Hiroshi Okochi6, Hiroko Ogata6, Masahide Ishizuka7, Sakae Toyoda8, Keita Yamada8, Naohiro Yoshida8,9,Yuji Zaizen10, Masao Mikami11, Hiroyuki Demizu12 and Yuichi Onda13

Abstract
A large quantity of radionuclides was released by the Fukushima Daiichi Nuclear Power Plant accident in March 2011,
and those deposited on ground and vegetation could return to the atmosphere through resuspension processes.
Although the resuspension has been proposed to occur with wind blow, biomass burning, ecosystem activities, etc.,
the dominant process in contaminated areas of Fukushima is not fully understood. We have examined theresuspension process of radiocesium (134,137Cs) based on long-term measurements of the atmospheric concentrationof radiocesium activity (the radiocesium concentration) at four sites in the contaminated areas of Fukushima as wellas the aerosol characteristic observations by scanning electron microscopy (SEM) and the measurement of thebiomass burning tracer, levoglucosan.
The radiocesium concentrations at all sites showed a similar seasonal variation: low from winter to early spring andhigh from late spring to early autumn. In late spring, they showed positive peaks that coincided with the wind speed
peaks. However, in summer and autumn, they were correlated positively with atmospheric temperature but
negatively with wind speed. These results differed from previous studies based on data at urban sites. The differenceof radiocesium concentrations at two sites, which are located within a 1 km range but have different degrees of
surface contamination, was large from winter to late spring and small in summer and autumn, indicating that
resuspension occurs locally and/or that atmospheric radiocesium was not well mixed in winter/spring, and it was
opposite in summer/autumn. These results suggest that the resuspension processes and the host particles of theradiocesium resuspension changed seasonally. The SEM analyses showed that the dominant coarse particles insummer and autumn were organic ones, such as pollen, spores, and microorganisms. Biological activities in forest
ecosystems can contribute considerably to the radiocesium resuspension in these seasons. During winter and spring,soil, mineral, and vegetation debris were predominant coarse particles in the atmosphere, and the radiocesium
resuspension in these seasons can be attributed to the wind blow of these particles. Any proofs that biomass(Continued on next page)* Correspondence: t.kinase.J52@gmail.com
1
Graduate School of Science and Engineering, Ibaraki University, 2-1-1
Bunkyo, Mito, Ibaraki 310-8512, Japan2
Hitachi Power Solutions Co., Ltd, 2-2-3 Saiwaicho, Hitachi, Ibaraki 317-0073,
Japan
Full list of author information is available at the end of the article© The Author(s). 2018 Open Access This article is distributed under the terms of the Creative Commons Attribution 4.0
International License (http://creativecommons.org/licenses/by/4.0/), which permits unrestricted use, distribution, andreproduction in any medium, provided you give appropriate credit to the original author(s) and the source, provide a link to
the Creative Commons license, and indicate if changes were made.

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

Page 2 of 17(Continued from previous page)burning had a significant impact on atmospheric radiocesium were not found in the present study.
Keywords: Radiocesium, Atmospheric radioactivity, Seasonal variation, Resuspension, Fukushima, Nuclear accident,
Bioaerosol, Spore, Mineral dust, Host particle

Introduction
In March 2011, abundant and various radionuclides
were released into the atmosphere (e.g., Chino et al.
2011) as a result of the nuclear accident at the Fukushima Daiichi Nuclear Power Plant (FDNPP) caused by
the Great East Japan Earthquake (Holt et al. 2012), andtheir amounts and diffusion have been of public concern because of their health impacts (Report: Working
Group on Risk Management of Low-dose Radiation
Exposure 2011). Radiocesium isotopes 134Cs and 137Cs
were two of the main radionuclides released from the
FDNPP accident (Steinhauser et al. 2014). As an example, the total amount of released 137Cs was estimated by Chino et al. (2011) to be 1.3 × 1016 Bq (13PBq). The 134Cs emission inventory should be almost
the same as that of 137Cs, i.e., 1.2 × 1016 Bq (12 PBq;
Steinhauser et al. 2014) because the activity ratio ofthese two isotopes was almost united in the case of the
FDNPP accident (e.g., Merz et al. 2013). There havebeen concerns that the deposited radiocesium will influence the contaminated area for an extended period,in contrast to the effects of radioiodine, because thehalf-lives of 134Cs and 137Cs are relatively long:
2.06 years and 30.17 years, respectively (Dietz andPachucki 1973; Unterweger 2002). Figure 1 shows thecontaminated deposition density map for 134, 137Cs (theunit of deposition density is Bq m− 2) measured from
October to November 2011 by the Ministry of Education, Culture, Sports, Science and Technology, Japan(MEXT) (2011). The radionuclides released by the
FDNPP were deposited on the ground, vegetation,
houses, and other objects, causing serious contamination of widespread areas, especially areas in the northwest of the FDNPP.
Although more than 5 years have passed since the
FDNPP accident, significant atmospheric radiocesium
activity concentrations (hereafter called the radiocesium concentration), on the order of approximately 10−4
Bq m− 3, have still been observed in the contaminated area (e.g., Ochiai et al. 2016). These concentrations are at least two orders of magnitude higher thanthose observed in East Japan before the FDNPP accident (~ 1 μBq/m3; Igarashi et al. 2015). Several studies
showed that this atmospheric radiocesium could not be
attributed to the direct emissions/leakage from the
FDNPP site (e.g., Igarashi et al. 2015; Kajino et al.
2016) but could be supplied by secondary emissions of

the deposited radiocesium, which is called resuspension
to the atmosphere (e.g., Igarashi 2009; Igarashi et al.
2015; Kajino et al. 2016). Igarashi et al. (2015) concluded that the delayed primary emissions from the
FDNPP could not be a major source of the current (tillMarch 2015) radiocesium in the atmosphere based on
the atmospheric observations in Tsukuba and the emission inventory data from Tokyo Electric Power Co.,
Inc. (TEPCO) (2012). Kajino et al. (2016) also showedthat resuspension could predominantly contribute to
the atmospheric radiocesium concentrations in the year
2013 using a 3D aerosol model simulation and theemission data from TEPCO.
The soluble form of primary radiocesium is carried bysubmicron particles (Masson et al. 2013), which would
most likely be non-sea salt sulfate (nss-SO2−
4 ), as suggested by Doi et al. (2013) and Kaneyasu et al. (2012). In
general, water-soluble cesium ions from SO2−
4 are considered to be adsorbed or attached to mineral particles
in a soil environment (Bostic et al. 2002; Dumat andStaunton 1999; Mukai et al. 2014). These mineral particles were shown to act as resuspended host particles in
an atmospheric environment by the previous investigations of the global fallout from atmospheric nuclear
weapon tests (e.g., Igarashi et al. 2005; Masson et al.
2010). Therefore, a major resuspension source has beenthought to be the suspension of contaminated dust particles by wind (Igarashi et al. 2016; Ishizuka et al. 2017;Sýkora et al. 2012), even for the case of the FDNPP accident contamination. In addition, a significant amount ofwater-soluble cesium ions has been absorbed by livingorganisms, such as mushrooms (Yoshida et al. 1994) andplants, and has been relocated within vegetation, frombacteria (Tomioka et al. 1992), through plant roots(Ehlken and Kirchner 2002), to pollen and then to honey(Barišić et al. 1992; Kanasashi et al. 2015). Such contaminated forest ecosystems could be another source forradiocesium re-emission into the atmosphere (also wecall “resuspension” in this study). The cedar pollen incontaminated forests contains significant amounts ofradiocesium. The highest total radiocesium concentration in Fukushima cedar pollen was reported to be−1
25.4 Bq g (dry weight) in 2014 (Forestry Agency, Japan(FAJ) 2015), suggesting that pollen emissions could still
be a candidate for the host particles of radiocesium
resuspensions, although pollen emissions occur during
limited durations of weeks or months each year

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

Page 3 of 17

Fig. 1 The four sampling locations in this study plotted over the deposition density map for 134,137Cs obtained by the Japanese government(Ministry of Education, Culture, Sports, Science and Technology, Japan (MEXT) 2011). The radius of dotted outer circle is 30 km

depending on plant species. Furthermore, open biomassburning of contaminated vegetation is also a candidate
for the radiocesium resuspension process (Bourcier et al.
2010; Igarashi 2009; Lujanienė et al. 2009; Sýkora et al.
2012; Yoschenko et al. 2006).
Many studies have been carried out to address the environmental radiocesium contamination resulting fromthe FDNPP accident, but the processes of radiocesium
resuspension have not yet been fully identified in thiscase. Knowledge of the host particles of such secondary
radiocesium emissions and their long-term variations islimited (e.g., Igarashi et al. 2015) because of the lack oflong-term and intensive studies in Fukushima. Ishizuka
et al. (2017) discussed the processes contributing to theradiocesium resuspension with soil dust particles fromthe surface based on measurements within the contaminated area in Fukushima. Their proposed scheme can beapplied to evaluate secondary emissions caused byaeolian resuspension of radioactive materials associatedwith mineral dust particles from the ground surface.
However, their study only focused on the resuspension
from winter to spring; the seasonal variations and otherprocesses of resuspension were not discussed. A recent

work by Kajino et al. (2016) has revealed that the forest
ecosystem could be another source of radiocesium during the warm seasons, in addition to that of dust suspension. These authors applied a novel scheme forradiocesium resuspension from the forest (as the green
fraction) that operates according to forest activity as
measured by the normalized difference vegetation index(NDVI) satellite data (Gutman and Ignatov 1998). However, Kajino et al. (2016) showed a model simulation application and did not reveal the actual host material ofradiocesium. Recently, Ochiai et al. (2016) added newobservational results of the monthly radiocesium concentrations (September 2012 to December 2014) at acontaminated area in western Tsushima that is near ourpresent observation site. These authors applied a twostage size separation (coarse and fine) in their monitoring and demonstrated that 137Cs activity concentration
enhancement during the warm season and the coarse
fraction (> 1.1 μm) of 137Cs exhibited seasonal changes
along with the concentration increases. These authorsinterpreted this seasonal trend via the relevance of theseasonal changes in the prevailing local winds and locations of the surface contamination. Although their

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

observational results agree well with our present results,
their explanation remains incomplete; no explanationsof the radiocesium host particles involved are given. It is
important to understand the contributions of each resuspension process in all seasons and to break thecurrent limitations of our understandings, especially inthe heavily contaminated areas in the Fukushima prefecture, to evaluate the transport/diffusion of radiocesium
from the contaminated areas to other areas. The purposes of this research were thus set as follows: (1) to
identify the seasonal variations of the concentrations ofatmospheric radiocesium via long-term monitoring, (2)to determine the spatial scales of the resuspension andredistribution of atmospheric radiocesium, and (3) to estimate the sources and host particles that contribute to
the resuspension.

Methods/Experimental
For the measurements of the radiocesium concentrations, we have carried out aerosol sampling at four sites(labeled herein as the FD, FC, GD, and GC sites), which
are located in heavily contaminated areas in the
Fukushima prefecture, as shown in Fig. 1. The areas ofthe sampling sites belong to typical Japanese villages/
towns, which are surrounded by forests in a mountainous area as is the traditional farming landscape in thisregion (e.g., Berglund 2008). The forests around thesampling sites are secondary forests surrounding farm
villages (Satoyama) and are not dense, primeval forests.
According to Berglund (2008), Satoyama is defined as“forests surrounding farm villages and managed by
farmers for different needs—timber for buildings, woodfor fuel and charcoal production, and leaf litter andtwigs used as fertilizer for crops, particularly in the rice
paddy fields situated in the lowlands.” Currently, theagricultural lifestyle has been changing greatly; however,major landscape changes have not occurred. Inhabitants
were evacuated from these contaminated areas by thetime that this manuscript was revised (up to March2017), so this area has been free from human activity,
except for decontamination and monitoring activities.
The sites labeled with the letters F and G are the forest
and open ground sites, respectively, and the sites labeledwith the letters D and C are sites where decontamination was conducted or not, respectively. The FD site in
the Yamakiya area, Kawamata-town, is located in a cedar
forest 35 km northwest of the FDNPP, and its contamination was the lowest of the four observation sites even
before decontamination because the radioactive plumewas blocked by hills.
We could start the monitoring at the FD site soonafter the FDNPP accident, i.e., since July 2011. The areaaround the FD site was decontaminated between AugustPage 4 of 172013 and October 2014. The dose rate monitoring by
the Nuclear Regulation Authority showed that the
average dose rate in this area had decreased from
0.90 μSv h− 1 in 2012 to 0.28 μSv h− 1 in 2014 mainly because of the decontamination. The FC, GC, and GD sites
are located in the Tsushima area in Namie-town, approximately 30 km northwest of the FDNPP. The GC
site was in a school playground (area is 10.7 × 103 m2),
and the aerosol sampling and meteorological observations here started in December 2012 (Ishizuka et al.
2017). The GD site was located in another school playground (area is 5.7 × 103 m2), within a 1 km distancefrom the FC and GC sites. Both sites are basically flatand originally paved by sand because of their gymnastic
uses. An area of approximately 5 × 104 m2, including theGD site, had been locally decontaminated during theperiod of December 2011 to February 2012, and the dose
rate at the GD site decreased from 9.60 to 2.8 μSv h− 1
because of the decontamination (Japan Atomic Energy
Agency (JAEA) 2012). The aerosol sampling at this site
started in November 2012. The FC site is in a broadleaved forest mixed with red pine trees adjacent to the
GC site, and the aerosol sampling at this site started in
January 2014. The Tsushima area, including the FC andGC sites, was not decontaminated during our observation period, except around the GD area, and their dose
rates were higher (by at least three to ten times) thanthose at the FD and GD sites. Details of the four sites
are summarized in Table 1.
High-volume air samplers (120SL, Kimoto, Japan, andHV-1000R/F, Sibata, Japan) were used for aerosol sampling (a flow rate of 0.7 m3 min− 1) with quartz-fiber filters (2500QAT-UP, Pallflex, USA). Six-stage cascade
impactors (TE-236, Tisch Environmental, USA)
mounted on other high-volume air samplers (a flow rateof 0.556 m3 min− 1) were also used to collect the aerosols
separately according to their aerodynamic diameters,with quartz-fiber filters (TE-230QZ, Tisch Environmental, USA and 2500QAT-UP, Pallflex, USA) applied as collection substrates. This impactor classifies particles aslarger than 10.3 μm (stage #1), 4.2–10.3 μm (stage #2),
2.1–4.2 μm (stage #3), 1.3–2.1 μm (stage #4), 0.69–1.3 μm(stage #5), 0.39–0.69 μm (stage #6), and smaller than
Table 1 Information about the observation locations andconditions at the four sites
Site-ID Latitude

Longitude

Land condition Decontaminatedor contaminatedFD

37° 35′ 07″ N 140° 41′ 29″ E Cedar forest

DecontaminatedFC

37° 33′ 41″ N 140° 46′ 06″ E Broad leaf
forestContaminatedGD

37° 33′ 38″ N 140° 45′ 37″ E School ground DecontaminatedGC

37° 33′ 44″ N 140° 46′ 07″ E School ground ContaminatedKinase et al. Progress in Earth and Planetary Science (2018) 5:12

Page 5 of 17

0.39 μm (backup filter) in diameter, in which the collection efficiency exceeds 50%.
The aerosol sampling period at the FD, FC, and GD
sites was mostly between 1 week and 1 month. At theGC site, short-term sampling (1–3 days) was carried out
using seven high-volume air samplers with timercontrols. The radioactivities of 134,137Cs of the aerosols
sampled in filters were measured at the MeteorologicalResearch Institute (MRI), Osaka University, Tokyo
Institute of Technology, and Ibaraki University using
coaxial-type Ge semiconductor detectors coupled withcomputed spectrometric analyzers. The radioactivities of134
Cs and 137Cs were identified and determined from the
gamma-ray peak intensities at 605 and 662 keV, respectively. Details of the detector models and the spectrometricanalyzers are summarized in Table 2. The radiocesium
concentration was calculated from the measured radioactivity and the total volume of the sampled air.
A scanning electron microscope (SEM; SU3500,
Hitachi High-Technologies Co., Tokyo Japan) with anenergy-dispersive X-ray spectroscope (EDS; X-max
50 mm, Horiba Ltd., Kyoto, Japan) was used to observe
the shapes and chemical compositions of the aerosols
collected in the above samplings to identify the major
host particles of the resuspended radiocesium.

Results and discussion
Seasonal and long-term variations of the atmosphericconcentrations of radiocesium activity

Figure 2 shows the time series of the radiocesium
concentrations of 134,137Cs obtained from the aerosol
sampling at the four sites. For the case of the samples
obtained using the impactor, the radiocesium concentration values were the sum of all stages. The radiocesium
concentration at the FD site showed a rapid exponential
decrease during the earlier period from July 2011 to
October 2011. After that, the decrease became slower.
Igarashi et al. (2015) showed that the direct emissions/leaks of radiocesium from the FDNPP site significantly
affected the radiocesium concentrations in 2011 and2012 (in the early period), when spikes occurred in theradiocesium concentrations monitored at the MRI, Tsukuba (approximately 170 km south–southwest from the
FDNPP), and that atmospheric radiocesium was mainly
supplied by resuspension after this period. It should be
noted here that radiocesium of the resuspension origin
is also subjected to transportation and diffusion as is theprimary emission.
The radiocesium concentrations at all of the four sites
showed similar seasonal variations, being low in winterand early spring and high in late spring, summer, andearly autumn, after November 2011. The annual mean,
maximum, and minimum values of the 137Cs concentrations measured at each site are summarized in Table 3.
Similar seasonal variations at all sites indicate that resuspension occurred with identical processes/sources inboth open areas and forests. This result is opposite tothat of the seasonal variation (high in winter and low insummer) found in urban areas by other studies (e.g.,
Igarashi et al. 2015; Sýkora et al. 2012; Watanabe 2014),
suggesting that the dominant resuspension processes/
sources are different because of the locations/surface
conditions of the study areas, specifically between theurban sites and the forest sites. The sampling sites inour study are located in low mountainous areas, where
the major land uses are agricultural fields (abandoned
because of the evacuation) placed along valleys and forests covering mountain slopes. Because it is a typicalsurface condition in heavily contaminated areas in
Fukushima as mentioned previously, the present seasonal trend is probably typical of these areas. Similarseasonal variations of radiocesium concentrations were
found in the contaminated area of the western part ofNamie (near the GC site) by Ochiai et al. (2016).
Between spring and autumn, two maxima were foundin May and from August to September, especially for the2014 results; Fig. 3 compares the temporal variations ofTable 2 Information about the instruments used in this study
Measured at

Detector model

Production

Spectrometric analyzer

Production

MRI

GEM50

Seiko EG&G

Multiport

Oxford-Tennelec

IGC60210

Princeton Gamma-Tech

EGPC40Canberra
MCA760092x

Seiko EG&G

MCA7600

Seiko EG&G

Osaka University

GEM-90205-P

Seiko EG&G

GC6020Canberra

GEM40

Ortec

GEM 80

Ortec

Tokyo Institute of Technology

GC3018

Canberra

DSA1000

Canberra

Ibaraki University

GC4020Canberra

Linx

Canberra

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

Page 6 of 17

Fig. 2 Activity concentrations of 134,137Cs (Bq m− 3) measured at the FD, FC, GD, and GC sites in the Fukushima prefecture in July 2011 to March2015. These results show high concentrations in late spring, summer, and early autumn and low concentrations in winter and early spring at allof the presented observation locations(a) the radiocesium concentrations at the GC site withthe 30 min averages of various meteorological parameters: (b) wind speed, (c) atmospheric temperature, and(d) precipitation. At the first maximum, occurringaround May, the wind speed peaks generally coincidedwith the radiocesium concentration peaks. During thesecond maximum, occurring around September, theradiocesium concentration and temperature had similarvariations, but the wind speeds were negatively correlated with these variables. These results indicate that theprocesses or sources of the resuspension are different inthese seasons: the physical processes (suspension bywind) are significant in late spring, and another process
is dominant in summer/autumn.

Seasonal variations of resuspended host particles foundby impactor sampling

To indicate the fractions of 137Cs captured on each stage
as well as on the backup filter in the cascade impactor

Table 3 Averages, maximums, and minimums of the 137Cs activity concentration (μBq m−3) at each site
Average
Location (sampling duration)
FD (July 2011–October 2014)

Year

Maximum

137

Cs Error Date (start)
m/d h:mMinimum
Date (end)m/d h:m137

Cs

Error Date (start)
m/d h:mDate (end)am/d h:m2012/1/6 11:16

137

Cs Error2011 1450

33.7

7/9 16:40

7/18 12:10

5600

66.4

12/9 12:47

2012 287

9.7

3/9 10:43

4/6 11:15

547

7.8

11/16 12:01 12/7 10:18

2013 12713.6

8/9 14:16

8/30 11:20

432

4.6

10/18 14:48 10/28 10:45

45.9

3.2

2014 115

8.3

6/29 9:41

7/4 10:30

202

6.4

10/3 10:27

38.7

4.6

10/20 11:07

186

11.0

41.2

2.0

FC (May 2014–January 2015)2014 53778.6

8/22 16:13

8/29 12:52

1620

843

12/26 15:41 2015 1/11 13:24 74.3

5.7

GD (November 2012–June 2015)2012 213

2.7

11/3 13:40

11/9 12:41

294

2.1

12/28 12:45 2013 1/11 13:46 130

1.42013 521

4.1

8/9 11:55

8/16 11:16

3690

8.0

12/13 12:04 12/23 12:08

72.0

3.0

2014 24326.4

9/19 11:16

10/3 10:59

684

94.6

12/26 11:28 2015 1/11 12:07 47.4

8.5

6/4 10:52

56.8

9.9

95.2

4.7

2015 123

18.7

5/4 11:26

225

31.9

4/25 10:41

GC (December 2012–November 2014) 2012 462

12.2

12/21 13:00 12/22 13:00 2550

28.2

12/19 13:00 12/20 13:00

2013 1170

19.2

8/14 13:00

8/15 13:00

60,400 132

11/12 13:00 11/13 13:00

68.9

11.6

2014 546

11.6

5/16 13:00

5/18 13:00

6750

1/15 13:00

48.4

7.8

a

In some cases, the sampling duration crossed over years

31.0

5/4 11:11

1/17 11:38

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

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Fig. 3 Time series of (a) radiocesium concentrations, (b) wind speeds, (c) temperatures, and (d) rain/snow precipitation. At the GC site, we cansee two maxima in the higher concentration seasons: in May and September. The first maxima occurred with wind speed peaks, and the second
maxima occurred with atmospheric temperature peaks

sampling, the measured values of the 137Cs radioactivity
on the filters 137Cs(i) were normalized as shown in Eq. 1:
137

Normalized 137CsðiÞ ¼ 137

CsðiÞ;CsðTotalÞ

ð1Þ

where (i) is the stage number, 137Cs(i) is the 137Cs radioactivity of the (i)-th sampling stage collection substrate,and 137Cs(Total) is the total of the 137Cs radioactivity of allof the sampling stage substrates, including the backup filter for each sampling period. The radioactivity of the
backup filter was also normalized to the normalized
137
Cs(Total) value as 137Cs(7). Figure 4 shows the time seriesof the 137Cs fraction of stage #2 (nominal aerosol diameterof 4.2–10.3 μm) and the backup filter (< 0.39 μm), where
the highest fraction was most often found. In the fractionof stages #1–5, the normalized 137Cs(1–5) showed similarseasonal variations among themselves, i.e., high insummer/autumn (June to November) and low in winter/spring (December to April or May). This result is
consistent with that of Ochiai et al. (2016); the coarse fraction (> 1.1 μm) of 137Cs exhibited similar seasonalchanges. In contrast, the fraction of the backup filter ofnormalized 137Cs(7) showed the opposite trend; it was highin winter/spring and low in summer/autumn. In this discussion, we omitted the 137Cs(6) values because they were
frequently lower than the detection limit. This seasonalvariation of the 137Cs(1–5) and 137Cs(7) fractions could be
attributed to a measurement artifact, i.e., the bounce

effect, rather than the actual size variations of the hostaerosols for the radiocesium resuspension. Bouncing ofparticles, particularly coarse ones, occurred in stages#1–6, especially during winter/spring, resulting in theconsiderable amount of coarse particles sampled at the
backup filter, although the filter was expected to capture only finer particles. Figure 5 shows a picture of atypical backup filter sample at the GC site in spring,showing many coarse particles, such as pollen and soil/
mineral particles. These coarse particles would nominally have been trapped in stages #1 and #2, but theybounced in these earlier stages and accumulated on the
backup filter. As indicated by the red rectangles, accumulations of coarse particles were found near the edgeof the backup filter, presumably because of the stagnation of air flow or narrower gaps between the filter andthe cover plate at the edge. This bounce effect was significant in winter and spring when the normalized
137
Cs(7) values were high. This finding enables us to discuss the seasonal differences of the host particles forthe radiocesium resuspension, although this finding
also makes discussing their size distributions difficult.
Figure 6a, b shows the SEM images of the stage #2
substrate samples obtained (a) on April 10–24, 2014
(NHVB-260424) and (b) on October 17–30, 2014
(NHVB-261030). The 137Cs(2) values were 44.8 ± 7.1 and28.3 ± 11.3 μBq m− 3, respectively, for the NHVB-260424
and NHVB-261030 substrate samples. As shown in
Fig. 6b (denoted by the red curve), the adhesion of

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

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Fig. 4 Time series of the normalized 137Cs(2) and 137Cs(7) measured at the (a) FD, (b) FC, and (c) GC sites. These results clearly showed theseasonal variation but did not accurately reflect the size distribution of radiocesium in the atmosphere because of the bouncing particles. The
amount of bouncing particles significantly changes the collection efficiency of the cascade impactor

organic matter on quartz fibers was frequently found ineach stage of substrates in the summer/autumn samples.
In contrast, such organic matter adhesion was not foundin winter/spring. We also found that most coarse particles could not be removed from the summer/autumn
sample substrates via soaking them in water, but that thecoarse particles could be easily removed from the winter/spring sample filters. The adhesive organic matter on

the quartz fibers probably worked as glue to fix the hostparticles of radiocesium to the sample substrates and thefilter, or the host particles themselves could have firmly
attached to the sample filter, reducing bouncing ofcoarse particles significantly in summer and autumn. In
contrast, the host particles did not firmly adhere to thefibers in winter and spring.
The spatial extent of the resuspension

The atmospheric activity concentrations at the GC andGD sites are compared by adopting their ratio, Rscale, as
follows:
137

Rscale ¼ 137

Fig. 5 Picture of a typical backup filter sampled at the GC site inspring (April 2014). We found many coarse particles, which include
pollen and coarse soil/mineral dusts, in the areas indicated by the
red rectangles

Cs‐GD
;
Cs‐GC

ð2Þ

where 137Cs-GC and 137Cs-GD denote the 137Cs concentrations observed at the GC or GD sites, respectively. There
are two important points for interpreting the Rscale values.
Although the GC and GD sites are situated very close together (about 0.75 km), the amount of radiocesium
remaining over the ground surface (0–5 cm depth) at theGC site was much higher than that at the GD site becauseof the local decontamination (approximately 50,000 m2)
around the GD site. By sampling the surface soil (0–5 cm
depth) at six points around the high-volume air samplers
at each site on June 4, 2015, their radiocesium

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

Fig. 6 SEM images of filters for stage #2 sampling at the GC site ina spring and b autumn. The adhesive organic matter that sticks to
the quartz fiber is surrounded by the red line. From themorphology, we suspect that the material was dense liquid when it
was suspended in air. This matter probably affects the collection
efficiency of the coarse particles on the filter

concentrations were measured. The dried soil samples(approximately 150 g) were put in U-8 plastic containers(AS ONE Corporation, Osaka, Japan), and their 137Cs
radioactivities were measured at Ibaraki University. The
average soil activity concentrations at the GC and GD
sites were 128 ± 3.9 and 2.17 ± 0.07 Bq g−1, respectively.
Considering quite similar ground surface conditions,
we can assume that the resuspension of 137Cs occurredsimilarly at both sites. In this case, if the resuspensionof 137Cs occurred very locally and/or the host particles
were not transported, the Rscale values would be approximately 0.017, which is the ratio of the GD surface
soil activity to that of GC. In contrast, when the resuspension occurred at larger spatial scales and/or the hostparticles could be distantly transported and well mixed,
Rscale could be nearly one. Therefore, Rscale is expected
to indicate the spatial extent of the resuspension, i.e.,
Rscale is determined by a combination of (1) the extent

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of those areas wherein similar resuspensions occur and(2) the transport distance of the host particles.
Figure 7a shows that the time series of Rscale and its
median values varied from 0.24 to 0.82 and were larger
than the ratio of the ground surface activities at GD to
GC (0.017), indicating that the activity concentration at
the GD site was not only governed by local resuspension
but also by one in the broader areas around the GD site.
These observations also indicate that the host particles
in the atmosphere were not completely mixed by theirtransport/diffusion, especially when Rscale was low.
Because the absolute value of Rscale probably depends onthe extent of the decontaminated area around the GD
site as well as on the ratio of the surface radioactivity
densities at both sites, comparisons of Rscale values atvarious places are meaningless. Therefore, the Rscalevalues indicate the relative variations of the emission
scales and the magnitudes of the transport/mixing at a
specific pair of sites.
Note that Rscale showed clear seasonal variations,such that smaller values occurred in winter/spring andlarger values occurred in summer/autumn. In addition,these observed variations were synchronized withthose of the fractions of activity of the aerosol particles
trapped at the stage #2 substrate and the backup filterin the impactor, as shown in Fig. 4. Figure 7b compares
the time series of the normalized 137Cs(7) (the red line)with that of the Rscale values, and Fig. 7c shows thecorrelation between Rscale and the normalized 137Cs(7).
The green lines in Fig. 7a show the median values(since the data quantities are not sufficient, we used
median values instead of average ones) of Rscale during
summer/autumn, which were 0.60 + 0.07/− 0.05 (central 50% range) in 2013 and 0.61 + 0.21/− 0.03 in 2014,
and the blue lines show those during winter/spring,which were 0.33 + 0.1/− 0.09 in 2012–2013 and 0.43 +0.11/− 0.09 in 2013–2014. Comparing Fig. 7a, b, theRscale and normalized 137Cs(7) values show nearly synchronized seasonal variations with each other. Fig. 7cshows an evident correlation between them. These results can be interpreted both by the transport of hostparticles and by the extent of the resuspension area.
The transport/mixing of resuspended host particles depends on the wind speed and particle deposition speed.
Considering that the average wind speed was lower insummer/autumn than in winter/spring, as shown in
Fig. 3b, the Rscale decrease in summer/autumn indicates that the deposition speed should decrease considerably as the properties of the particle change (i.e.,changes in the size and mass density or both of thehost particle). In addition, if the resuspension activity
in the areas surrounding the GD sites was similar tothat around the GC site in summer/autumn and was
lower in winter/spring, the observed seasonal

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

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Fig. 7 Time series of a Rscale and b normalized 137Cs(7) at the GC site and c the correlation plot of Rscale and the normalized 137Cs(7). Green andblue curves in a are the medians of Rscale in summer/autumn and winter/spring, respectively. The error bars for each curve in a and b show therange of central 50%

variations could have occurred. However, because theseasonal variations of the Rscale values were synchronized with the normalized 137Cs(7), this seasonal variation of the resuspension scale can (at least partly) be
attributed to the seasonal changes of the host particle
materials.

Possible host particles for radiocesium resuspension

Radiocesium was primarily emitted in the form of submicron water-soluble aerosols during the FDNPP accident (e.g., Kaneyasu et al. 2012). Another type ofprimary radiocesium emission has been reported toconsist of glassy, water-insoluble particles with diametersof approximately 2 μm (Adachi et al. 2013; Abe et al.
2014; Yamaguchi et al. 2016) and with high specific activity (Igarashi et al. 2014). These particles are stable atleast for a few years in the environment and have recently been found at the ground surface (Satou et al.
2016). If such particles are scattered by wind, they would
contribute to atmospheric radiocesium. However, in thepresent observations, we could not find serious concentration increases due to such insoluble radiocesiumbearing particles.
Studies performed after the Chernobyl accident
showed that biomass burning was a plausible

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

resuspension process of radiocesium in the contaminated area (Bourcier et al. 2010; Igarashi 2009; Lujanienė
et al. 2009; Yoschenko et al. 2006). Thus, this study
tested the hypothesis of resuspension due to biomassburning. Levoglucosan is known to be an organic tracerof biomass burning (Simoneit et al. 1999), and the positive correlation of radiocesium with levoglucosan indicates the contribution of biomass burning toradiocesium re-emission (Bourcier et al. 2010). In Marchand May, biomass burning usually increases in the
Fukushima prefecture (Local government of Fukushima
prefecture 2014). A forest fire occurred on March 17,
2013, at Ootaki, approximately 28 km away from the observation sites (Real time disaster information of
Fukushima). We compared the concentration of 137Cs
activity with that of levoglucosan in the air. Figure 8
shows the time series of (a) the concentration of the137
Cs activity and (b) that of levoglucosan at GC. In
addition, we also compared them with water-soluble organic carbon (WSOC), which is not a tracer of biomassburning but is emitted from various organic aerosol
sources. Although the levoglucosan and WSOC showedsimilar trends (except for short-term spike changes),
there was little correlation between the concentrationsof 137Cs activity and the levoglucosan and WSOC concentrations in the present work. We focused on theevents of the biomass burning on March 17, 2013, andthe singular increasing data on March 25. The forest fireon March 17 did not affect the levoglucosan and WSOC
concentrations at this observation site. Figure 9a, b illustrates the maps showing the air mass forward and backward trajectories calculated using the NOAA HYSPLIT
model (Stein et al. 2015) during the forest fire event periods. Figure 9a shows the forward trajectories from thefire site on March 17 at 16:00 (LT) to March 18 at 13:00.

Page 11 of 17

The air mass did not pass over the observational sites;
instead, it went north. Therefore, the increased concentration of 137Cs activity on March 17 was not attributable to the biomass burning event. For the event on
March 25, with an increased levoglucosan concentration,
the backward trajectories from the measurement site
GC show two origin areas (Fig. 9b, which is calculated
from March 24 at 13:00 (LT) to March 25 at 13:00 (LT)):the Sea of Japan and the mountainous areas to thenorthwest which is the local contaminated areas. In
these areas, there was no official information about thefire disasters at that time, although the possibility ofsmall fires is not fully neglected. Overall, the present results indicate that biomass burning was not a dominantsource of the resuspension during our observational
period at the sites.
The elements in the aerosol samples taken at the GC
and GD sites were analyzed quantitatively using SEMEDS, showing that the major components were different
in winter/spring and summer/autumn. Figure 10a showsa SEM image of the sample filter obtained at the GD sitein July 2015 (summer), and Fig. 10b–e are the element
map images of C, Fe, Al, and Na, respectively, in thesame viewing field with Fig. 10a. Numerous spots of C
indicate that carbonaceous aerosols were composed ofmajor coarse particles during summer/autumn, and theAl, Fe, and Na spots probably represent that soil/mineraldust, metal, and (sea) salt particles were rare. We tried
to identify these carbonaceous particles as some bioaerosol particles on the basis of their sizes and shapes on the
SEM image (Carrera et al. 2007; Chaturvedi et al. 1998;
Healy et al. 2012; Laucks et al. 2000). Figure 11a shows
an example image of these organic particles in a wide
field of view with close-up images of pollen (Fig. 11b), a
spore (Fig. 11c), and microorganisms (indicated by theFig. 8 Time series of a the radiocesium concentrations and b the concentrations of levoglucosan and WSOC in the atmosphere at GC

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

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Fig. 9 Maps show the forward and backward trajectories calculated with NOAA HYSPLIT model (Stein et al. 2015) during the forest fire event.
a The forward trajectories from Ootaki (forest fire area), which are calculated from March 17 at 16:00 (LT) to March 18 at 13:00 (LT). b Backward
trajectories from the measurement site (GC), which are calculated from March 24 at 13:00 (LT) to March 25 at 13:00 (LT)red circle in Fig. 11d) at the GD site. However, it wasdifficult to distinguish spores and microorganism particles only from the SEM images in the present work. We
separately counted pollen and microorganism/spores in
the SEM images at × 500 and × 1000 magnifications, respectively, of the four sample filters obtained at the GD
site in summer (July and August 2014) and winter(December 2014 and January 2015). To compare themagnitudes of these particle counts, the count numbersof the four filter samples were normalized by the sampling air volume and were scaled by the values for theJanuary sample. Figure 12 shows the scaled particle
counts of pollen and microorganism/spores with their
Poisson standard deviations and compares them withthe concentrations of 137Cs activity. Because so many
particles had accumulated and overlapped with eachother on the sample filters, it was difficult to count theseparticles precisely. Thus, the absolute values were nothighly accurate, and the counts showed only relative
amounts. The number of microorganisms/spores as wellas that of pollen was much larger in summer than thosein winter. This result indicates that these bioaerosol
emissions can significantly contribute to the resuspension process. Several studies suggested that 137Cs processed by the ecosystem can be re-emitted into theatmosphere as bioaerosols. For instance, Kuwahara et al.
(2005) showed that mushrooms and microorganisms
could absorb and accumulate 137Cs from the soil.
Furthermore, Kanasashi et al. (2015) showed that theacropetal translocation of 137Cs occurred in Japanesecedars and that their pollen contained 137Cs. Several
studies have described the processes of bioaerosol

emissions and the seasonal variations in the bioaerosol
concentrations and residence time. For instance, Jonesand Harrison (2004) showed that bacteria are emitted
from nearly all surfaces, such as the ground, vegetation,and so on. Burrows et al. (2009a, 2009b) and Schumacher
et al. (2013) described an increase in the bioaerosol concentrations and an extension of the residence time during
the warm season because of the greater turbulence andvertical mixing. The results in these studies are consistent
with our results in that bioaerosols, such as microorganism, spores, and pollen, which are actively emitted fromnearly all surfaces in the contaminated areas in summer/
autumn, would increase 137Cs in the atmosphere. The
Rscale values approached unity, as shown in Fig. 7, in theseseasons, which can be attributed to longer bioaerosol residence times, which would result in well-mixed atmospheric 137Cs. In this study, the sampling sites are locatedin low mountainous areas and are surrounded by forests,
and thus, it is natural that the increases of bioaerosols significantly contribute to resuspension, unlike in urban
areas. Kajino et al. (2016) simulated resuspension fromthe soil (Ishizuka et al. 2017) and that from vegetation (details were showed in the “Introduction” section), succeeding in reconstructing the 137Cs activity concentrationlevels that we observed. In their results, most of atmospheric 137Cs in summer/autumn was supplied by biogenic
emissions, which is consistent with our results. Thesestudies show that the enhancement of ecosystem activity
in summer/autumn could allow increased emissions oflarger amounts of pollen, spores, and microorganisms
from contaminated forests than those in winter, resulting
in a more active resuspension of 137Cs.

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

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Fig. 11 Examples of organic particles in the samples obtained in
September 2014 at the GD site. a A SEM image of the wide visible
range. b Pollen. c Spore. d Microorganisms (surrounded by red circles)

Fig. 10 The SEM image and elemental mapping images of thesample filters obtained at the GD site in July 2015. a The SEM image,(b) the elemental map of C, (c) the elemental map of Fe, (d) theelemental map of Al, and (e) the elemental map of Na. The
characteristic organic particles commonly exist in the filter samples

Figure 13a–e shows the SEM and element map images, which are similar to Fig. 10a–e, of the sample filters obtained at the GD site in January 2015. In contrast
to Fig. 10, Fig. 13 shows the soil/mineral dust particles
indicated by the numerous Al spots that were thedominant coarse aerosols observed in winter/spring. Although some C spots were detected, they were mostly
debris from vegetation rather than from living bioaerosols because these particles have different shapes andsizes than pollen, spores, and microorganisms. The
smaller values of Rscale in Fig. 7 suggest that resuspension would occur from local surfaces, such as the playground surface, and that the deposition speeds of thehost particles were high in winter/spring. Because there
were no significant sources of metal and salt particles
around the observational sites, these particles were probably emitted by sources far from the observational sites(Igarashi 2009) and were unlikely to be major hostparticles. On the basis of these results and considerations, the soil and mineral dust particles from nearby
ground surfaces as well as the debris from the vegetation
were the most likely host particles of the 137Cs resuspension in winter/spring.
Many studies have indicated that 137Cs is firmly
adsorbed to clay mineral particles (e.g., Dumat andStaunton 1999; Igarashi et al. 2005; Masson et al.
2010; Mukai et al. 2014; Ishizuka et al. 2017), and thetransport of mineral dust particles by wind is considered to be a significant resuspension process. As
shown in Fig. 3, we can see a positive correlation between the radiocesium concentrations and windspeeds in spring, indicating that the soil/mineral dust,
pollen, forest debris, etc., that were suspended bywind blow were plausible host particles of the resuspension. Ishizuka et al. (2017) used the winter/spring
samples obtained with the cascade impactor at theGC site, which was also used in this study, and evaluated the relation between the soil/mineral dust
particles and activity concentrations of 137Cs during
winter/spring; then, they compared these relationships with those estimated using a size-resolved onedimensional resuspension scheme, and theirconclusion is consistent with our results.

Conclusions
To understand the resuspension process of radiocesiumin areas heavily contaminated by the FDNPP accident,

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

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Fig. 12 The normalized particle count of the bioaerosols using SEM (in relative values, the error bar shows the Poisson standard deviation) andatmospheric concentrations of 137Cs activity sampled at the GD site during summer (July and August 2014) and winter (December 2014 andJanuary 2015). The counts of microorganisms include those of spores (see the text)

we conducted long-term measurements of atmospheric radiocesium with high-volume air samplers
and impactors in the Tsushima and Yamakiya areas of
the Fukushima prefecture. We analyzed the temporalvariations of the radiocesium concentrations and examined the particles on the sample substrates of the
impactor and its filters using SEM-EDS. We also introduced R scale, which is the ratio of the 137Cs air concentrations from two close sites with similar land
uses but different contamination levels, in order toestimate the spatial extent of the resuspension. The
main finding in this study is that the seasonal differences of both the radiocesium resuspension processes
and the host particle material can be described asshown below:
1. In summer and autumn, the atmosphericradiocesium activity concentration was higher,
and both the higher normalized 137Cs(1–5) values
and SEM-EDS analysis results indicated thatcoarse organic particles, such as spores, pollen,
and microorganisms, with adhesive organic matter were the dominant host particles of radiocesium resuspension. The larger values of Rscale inthese seasons are consistent with this result, considering that the deposition velocity of the lightbiogenic organic particles was probably low andthat the resuspended radiocesium on them should
be well mixed. These organic particles could beemitted from forest ecosystems.

2. In winter and early spring, the atmosphericradiocesium activity concentrations were lower, andthey increased in May. The higher normalized
137
Cs(7) values and the SEM-EDS analysis results
showed that the dominant coarse particles were dust
(soil and mineral) and vegetation debris and thatthese were the dominant host particles for radiocesium resuspension. The temporal variations of theradiocesium concentrations showed similar variations to those of wind speed, suggesting that windsuspension of contaminated particles from theground surface is considered as the main resuspension process in these seasons. The smaller Rscalevalues in these seasons are still consistent with thisresult, considering that the deposition velocity of thedust particles with higher density was probably highand that the resuspended radiocesium on themshould be less mixed.
3. No significant correlation between theconcentrations of levoglucosan and WSOC and thatof radiocesium was found in spring, indicating thatbiomass burning was not a significant resuspension
process of radiocesium.
4. Although many studies have suggested that soil/
mineral dusts and bioaerosols could be major hostparticles of 134, 137Cs, we first identified the seasonalvariations of the dominant resuspension processes/
host particles. Further research on bioaerosols, suchas working to identify specific bioaerosols andevaluating the quantitative relationship between

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

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for the periodical maintenance of the measurement stations in the field;
members of Ibaraki University and the College of Engineering of Ibaraki
University for their maintenance of the tools after sampling and work thesetting field instruments; Drs. T. Sekiyama (MRI), M. Kajino (MRI), Y. Hatano(University of Tsukuba), and T. Maki (Kanazawa University) for their
discussions; and the many people who cooperated this study. We also
express our gratitude to the local governments of Kawamata and Namie aswell as to Fukushima prefectural government, who kindly offered thismonitoring opportunity.
Funding
This work was financially supported by MEXT KAKENHI (a Grant-in-Aid for
Scientific Research on Innovative Areas under the A01–1, A01–2 and A04–8
research teams in the “Interdisciplinary Study on Environmental Transfer ofRadionuclides from the Fukushima Daiichi NPP Accident”; the grant numbers
are 24110002, 24110003, and 24110009, respectively) and the MEXT JapaneseRadioactivity Survey at the MRI. Additionally, this study was partly supported bya grant from the Asahi group foundation at Ibaraki University. Thus, this study
was discussed and supported by the workshop of ISET-R for writing articles.
Authors’ contributions
TK collected and analyzed the samples and summarized these data in thisarticle. KK planned and carried out the measurements as well as the samplings.
YI is the leader of our project team, helped in writing the manuscript, andprovided important discussions for this article. KA and YZ supported theidentification of the aerosols by SEM-EDS. KN, AS, ST, KY, and NY analyzed theradioactivity of filter samples. HOk and HOg analyzed the levoglucosan and
WSOC of filter samples. MI and HD supported the maintenance ofmeasurements in the field. MM and YO provided important suggestions for thisarticle. All authors read and approved the final manuscript.

Fig. 13 The SEM and elemental mapping images of the sample
filter obtained at the GD site in January 2015. a The SEM image, (b)the elemental map of C, (c) the elemental map of Fe, (d) theelemental map of Al, and (e) the elemental map of Na. Major coarse
particles include Al (clay mineral) and Na (salt or mineral)bioaerosols and radiocesium, is required. In the near
future, we will include more observations of
bioaerosols to reveal more detailed radiocesium
resuspension processes. A flux measurement ofthese particles would be necessary to quantitatively
evaluate the radiocesium resuspension rate.
Abbreviations
EDS: Energy-dispersive X-ray spectroscopy; FAJ: Forestry Agency Japan;
FDNPP: Fukushima Daiichi Nuclear Power Plant; HYSPLIT: The Hybrid Single
Particle Lagrangian Integrated Trajectory Model; JAEA: Japan Atomic Energy
Agency; MEXT: Ministry of Education, Culture, Sports, Science andTechnology, Japan; MRI: Meteorological Research Institute; NDVI: Normalized
Difference Vegetation Index; NOAA: National Oceanic and Atmospheric
Administration; SEM: Scanning electron microscopy; TEPCO: Tokyo ElectricPower Company; WSOC: Water-soluble organic carbon
Acknowledgements
We acknowledge T. Kimura, K. Inukai, K. Kamioka (Atox Co. Ltd.), S. Kakitani(Osaka University), and N. Hayashi (Ibaraki University) for their sample
preparation and measurements of radioactivity; T. Kanari (Green Blue Co. Ltd.)

Authors’ information
TK studied atmospheric environments, especially focusing on black carbon(BC) aerosols, and received his PhD in atmospheric chemistry from Ibaraki
University through the study of Antarctic aerosol science. He has beenworking at Hitachi Power Solutions Co., Ltd. since 2016 and wrote this article
while working. In his studies, he participated in the project of the workshop
of Interdisciplinary Study on Environmental Transfer of Radionuclides fromthe Fukushima Daiichi NPP Accident (ISET-R) to shed light on theenvironmental transfer of radionuclides and to contribute to the revival of
the affected areas in the Fukushima prefecture.
KK received his PhD from Tokyo University in 1990 and worked at Tokyo
University until 2002 as a research assistant. He moved to Ibaraki University
in 2002. His studies focus on the atmospheric O3 and NO2, the effects of BC
on global warming, and the resuspension processes of radiocesium.
YI received his PhD from the University of Tsukuba in 1987. From 1987 to
1991, he worked at the former National Institute of Radiological Sciences andmoved to the MRI in 1991. His current interests are atmospheric aerosols;
their precursors, including Asian dust and PM2.5; their possible influences on
climate and general environmental changes; and other related phenomena.
KA received his PhD from Kobe University in 2005, worked at Arizona State
University between 2005 and 2011 as a postdoctoral/faculty research
associate and is currently studying atmospheric aerosols at the MRI.
KN received his PhD from Osaka University in 2007 and worked at Osaka
University in 2008. He moved to JAEA and worked as a research scientist
between 2009 and 2012; then, he returned to Osaka University. His main
research field is nuclear- and radiochemistry.
AS received his PhD from Osaka University in 1985 and then worked at theJapan Society for the Promotion of Science. He moved to Nagoya University,where he worked between 1987 and 1998. Then, he moved to Kyoto
University in 1998, before moving to Osaka University in 1999. His main
research field is nuclear- and radiochemistry.
HOk worked at Kanagawa University between 1991 and 1999 and received
his PhD from Tokyo University in 1997 while he was working. From 1999 to
2003, he worked at the University of East Anglia, UK. He moved to the Tokyo
Metropolitan Institute of Technology, where he worked from 2003 to 2005.
Then, he moved to Waseda University. His studies are focused on theenvironmental science topics of the atmosphere, hydrosphere andecosystems of the forest, and on the development of a technique forenvironmental remediation via soft-chemical methods.

Kinase et al. Progress in Earth and Planetary Science (2018) 5:12

HOg received her PhD from the Prefectural University of Kumamoto in 2011.
She worked at Waseda University between 2011 and 2016, and then, she
moved to Sibata Scientific Technology Ltd. Her main research field is
atmospheric aerosols.
MI received his PhD from Osaka University in 2000 and then worked at theFaculty of Systems Engineering, Wakayama University from 2000 to 2005. He
moved to the Faculty of Engineering, Kagawa University. His studies are
centered on the resuspension processes of radionuclides with dust particles
as forced by wind, the processes of the suspension of Asian dust, andresearching the hydrological environment and its processes.
ST received his PhD from the University of Tokyo in 2007 and started his
study on the development of analytical methods for isotopocules, i.e.,
isotopically substituted molecules, of atmospheric nitrous oxide (N2O) at
Nagoya University as a postdoctoral fellow of the Japan Science andTechnology Corporation. After moving to the Tokyo Institute of Technology
in 2008, he pursued an investigation of the geochemical cycle of N2O and iscurrently expanding his research interests into other trace gasses andaerosols as an associate professor.
KY received his PhD from Tokyo Metropolitan University in 1997. He has
worked at the Tokyo Institute of Technology since 2000. He studies therelationship between radionuclides and organics in aerosols.
NY received his PhD from the Tokyo Institute of Technology in 1984. He waswith the Mitsubishi-Kagaku Institute of Life Sciences, Toyama Univ., Nagoya
Univ., and has been a full professor in charge of two Environmental Departments of the Tokyo Institute of Technology since 1998 and is a Principal Investigator at the Earth-Life Science Institute. He studies “isotopomers:
isotopic substituted molecules” as powerful tracers that can help to reduce
uncertainties in the biological, chemical, physical, and anthropogenic sourceand sink processes of environmental molecules, especially for global
warming-related species.
YZ graduated the Meteorological College in 1984 and worked at theNagasaki Marine Observatory, Japan Meteorological Agency from 1984 to
1989. He has been studying atmospheric aerosols at the MRI.
MM received his PhD from Tohoku University in 1996 and has worked at the
MRI from 1985 to 2015. He held the position of Senior Researcher for
Research Affairs at the Institute. Currently, he is working at the Japan
Meteorological Business Support Center.
HD graduated the Faculty of Maritime Sciences of Kobe University in 2005
and worked at a general company of the manufacturing industry. In 2010,
he moved to the Technical Services Division of the College of Engineering,
Ibaraki University. He works as an engineer providing the technical support
at Ibaraki University and helped with our research.
YO graduated and received his PhD from the University of Tsukuba in 1990,
working at Nagoya University from 1992 to 1999. Currently, he is working at
the University of Tsukuba, where he researches hydrology and theenvironmental dynamics of radionuclides as well as being the PI of thewhole ISET-R project.
Competing interests
The authors declare that they have no competing interests.

Publisher’s NoteSpringer Nature remains neutral with regard to jurisdictional claims inpublished maps and institutional affiliations.
Author details
1
Graduate School of Science and Engineering, Ibaraki University, 2-1-1
Bunkyo, Mito, Ibaraki 310-8512, Japan. 2Hitachi Power Solutions Co., Ltd, 2-2-3
Saiwaicho, Hitachi, Ibaraki 317-0073, Japan. 3College of Science, Ibaraki
University, 2-1-1 Bunkyo, Mito, Ibaraki 310-8512, Japan. 4Atmospheric
Environment and Applied Meteorology Research Department,
Meteorological Research Institute, 1-1 Nagamine, Tsukuba, Ibaraki 305-0052,
Japan. 5Graduate School of Science, Osaka University, 1-1, Machikaneyama,
Toyonaka, Osaka 560-0043, Japan. 6Faculty of Science and Engineering,
Waseda University, 3-4-1 Okubo, Shinjuku, Tokyo 169-8555, Japan. 7Faculty ofEngineering, Kagawa University, 2217-20 Hayashi-cho, Takamatsu, Kagawa
761-0396, Japan. 8Department of Chemical Science and Engineering, Schoolof Materials and Chemical Technology, Tokyo Institute of Technology, 4259
Nagatsuta, Yokohama, Kanagawa 226-8502, Japan. 9Earth-Life Science
Institute, Tokyo Institute of Technology, 2-12-1 Ookayama, Meguro, Tokyo

Page 16 of 17

152-8551, Japan. 10Forecast Research Department, Meteorological ResearchInstitute, Japan Meteorological Agency, 1-1 Nagamine, Tsukuba, Ibaraki
305-0052, Japan. 11Japan Meteorological Business Support Center, 3-17
Nishikicho, Kanda, Chiyoda, Tokyo 101-0054, Japan. 12College of Engineering,
Ibaraki University, 4-12-1 Narusawa, Hitachi, Ibaraki 316-8511, Japan. 13Center
for Research in Isotopes and Environmental Dynamics, University of Tsukuba,
1-1-1 Tennodai, Tsukuba, Ibaraki 305-8577, Japan.
Received: 22 August 2016 Accepted: 9 February 2018

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Geochemical Journal, Vol. 52, pp. 137 to 143, 2018

doi:10.2343/geochemj.2.0514

Analysis of two forms of radioactive particles emitted during the early stagesof the Fukushima Dai-ichi Nuclear Power Station accident
YUKIHIKO SATOU,1* K EISUKE SUEKI ,1 K IMIKAZU SASA ,1 HIDEKI YOSHIKAWA,2 SHIGEO NAKAMA,2HARUKA MINOWA,3 YOSHINARI A BE,4 IZUMI NAKAI,4 TAKAHIRO ONO ,4 KOUJI ADACHI5 and YASUHITO IGARASHI5
1

Graduate School of Pure and Applied Sciences, University of Tsukuba, 1-1-1 Tennodai, Tsukuba, Ibaraki 305-8577, Japan
2
Fukushima Environmental Safety Center, Japan Atomic Energy Agency,
10-2 Fukasaku, Miharu, Tamura-gun, Fukushima 963-7700, Japan
3
Radioisotope Research Facility, Jikei University School of Medicine, 3-25-8, Nishi-shimbashi, Minato, Tokyo 105-8461, Japan
4
Department of Applied Chemistry, Faculty of Science, Tokyo University of Science,
1-3 Kagurazaka, Shinjuku-ku, Tokyo 162-8601, Japan
5
Atmospheric Environment and Applied Meteorology Research Department, Meteorological Research Institute,
1-1 Nagamine, Tsukuba, Ibaraki 305-0052, Japan(Received August 1, 2017; Accepted December 11, 2017)
Two types of radioactive particles were isolated from environmental samples collected at various distances fromthe Fukushima Dai-ichi Nuclear Power Station. “Type A” particles are 2–10 mm in diameter and display characteristic Cs
X-ray emissions when analyzed using energy-dispersive X-ray spectrometry (EDS). “Type B” particles are considerablylarger, up to 400 mm in diameter, with Cs concentrations too low to be detectable with EDS. These larger particles wereisolated from the region north of the nuclear reactor site, which was contaminated on March 12, 2011. The specificactivity of Type B particles is much lower than Type A, and the mean 134Cs/137 Cs ratios are ~0.93 and 1.04, respectively.
The Type B ratio indicates that power station Unit 1 is the source, implying that these larger radioactive particles weredischarged on March 12. This study found that different type of radioactive particles were released not only on March 15
but also on March 12.
Keywords: Fukushima Dai-ichi Nuclear Power Station accident, radioactive particles, caesium isotopic ratio, hydrogen explosion, specific radioactivity

The elements Fe, Zn, and Pb were also present. It has
since been reported that very similar micron-scale spherical particles have been found 20 km northwest and 3 km
south of the F1NPS (Satou et al., 2016; Furuki et al.,2017). Similar particles were isolated from dust on nonwoven fabric cloth and needles of Japanese cedar(Yamaguchi et al., 2016; Kogure et al., 2016). This revealed that such Cs-bearing particles were widely dispersed within the Fukushima region. Although their origin could be attributable to any of the various release
events that occurred at the F1NPS, the exact source ofthe radionuclides is unclear. In this study, roadside soil
and dust samples from the immediate vicinity of the
F1NPS were analyzed using scanning electron microscopy(SEM), EDS, synchrotron radiation micro-beam X-ray
fluorescence (SR- m -XRF) analysis, and gamma
spectrometry, with the aim of attributing sources to them.

INTRODUCTION
Large quantities of radioactive materials were emitted to the environment during the March 2011 Fukushima
Dai-ichi Nuclear Power Station (F1NPS) accident(Kinoshita et al., 2011). Radioactive Cs is one of the mostprevalent sources of radioactive pollution and is the mostwidely studied radionuclides dispersed into the environment as a result of the accident. Radioactive particles,
which are named Cs-bearing radioactive particles, werediscovered in air dust samples collected from 9 pm on
March 14 to 9 am on March 15 as far away as Tsukuba,
170 km south of the F1NPS (Adachi et al., 2013). Cs concentrations in the particles were sufficiently high foranalysis by energy-dispersive X-ray spectrometry (EDS).
*Corresponding author (e-mail: satou.yukihiko@jaea.go.jp)*Present address: Collaborative Laboratories for Advanced
Decommissioning Science (CLADS), Japan Atomic Energy Agency(JAEA), 790-1 Otsuka, Motooka Tomioka, Futaba, Fukushima 9791151, Japan.

MATERIALS AND METHODS
Sampling location
Soil and dust samples were collected within the highly

Copyright © 2018 by The Geochemical Society of Japan.

137

Fig. 1. Sampling locations. “Satou et al.” indicates the location of the fragmented and spherical Cs-bearing particles studied ofSatou et al. (2016). “F1NPS” indicates the center of the Fukushima Dai-ichi Nuclear Power Station.

contaminated zone around the F1NPS from June 2013 toJanuary 2015, at the locations indicated in Fig. 1 and Table 1. To distinguish between emissions from different
sources, samples were collected along two transects extending away from the F1NPS. Contamination on thenorthern sampling line is understood to have been generated by the hydrogen explosion that occurred within Unit
1 on March 12, 2011 (Satou et al., 2015; Chino et al.,2016). The method used to sample core soil follows Satou
138 Y. Satou et al.

et al. (2016). Surface soils were collected by shovel anddust samples were picked up by hand.
Particle extraction
The radioactive particles were extracted from the soil
and dust samples using methods reported in previous studies (Adachi et al., 2013; Satou et al., 2016). Autoradiography, combined with a phosphorus imaging plate andlaser readout system, was used to locate radioactive par-

Table 1. Details of the sampling locations
Sample ID

Sampling date Location

Longitude

Latitude

Sample type

NM-06
NM-14
FT-03
FT-13
CF-01

13/06/2013
01/08/2013
12/10/2014
28/11/2014
23/01/2015

37.477137.4651
37.471437.417837.4379140.9857
140.9272
140.9998
140.9999
141.0223

Core soilCore soil
Bulk soil
Bulk soil
Dust

Paddy field
Paddy field
Wasteland
Wasteland
Pavement

Fig. 2. Gamma-ray spectra of representative particles. Radionuclides are labelled on their main energy peaks.

ticles within the samples, with an exposure time of only
five min due to the high specific activity of the former.
After the sample contents were sorted, radioactive particles were isolated by hand under an optical microscopeusing a micro-manipulator system. Radioactivity wasconfirmed using a Geiger-Müller survey meter and subsequent more accurate gamma radiation measurements.
Analysis of radionuclides
Particles were analyzed for gamma-emitting
radionuclides using high-resolution gamma spectrometry
with a hyper-pure germanium detector. The detector efficiency was calibrated using standard point 134Cs and 137Cs
gamma sources (<1 mm diameter) from the Japan Radioisotope Association.
Particle observation and analysis
Backscattered electron imaging (BEI) with an SEM
and associated EDS instruments (SEM: SU 3500, Hitachi

High-Technologies Co., Tokyo, Japan; EDS: X-max 50
mm, Horiba Ltd., Kyoto, Japan) were used to study particle surfaces and analyze their compositions. In cases
where Cs gamma emissions were detected but the concentration was below EDS detection limits, SR-m-XRF
was performed as a 37.5 keV monochromatic X-ray beam
enables detection of Cs K-lines without interference fromother elements. The technical procedures followed are as
described in Abe et al. (2014) and Ono et al. (2017).
RESULTS AND DISCUSSIONRadioactivity and morphology
A total of 19 particles were extracted from the soil
and dust samples, and all were stable despite exposure toenvironment for more than four years after the accident.
134
Cs and 137Cs were detected in all particles. Particle CF01#01 contained >20 kBq 137Cs and 300 Bq 125Sb; bothof these are volatile elements released from nuclear fuelTwo different types of radioactive cesium particle 139

Fig. 3. Backscatter electron images of representative radioactive particles from the environmental samples collected around the
Fukushima Dai-ichi Nuclear Power Station.

Fig. 4. Elemental dispersive X-ray spectra of representative particles. Note the absence of Cs peaks in the Type B particles.

(Pontillon and Ducros, 2009) (Fig. 2). Other radionuclides
such as 60Co and 103Ru, which were found in hot fuel particles around the Chernobyl Nuclear Power Station(ChNPS), were not detected (Devell et al., 1986).
BEI images of several representative particles areshown in Fig. 3. Particles isolated from the northern area(samples CF-01, FT-03, and NM-06) are much larger than
140 Y. Satou et al.

those from the western region (FT-13 and NM-14), withmaximum dimensions of up to 400 mm. Both sphericaland fragmented particles were found at location FT-03, 5
km north of the F1NPS. Similar observations were reported by Satou et al. (2016). Particles are evidently not
sorted by shape during transport and deposition close to
the reactors, with various forms being deposited.

Fig. 5. Micro-beam maps of elemental distributions, showing uneven patterns. Data reconstructed from Ono et al. (2017).

Table 2. Characteristics of Type A and Type B particles
Particle type
Characteristic
Size distribution (observed)
134

Fig. 6. Scatter-plot of 134Cs/137 Cs ratio vs 137 C content (including data from Satou et al. (2016)), showing the separationof particles into distinct northern and western groups.

Elemental composition of particles
Figure 4 shows the EDS spectra, obtained by SEMEDS, of representative particles from the northern area
and western region. The compositions of each particle
are very similar, primarily silicate, which indicates that
unlike in the ChNPS accident, they are not fuel particles(Salbu and Kerkling, 1998). The elemental compositionsof particles from the western region are very similar to
Cs-bearing particles found 20 km northwest and 170 km
south of the F1NPS (Adachi et al., 2013; Satou et al.,2016), with Fe and Zn also being detectable from theircharacteristic X-rays. Particles from the northern area areslightly different, containing Al, Sn, Cl, Fe, and Mg, withAB

1-10 m m

70-400 m m
0.93
Antimony-125 (1 2 5Sb)
limited (North)
March 12, 2011
Unit 1Low
synchrotron X-ray

Cs/1 3 7Cs (mean)
Other radionuclides
Distribution
Suspected emission date
Source reactor(s)Specific radioactivity
Cs detection method of

1.04
N/A
wide
March 15, 2011
Unit 2 or 3
high
SEM-EDS

Elements contained
Elements in common

Fe, Sn, Cl

Na, Mg, Ca, Ba
Si, O, Zn

Cs concentrations that were too low for X-ray detection.
SR-m-XRF mapping was applied to determine the distribution of Cs and other heavy elements in individual
particles. A representative result is shown in Fig. 5. Detailed SR-m-XRF results have been published previously(Ono et al., 2017). Particles from the western region,
which are similar to Cs bearing particles, show uniform
distribution of Cs and other elements (Satou et al., 2016).
In contrast, SR-m-XRF results of particles from the northern area show an uneven distribution of elements. This
result indicates that based on Cs distribution, there are
two types of radioactive Cs-bearing particles in the vicinity of the F1NPS site.
Sources of particles
It is possible that the sampling locations were contaminated with radioactive materials from different reactor units. Other studies have shown that the northern and
Two different types of radioactive cesium particle 141

Fig. 7. Scatter-plot of 137Cs content vs particle volume. The “170 km south” and “20 km northwest” labels indicate particles
collected by Adachi et al. (2013) and Satou et al. (2016), respectively.

western areas were contaminated by Unit 1 and Units 2
or 3, respectively (Satou et al., 2015; Chino et al., 2016).
The Cs isotopic ratios (134Cs/ 137Cs), which were decay
corrected as of March 11, 2011, form two clearly definedgroups of particles, as shown in Fig. 6: particles in thelower group were extracted from the northern area (samples NM-06, FT-03, and CF-01), while those in the uppergroup were from the western region (including samples
from 20 km northwest; Satou et al., 2016). The 134Cs/
137
Cs ratios of 0.93 and 1.04 (the lower and upper groups,
respectively, in Fig. 6) that allowed us to distinguish theparticles from Unit 1 and Units 2 or 3, respectively, aresimilar to those obtained using ORIGEN-II performed
with JENDL data (Nishihara et al., 2012).
A particle isolated from sample CF-01 contained 300
Bq of 125Sb, which indicated that the temperature of nuclear fuel was high enough to produce Sb vapor (Pontillon
and Ducros, 2009) and that the surrounding environment
was hot enough for the Sb to be transported to the place
where the particle was formed. One of the characteristics
of Unit 1 was the release of radioactive materials fromthe reactor at a high temperature. Furthermore, as the dispersion of large particles over significant distances requires considerable energy, meaning that such particles
detected in the northern area, within 15 km of the F1NPS,
would have been released from the F1NPS during thehydrogen explosion in Unit 1 on March 12.
142 Y. Satou et al.

Particles collected in the northern area are thus considered to have originated from Unit 1.
In contrast, the Cs isotopic ratios in particles from thewestern region are similar to those present in Units 2 and3 at the time of the accident, although it is not possible todistinguish between these two units as they were very
similar (Unit 2, 1.08; Unit 3, 1.05; Nishihara et al., 2012).
A more accurate method to measure radioactive Cs in theparticles was required to enable a distinction and distinctly
identify the source of particles in the western region.
Based on the results of the present study, two types of
Cs-bearing particles, Cs isotopic ratio and minor-element
content, are defined on the basis of sampling location:
Type A, from samples FT-13 and NM-14, and Type B,
from samples CF-01, FT-03, and NM-06. The characteristics of the two types are summarized in Table 2. Type A
particles are similar to the Cs-bearing particles reported
in previous studies (Adachi et al., 2013; Satou et al., 2016;
Yamaguchi et al., 2016; Furuki et al., 2017).
Particle-specific activity
Specific activity varies between particles, as thescatterplot of particle volume vs 137Cs activity shown in
Fig. 7 (the volume of fragmented particles was calculatedusing the equivalent diameter). There is little variation in
specific activity for Type A particles, with significant
correlation between activity and volume (r2 = 0.75), simi-

lar to the results of a previous study (Satou et al., 2016).
Type B particles are large in size and have a high Cs activity but at a lower specific activity than Type A particles. Type B particles have significant variability and poorcorrelation between activity and volume (r2 = 0.10), which
explains why the Cs was not detected in EDS X-rays.
Particle-specific activity can thus be used to indicate thepossibility of different generation and release processes
at F1NPS thus providing an additional means of classifying particles.
CONCLUSION
This study identified two types of radioactive particles, Type A and Type B, which are both composed mainlyof silicate but with contain different compositions of otherelements and have different 134 Cs/137Cs ratios and specific activities. Type B particles were isolated from thearea north of the F1NPS, which was contaminated on
March 12, 2011. Their Cs isotopic ratio is consistent withthe theoretical ratio of F1NPS Unit 1, which suggests thatthe particles were emitted on March 12, 2011. This result
indicates that radioactive particles were emitted to theenvironment not only on March 15, but also on March
12. In addition, radioactive particles emitted during the
F1NPS accident are markedly different from those released from the ChNPS, and persist in the environment
five years after the event.
Acknowledgments—The authors are indebted to Tsuyoshi
Iijima, Kenso Fujiwara, Terumi Dohi, Tetsuya Matsunaka,
Tsutomu Takahashi, and Nao Shibayama for their help withsample collection. We also express thanks to Yushin Iizawa at
Tokyo University of Science for assistance with the analyses inSPring-8. In addition, the authors are grateful to David A.
Richards and Peter G. Martin at the University of Bristol for
discussion and English improvements.
This study was supported by MEXT/JSPS KAKENHI grants(numbers 24110003, 24110006, and 15H02340) and by a JAEAAcademia joint study.
The synchrotron radiation experiments were performed at
the BL37XU of SPring-8 with the approval of the Japan Synchrotron Radiation Research Institute (JASRI) (proposal numbers 2014B1493, 2015A1895, 2015B1828, and 2016A1705).

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Two different types of radioactive cesium particle 143ö÷øùúû÷ÿýûùþÿ0÷ÿ

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Anthropocene 14 (2016) 71–76

Contents lists available at ScienceDirect

Anthropocene
journal homepage: www.elsevier.com/locate/ancene

Short communication

First successful isolation of radioactive particles from soil near the
Fukushima Daiichi Nuclear Power Plant
Yukihiko Satoua,* , Keisuke Suekib , Kimikazu Sasab , Kouji Adachic , Yasuhito IgarashicaGraduate School of Pure and Applied Sciences, University of Tsukuba, 1-1-1 Tennodai, Tsukuba, Ibaraki 305-8577, Japan
Faculty of Pure and Applied Sciences, University of Tsukuba, 1-1-1 Tennodai, Tsukuba, Ibaraki 305-8577, Japanc
Atmospheric Environment and Applied Meteorology Research Department, Meteorological Research Institute, 1-1 Nagamine, Tsukuba, Ibaraki 305-0052,
Japanb

A R T I C L E I N F O

Article history:
Received 25 November 2015
Received in revised form 27 April 2016
Accepted 8 May 2016
Available online 10 May 2016
Keywords:
Fukushima Dai-ichi Nuclear Power Plant
accident
Silicate compounds
Radiocaesium-bearing particle

A B S T R A C T

The Fukushima Daiichi Nuclear Power Plant (F1NPP) accident in 2011 released radionuclides into theatmosphere in both aerosol and gaseous form. Subsequent studies of contamination in the environment
have focused on the bulk radioactivity in samples. Comparatively little is known about the relative
contribution and patterns of soluble versus particulate deposition of the radionuclides.
We investigated a sample of heavily contaminated surface soil from a site 20 km northwest of the
F1NPP and isolated four radioactive particles from the surrounding soil. These particles had a maximum
particle area equivalent diameter of 6.4 mm and a maximum 137Cs radioactivity of 67.5 ! 0.1 Bq per
particle. They were larger than the particles identiﬁed in aerosol samples shortly after the accident at alocation 170 km southwest of the F1NPP. Two of the particles were spherical and two were fragmental.
Silicates were a major component of the Fukushima radioactive particles. These characteristics clearly
differ from the so-called hot particles observed at the Chernobyl accident in 1986. Clarifying the physicaland chemical properties of the radionuclides released from the F1NPP accident is important for assessing
the potential long-term impacts to humans.
ã 2016 Elsevier Ltd. All rights reserved.

1. Introduction
The Fukushima Daiichi Nuclear Power Plant (F1NPP) accident,
caused by tsunami damage following the Great East Japan
Earthquake on 11 March 2011, released enormous amounts ofradioactive material into the atmosphere. Radiocaesium (Radioactive Cs) was one of the most abundant radionuclides released into
the environment during the F1NPP accident, with an estimatedradioactive Cs release on the order of 9–36 PBq (Terada et al., 2012;
Stohl et al., 2012; Katata et al., 2015; Winiarek et al., 2012).
Understanding the physical and chemical properties of the emitted
radioactive caesium is important to accurately evaluate thepossible impacts to human health and to assess the long-term
distributions of these particles after deposition in residential areas,
agricultural ﬁelds, mountains, and aquatic environments. In theearly stages of the F1NPP accident, radioactive Cs-bearling
particles 2–10 mm in diameter was dominant (Doi et al., 2013;
Miyamoto et al., 2014). In addition, radioactive Cs were transported* Corresponding author.
E-mail address: yukihiko@ied.tsukuba.ac.jp (Y. Satou).
http://dx.doi.org/10.1016/j.ancene.2016.05.001
2213-3054/ã 2016 Elsevier Ltd. All rights reserved.

from catchment to downstream as suspended matters (Chartin
et al., 2013; Lepage et al., 2016). And, it was stored in river basin as
sediment (Kitamura et al., 2014). However, the chemical forms ofradioactive Cs were unclear at that time.
To investigate the chemical forms of radioactive Cs, Adachi et al.
(2013) isolated radioactive water-insoluble microspheres fromatmospheric samples collected on 14 and 15 March in Tsukuba,
which is approximately 170 km southwest of the F1NPP. The
radioactive particles were isolated from the aerosol samples usinga scanning electron microscope equipped with an energy
dispersive X-ray spectrometer (SEM–EDS). The particles contained
Cs, O, Fe, and Zn and were several micrometres in diameter; the137
Cs activity of the particles was approximately 2–3 Bq per
particle. These particles were collected from the plume that was
released on 14 and 15 March, but they were not observed in aerosol
samples from other plumes, such as one that occurred 20–21
March 2011. Abe et al. (2014) used state-of-the-art synchrotron
X-ray spectroscopy to examine the chemistry of the Cs-bearing
particles isolated by Adachi et al. (2013). They concluded that theparticles: (1) contained elements derived from nuclear ﬁssion
processes and from nuclear reactor and fuel materials; (2) were
amorphous; (3) were highly oxidized; and (4) consisted of glassy

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Y. Satou et al. / Anthropocene 14 (2016) 71–76

spherules formed from a molten mixture of nuclear fuel andreactor material. Yamaguchi et al. (2016) recently found similar Csbearing particles from deposits on non-woven fabric cloth and on aneedle of Japanese cedar in Fukushima and showed their internal
compositions using a focused ion beam and a transmission
electron microscope. In contrast, Kaneyasu et al. (2012) showedthat radioactive Cs was hosted in sulfate aerosols, as determined bya particle size analysis of the radioactive material and measurements of non-radioactive aerosol chemical components in April–
May 2011. Kaneyasu et al. (2012) argued that the Cs-bearing sulfate
aerosols were submicron in size and soluble in water; suchparticles would behave differently than the Cs-bearing particles
reported by Adachi et al. (2013). Therefore, it is thought that thephysical and chemical forms of the radioactive Cs emitted during
the F1NPP accident, as well as their release times, varied
considerably. In particular, soluble radioactive Cs has been
assumed to distribute rather uniformly in the dose calculationin human body. However, the presence of high concentrated
insoluble radioactive Cs-bearing particle in the early F1NPP plume
could result in underestimation of internal exposure in humanbody who had been exposed to such plume. For instance, effective
dose coefﬁcients for inhalation become the larger if radioactive Cs
is the less soluble (e.g. ICRP Publication 119).
The radioactive solid particles observed at nuclear power plant
accidents, nuclear tests, operations of nuclear fuel reprocessing
factory, and other nuclear disasters (Salbu, 2011). For example,
radioactive particles so-called hot-particles were observed aroundthe Chernobyl Nuclear Power Plant site (Sandalls et al., 1993). Mostof the radioactive particles were ﬁne particles of uranium oxide.
The Chernobyl accident happened phreatic explosion, and nuclear
reaction resulting ﬁre (Tcherkezian et al., 1994). Windscale
accident in 1957 also happened core ﬁre, which resulting inemission of radioactive particles from the stack to the atmosphere(Salbu et al., 1994). Emissions of radioactive particles were
anticipated in the F1NPP accident (Salbu, 2011). However, a coreﬁre and phreatic explosion did not occur in the F1NPP accident. The
explosions of unit 1, 3, and 4 were hydrogen explosions (TEPCO,
2012). The unit 4 hydrogen explosion occurred in the reverse ﬂow
of hydrogen generated in unit 3. In addition, in the Cs-bearing
particles isolated by Adachi et al. (2013) and Abe et al. (2014) onlyisotopes of radioactive Cs were detected by activity measurements.
Other radionuclides (e.g. Co, Ru, Ag, Sn, Sb, and Am) were notdetected. Thus, it is suspected that radioactive particles werereleased directly from the reactor, and further investigation of theCs-bearing particles is necessary to reveal their chemical andphysical features. It is also necessary to identify the differences
between the Fukushima radioactive particles and those in otherincidents to gain basic scientiﬁc information useful for efforts suchas environmental and health impact assessment, decontamination,and decommission.
The Cs-bearing particles have not been found from aerosol
samples from known plumes in Fukushima Prefecture, which is
closer to the release site and the location of the most severe
contamination. In contrast, studies applying autoradiography haveshown high-intensity spots of autoradiographic images in samples
of soil (Itoh et al., 2014; Niimura et al., 2015; Satou et al., 2015;
Yanaga and Oishi, 2015), forest leaves, vegetables, and dust frominside a building 40 km south of the F1NPP (Kashimura et al., 2013;
Shibata et al., 2013; Tanaka et al., 2013). These results imply thepresence of radioactive particles. None of the studies, however,
have conﬁrmed the chemical and physical properties of Cs-bearing
radioactive particles by isolating them.
Heavy contamination by radionuclides occurred to the area
northwest of the F1NPP on the same day (Katata et al.,2015).
Within 30 km-range from the F1NPP site is about the same range ofmuch of the radioactive particles (so-called hot particles) were

found in the case of the Chernobyl accident (Sandalls et al., 1993).
The particles had different characteristics, depending on thesampling locations and the distance from the accident site (Salbu
and Kerkling, 1998). Therefore, the radioactive particles obtained
closer to the nuclear power plant would be expected to give evenmore information about the accident. Thus, we aim to investigate
the radioactive particles’ distribution in the northwest area of the
F1NPP which has not been made hitherto. We addressed thefollowing questions. First, did radioactive particles come from the
F1NPP? Second, did spot type contaminations come fromCs-bearing particle?
2. Materials and methods
2.1. Sample collection
A soil sample was collected at a location approximately 20 km
northwest of the F1NPP on 12 June 2013, twenty-seven months
after the accident. We collected ﬁve surface samples of soil from alocation at 37" 330 3000 N, 140" 490 3800 E (Fig. 1), inside the exclusionzone located northwest of the F1NPP. This zone is distinguished byhigh air dose rates, based on airborne monitoring by the Ministry of
Education, Culture, Sports, Science and Technology (MEXT). The
location was the highest air dose rate spot within our entry
permission. This land, which was used as a paddy ﬁeld before theaccident, has been abandoned and vegetation has therefore
regrown. We sampled the surface soil (0–5 cm) with a 5-cmdiameter core sampler (DIK-115B, Daiki-Rika, Saitama, Japan). The
soil samples were bulked and homogenised in a plastic bag (Ondaet al., 2015) and then transported to the laboratory and stored atambient temperature.
2.2. Particle isolation and analysis
The radioactivity of the bulk soil sample was measured with ahigh-purity germanium (HPGe) semiconductor detector (GEMtype, Seiko EG&G, Tokyo, Japan). After the radionuclide determination, the soil sample was packed into a polyethylene bag for
imaging plate (IP) autoradiography using a BAS IP MS 2025 (highsensitivity model, GE Healthcare Japan, Tokyo, Japan) and a digital
imager (BAS-1800II, Fujiﬁlm, Tokyo, Japan). Because high concentrations of radioactive Cs were detected in the soil sample, a
relatively short IP exposure time (5 min) was applied to the sample
to detect only relatively strong radioactive emissions fromparticles; such emissions produce distinctive spots on the
autoradiograph. Soil near high-strength spots detected by IP was
separated from the surrounding soil for further analysis. The IP and
separation procedures were repeated on increasingly smaller
samples until relatively few soil particles remained. The selectedsoil particles were collected on sticky carbon tape and theexistence of gamma emitter activity in the g-spectrum, wasdemonstrated by using an HPGe detector.
We then cut the carbon tape into successively smaller pieces toisolate any radioactive particles from non-radioactive soil particles
using a technique developed in previous studies (Adachi et al.,2013; Abe et al., 2014). Backscattered electron images are very
useful in searching for radioactive particles containing elementswith high atomic number, such as Fe, Zn, Pb, and Cs, from carbon
tape (Salbu et al., 1994). Finally, we obtained small pieces of carbon
tape containing radioactive particles and analysed these particles
using SEM–EDS (SEM: SU 3500, Hitachi High-Technologies Co.,Tokyo, Japan; EDS: X-max 50 mm, Horiba Ltd., Kyoto, Japan) toobserve the particle morphology, measure the particle size, anddetermine the composition. In some cases, particles on carbon tapewere transferred to Gecko Tape (Nitto Denko Co. Ltd., Japan), which
is of high purity elemental carbon, using a microscope (BXiS series,

Y. Satou et al. / Anthropocene 14 (2016) 71–76

73

Fig. 1. Left: Map showing the location of the Fukushima Daiichi Nuclear Power Plant (F1NPP) on the east coast of Japan and the sampling site (solid dot), located #20 km
northwest of the FNDPP. Right: Map of the area northwest of the F1NPP (the location of the mapped area is indicated by the black square in the left panel), showing the air dose
rate on 5 November 2011 (as determined by MEXT) and the location of the soil sampling site (black dot); the sample was collected from the most highly contaminated zone,i.e., the so-called difﬁcult-to-return zone.

Olympus, Tokyo, Japan) and a micromanipulator (Quick Pro, Micro
Support Co., Ltd., Shizuoka, Japan) equipped with tungsten
needles.
3. Results
3.1. Particle ﬁgures and radioactivity
The sampling location (shown in Fig. 1) received radionuclide
deposits on 15 March 2011, mainly via wet processes (Kinoshita
et al., 2011; Tsuruta et al., 2014). The air dose rate at the time ofsample collection was 22.3 mSv/h, measured 1 m above the groundsurface. The bulk density of soil was 1.24 g/cm3, which is within therange of the values reported for soils in Fukushima (Kato et al.,2012; Honda et al., 2015; Matsunaka et al., 2016). Radionuclides inthe soil sample were measured by gamma spectrometry using ahigh-purity germanium semiconductor detector. On the basis ofthese measurements and a correction for radioactive decay, we
calculated that the amount of 134Cs deposited from the atmospheric plume(s) was 27.07 ! 0.06 MBq/m2 (86,900 ! 200 Bq/kg)
and that of 137Cs was 27.02 ! 0.01 MBq/m2 (86,720 ! 40 Bq/kg). The
activity ratio of 134Cs/137Cs was 1.00 (decay corrected as of 11
March 2011) which was consistent with the results of a large-scale
ground survey performed by the Japanese government (Saito et al.,2015).
IP autoradiography images of the soil sample showed that
contamination by radionuclides was relatively uniform, althoughmany spot-type contaminations were apparent (Fig. 2a). We
isolated four radioactive particles from such spot-type contamination positions in the sample (Fig. 2b). Compared to the analysisof radioactive Cs adsorbed on clay minerals, the present radioactive
particles were detected within #10 min by using IP, so that the
speciﬁc activity seems stronger than those found by Mukai et al.
The sizes and morphologies of the particles varied: the maximum
area equivalent diameter was 6.4 mm (particle B), and the particles
were either spherule or fragmental in shape. The radionuclides

134

Cs and 137Cs were detected by gamma spectroscopy in all of theparticles, with particle B showing the highest concentrations(Fig. 2c); namely 69.5 ! 0.5 Bq for 134Cs and 67.1 ! 0.1 Bq for 137Cs(values corrected for decay). Although the particles’ total
radioactive Cs activity was less than 1/1000 of the total unit soil
activity (Bq/kg), there were many more radioactive particles in thesample (see Fig. 2a). The activity ratio of 134Cs/137Cs was
1.03 ! 0.01, which is consistent with values for the bulk soil
sample. However, no other gamma emitters were detected in theparticles (Fig. 2c). The relationship between particle volume and137
Cs concentration is proportional (Fig. 3), suggesting that theseparticles have the same source.
3.2. Particle compositions
The elemental constituents of the particles were measured withan EDS (Fig. 4a). All particles showed X-ray spectra of Si, O, Fe, andZn, as well as Cs. Although X-ray peaks for Al were observed in allparticles, the peaks could not be separated from those of the
background sample, and we therefore could not conﬁrm its
presence in the individual particles. An element map of particle B isshown in Fig. 4b. Uniform distributions of Cs, O, Fe, Zn, and Si wereobserved within the particles (Fig. 4b). Approximately 80% wascomposed of Si and O, within each particle followed by K, Fe, andZn.
4. Discussion
The particles, collected 27 months after the accident, werelarger and more radioactive than those analysed in previous
studies (Adachi et al., 2013; Abe et al., 2014; Yamaguchi et al.,2016). In particular, particle B had approximately 20 times thevolume of the particles observed in aerosol samples collected in
Tsukuba (Adachi et al., 2013), which is located approximately
170 km southwest of the F1NPP. Because of the high dry deposition
velocity of the present particles (#10 cm/s for a 6 mm particle;

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Y. Satou et al. / Anthropocene 14 (2016) 71–76

Fig. 2. Distribution of radioactive spots of different types in the soil sample andSEM images of the isolated radioactive particles. (a) Autoradiography image
obtained by the imaging plate (IP) technique. The square frame represents theimaging plate and the grey colour distributed over the entire soil sample indicatesthat the soil was uniformly contaminated with radioactive materials. Black spots
indicate aggregates of radioactive atoms, which occur throughout the soil sample.
The black arrow indicates particle B. (b) SEM images of four radioactive particles(arrows) isolated from the soil sample. The particles are either spherules orfragmental in shape. Area equivalent diameters of particle A, B, C, and D are 4.5, 6.4,
4.0 and 3.0 mm, respectively. (c) Gamma ray spectrum of particle B.

Seinfeld and Pandis, 2006), it is possible that particles of the size
collected in this study deposited rapidly and were therefore
conﬁned to areas near the source.
The speciﬁc radioactivity of 137Cs within the present radioactive
particles is consistent with those in the previous studies from the

F1NPP by Adachi et al. (2013) and Abe et al. (2014)(=2 $ 1016 $ [volume]1.43; Fig. 3). By contrast, the speciﬁc radioactivity of hot particles observed in the Chernobyl accident(=4 $ 107 $ [volume]0.74; Tcherkezian et al., 1994) is smaller thanpresent particles from the F1NNP, indicating that the radioactive
particles from the F1NNP accident have higher radioactive Csconcentration within the particle.
In addition, the diameters of the present radioactive particles
range from 2 to 6.7 mm, which is consistent with the particle size
distribution of aerosol containing the F1NPP 137Cs (Doi et al., 2013;
Miyamoto et al., 2014) and are within the range of PM10 and PM2.5(Davidson et al., 2005), which are inhalable to human body.
Particles C and D were spherules, suggesting that they formed
from liqueﬁed materials (Adachi et al., 2013). During the F1NPP
accident the containment vessel pressure was observed tosuddenly drop in each of the reactor units, suggesting that rapid
cooling could have occurred following an adiabatic explosion
during the exhaust process. Such rapid cooling offers a simple
mechanism for the liquid-to-solid micro-particle conversion. In
contrast, particles A and B were fragmental, suggesting that they
were derived from larger fragments that had fractured in the soil
during transportation or release.
The particles we collected contained Si, O, Fe, Zn, and Cs, whichis consistent with the composition of previously reported particles(Adachi et al., 2013; Abe et al., 2014), although those studies could
not conﬁrm the presence of Si in the particles because Si waspresent in the substrate materials used in the previous analyses.
We avoided this limitation by using a Gecko Tape substrate (Maeno
and Nakayama, 2009). The result is consistent with the result by
Yamaguchi et al. (2016) who showed the presence of Si using atransmission electron microscopy and EDS. Possible sources for the
Si include silicate products that were used for the insulation
in/around the reactor unit and around steam lines in thecontainment building of the F1NPP (Nuclear Safety Commission,
Japan, 2006) as well as in the concrete of the containment vessel(Abe et al., 2014). We developed this novel point by vigorously
surveying the material used for the reactor core and peripherals aswell as rich in silicate.
The water solubility of these particles has important implications for human health, environmental impact, transportation, anddeposition (Adachi et al., 2013). In the Chernobyl accident, highly
radioactive particles were observed (Sandalls et al., 1993; Salbu
et al., 1994), and they gradually collapsed and became soluble inwater as time elapsed (Kashparov et al., 2004). The dissolution
processes from radioactive particles contributed soluble radionuclides to the terrestrial environment. The dissolution ratedepends on the soil pH (Kashparov et al., 2004). The radioactive
particles in the soil sample remained stable or quasi-stable in theenvironment for 27 months, in a location that had experienced
more than 3000 mm of precipitation during that time (Japan
Meteorological Agency, 2015). Thus, the particles appear to exhibit
low solubility in water and are stable in the ambient conditions of
the environment where they were deposited. We could argue that,
although the particles may have lost some components or been
totally or partially oxidised, and although the molecular composition may have changed, these particles have remained in theambient environment while maintaining their shape. Thus, longterm water solubility is recommended to be investigated todetermine the corrosion rate of these Cs-bearing particles in thesoil environment.
Decontamination of radioactive Cs from living areas are
emerging issue in the area and large amount of contaminatedsoil and materials are being generated continuously. Volume
reduction of the contaminants is pressing issue. In contrast,
understanding of chemical form of radioactive Cs is important toremove radioactive Cs from contaminant. We have demonstrated

Y. Satou et al. / Anthropocene 14 (2016) 71–76

5. Conclusions

103

Bq = 2×1016×[volume]1.43

101

137

Cs Bq/particle

102

100

Bq = 4×107×[volume]0.74
Chernobyl particles
10-1 -12
10

75

10-11

10-10

10-9

Volume of particle/cm3
Fig. 3. Relationship between the volume of particles and the radioactivity of 137Cs.
The dashed line shows speciﬁc radioactivity of hot particles in the Chernobyl
nuclear accident in 1986. Data were reconstructed from Tcherkezian et al. (1994).

The radioactive particles we collected from the polluted soil
were larger and more active than Cs-bearing particles analysed byAdachi et al. (2013), Abe et al. (2014) and Yamaguchi et al. (2016),
but their constituents were similar to, those of the reported aerosol
particles. In addition, the present study demonstrated that silicatesare one of the major components in the radioactive particle, andthat the particle characteristics were different from those of
so-called hot particles observed at the Chernobyl accident in 1986.
We suggest that the present radioactive particles in the soil weredischarged to the atmosphere during the same F1NPP emission
event on 14 and 15 March 2011 with the source of previously
reported Cs-bearing solid particles. Our results provide information on the temperature of materials ejected during the accident,
some of which appear to have been molten.
Our ﬁndings are useful to understand environmental dynamics
studies of radioactive materials from the Fukushima accident. In
addition, our ﬁndings have implications for the removal of theradioactive materials from the living areas and understanding ofprocess of the accident. We suggest that additional investigations
be conducted to analyse more particles from various distancesfrom the F1NPP to elucidate the particle size–distance relationship,
to determine the processes of synthesis of the particles in thereactor cores and peripherals, and to evaluate the effect ofweathering on the particles in the environment.
Acknowledgements
The authors thank the mayor of Namie, Fukushima Prefecture,for permission to enter the exclusion zone to collect the samples.
We are also indebted to Tetsuya Matsunaka, Tsutomu Takahashi,
and Nao Shibayama for their help with sample collection. Takuya
Yumine and his colleagues (Nitto Denko Co. Ltd.) provided the
Gecko Tape. In addition, we acknowledge Anne Chin for editorialadvices, and three anonymous reviewers for constructive comments and suggestions. This study was supported by MEXT/JSPS
KAKENHI grants (numbers 24110003, 24110006, 15H02340, and
256895). Y.S. received a Grant-in-Aid for JSPS Fellows.
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OPEN

received: 03 June 2015
accepted: 06 January 2016
Published: 03 February 2016

Internal structure of cesiumbearing radioactive microparticles
released from Fukushima nuclear
power plant
Noriko Yamaguchi1,*, Masanori Mitome2, Akiyama-Hasegawa Kotone2, Maki Asano1,3,
Kouji Adachi4 & Toshihiro Kogure5,*
Microparticles containing substantial amounts of radiocesium collected from the ground in Fukushima
were investigated mainly by transmission electron microscopy (TEM) and X-ray microanalysis withscanning TEM (STEM). Particles of around 2 μm in diameter are basically silicate glass containing Fe
and Zn as transition metals, Cs, Rb and K as alkali ions, and Sn as substantial elements. These elements
are homogeneously distributed in the glass except Cs which has a concentration gradient, increasingfrom center to surface. Nano-sized crystallites such as copper- zinc- and molybdenum sulfide, andsilver telluride were found inside the microparticles, which probably resulted from the segregationof the silicate and sulfide (telluride) during molten-stage. An alkali-depleted layer of ca. 0.2 μm thickexists at the outer side of the particle collected from cedar leaves 8 months after the nuclear accident,
suggesting gradual leaching of radiocesium from the microparticles in the natural environment.
Although almost five years have passed since the accident of Fukushima Daiichi Nuclear Power Plant (FDNPP),
radioactive contamination in the surrounding area is still a serious problem in Japan. The largest radionuclide
deposition event occurred on March 15–16 and the second largest on 21–23, 2011. Wet deposition was a major
source of radiocesium contamination of terrestrial environment1, while contribution of dry deposition was larger
near the FDNPP2. In order to understand and predict the fate of radioactive materials contaminating the terrestrial environment, it is important to clarify the physicochemical properties of the deposited materials. Fromprevious cases of radionuclide release, it is known that the chemical species of released radiocesium is monovalent cation (Cs+) which is soluble3. Deposition of radiocesium as insoluble particles has also been pointed out.
Autoradiography analyses using imaging plate (IP) showed spots of particulate materials on plant tissues collectedfrom Fukushima4–6. On the aerosol filter collected from March 14–15, 2011 in Tsukuba, 170 km south-southwestof FDNPP, Adachi et al.7 discovered spherical particulate radiocesium of 2.0–2.6 μ m in diameter, with particles
insoluble in water having a glass-like structure8. These microparticles contain several fission products of U-235
other than radiocesium, and Fe and Zn which are also used in nuclear reactors8. Hence, they were considered tobe released directly from nuclear reactors.
Kaneyasu et al.9 suggested that vaporized radiocesium was transported with sulfate aerosol in the air, dissolved
to cloud droplets and fell as rain. On the aerosol filter collected on March 20–21, 2011, rainy days in Tsukuba,
the majority of radiocesium was in water-soluble form7. Such water-soluble radiocesium that reached the groundsurface as a solute was fixed to soils, especially to clay minerals10. In the terrestrial environment, the majority ofradiocesium is present in solid form regardless of the initial form of deposition. However, compared to clay minerals originally contaminated by soluble radiocesium in soil, the solid radiocesium, which was initially depositedas radioactive microparticles, had stronger radioactivity. Although the contribution or percentage of such radioactive microparticles in the contamination level of Fukushima has not been evaluated, its influence on human1

National Institute for Agro-Environmental Sciences, Kannondai, Tsukuba 305-0864, Japan. 2National Institutefor Materials Science, Namiki, Tsukuba 305-0044, Japan. 3Graduate School of Life and Environmental Sciences,
University of Tsukuba, Tennodai, Tsukuba, 305-8572, Japan. 4Meteorological Research Institute, Nagamine, Tsukuba,
305-0052, Japan. 5Graduate School of Sciences, The University of Tokyo, Bunkyo, Tokyo 113-0033, Japan. *Theseauthors contributed equally to this work. Correspondence and requests for materials should be addressed to N.Y.
(email: nyamag@affrc.go.jp) or T.K. (email: kogure@eps.s.u-tokyo.ac.jp)
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Figure 1. (a) Bright-field (BF) image of NWC-1. The opaque material outside the microparticle is tungsten(W) and copper (Cu) deposited in the FIB process. (b) Electron diffraction (ED) pattern from the particle.
(c,d) EDS spectrum acquired from almost the whole area of the particle for the energy range of (c) 0–11 keV
and (d) 10–30 keV. (e) BF image of CB-8. The opaque material outside the sphere is W and the thin transparent
material is Kapton tape. (f) ED pattern from the particle.

health may be serious in terms of its intense radioactivity. Moreover, the structural detail of the microparticles
may give insights into the state of the broken reactor and fuel debris.
In the present study, we investigated radioactive microparticles, similar to those reported by Adachi et al.7, butcollected from the ground, by observing their internal structure with transmission electron microscopic (TEM)
techniques.

Results

Structure and composition of Cesium-bearing radioactive microparticles.

Cesium-bearing radioactive microparticles that had been deposited on non-woven fabric cloth (NWC-1) and on a needle of Japanesecedar (Cryptomeria japonica) (CB-8) were investigated. They were in the field for five and eight months, respectively, until sampling. Scanning electron microscope (SEM) images of NWC-1 of the whole microparticle before
preparing thin sections for TEM analyses; and elemental composition of the whole particle determined by synchrotron radiation microbeam X-ray fluorescence (SXRF) are shown in Supplementary Figs S1 and S2 online,
respectively. The activities of 137Cs for the NWC-1 and CB-8 were 5.04 ±  0.472 and 3.14 ±  0.178 Bq, respectively.
Bright-field (BF) images and selected-area electron diffraction (SAED) patterns recorded in TEM from thewhole area of NWC-1 and CB-8 are shown in Fig. 1, along with the energy-dispersive X-ray spectrum (EDS)
acquired in TEM from NWC-1. Preparation of thin specimens from NWC-1 microparticle by focused-ion-beam(FIB) process was successful only for the upper half of the sphere. The BF contrast of NWC-1 (Fig. 1a) was almost
uniform except for two dark nanoparticles inside the sphere. These nanoparticles will be described in a later subsection. The SAED pattern from the whole particle consists of only a halo (Fig. 1b) and the EDS from the whole

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particle (Fig. 1c,d) mainly consist of peaks of Si and O (Cu is from the supporting mesh), indicating that the particle was basically silicate glass. Additionally, Cl, K, Cs, Fe, Zn, Rb, and Sn were definitely identified. The presence ofNa was not confirmed since its peak overlapped with the L-peak of Zn. These elements were previously reported7,8
except for K. EDS Semi-quantitative analysis for almost the entire area of the TEM specimen, without considering
the absorption effect estimated the glass composition (wt.%) to be SiO2; 69.3, K2O; 1.9, Fe2O3; 8.6, ZnO; 11.0,
Rb2O; 1.3, SnO2; 1.4, Cs2O; 3.4 with a small amount of Cl (1.4 wt.%). The valence states of cations were assumed to
be in line with results in an X-ray absorption spectroscopic study8. Rubidium, Cs and Sn are the fission productsof U-235. Tin is also used for fuel cladding11. Iron is used as steel for reactor pressure vessels, and Si and O aremain components of concrete at the bottom of the containment vessel, where the melt-down fuel or core debris
is thought to exist. Zinc had been added to the primary cooling water12. Potassium and Cl may originate fromconcrete and/or seawater, which was used to cool down the reactor.
The BF TEM image of CB-8 showed a two-layer structure having an inner core with darker contrast, an outercrust with a lighter one, and a small dark nanoparticle near the center (Fig. 1e). The thickness of the crust wasaround 0.2 μ m. SAED pattern indicated that both are amorphous (Fig. 1f). Locally, a bubble-like structure wasobserved at a certain radius in the crust, or the outside part was peeled off at the radius (Fig. 1e). The chemicalcomposition of the entire area of the specimen was SiO2; 73.3, K2O; 1.4, Fe2O3; 7.2, ZnO; 11.4, Rb2O; 1.2, SnO2;
1.5, Cs2O; 3.3 and Cl; 0.7, which is similar to that of NWC-1.

Distributions of elements in microparticles. Elemental maps from NWC-1 using STEM-EDS are
shown in Fig. 2a. Brighter color indicates higher concentration of the elements. All elements were almost uniformly distributed in the particle with Cs the only exception. Cesium concentration was lower near the center
than the outside of the particle. The ratio of the amounts between the center and outside is around two. It is
possible that vapor-phase cesium in the reactor was absorbed to the molten silicate microparticles, from theirsurface. The low diffusion velocity of cesium in the glass may have left the concentration gradient. On the contrary, elemental mapping for CB-8 (Fig. 2b) indicates that K, Rb and Cs were distinctively depleted in the crust
but concentrated in the vicinity of the surface. Although the reason for the high concentration at the surface is
not certain, it might be an artifact by diffusion of alkali cations caused by electron-beam radiation in the analysis.
Chlorine was concentrated in the crust.
Crystalline nanoparticles in the glass. Two nanoparticles with dark contrast were observed in the TEM
image of NWC-1 (Fig. 1a). Although EDS spectrum only from the particles cannot be obtained since they arestill buried in the glass in spite of the thinning by FIB, S was distinctively identified from the particles (Fig. 3a).
Elemental mapping by STEM-EDS indicated enrichment of Cu, Zn and Mo at the nanoparticles along with S,
indicating that they are sulfide. The possibility of sulfate was excluded because oxygen is deficient in the nanoparticles in the oxygen map (Fig. 3a). Moreover, the locations of Cu, Zn and Mo do not overlap within thenanoparticles, suggesting that these metal elements form different sulfide phases in the nanoparticle (Fig. 3a).
SAED from the left particle showed a single-crystal diffraction pattern (Fig. 3b), which can be explained by the
crystallographic parameters for a high-temperature polymorph of digenite (Cu2−xS). The origin of Cu and S may
be minor elements in the concrete. On the other hand, Ag and Te were distinctively detected by EDS in TEM
from the nanoparticle in CB-8 (Supplementary Fig. S3 online), indicating the particle to be silver telluride. Bothelements can be fission products of U. Diffraction spots were observed in the SAED pattern from the particle,
indicating that it is crystalline but the phase could not be identified. These sulfide and telluride were probably
segregated and crystallized in the microparticles at the molten state.

Discussion

Our most significant finding is that the matrix of the Cs-bearing microparticles is silicate glass, based on the
TEM-EDS analysis with FIB sample preparation. Previous studies suggested that Fe, Mo, Sn and Zn in the
Cs-bearing microparticles had a similar X-ray absorption near-edge structure to those composed of glass8, however the presence of Si in the microparticles has not been verified7,8. It is probable that the high-temperature
melt-down fuel from the reactor came into contact with and melted the concrete, and then splashed microparticles of silicate melt, which were solidified by cooling to form silicate glass in the atmosphere. However, there areseveral questions with respect to the selection of the constituting elements in the glass from various ones in thereactor. For instance, Ca which is one of the major elements in concrete, was almost absent in the microparticles
of NWC-1. Since TEM observed only a small portion of the microparticles, by making them thin using FIB, there
may have been other elements in the microparticles, for instance, as a form of chalcogenide nanoparticle.
The next important finding is the alkali-depleted crust in CB-8 microparticle. This is probably the result of
elution of alkali ions by contact with acidic solution in the field, commonly observed in silicate glass13. On theother hand, such alkali-depleted crust was not observed in NWC-1. This may be attributed to the different environments of the two microparticles after release from the nuclear plant. NWC-1 was on non-woven cloth andCB-8 on a cedar leaf before it was collected. It is well-known that silicate glass elutes alkali components fromtheir surface by ion-exchange with proton or hydronium ions to form an alkali-leaching layer on the surface if
pH of reacting solution is low, whereas the silicate framework of the glass itself is dissolved with high-pH solution13,14. Coniferous forest canopy induces acidic condition due to ammonia uptake, nitrification and leachingof plant-derived acid15. It was likely that CB-8 deposited on the cedar leaf had been in acidic conditions, which
derives alkali depletion in the crust, for eight months. The finding of the alkali-depleting crust on the surface of
the Cs-bearing radioactive microparticle indicates that radiocesium in the particles can be released by “weathering” of the glass in natural environments, and considering its small size, duration for the total release of theradioactive cesium from the particles is probably not long, from several years to a few decades, though it willstrongly depends on the environment.
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Figure 2. (a) Element maps of NWC-1 acquired by STEM-EDS with a STEM Dark-field (DF) image at thesame area, and (top) TEM Bright-field image in which the rectangle indicates the area analyzed by STEM-EDS.
The thin area around the top of the particle which was not observed in Fig. 1a was formed by a further thinning
process by FIB. (b) Element maps and images of CB-8 with the same framing as in (a).

In order to investigate the dissolution rate and detailed Cs-leaching properties of the Cs-bearing radioactive
microparticles, a leaching experiment should be conducted as a function of temperature and pH. However, collecting and isolating the Cs-bearing microparticles is time-consuming and it is difficult to obtain a large enoughnumber of Cs-bearing microparticles to investigate dissolution properties. Alternatively, synthesized silicate glasswith the same composition as the microparticles presented in this study may help to obtain information on thefate of Cs-bearing radioactive microparticles. However, in our preliminary experiment, we were unable to makeuniform glass with the same composition at present. This is probably due to the liquid immiscibility at the composition. A solution to this problem is now being considered.
There are mainly two types of solid-phase radiocesium in the terrestrial environment affected by FDNPP accident; that fixed to clay minerals in the soil via wet deposition and that contained in the microparticles of silicate
glass flown directly from the nuclear reactors. The radioactivity of the former is rather in low-density but distributed widely, therefore it is a major source of external radiation from ground. Part of the radiocesium adsorbed on
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Figure 3. (a) Element maps and STEM bright-field (BF) image of the two nanoparticles inside NWC-1.
The circular contrasts outside the nanoparticles in the BF image are radiation damage in glass formed by theelectron probe for chemical analyses. (b) (Left) ED pattern from the left nanoparticle. (Right) Drawing of ED
pattern for the high-temperature polymorph of digenite (Cu2−xS) along < 121> .

the soil clay is desorbed and transferred to crops thereby causing internal radiation. On the other hand, contribution of the microparticles to the air radiation is most likely not significant, but their radiation density is very high,which is particularly problematic for organisms including humans if the microparticles are inhaled or ingested.
The plant availability of radiocesium in the microparticles should depend on its solubility. Consequently, further
research on this material should be carried out as soon as possible.

Methods

Samples. Radioactive microparticles attached on non-woven fabric cloth and needles of Japanese cedar, collected from Fukushima, were identified by autoradiography followed by point-by-point analyses with scanningelectron microscope (SEM; Hitachi high-Technologies SU3500) equipped with an energy dispersed X-ray spectrometer (EDS; Horiba X-max 50 mm). The identified particles were further determined by SXRF with excitation
X-ray energy of 37.5 keV to ensure the detection of Cs-K line from the particle. The activities of 137Cs was determined by a Ge detector (GCW2523S Canberra, USA). Detailed methods to identify the Cs-bearing microparticles are described in the supplementary information.
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TEM Analyses.

Cross-sectional thin TEM specimens were prepared from radioactive microparticles using
a focused ion beam (FIB) instrument with micro-sampling system (Hitachi FB-2100) as described in the supplementary information. Then specimens were initially examined using a TEM (JEOL JEM-2010UHR) operated at
200 kV with an EDS analyzer system (JEOL JED-2200). Elemental mapping in the microparticles and quantitative
analyses were performed using a JEOL JEM-3100FEF operated at 300 kV in the STEM mode, with an EDS analyzer system (Thermo Fisher Scientific NORAN System SIX). Finally, elemental maps for nanoparticulates inside
the microparticles were acquired using a JEOL JEM-2800 operated at 200 kV with double wide-area (0.95 sr.)silicon drift detectors (SDD) for EDS analyses.

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Acknowledgements

Part of this work was supported by a Grant-in-Aid for science research (15H04222 and 24110003) by JSPS, Japan.
This work was also supported by 2014 JAEA contract research for Fukushima environment recovery, “Studyon Cs adsorption and desorption process on clay minerals” and by MAFF contract research “Development of
Decontamination technologies for radioactive substances in agricultural land”. The synchrotron radiation
experiment was performed under approval of the SPring-8 Program Advisory Committee (Proposal Number:
2014B1282). The authors wish to thank Dr. M. Kobayashi for providing the Japanese cedar samples, Ms. E. Fujii
for preparing TEM specimens, Dr. T. Hara for providing the opportunity to use the STEM instrument, Prof. Clive
Langham for English edition, and Dr. Y. Igarashi for comments on the manuscript.

Author Contributions

N.Y. and T.K. designed the research and wrote the paper; N.Y. performed the IP, SEM and SXRF analyses; M.A.
performed SEM and SXRF analyses; M.M. and K.A.-H. performed TEM analyses; K.A. performed SEM analyses;
T.K. performed SEM and TEM analyses and supervised the study; all authors contributed to the manuscript. N.Y
and T.K. contributed equally to this work.

Additional Information

Supplementary information accompanies this paper at http://www.nature.com/srep
Competing financial interests: The authors declare no competing financial interests.
How to cite this article: Yamaguchi, N. et al. Internal structure of cesium-bearing radioactive microparticles
released from Fukushima nuclear power plant. Sci. Rep. 6, 20548; doi: 10.1038/srep20548 (2016).
This work is licensed under a Creative Commons Attribution 4.0 International License. The images
or other third party material in this article are included in the article’s Creative Commons license,
unless indicated otherwise in the credit line; if the material is not included under the Creative Commons license,
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6

Atmos. Chem. Phys., 16, 13149–13172, 2016www.atmos-chem-phys.net/16/13149/2016/
doi:10.5194/acp-16-13149-2016© Author(s) 2016. CC Attribution 3.0 License.

Long-term assessment of airborne radiocesium after the Fukushima
nuclear accident: re-suspension from bare soil and forest ecosystems
Mizuo Kajino1,2 , Masahide Ishizuka3 , Yasuhito Igarashi1 , Kazuyuki Kita4 , Chisato Yoshikawa5 , and Masaru Inatsu6
1 Meteorological Research Institute (MRI), Japan Meteorological Agency (JMA), Tsukuba, Ibaraki, Japan
2 RIKEN Advanced Institute for Computational Science (AICS), Kobe, Hyogo, Japan
3 Faculty of Engineering, Kagawa University, Takamatsu, Kagawa, Japan
4 Faculty of Science, Ibaraki University, Mito, Ibaraki, Japan
5 Japan Agency for Marine-Earth Science and Technology (JAMSTEC), Yokosuka, Kanagawa, Japan
6 Faculty of Science, Hokkaido University, Sapporo, Hokkaido, Japan

Correspondence to: Mizuo Kajino (kajino@mri-jma.go.jp)
Received: 29 March 2016 – Published in Atmos. Chem. Phys. Discuss.: 25 April 2016
Revised: 26 July 2016 – Accepted: 19 September 2016 – Published: 27 October 2016
Abstract. The long-term effect of 137 Cs re-suspension fromcontaminated soil and forests due to the Fukushima nuclear
accident has been quantitatively assessed by numerical simulation, a field experiment on dust emission flux in a contaminated area (town of Namie, Fukushima prefecture), andair concentration measurements inside (Namie) and outside(city of Tsukuba, Ibaraki prefecture) the contaminated area.
In order to assess the long-term effect, the full year of 2013
was selected to study just after the start of the field experiments. The 137 Cs concentrations at Namie and Tsukuba
were approximately 10−1 –1 and 10−2 –10−1 mBq m−3 , respectively. The observed monthly median concentration at
Namie was 1 to 2 orders of magnitude larger than that at
Tsukuba. This observed difference between the two sites wasconsistent with the simulated difference, indicating successful modeling of 137 Cs re-suspension and atmospheric transport. The estimated re-suspension rate was approximately
10−6 day−1 , which was significantly lower than the decreasing rate of the ambient gamma dose rate in Fukushima prefecture (10−4 –10−3 day−1 ) as a result of radioactive decay,
migration in the soil and biota, and decontamination. Consequently, re-suspension contributed negligibly in reducing
ground radioactivity. The dust emission model could reproduce the air concentration of 137 Cs in winter, whereas thesummer air concentration was underestimated by 1 to 2 orders of magnitude. Re-suspension from forests at a constant rate of 10−7 h−1 , multiplied by the green area fraction,
could explain the air concentration of 137 Cs at Namie andits seasonal variation. The simulated contribution of dust resuspension to the air concentration was 0.7–0.9 in the coldseason and 0.2–0.4 in the warm season at both sites; theremainder of the contribution was re-suspension from forest. The re-suspension mechanisms, especially through theforest ecosystems, remain unknown. This is the first study
that provides a crude estimation of the long-term assessmentof radiocesium re-suspension. Additional research activities
should investigate the processes/mechanisms governing there-suspension over the long term. This could be achieved
through conducting additional field experiments and numerical simulations.

1

Introduction

The Fukushima Daiichi Nuclear Power Plant (FDNPP) accidentally released nuclear fission products into the atmosphere and the ocean environment following the catastrophic earthquake and tsunami that occurred in March 2011.
The accident caused serious contamination of the soil over
the Tohoku region (northeastern part of Japan, including
Fukushima and Miyagi prefectures) and the Kanto region(eastern part of Japan, including Ibaraki, Tochigi, Gunma,
and Chiba prefectures; NRA, 2012). Since then, a number ofstudies have been conducted, particularly during the months
that followed the accident. These assessments include primary emission estimations (Chino et al., 2011; Danielache

Published by Copernicus Publications on behalf of the European Geosciences Union.

13150M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium

et al., 2012; Stohl et al., 2012; Terada et al., 2012; Katata
et al., 2012a, b, 2015; Winiarek et al., 2012, 2014; Hirao
et al., 2013; Saunier et al., 2013; Yumimoto et al., 2016;
Danielache et al., 2016), field observations (ground aerosol
sampling: Masson et al., 2011, 2013; Kaneyasu et al., 2012;
Adachi et al., 2013; Tsuruta et al., 2014; Igarashi et al., 2015;
Oura et al., 2015; aircraft measurements: NRA, 2012; andmeasurements carried out on foot: Hososhima and Kaneyasu,
2015), and numerical simulations of transport and depositions (deterministic simulation: Chino et al., 2011; Morino etal., 2011; Yasunari et al., 2011; Stohl et al., 2012; Terada etal., 2012; Katata et al., 2012a, b; Winiarek et al., 2012, 2014;
Hirao et al., 2013; Saunier et al., 2013; Katata et al., 2015;
Yumimoto et al., 2016; Danielache et al., 2016; deterministic
simulation with sensitivity runs: Morino et al., 2013; Adachi
et al., 2013; Groëll et al., 2014; Saito et al., 2015; Sekiyama
et al., 2015; Quérel et al., 2015; uncertainty modeling andprobabilistic forecast: Girard et al., 2016; Sekiyama et al.,2016; and multi-model intercomparison and multi-model ensemble analysis: SCJ, 2014; Draxler et al., 2015; Kristiansen
et al., 2016). The targeted radionuclides were species withboth short and long half-lives: 99 Mo–99m Tc (half-life 65.9–6 h), 129m Te (33.6 days), 131 I (8.02 days), 132 Te–132 I (3.2days–2.3 h), 134 Cs (2.07 years), 136 Cs (13.2 days), 137 Cs(30.1 years), 133 Xe (5.2 days), and 35 S (87.5 days).
In contrast, there have been few studies on the long-term(more than 1 year) quantitative assessment of radioactivity
in the atmosphere associated with the Fukushima accident(Igarashi et al., 2015; Ishizuka et al., 2016; Kinase et al.,2016). More than 100 000 people were evacuated (METI,2012), and most have still not been able to return to theirhomes and the public remains anxious about the safety of
the affected areas. Radionuclides with long half-lives such as
134 Cs (2.07 years) and 137 Cs (30.1 years) are of particularconcern (Evangeliou et al., 2014). Following the Chernobyl
accident there were several studies on the re-suspension andlong-term assessment of these radionuclides, such as Holländer and Garger (1996), Garger et al. (1998), Hatano andHatano (2003), and Garger et al. (2012). For example, Garger
et al. (2012) estimated the re-suspension “descending trend”
as having a half-life of 300 days based on the surface activity
concentration of 137 Cs. In the case of the Fukushima accident, Igarashi et al. (2015) estimated the half-reduction timeby fitting multi-component exponential functions based on
the 137 Cs concentration at the Meteorological Research Institute (Tsukuba) as 5.9 days, 16 days, and 1.1 years. Theseestimates were based on the trend in the observed surface airconcentrations of 137 Cs, and thus the contributions from advection, diffusion, emission, and deposition terms were notquantified.
There are thousands of monitoring posts situated in thecontaminated area in Fukushima prefecture to measure theambient gamma dose rate. However, the data cannot be used
for evaluating internal exposure: evaluation of internal exposure requires direct measurement of the surface air acAtmos. Chem. Phys., 16, 13149–13172, 2016tivity concentration. There are only a few observation sitesthat continuously measure the concentration of radiocesium(e.g., Igarashi et al., 2015; Ishizuka et al., 2016; Kinase et al.,2016). To assess the spatial distribution of the internal exposure hazard, 3-D numerical simulation is necessary to interpolate values for unmonitored locations. The numerical simulation requires emission flux as a boundary condition. However, the mechanism and thus the radioactivity flux associated with the re-suspension of 137 Cs were unknown, despiteextensive efforts based on field observations (e.g., Igarashi etal., 2015; Ishizuka et al., 2016; Kinase et al., 2016).
Garger et al. (2012) summarized the re-suspension sources
following the Chernobyl accident as (1) dust emission,(2) human activity in fields as well as on roads and construction sites, (3) forest fires, and (4) emissions from thepower plant (i.e., opening of the Chernobyl sarcophagus).
Re-suspension sources (1) (Ishizuka et al., 2016) and (4), i.e.,
additional emissions from the reactor buildings of FDNPP(TEPCO, 2012, 2013, 2014a, b, 2015), were considered in
the present study. With respect to source (2), since the townof Namie is located in the evacuation zone, human activity has been extremely limited except for decontaminationrelated work. Although the decontamination-related work
could be a potential source of re-suspension, it is not considered in the current simulation as the re-suspension flux has
been hardly quantified. As to source (3), there is a low risk offorest fires in Japan given the high humidity, although someopen biomass burning remains possible. Kinase et al. (2016)
found no increase in 137 Cs activities when the concentration of levoglucosan, a marker of biomass burning, was increased, and thus re-suspension due to biomass burning was
not considered in the present study. In addition to the four
sources of 137 Cs from the Chernobyl accident, re-suspension
from terrestrial biota was considered as suggested by Kinase
et al. (2016). They found substantial amounts of bioaerosols
upon scanning electron microscopy samples collected in thesummer, when 137 Cs concentration was high.
In the present study, the long-term effect of radiocesium
re-suspension from contaminated soil and terrestrial biota
was quantitatively assessed using 3-D numerical simulation,a field experiment on dust emission flux in a contaminatedarea (Namie, Fukushima prefecture), and air concentration
measurements conducted inside (Namie) and outside (cityof Tsukuba, Ibaraki prefecture) the contaminated area. Even
though the re-suspension mechanisms still remain unknown,
by utilizing the observational data both inside and outside,
together with the transport model, we aimed to provide as
robust an analysis as possible of the re-suspension, transport,
and re-deposition of 137 Cs over the Tohoku and Kanto regions of Japan.

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Figure 1. (a) The model domain showing the model terrestrial elevation, observation sites, and other locations described in the study. The
linear distances from FDNPP to Namie (Tsushima district, town of Namie), Tsukuba and Tokyo are approximately 30, 170, and 220 km,
respectively. The numbers denote prefectures: 1, Miyagi; 2, Fukushima; 3, Ibaraki; 4, Tochigi; 5, Gunma; 6, Saitama; 7, Chiba; and 8, Tokyo.
(b) Fukushima prefecture and (colored circles) the locations (village, town, or city name) of monitoring posts used in this study (see Fig. 14).
The red letters in both (a) and (b) denote the name of the area based on geographical features: A, Nakadori valley; B, Abukuma highland;
and C, Hamadori coastal area.

2

Numerical simulation

A brief description of the numerical method, such as the processes considered in the model and simulation settings, ispresented in this section, and detailed model formulations aredescribed in Appendix A. Because the schemes and assumptions regarding the emissions are key to the current study,
they are described in detail in the following subsections.
2.1

Lagrangian model and simulation settings

Figure 1 shows the domain of the Lagrangian model (LM)with model terrestrial elevations, covering 138–143◦ E and
34–39◦ N. The model domain covers the southern part of the
Tohoku region (the northern mountainous part of the domain,
including Yamagata, Miyagi, and Fukushima prefectures),and includes the FDNPP and highly polluted areas such as
the habitation-restricted zone (HRZ; 20–50 mSv yr−1 ) anddifficult-to-return zone (DRZ; > 50 mSv yr−1 ; METI, 2012).
It also covers the Kanto region (or Kanto Plain, the largest
plain in Japan, approximately 120 km × 120 km), a highly
populated region that includes low to moderately polluted
areas such as Tokyo, Gunma, Tochigi, Ibaraki, Saitama, andChiba prefectures.
The LM considers horizontal and vertical diffusion and advection, gravitational settling, dry and wet depositions, andwww.atmos-chem-phys.net/16/13149/2016/

radioactive decay. It uses simple parameterizations for dry
and wet deposition schemes, and it can be driven by meteorological analysis datasets so that it does not require a meteorological model to predict detailed meteorological fields
and variables. The model was designed to be easily handled
and computationally efficient so that non-specialists of numerical simulations can conduct long-term assessments ofatmospheric diffusion problems using their desktop or laptop computers. The LM was designed for rough budget estimates, as presented in the current study, or for sensitivity
analyses using a large number of parameters (e.g., Groëll
et al., 2014; Quérel et al., 2015; Girard et al., 2016), rather
than for process-oriented analysis (e.g., Morino et al., 2013;
Katata et al., 2015) or sensitivity analyses of the physicaland chemical parameters of aerosols (Adachi et al., 2013).
Details of each process and parameter are described in Appendix A. Statistical error of a Lagrangian simulation isinversely proportional to the square of the number of Lagrangian particles (LPs). The statistical accuracy of the current simulation setting is discussed in Appendix B.
The Grid Point Value-Mesoscale Model (GPV-MSM) of
the Japan Meteorological Agency (JMA) was used for meteorological analysis to calculate the transport of LPs. It covers 120–150◦ E and 23–47◦ N and provides 3-hourly and 16
pressure levels of 3-D meteorological variables, from 1000 to
100 hPa, with a horizontal grid resolution of approximately
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Table 1. The observation sites and monitoring posts used to provide data for this study.
Name

Location

Description

140.7683◦ E, 37.5621◦ N
140.1254◦ E, 36.0551◦ N

Namie High School, Tsushima Campus1
Meteorological Research Institute

140.9969◦ E, 37.4163◦ N
140.9296◦ E, 37.4665◦ N
140.7385◦ E, 37.6772◦ N
140.6979◦ E, 37.5836◦ N
140.4765◦ E, 37.6870◦ N
140.1904◦ E, 37.1241◦ N

Ottozawa 3 Community Center2
Omaru Multipurpose Community Center3
Iitate Junior High School4
Yamakiya Otsu 8 Community Firehouse5
Fukushima-Minami Fire Department
Takayama-Kita Park

Observation sites
Namie (Tsushima)
Tsukuba
Monitoring posts
Okuma
Namie (Omaru)
IitateKawamata
Fukushima
Shirakawa1 Original location (now moved to city of Nihonmatsu). 2 Ottozawa San-ku Chiku Shukaijo (in Japanese). 3 Omaru

Tamokuteki Shukaijo (in Japanese). 4 Original location (now moved to city of Fukushima). 5 Yamakiya Otsu Hachi-ku
Community Shoubou Center (in Japanese).

11 km (1 longitude = 0.125◦ and 1 latitude = 0.1◦ ) and surface variables at twice the resolution of that for the 3-D variables (1 longitude = 0.0625◦ and 1 latitude = 0.05◦ ). In thesimulation, the whole model domain where LPs can travel is
138–143◦ E, 34–39◦ N and from ground surface to 500 hPa.
For output of the model results, LP fields are converted toEulerian concentration (Bq m−3 ) and deposition (Bq m−2 )fields in the same horizontal space as the 3-D variables but
are vertically allocated from the ground surface to an altitudeof 1 km at 100 m intervals. The observed surface air concentration was compared with the simulated mean concentration
at 0–100 m above ground level (AGL).
2.2

Re-suspension from bare soil

Ishizuka et al. (2016) developed a re-suspension scheme forradiocesium from bare soil based on measurements on theschoolyard at Namie High School, Tsushima Campus (denoted as Namie (Tsushima) in Table 1 and Fig. 1), in the
DRZ.
Fsoil = p20µm FM (1 − fforest )B5mm (t),(1)

is the 137 Cs dust re-suspension flux from soil

where Fsoil(Bq m−2 s−1 ), p20µm is the surface area fraction of dust
smaller than 20 µm in diameter against soil containing a maximum size of 2 mm particles and varies depending on soil
texture (1.3 × 10−8 for sand, 0.19 for loamy sand, 0.45 forsandy loam, and 0.80 for silt loam), FM is the total dust
mass flux (kg m−2 s−1 ), fforest is the forest area fraction, andB5 mm (t) is the specific radioactivity of surface soil (from thesurface to a depth of 5 mm; Bq kg−1 ) as a function of timesince March 2011. The formula is based on the assumptionthat dust particles smaller than 20 µm in diameter originating from the surface soil and to a depth of 5 mm were suspended and transported through the atmosphere. The surface
area fraction was used for p20µm based on the assumption that
Atmos. Chem. Phys., 16, 13149–13172, 2016

radiocesium is bound to the surface of soil particles (Evrard
et al., 2015). FM is formulated as being proportional to the
cube of the friction velocity u∗ (m s−1 ) as described by Loosmore and Hunt (2002) and was applied to the dust emission:
FM = 3.6 × 10−9 u3∗ .

(2)

Since u∗ is not available in GPV-MSM, u∗ was estimatedusing a wind speed at 10 m AGL by assuming neutral stratification conditions.
B5 mm (t) was derived from the combination of Bobs , theobserved horizontal distribution of 137 Cs deposition obtained
from an airborne radiological survey (NRA, 2012; Bq m−2 ),and r5 mm , the surface soil activity ratio of 0–5 mm to 0–5 cm
obtained from a vertical profile measurement of 137 Cs in the
ground soil at Namie High School (i.e., 0.57 Bq Bq−1 ), as
B5 mm (t) =Bobs r5 mm Rdecay (t)5 × 10−3 ρb, soil,

(3)

where ρb, soil is the bulk density of soil particles per unit volume in the ground space (kg m−3 ) obtained from the porosity(0.4 m3 m−3 ) and the density of dust particles (2650 kg m−3 ).
For Rdecay , which is the decreasing ratio of activity in the
ground to its initial state (Rdecay (t = 0) = 1), only radioactive
decay was considered for the re-suspension calculation. The
decreasing rate due to other processes such as land surface
processes (e.g., runoff, erosion, percolation and all the processes resulting in migration of radiocesium in the soil andbiota; Evrard et al., 2015; Matsuda et al., 2015) and decontamination were not considered here. Furthermore, suppression of dust emission due to soil moisture and snow cover
was not considered. Therefore, it should be noted here that
Fsoil in Eq. (1) is considered as the upper boundary of 137 Cs
re-suspension flux from surface soil. Effects such as land surface processes, decontamination, and dust emission suppression due to snow cover are extensively discussed in Sect. 5.3
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M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium

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Figure 2. The areal fractions of (a)–(c) soil texture and (d)–(f) land use category used for the boundary conditions of the simulation.

using ambient gamma dose rate measurements obtained by
the monitoring posts in Fukushima prefecture.
Equation (1) is a function of soil texture. The areal
fraction of soil texture of the model grid was obtained
from the database of the advanced research Weather Research and Forecasting model version 3 (WRFV3; Skamarock et al., 2008). Sixteen categories of soil texture (Miller and White, 1998) with a 30 arcsec resolution dataset can be obtained from the web after subscription at http://www2.mmm.ucar.edu/wrf/users/download/get_sources_wps_geog.html and were re-categorized into theabove-mentioned four categories (i.e., sand, loamy sand,sandy loam, and silt loam) and interpolated to the LM resolution (∼ 11 km) as shown in Fig. 2a–c. Note that the loamy
sand fraction is not presented because it is zero for the entire
domain. The parameter fforest (Fig. 2d) was also obtained
from the database of WRFV3 and was calculated based on
the 24 United States Geological Survey (USGS) land use categories, which are constant over time. The land use category
dataset can also be obtained from the above website.
Ishizuka et al. (2016) validated their dust emission module
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tration of 137 Cs at Namie in the winter. After applying themodule to our 3-D simulation, we found that the air concentration at Namie was underestimated by about factor of 5 forthe same period. The module was formulated based on physical parameters (such as u∗ ) but contains parameters obtained
at a single location (such as r5 mm and ρb, soil ) and under afixed atmospheric condition (Ishizuka et al., 2016), whereas
ideally parameters in Eqs. (1)–(3) should have consideredvariations among locations and atmospheric conditions forthe 3-D simulation. We simply multiplied the dust emission
flux by 5 after adjusting the simulation results against the observed concentration of 137 Cs at Namie in the winter. This is
one of the simplest top-down approaches for adjusting theemission flux according to the air concentration. The module
requires improvement in the future as more reliable parameters become available for various conditions and locations.
2.3

Re-suspension from the forest ecosystems

The re-suspension mechanism of radiocesium from land
ecosystems remains unknown. Kinase et al. (2016) foundsubstantial amounts of bioaerosols (rather than mineral dust
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particles) in samples collected for scanning electron microscopy in the summer, when the 137 Cs concentration was
high. This does not prove that the bioaerosol was carrying
radiocesium but that it could be a potential carrier. The behavior of Cs in the environment can be inferred by analogy with K, a congener of Cs. Potassium is a necessary
and abundant element in plants and circulates between land
ecosystems. The addition of potassium fertilizer to a rice
field in Fukushima significantly reduced the Cs content of
the rice (Ohmori et al., 2014). Substantial amounts of Ksalt-rich particles, possibly emitted by active biota such as
plants and fungi, and coated with secondary organic aerosols,
were observed in pristine Amazonian rainforest (Pöhlker etal., 2012). The major areal fraction of the contaminated areain Fukushima is covered by biota-rich mountain forests. Despite the differences in plant species and locations, it is plausible that water-soluble radiocesium circulating in the biota
and soil in the forests was somehow re-emitted into the atmosphere and contributed to the surface air concentration. The
re-suspension from the forest ecosystem was simply formulated as follows:
Fforest = fforest fgreen rconst Bobs Rdecay (t),(4)

where Fforest is the 137 Cs re-suspension flux from forest (Bq m−2 s−1 ); fforest is the forest area fraction; fgreen
is the monthly green area fraction; rconst is the constantre-suspension coefficient (s−1 ); and, as in Eq. (3), Bobs
and Rdecay are the observed 137 Cs deposition (NRA, 2012;
Bq m−2 ) and the decreasing ratio of activity in the ground,respectively. rconst is a tunable parameter to adjust the simulated air concentration of 137 Cs to that observed. In the current study, rconst is set to 10−7 h−1 by adjusting the simulation data using the observed 137 Cs concentration at Namie inthe summer, when the re-suspension from soil was negligible due to the higher soil moisture content (following considerable rain) and lower wind speed. As with re-suspension
from bare soil, only radioactive decay was considered for
Rdecay and the other processes were not considered. The parameter fgreen was obtained from the database of WRFV3
and was originally derived from satellite Advanced VeryHigh Resolution Radiometer (AVHRR) normalized difference vegetation index (NDVI) data (Gutman and Ignatov,1998). Whereas fforest remains constant, the monthly averaged fgreen was used in order to reflect seasonal changes in
the activity of the biota.
2.4

Emission from FDNPP (primary emission,
additional emissions from the reactor buildings,
and unexpected re-suspension associated withdebris removal operations)Figure 3. Monthly mean emission flux of radiocesium released
from the reactor buildings of FDNPP from October 2011 to August
2015 as estimated by TEPCO (TEPCO, 2012–2015).

zontal distribution of 137 Cs deposition of the airborne radiological survey (NRA, 2012; as shown later in Fig. 4a) andsurface air concentrations measured at Tsukuba (Fig. 1a).
We selected this inventory because it is JAEA’s most up-todate version. Based on an integrated understanding of environmental radioactivity, atmospheric dispersion, and the nuclear reactors, the JAEA team has carefully established a series of inventories for about 5 years, starting with Chino etal. (2011), followed by Katata et al. (2012a, b) and Terada
et al. (2012), and finally the current inventory (Katata et al.,2015), which is substantially improved compared to the previous versions.
Ongoing emissions during the study analysis period after the emergency situation, i.e., January to December 2013,
were obtained from the Tokyo Electric Power Co. Inc.
(TEPCO) monthly mean emission flux from the reactor
buildings (TEPCO, 2012, 2013, 2014a, b, 2015). Because
only the sum of 134 Cs and 137 Cs was provided, the fractionsof these two isotopes were calculated based on their halflives and the assumption that their activities were equal in
March 2011 (e.g., Katata et al., 2015), as shown in Fig. 3.
The values range from 105 to 107 Bq h−1 ; however, for simplicity we set a constant value of 106 Bq h−1 in the current
simulation.
In August 2013, unexpected re-suspension associated withdebris removal operations was reported by TEPCO (2014c)
and NRA (2014) and the gross amount was 1010 –1011 Bq of
137 Cs (TEPCO, 2014c; NRA, 2014; Steinhauser et al., 2015).
The impact of this unexpected re-suspension is briefly discussed in Sect. 5.2 along with an additional finding, but thisemission was not considered in the present LM simulation. In
this study we focused on the ongoing and continuous emission, mostly from the natural environment, that is difficult tocontrol.

3
The Japan Atomic Energy Agency (JAEA)’s latest estimateof the primary emission from FDNPP, Katata et al. (2015),
was applied for the emergency situation of March 2011 toevaluate the performance of the LM model against the horiAtmos. Chem. Phys., 16, 13149–13172, 2016

Field observations

Details of the surface air activity concentration measurement
techniques can be found in Ishizuka et al. (2016) and Kinase et al. (2016) for Namie and Igarashi et al. (2015) for
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M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium
Tsukuba. At the both sites, 134 Cs and 137 Cs surface air concentrations were measured. The sampling intervals were 1–
2 days at Namie and 1 week at Tsukuba for the analysis period of this study, the year 2013. The observations at Namie
started on 14 December 2012, while those at Tsukuba started
on 31 March 2003, before the FDNPP accident. In March
2011, the sampling interval was increased at Tsukuba to 6 h–
1 day and the data for these periods were used for the validation of LM and its parameters, as presented in Sect. 4.1.
The Namie site is located on the schoolyard of Namie HighSchool, Tsushima Campus, in the Tsushima district of Namiein Fukushima prefecture, as shown in Table 1 and Fig. 1. The
town of Namie extends from the Hamadori coastal area (denoted as C in Fig. 1) to the Abukuma highland area (B in
Fig. 1). There are two sites in Namie used in the study. In order to distinguish the Tsushima Campus site in the highlandarea from the monitoring post site located in Omaru districtin Namie in the coastal area, the Tsushima site is referred toas Namie (sometimes as Namie (Tsushima)) and the Omaru
site is referred to as Omaru (sometimes as Namie (Omaru))
throughout the manuscript. The Tsukuba site is located on
the premises of the Meteorological Research Institute (Table 1 and Fig. 1a).
Namie (Tsushima) was located in the DRZ(> 50 mSv yr−1 , ∼ 9.55 µSv h−1 )1 and the observed 137 Cs
deposition amount was 2300 kBq m−2 (NRA, 2012). The
ambient gamma dose rate was 11.2 µSv h−1 on 1 April
2012 at the site and had dropped to 4.8 µSv h−1 on 16
February 2016 and at the HRZ level (20–50 mSv yr−1 ,3.85–9.55 µSv h−1 ). Tsukuba is located approximately
170 km southwest of FDNPP. The observed 137 Cs deposition
amount was 21 kBq m−2 (NRA, 2012), 2 orders of magnitude lower than at Namie, and the dose rate has remained
below 0.1 µSv h−1 since 2012.
4

Results

Section 4.1 presents a validation of the LM model and the optimization of the model deposition parameters by using airborne observations (NRA, 2012) and the emission inventoryof Katata et al. (2015) for the emergency situation of March
2011. Using the optimized ranges of model parameters validated in Sect. 4.1, the simulated re-suspension of 137 Cs fromsoil and forest, as well as emission from FDNPP, is presented
in Sect. 4.2, and the budgets for re-suspension, transport, andre-deposition are presented in Sect. 4.3.

1 Calculated by the following equation: Annual radiation exposure = (ambient dose rate – background dose rate(= 0.04 µSv h−1 )) × (8 h + 0.4 × 16 h) × 365 days.

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4.1

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Model and parameter validation for the emergency
situation (March 2011)

Figures 4 and 5 show the observed and simulated distribution of 137 Cs deposition in March 2011, and the scatterplot
comparing the observational and simulation results. In thesimulation shown in the figures, the “reference” sets used for
dry and wet deposition parameters, namely the collection efficiency of aerosols using hydrometeors Ec (Eq. A1) and the
dry deposition velocity over land vd (Eq. A3), were 0.04 and0.1 cm s−1 , respectively.
Since LM uses simple parameterizations for dry and wet
deposition, as described in Appendix A, sensitivity tests wereconducted for model validation, together with optimizationof the deposition parameters. Table 2 summarizes the rangesof the deposition parameters for the sensitivity tests and theresults of the 137 Cs budget and statistical measurements between the observations and the simulation.
The parameter Ec was 0.05 for the JMA dispersion model(JMA, 1998) but the targeted species are different. For example, volcanic ash (particles larger than 1 µm in diameter) used
for the JMA model was generally larger than the aerosols carrying 137 Cs (around 1 µm in diameter observed in the downwind area, Tsukuba; Kaneyasu et al., 2012; Adachi et al.,2013). Since the inertia of these smaller 137 Cs particles islikely to be smaller than that for volcanic ash, Ec could besmaller. The range of Ec was set as 0.02–0.06. The dry deposition velocity vd was selected as 0.1 cm s−1 for 137 Cs inFuruno et al. (1999). The range for vd was set as 0.05 to 0.15
in the present study.
The emission inventory of Katata et al. (2015) amounted
to 14.1 PBq from 12 March to 1 April 2013. The simulated
deposition over the model domain (138–143◦ E, 34–39◦ N)
ranged from 3.4 to 4.7 PBq, which is approximately 24–33 %of the emission from the FDNPP. Sixty percent of the total deposition occurred over land, for a total of 2.0–2.8 PBq,
which is close to the observed value of 2.68 PBq, and the observed value is within the range of the sensitivity runs. Statistical measures such as the fractional bias FB, the correlation coefficient R (linear vs. linear), and FAx (fraction of thesimulated values within a factor of x) are listed in Table 2.
To find better combinations of (or to optimize) the dry andwet deposition parameters, sensitivity runs were screened
based on the criteria FA10 > 0.9, FA5 > 0.7, R > 0.75, andan absolute value of FB < 10 %. Only one combination (Ec ,vd ) = (0.04, 0.1 cm s−1 ) satisfied the criteria, and thus thisis referred to as the “reference” parameters. To evaluate
the sensitivity (or uncertainty) of the re-suspension simulation for 2013 due to the deposition parameters, the range ofthe combination of parameters was set as (Ec , vd ) = (0.03–
0.05, 0.05–0.1 cm s−1 ) around the reference parameters (referred to as the “optimized range”) by excluding the parameters with the worse performances. The ranges of the statistical measures of the optimized runs are listed in Table 2.
FB, R, FA2, FA5, and FA10 after the optimization had theAtmos. Chem. Phys., 16, 13149–13172, 2016

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Table 2. 137 Cs budget and statistical analysis for the comparison of observed and simulated deposition data for March 2011.
Eca
(–)

vdb
(cm s−1 )

c
Dall
(PBq)

d
Dland
(PBq)

FBe
(–)

Rf
(–)

FA2g
(–)

FA5h
(–)

0.02–0.06

0.05–0.15

3.4–4.7

2.0–2.8−0.25–
0.00050

0.73–0.78

0.25–0.30

0.63–0.77

0.04
0.03–0.05

0.10
0.05–0.10

4.2
3.7–4.3

2.5
2.2–2.6−0.062−0.18 to−0.0360.75
0.74–0.77

0.28
0.26–0.30

0.74
0.68–0.74

Sensitivity test
Range
Optimization used for the
re-suspension analysis for 2013
Reference
Optimized range
Reference values
Observed deposition over land
Dobs (NRA, 2012)
Emission amount(Katata et al., 2015)

2.68 PBq
14.1 PBq

a Collection efficiency; see Eq. (A1) b Dry deposition velocity over land; see Eq. (A3). c Simulated deposited amount over the whole model domain. d Simulated deposited
amount only over land. e Fractional bias between Dland and Dobs . f Correlation coefficient between each grid cell of the observed and simulated deposition (linear vs. linear).
g Fraction of simulated values within a factor of 2 of the observed values. h Fraction of simulated values within a factor of 5 of the observed values. f, g, h Compared only at grids
where the observed values are greater than 10 kBq m−2 .

Figure 4. (a) Aircraft observation and (b, c) simulation of 137 Cs deposition depicted (b) only over land and (c) for the whole domain(kBq m−2 ). The observation was interpolated to the model grid (1 longitude = 0.125◦ and 1 latitude = 0.1◦ ). A decay correction for theobservation was made for March–May 2012, varying depending on the region. The simulation provided a cumulative value from 12 March
to 1 April 2011. The total activities are presented as numbers. The color scales are the same for (a)–(c) but observed values are not depicted
below the detection limit, 10 kBq m−2 . The deposition parameters of the simulation in the figure are Ec = 0.04 and vd = 0.1 cm s−1 .

ranges −0.18 to −0.036, 0.74–0.77, 0.26–0.30, 0.68–0.74,
and 0.91–0.92, respectively. These statistical measures werecomparable to those reported in previous multi-model comparison studies (R: 0.27–0.85; FB: −0.84–0.56; and FA2:
0.14–0.57, in SCJ, 2014; Draxler et al., 2015). The current
model is thus shown to be sufficiently credible for the budget analysis in this study, despite the simple parameterizationand the low resolution in space (∼ 11 km) and time (3 h).
In agreement with many previous studies, the simulated
contribution of wet deposition was larger than that of dry
deposition: the ratio of the amount of dry to wet deposition
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ranged from 0.12 to 0.23 for the optimized parameter ranges,
indicating that the results were less sensitive to the dry deposition parameter. Generally speaking, R became higher as
Ec became lower, whereas FAx became higher as Ec became
higher for the various ranges of the sensitivity tests. Therefore, lower Ec did not meet the criteria of FAx and higher
Ec did not meet the criteria of R. Consequently, after the optimization, the maximum values of the statistical measures
were lower but the minimum values became higher, indicating that the optimization was successful in excluding the parameters with the worse performances (rather than selecting
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using the optimized parameter ranges Ec = 0.03–0.05 andvd = 0.05–0.1 cm s−1 . Due to differences in the parameters
used, the surface air concentration could vary by approximately 1 order of magnitude in transported plumes that experienced wet scavenging along their pathway.
4.2

Figure 5. Scatterplot between the observational data and the simulation results for 137 Cs deposition (kBq m−2 ). The deposition
parameters of the simulation in the figure are Ec = 0.04 andvd = 0.1 cm s−1 . N indicates the number of samples, and the statistical measures R, FA2, and FA5 are described in Table 2.

Figure 6. Time series of the surface activity concentration of (black)observed and (red) simulated 137 Cs concentrations at Tsukuba from
12 March to 1 April 2011 (Bq m−3 ). The red shaded areas indicate
the range of 137 Cs concentrations obtained when the simulations
were run using the optimized parameter ranges Ec = 0.03–0.05 andvd = 0.05–0.1 cm s−1 .

the best parameters). It should be noted here that the optimized deposition parameters are not necessarily physicallyvalid but rather give simulation results consistent with theavailable observations. The results presented in this section
indicate that the current LM simulation with these optimized
parameters has the potential to reproduce consistent featuresof the radiocesium budget over the Tohoku and Kanto regionsof Japan.
Figure 6 shows the temporal variation in simulated (red)and observed (black) 137 Cs concentrations at Tsukuba inMarch 2011. The model reproduced the three major plumes
arriving at Tsukuba on 15–16, 20–21, and 28–30 March; aplume on 23 March was not observed but only appeared in
the simulation. The red shaded areas indicate the range of
137 Cs concentrations obtained when the simulations were run
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Re-suspension in 2013

Figure 7 shows the observed and daily mean simulated (withthe reference parameters) surface air concentrations at Namieand Tsukuba for the year 2013. The red, green, and bluelines indicate re-suspension from soil, re-suspension fromforest, and emission from the FDNPP reactor buildings, respectively. Scatterplots between the observed and the simulated total concentrations are also shown in the figure and thestatistical measures between them are listed in Table 3. The
correlation coefficients (linear vs. linear) are low at the bothsites, but the simulated and the observed medians are close toeach other and FAx shows high values at the both sites. Consequently, the current simulation is credible enough for thequantitative assessment of the annual radiocesium budget.
As discussed in Sect. 2.2, note that the re-suspension flux
due to the dust emission module (Ishizuka et al., 2016) ismultiplied by 5 in this study in order to adjust to the observation level at Namie in the cold season (January to March,
October to December). Also note that the re-suspension coefficient rconst in Eq. (4) was set as 10−7 h−1 in order to adjust to the observation level at Namie in the warm season(May to September). The emission flux reported by TEPCO
varied from 105 to 107 Bq h−1 during the study period but
we set it to 106 Bq h−1 for simplicity. We did not try to precisely adjust rconst to the observation by, for example, using inverse modeling, and instead we simply multiplied bypower-of-10 values (except for dust emission, which is five),
which are constant in time and space because (1) this facilitated straightforward interpretation of the simulation resultsby keeping the simulated variation solely originating fromthe variation in boundary conditions (i.e., emission and meteorology), and therefore (2) this provided simple but useful clues for understanding the re-suspension mechanisms,
which remain unknown.
Using the dust emission module (which has a physicalbasis), 137 Cs flux re-suspended from soil could account for
the level of the observed surface air concentration of 137 Cs
at Namie in the cold season. Under the influence of thenorthwesterly winter monsoon, the surface wind speed ishigh over the contaminated area compared to the summer(as shown later in Fig. 10). Note that the flux might be a
maximum estimate since it does not consider land surface
processes (such as soil moisture, snow cover, or migrationof 137 Cs in the soil and biota) and decontamination, which
could reduce the 137 Cs re-suspension flux. In contrast, inthe warm season, the estimated flux significantly underestimated the observation by 1 to 2 orders of magnitude dueto the weak surface wind, indicating that the dust emission
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Table 3. Statistical analysis for the comparison of observed and simulated 137 Cs air concentration at Namie and Tsukuba for 2013.
Site

N
(–)

Median, obs.
(mBq m−3 )Median, sim.
(mBq m−3 )

R
(–)

FA2
(–)

FA5
(–)

FA10
(–)

Namie
Tsukuba

311
74

0.71
0.022

0.69
0.024

0.12
0.19

0.67
0.70

0.94
0.95

0.99
1.00

Figure 7. Time series of the surface air concentration of (black) observed 137 Cs and (colors) simulated daily 137 Cs levels on the left andscatterplots between observed and simulated total (i.e., sum of colors) 137 Cs at (a) Namie and (b) Tsukuba. The colored lines indicate
simulated 137 Cs concentrations due to (red) re-suspension from soil using the scheme given in Ishizuka et al. (2016) (five times), (green)
re-suspension from forest with a re-suspension rate of 10−7 h−1 , and (blue) emission from the FDNPP reactor buildings with a constantemission rate of 106 Bq h−1 . The deposition parameters of the simulation in the figure are Ec = 0.04 and vd = 0.1 cm s−1 .

process may not be the sole process involved in sustaining
the air concentration of 137 Cs during this period. Introducing
the 137 Cs re-suspension component from forest with a resuspension coefficient of 10−7 h−1 and a monthly variation in
the green area fraction (derived from NDVI) could quantitatively account for the observed air concentration together
with its seasonal variation at Namie. Even though both thesimulated re-suspension from soil and forests reproduced thequantity and seasonal variation in the background concentration (in other words, concentrations originating from continuously presenting emissions, such as natural emissions, andnot accidental ones) at Namie, sporadic peak events, such as
the daily mean 137 Cs concentration exceeding 10 mBq m−3
as observed in June and August at Namie, were not simulated. Some specific re-suspension events might occur fromthe highly contaminated areas such as the premises of FD-

Atmos. Chem. Phys., 16, 13149–13172, 2016

NPP (e.g., debris removal operations) or very close to FDNPP on these days, as indicated later in Sect. 5.2. The simulated 137 Cs concentrations due to the monthly mean emission
from the reactor buildings (i.e., 106 Bq h−1 ) significantly underestimated the observed concentration by more than 3 orders of magnitude at Namie and by 2 orders of magnitude
at Tsukuba. Even the maximum estimate of 107 Bq h−1 does
not reach the observed level. The emission from FDNPP may
not have been the sole process sustaining the air concentration of 137 Cs in 2013, supporting the discussion in Igarashi
et al. (2015), which concluded that direct emission from the
FDNPP played a minor role in the observed atmospheric radiocesium concentrations over Tsukuba during 2013–2014.
The observed air concentration of 137 Cs at Tsukuba was
about 1 to 2 orders of magnitude lower than that at Namie.
The simulated difference between the two sites inside andwww.atmos-chem-phys.net/16/13149/2016/

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Figure 8. Annual total amounts of (a) the re-suspension and (b) the re-deposition of 137 Cs. The total activities are presented as numbers.
Fractions of the re-suspension and the re-deposition to the observed 137 Cs deposited amounts are also shown in (c) and (d), respectively. The
deposition parameters of the simulation in the figure are Ec = 0.04 and vd = 0.1 cm s−1 .

outside of the contaminated areas was consistent with theobserved difference. This finding indicates that the current
LM simulation provided consistent features of re-suspension,transport, and re-deposition in the Tohoku and Kanto regionsof Japan of 137 Cs originating from Fukushima.
4.3

Budget analysis

Figure 8 illustrates the simulated (with the reference parameters) annual amounts of total re-suspension and re-depositionof 137 Cs, together with their ratios to the observed deposition (Fig. 4a). The simulated areal total re-suspended amountwas 1.28 TBq, which was equivalent to 0.048 % of the total deposited amount, 2.68 PBq. The areal total re-deposited
amount (with the reference parameters) was 0.22 TBq (0.18–
0.23 TBq for the optimized range of the deposition parameters), corresponding to approximately 17.2 (14.1–18.0) % of
the re-suspended amount deposited mainly in the Tohoku region, with the remainder being transported out of the model
domain. Therefore, the regional mean rate in the decreasewww.atmos-chem-phys.net/16/13149/2016/of the land surface 137 Cs concentration due to re-suspension
was estimated to be 0.040 (0.039–0.041) % yr−12 , equivalent
to 1.1 (1.06–1.12) × 10−6 day−1 . The spatial distribution of
the re-suspension and re-deposition ratio to the primary deposition ranged from 0.01 to 0.3 and 0.001 to 0.03 %, respectively. The spatial distribution of the land surface 137 Cs deposition decay due to re-suspension ranged from 2.7 × 10−7 to
8.2 × 10−6 day−1 . Re-suspension therefore had a negligible
effect on reducing land surface radioactive contamination.
5

Discussion

Seasonal variation in the surface activity concentration andits source contributions are extensively discussed in Sect. 5.1.
A possible source of the observed sporadic peak events,
which could not be reproduced by the simulation, is dis2 The amount re-suspended, excluding re-deposition (1.28 TBq
minus 0.22 (0.18–0.23) TBq) for the year 2013, divided by the total
deposited amount of 2.68 PBq.

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13160M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium

Figure 9. Time series of observed (black) and simulated (colors) 137 Cs surface air concentration at Namie (a, c) in winter from January to
March and (b, d) in summer from June to August 2013. The colors of the lines are the same as in Fig. 7 but the time variation is hourly.
The shaded areas indicate the range of 137 Cs concentrations obtained when the simulations were run using the optimized parameter ranges
Ec = 0.03–0.05 and vd = 0.05–0.1 cm s−1 .

cussed in Sect. 5.2. The effects of other processes that were
not considered in the model, such as land surface processes
and decontamination, are discussed based on the dose ratemeasurements from the monitoring posts in Fukushima in
Sect. 5.3. Future issues are summarized in Sect. 5.4.
5.1

Seasonal variation and source contribution

The discussion in this section expands on that in Sect. 4.2.
Figure 9 shows the same temporal variation as Fig. 7 but for
simulated (using the optimized ranges of parameters) resultsfor 137 Cs from dust and FDNPP in winter (January to March)and from forest and FDNPP in summer (June to August).
In the winter, the simulated trend for dust agreed well withthe observed trend (Fig. 9a), and the surface air concentration
during this period was positively correlated with the surface
wind speed in both the simulation (R = 0.88 on an hourly basis) and the observations (R = 0.32 on a daily basis). There
was a sporadic peak in the observational data of 6.7 mBq m−3
from the 17 March at 13:00 local time (LT) to 18 March at
13:00 LT that could not be reproduced by the dust module,
and this peak coincided with a plume arriving from FDNPP,
as shown in Fig. 9c. Therefore, the observed peak could be
accounted for by specific re-suspension events on the order of
109 Bq h−1 (the left axis divided by the right axis multipliedby 106 Bq h−1 in Fig. 9c) if they occurred on the premisesof FDNPP or close to the area. There are also two events exceeding 2 mBq m−3 , one in January and another in February.
It is unlikely that the two peaks originated from the directemission from FDNPP and likely that they originated fromthe dust emission because the observed peaks coincided withthe simulated dust peaks (Fig. 9a) and not with the simulated
peaks due to the FDNPP emission (Fig. 9c).
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In the summer, the simulated quantity as well as the variation in the forest data agreed well with the observed data
(Fig. 9b). Because there is only monthly variation in the simulated emission, the simulated daily trend solely originatedfrom variations in the meteorological parameters (wind field,
turbulent mixing, and wet scavenging). A significant peak of
60.4 mBq m−3 is observed from 14 August at 13:00 LT to 15
August at 13:00 LT. This observed level was approximately
2 orders of magnitude larger than the simulated level and 1
to 2 orders of magnitude larger than the observed level forthe other days in this period. Therefore, continuous emission
such as re-suspension from forest is less likely to be the origin of the peak. Because the observed peak and the simulated
peak of 137 Cs from FDNPP coincided (Fig. 9d), the observed
level could be accounted for by specific re-suspension eventson the order of 1010 Bq h−1 either on the premises of FDNPP or close to the area. There have been several argumentsthat the observed peaks in August 2013 were associated withdebris removal operations at FDNPP, and this is discussed
separately in Sect. 5.2.
Figure 10 illustrates the seasonal mean surface wind vector and surface air 137 Cs concentration (simulated using thereference parameters) due to (a) dust re-suspension in thewinter and (b) forest re-suspension in the summer. Due tothe prevailing northwesterly winter monsoon, 137 Cs was carried southeastward in the winter. In the summer, under the
influence of the Pacific high-pressure system, 137 Cs was carried inland. The monthly mean wind speed is high in winterand low in summer. The upper panels of Fig. 11 illustrate theobserved and simulated (using the optimized ranges of parameters) total (from soil, forest, and FDNPP) 137 Cs concentration at Namie and Tsukuba. The time resolutions of the

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Figure 10. Seasonal mean surface (10 m above ground level) wind vector and 137 Cs surface concentration (a) due to dust re-suspension
in winter from January to March and (b) due to forest re-suspension in summer from June to August. The deposition parameters of thesimulation in the figure are Ec = 0.04 and vd = 0.1 cm s−1 .

simulation are daily for Namie and weekly for Tsukuba to
be consistent with the sampling intervals of the two respective sites. The simulation successfully reproduced the quantity and variation in the observed background concentration
at Namie and Tsukuba but could not reproduce the sporadic
peak events observed at Namie, as discussed above. The simulation also significantly underestimated the observations at
Tsukuba from January to March, 2013. Due to the northwesterly monsoon (Fig. 10a), there was less air mass transported from FDNPP to Tsukuba in the winter and therefore
this underestimation is probably due specifically to underestimation of the simulated re-suspension around Tsukuba.
The lower panels of Fig. 11 show the relative contributions of
137 Cs from soil and forests at Namie and Tsukuba. The contribution from FDNPP was negligible throughout the year. At
both sites, the contribution from dust was high (0.7–0.9) inthe cold season and low (0.2–0.4) in the warm season due tohigher surface wind speed in the cold season.
Figure 12a shows the observed and simulated (with thereference parameters) monthly Namie to Tsukuba 137 Cs concentration ratios. The mean concentration ratio exceeded 100
in June and 200 in August due to the sporadic peak events.
The monthly median would be relevant for comparing thebackground observation with the simulation results by considering only continuous emission sources. The values ofthe simulated concentration ratio and its seasonal variationagreed fairly well with the observed monthly median ratio: the observed and simulated annual means were 38.9 andwww.atmos-chem-phys.net/16/13149/2016/

31.1, respectively. Figure 12b shows the monthly mean simulated re-suspension source area contributions to the 137 Cs
air concentration at Namie and Tsukuba. The re-suspension
source area is defined as the model grid where the observed
deposition amount exceeded 300 kBq m−2 (Fig. 4a) and includes the Namie grid (2300 kBq m−2 ). Eighty to 90 % of the
137 Cs air concentration at Namie originated from the source
region, and there was no clear seasonal variation in the value.
In contrast, Tsukuba is characterized as a downwind region
and there was clear seasonal variation in the source contribution ratio: high in summer and low in winter, due to thesummer and winter monsoons, as discussed above. Nonetheless, the highest value at Tsukuba was 0.4 in July, and somore than half of the 137 Cs concentration at Tsukuba originated locally or from areas other than the contaminated regions throughout the year. As shown in Fig. 11a and b, thevariability in the simulated concentration at Tsukuba due touncertainty in the deposition parameters was much larger
than that at Namie. The differences in the variability indicated that the Namie and Tsukuba sites can be characterized
as the source area and the downwind area, respectively: as thetime required for the plume to move from the emission site tothe observation site increases, the variability becomes larger
due to the increased chance for the plume to experience dry
and wet scavenging.

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Figure 11. Time series of (a, b; black) the observed and (purple) the simulated surface 137 Cs concentration due to total re-suspension and (c,d) the relative contribution of (red) dust and (green) forest re-suspension to the 137 Cs concentration at (a, c) Namie and (b, d) Tsukuba. The
shaded areas of (a) and (b) indicate the range of 137 Cs concentrations obtained when the simulations were run using the optimized parameter
ranges Ec = 0.03–0.05 and vd = 0.05–0.1 cm s−1 .

Figure 12. Monthly mean (a) observed mean, observed median and simulated Namie to Tsukuba 137 Cs concentration ratio and (b) simulated
re-suspension source area (where the observed deposition amount > 300 kBq m−2 ) contributions to 137 Cs air concentration at Namie andTsukuba. The deposition parameters of the simulation in the figure are Ec = 0.04 and vd = 0.1 cm s−1 .

5.2

Possible source of sporadic peak events

There have been several scientific studies and governmental reports on the unexpected re-suspension from FDNPP in
August 2013. The high-dose-rate alarm was activated on August 19 within the premises of FDNPP associated with thedebris removal operation. Matsunami et al. (2016) related
the radiocesium contamination of brown rice in Fukushima
in 2013 to this operation, whereas MAFF (2015) denied
any association. The NRA estimated the 137 Cs emission rateduring the debris removal operation as 6.7 × 1010 Bq h−1
and the cumulative amount as 1.1 × 1011 Bq (NRA, 2014).
Atmos. Chem. Phys., 16, 13149–13172, 2016

TEPCO (2014c) estimated the emission rate during the operation as 5.8 × 1010 –1.2 × 1011 Bq h−1 and the cumulative
amount as 1.3–2.6 × 1011 Bq. Steinhauser et al. (2015) estimated the gross amount as 2.8 × 1011 Bq using measurements of weekly air filter sampling and monthly deposition,
and a numerical simulation. Their estimates are similar inmagnitude to our estimate (1010 Bq h−1 ; see Sect. 5.1) but
the dates are different: our observed peak was earlier than
the reported removal operation.
Our daily sampling showed a peak concentration(60.4 mBq m−3 ) from 14 August at 13:00 LT to 15 August at
13:00 LT before the reported operation but did not detect highwww.atmos-chem-phys.net/16/13149/2016/

M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium

Figure 13. 48 h forward trajectory (statistical locations of LPs
within 1 km AGL) predicted by the LM starting at (a) 12:00 LT and(c) 15:00 LT on 14 August and (b) 09:00 LT and (d) 15:00 LT on 19
August. Blue lines indicate median locations of LPs at 1 h and 1-day
intervals as blue and red dots, respectively. Sky-blue lines indicate
17th and 83rd percentile locations of LPs and red dashed circles
indicate areas containing 66 % of the LPs.

concentrations in the 19 August (0.33 mBq m−3 for August
18 at 13:00 LT to 19 August at 13:00 LT and 1.2 mBq m−3
for 19 August at 13:00 to 20 August at 13:00 LT). Figure 13 shows the forward trajectories predicted by the LM(statistical locations of LPs) starting from FDNPP on 14
August (left) and 19 August (right). The sky-blue lines
and red dashed circles indicate areas containing approximately two-thirds of the LPs within 1 km AGL: the extent
of the area reflects horizontal and vertical atmospheric diffusion. The highest dose rate peaks were observed from
13:50 to 14:10 LT on 19 August at 2.8–8.3 km north andnorth-northwest of FDNPP on the leeward side, as reported
by Fukushima prefecture (https://www.pref.fukushima.lg.jp/
download/1/20130827moni.pdf.pdf). The forward trajectories on 12 August indicated that plumes during the debris removal operation traveled north-northwest to north of FDNPP(Fig. 13b and d), rather than towards the west-northwest,
where the Namie site is located. On 14 August, on the other
hand, plumes were transported towards the west (startingat 12:00 LT, Fig. 13a), and then to the north (starting at
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Figure 14. Daily mean gamma dose rate at the six monitoring sites:(a; red) Okuma and (orange) Namie (Omaru district) in the coastal
area, (b; blue) village of Iitate and (sky blue) town of Kawamata
in the Abukuma highland area, and (c; green) cities of Fukushimaand (greenish yellow) Shirakawa in the Nakadori valley area of
Fukushima prefecture as presented in Fig. 1b for two fiscal years(from April 2012 to March 2014). The first-order decreasing ratesof the least-squares approximation over the period of no snow cover(May to October, 2012 and 2013) is also presented, along with theradioactive decay (134 Cs plus 137 Cs, by assuming the same amountof activity on 12 March 2011) in April 2012 and March 2013.

15:00 LT, Fig. 13c) due to fast changes in wind direction, resulting in the simulated peak concentration shown in Fig. 9dduring this period. Our simulation and observations together
indicated that the same order of magnitude of 137 Cs emission
occurred on August 14–15 and on 19 August. Alarm activation was not reported on August 14–15, but debris removal
operation was also conducted on August 14 and 16 (MAFF,
2015).

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5.3

M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesiumLand surface processes, decontamination, and dust
emission suppression due to snow cover

Figure 14 presents the time series of ambient gamma dose
rates measured at the monitoring posts in Fukushima prefecture indicated in Fig. 1b. The data were obtained fromthe Nuclear Regulation Authority (NRA), Japan, website(http://radioactivity.nsr.go.jp/map/ja/index.html). A total ofsix monitoring posts, two from each of the three geographical areas (Hamadori coastal area, Abukuma highland area,
and Nakadori valley area), were selected. There are tens tohundreds of monitoring posts in each municipality (village,
town, and city) in Fukushima prefecture. We selected a monitoring post in each municipality by applying the following
conditions: a post showing the highest dose rate of all posts inthe municipality at the time of downloading (around 11:00,
28 December 2015), data are available since 1 April 2012,
and the instruments are situated 100 cm above the ground.
In April 2012, the government of Japan designatedthe evacuation-directed zones as a difficult-to-return zone
(DRZ; > 50 mSv yr−1 ; 9.55 µSv h−1 ), a habitation-restrictedzone (HRZ; 20–50 mSv yr−1 ; 3.85–9.55 µSv h−1 ), and azone being prepared to have the evacuation directive lifted(< 20 mSv yr−1 ; 3.85 µSv h−1 ) (METI, 2012). The two sites
in the Hamadori area, Okuma and Omaru, have been designated DRZ (13.6 and 11.8 µSv h−1 on 28 December 2015).
The dose rates at the two sites in the Abukuma area, Iitate andKawamata, dropped below the HRZ level during the analysis period in this study (1.17 and 0.521 µSv h−1 on 28 December 2015). The dose rates in the Nakadori area are below
1 µSv h−1 (0.242 and 0.201 µSv h−1 on 28 December 2015).
The dose rate significantly dropped when the ground
was covered with snow, in January and December 2013 inHamadori and Nakadori, and from January to early Marchand December 2013 in Abukuma (the elevation of Abukuma
is 500–1000 m and higher than Hamadori and Nakadori).
Snow cover suppresses re-suspension due to dust emission.
Namie (Tsushima) is located in Abukuma and the ground
was covered with snow until early March (Ishizuka et al.,2016). The observed air concentration of 137 Cs at Namie(Tsushima) was correlated with the wind speed in the winterfrom January to March, indicating that re-suspension during
the period was mechanically induced. In the winter, dust resuspension from outside Abukuma, such as from Hamadori
and Nakadori, or from land surface where the snow cover waspartly melted due to solar radiation, might be the dominantsource contributing to the 137 Cs surface air concentration at
Namie (Tsushima) when the ground was covered with snow.
The first-order decreasing rates fitted by the least-squares
approximation for the period without snow cover, May to
October 2012 and 2013, are presented in Fig. 14. The rates
ranged from 5.2 to 12.1 × 10−4 day−1 . The monthly mean
radioactive decay rates of total radiocesium (134 Cs + 137 Cs),
determined by assuming that the activities of 134 Cs and
137 Cs were equivalent in March 2011, were 4.2 × 10−4 andAtmos. Chem. Phys., 16, 13149–13172, 2016

3.0 × 10−4 day−1 in April 2012 and March 2014, respectively, due to the difference in half-life of 134 Cs and 137 Cs(2.07 and 30.1 years, respectively). By assuming that the
gamma dose rate primarily originated to radiation from theland surface radiocesium, the radioactive decay accounted
for 35–50 % of the decreasing rate of total ground radioactivity; the exception was Shirakawa, where radioactive decay
accounted for 55–80 % of the decrease. In other words, 50–
65 % of the ground radioactivity decrease was likely due toland surface processes, decontamination, and re-suspension
to air. As discussed in Sect. 4.3, the estimated decreasingrate due to re-suspension was 2.7 × 10−7 –8.2 × 10−6 day−1 ,which is 2 to 3 orders of magnitude smaller than the decreasing rates due to the other processes (10−4 –10−3 day−1 ).
It is difficult to distinguish the contributions of land surface processes and decontamination. By subtracting the radioactive decay rate (3.0–4.2 × 10−4 day−1 ) and the decreasing rate due to re-suspension (2.7 × 10−7 –8.2 × 10−6 day−1 )from the gross decreasing rate (5.2–12.1 × 10−4 day−1 ), theestimated decreasing rates due to land surface processes anddecontamination ranged from 1.0 to 7.9 × 10−4 day−1 . Matsuda et al. (2015) summarized the depth profiles of radiocesium in soil at more than 80 locations in Fukushima, including Hamadori, Abukuma, and Nakadori. They foundthat the radiocesium levels have been slowly migrating
downward, with rates ranging from 1.7 to 9.6 kg m−2 yr−1
(equivalent to 1.1–6.0 mm yr−1 for a dust particle densityof 2650 kg m−3 and a porosity of 0.4 m3 m−3 , for example).
Evrard et al. (2015) summarized that significant transfer ofparticulate-bound radiocesium occurs during major rainfall
and runoff events (e.g., typhoons and spring snowmelt). Together with the relaxation depth–dose rate relationship provided by Saito and Petoussi-Hess (2014), the decreasing rate
due to land surface processes such as downward migration,
runoff, and erosion could be quantified and thus the decontamination effect could be separately extracted.
5.4

Future issues

Issues that remain to be resolved in future research aresummarized in this section. First of all, re-suspension frombiota could be predominant in the warm season, but the resuspension sources as well as mechanisms remain essentially unknown. Further study is needed to understand themechanism based on field experiments and numerical simulations. The current estimation could account for the measured background concentration (0.1–1 mBq m−3 ) but could
not reproduce the observed sporadic peak concentration (1–10 mBq m−3 ) at the Namie site. Further study is needed
to identify the cause. The dust flux module has been validated at a single location. The module needs to be improved to be applicable to various land use and soil texture conditions. The decontamination may reduce resuspension afterward, whereas the resuspension may occur duringdecontamination-related work. This effect should be evaluwww.atmos-chem-phys.net/16/13149/2016/

M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium
ated in the future. The current estimation was based on a single model simulation. Variability in multi-model simulations
is rather large (SCJ, 2014; Draxler et al., 2015) and therefore
multi-model assessment will be indispensable for long-term
re-suspension analysis.
6

4. The contribution from additional emission from the reactor buildings of FDNPP (106 Bq h−1 ) was negligible
throughout the year and underestimated the observed air
concentration by 2 to 3 orders of magnitude at both observation sites.
5. At Namie and Tsukuba, the simulated contribution ofre-suspension from soil was high (0.7–0.9) in the coldseason and low (0.2–0.4) in the warm season; the remaining contribution was from forest and was low in
winter and high in summer. The contribution of the resuspension from the source area (where the aircraftobserved deposition exceeded 300 kBq m−2 ) to the airconcentration at Namie was 0.8–0.9 throughout theyear, while that at Tsukuba varied from 0.1 to 0.4, andwas high in the summer and low in the winter.

Conclusions

The long-term effect of 137 Cs re-suspension from contaminated soil and biota due to the Fukushima nuclear accident
has been quantitatively assessed using a numerical simulation, a field experiment on dust emission in the contaminated area (Namie, Fukushima), and air concentration measurements inside (Namie) and outside (Tsukuba, Ibaraki) of
the area. The re-suspension mechanism remains unknown.
We therefore utilized the observational data obtained both
inside and outside the contaminated area, together with atransport model, to provide a robust budget analysis of there-suspension, transport, and re-deposition of 137 Cs in theeastern part (the Tohoku and Kanto regions) of Japan. Ourfindings are summarized as follows:

6. The simulated annual total re-suspended amount for
the whole region was 1.28 TBq, equivalent to 0.048 %of the aircraft-observed total deposited amount of
2.68 PBq. The total re-deposition was 0.18–0.23 TBq,
equivalent to 14.1–18.0 % of the total re-suspended
amount: the rest of the 137 Cs was transported out of themodel domain. The spatial distribution of the decreasingrate of land surface 137 Cs due to re-suspension ranged
from 2.7 × 10−7 –8.2 × 10−6 day−1 .

1. Optimization of the deposition parameters of the LM for
simulating the emergency situation of March 2011, using aircraft observation data (NRA, 2012) and the prescribed emission inventory (Katata et al., 2015), provided 0.1 (0.05–0.1) cm s−1 for a dry deposition velocity over land and 0.04 (0.03–0.05) for a hydrometeor
collection efficiency for aerosols. The optimized (or validated) ranges of the deposition parameters were applied to long-term re-suspension assessment for the year2013.
2. Using the dust emission module (Ishizuka et al., 2016),
which was developed based on physical parameters,
simulated 137 Cs re-suspension from soil multiplied by
5 accounted for the observed 137 Cs surface air concentration measured at Namie during only the cold season;
the module underestimated the 137 Cs concentration by
1 to 2 orders of magnitude in the warm season.
3. Introducing re-suspension from forest using a constant re-suspension coefficient of 10−7 h−1 and monthly
green area fraction could quantitatively account for theobserved concentration together with its seasonal variation.

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13165

7. The first-order decrease rate of the ambient gamma
dose rate in Fukushima prefecture ranged from 5.2 to12.1 × 10−4 day−1 . By subtracting the radioactive decay rate of 3.0–4.2 × 10−4 day−1 , the ground radioactivity decay due to land surface processes, decontamination, and re-suspension was found to range from 1.0
to 7.9 × 10−4 day−1 . The estimated re-suspension rate
was 2 to 3 orders of magnitude lower than the decrease in rate due to the other processes, showing thatre-suspension contributed negligibly towards reducing
ground radioactivity.
7

Data availability

All data are available upon request to the authors.

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13166

M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium

Appendix A: Model description
The current study employs a Lagrangian type model for thesimulation of emission (either point sources or areal sources),
horizontal and vertical diffusion and advection, gravitational
settling, dry and wet depositions, and radioactive decay inthe air. As described in Sect. 2.1, the current Lagrangianmodel (LM) uses simple parameterizations for dry and wet
deposition schemes for computational efficiency, so longterm assessment and parameter sweep experiments are easily feasible. The source code for the model is open with the
BSD 3-Clause License and is available on the web (https://ebcrpa.jamstec.go.jp/isetr_a01-1/, in Japanese).
The coordinate system of the model is horizontal for longitude and latitude and vertical for pressure level, consistent with meteorological analysis data commonly used. The
model can be driven only by fundamental meteorologicalparameters such as temperature, humidity, 3-D wind field,
geopotential height, and surface precipitation provided bymeteorological analysis data such as GPV-MSM. The model
does not need to drive meteorological models to predict detailed meteorological variables such as cloud microphysics,
turbulence quantities, and surface variables. Since the temporal and spatial resolution of the meteorological analysis is
not very high (e.g., 3 h and ∼ 11 km, respectively, for GPVMSM), linear interpolation is conducted in time and space.
Alternatively, higher temporal and spatial resolution can beachieved by using a meteorological model. Furthermore, although currently not implemented, detailed variables predicted by a meteorological model can be used for more accurate predictions of turbulent diffusion, surface flux, and dry
and wet deposition.
In the LM model, LPs are released constantly in time but
the initial activity of LPs (Bq LP−1 ) differs according to theemission flux (Bq h−1 ). The initial positions of LPs were randomly distributed within a fixed volume (or line) of plumecentered at a point emission source such as FDNPP for the
primary emission case simulation, or randomly distributed
within a horizontal model grid for the areal emission cases(such as re-suspension from soil and forest). LPs do not disappear unless transported across lateral and upper boundaries
or if they reach the surface layer due to gravitational settling (technically, gravitational settling velocity in the surface layer is included in the dry deposition velocity). The
other processes, such as dry deposition, wet deposition andradioactive decay, do not decrease the number of LPs but dodecrease the radioactivity carried by LPs because LPs represent an air mass rather than an actual particle, except in
the case of gravitational settling. The lowest level permitted for the position of LPs is set as 2 m AGL and LPs going
down across the level due to vertical turbulent motion willrebound at the level and go up. An LP whose radioactivity
is smaller than a preset value, i.e., 10−10 Bq, due to deposition or radioactive decay will disappear from the computation to maintain computational efficiency, since the cost of
Atmos. Chem. Phys., 16, 13149–13172, 2016

the computation is proportional to the number of LPs in themodel domain. To output the model results, the LP fields areconverted to Eulerian concentration (Bq m−3 ) and deposition(Bq m−2 ) fields on a prescribed coordinate system of grids.
In Lagrangian type models, the spatial resolution of tracer
emission, concentration, and deposition fields can be set independent of each other and with the spatial resolution of meteorological fields. In the current implementation of the LM,
the coordinate systems of meteorological fields and radioactivity fields are horizontally in common but vertically different: meteorological fields are provided on isobaric levels,
whereas radioactivity fields are calculated on height AGL.
The horizontal and vertical diffusion calculation followed JMA (2008), using the horizontal diffusion schemeof Uliasz (1990) with a constant horizontal diffusivity of
5.864 × 104 m2 s−1 and using a vertical diffusivity calculated
based on Louis et al. (1982; see Eqs. 8.1.8 through 8.1.15 of
JMA, 2008, for details).
The wet scavenging rate 3wet (s−1 ) is simply parameterized as a function of the surface precipitation rate P(mm s−1 ) as
3wet =

3 Ec (am , rm )
P,
4
am

(A1)

where Ec is the collection efficiency of aerosols by the hydrometeor, and am and rm are the mean radii of the hydrometeor and aerosols, respectively (JMA, 2008). Empirically, am
is characterized by P as
am = 0.35P 0.25 .

(A2)

JMA (2008) uses 0.05 for Ec . In the current study, instead of
explicitly predicting Ec , its range was set for the sensitivity
tests as listed in Table 2.
Conceptually, Eq. (A1) is the formulation for the washout
process, i.e., the collection of aerosols by the settling hydrometeor particles such as rain and snow. am and Ec should
differ for rain and snow, but common parameters are usedin the current simulation. Also, Eq. (A1) is not applicablefor the rainout process, since this process-cloud condensation nuclei or ice nuclei activation and deposition via subsequent cloud microphysical processes is totally different fromthe washout process. Because meteorological models were
not utilized in this study and thus only relative humidity andsurface precipitation rate are available and no cloud microphysical information (such as hydrometeors mixing ratio in
each model grid) is available, Eq. (A1) is applied for all theLPs located above the grid with P . In order to partly accountfor the rainout process, Eq. (A1) is not applied to LPs in agrid, where the relative humidity is lower than the minimum
value, set as 95 % in the simulation.
The dry deposition velocity vd (m s−1 ) of aerosols (orgases) is conventionally formulated, using an electrical analogy, as an inverse of the summation of resistances (s m−1 )representing turbulent diffusion in the surface layer, Brownian diffusion (or molecular diffusion for gases), interactionwww.atmos-chem-phys.net/16/13149/2016/

M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesiumwith the land surface (soil, water, and vegetation), and gravitational settling for aerosols (e.g., Wesely and Hicks, 2000).
Therefore, vd is a function of height as well as of turbulent
flux and surface conditions. Nevertheless, vd is set as constant in the simulation, but the height dependency of vd isconsidered in the dry scavenging rate 3dry (s−1 ), following
Furuno et al. (1999) as


z
2
1−
vd ,
(A3)
3dry =
zsrf
zsrf
where z is the height of the LP (m AGL) and zsrf is the surface
layer height set as 100 m AGL in the study. Instead of explicitly predicting vd , its range was set at around 0.1 cm s−1 , atypical speed for a range of aerosols around 1 µm in diameter,
for the sensitivity tests, as listed in Table 2. The value of vd
is applied over land, whereas vd over the ocean is multipliedby 0.1, because vd over a flat surface is approximately 1 and2 orders of magnitude smaller than vd over short vegetationsuch as grass and tall vegetation such as forest, respectively(e.g., Petroff and Zhang, 2010).

Because the statistical error of Lagrangian simulation is inversely proportional to the square of the number of LPs, thestatistical accuracy of the current simulation setting was evaluated using the following measures (relative errors of quantities of the sensitivity runs to those of the reference run):|Csens (x, y) − Cref (x, y)|,Cref (x, y)|Dsens (x, y) − Dref (x, y)|
Edep (x, y) =,Dref (x, y)

(B1)
(B2)

where x and y indicate grid points on the longitudinal and latitudinal axes, respectively. Csens and Dsens indicate temporalmean surface concentrations (Bq m−3 ) and temporal cumulative depositions (Bq m−2 ) of the sensitivity runs, respectively. Cref and Dref are the same as Csens and Dsens but for
the reference run. Econ and Edep were sampled only at grids
where Cref (x, y) and Dref (x, y) are greater than their areal
mean values, respectively.

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Point source case

The number emission rate of LPs, NLP , was set as 32 000 h−1(i.e., NLP_ref ) for a point source emission case such as the primary emission in March 2011 and additional emission fromthe reactor buildings in 2013. The median values together
with the 25th and 75th percentile values of Econ and Edepof the sensitivity runs (sensitivity to deposition parametersand sensitivity to NLP ) against the reference run are listedon the rows of the top half of Table B1. Both Econ and Edepof NLP_ref × 4 were significantly lower than those for the deposition parameters sensitivity run. This result indicates that
32 000 h−1 for NLP was sufficient to allow a statistically significant simulation for the purpose of this study, as shown in
Figs. 4 and 5: the difference in concentration and deposition
due to the deposition parameters was much larger than the
difference due to model uncertainty in NLP . Econ and Edep of
NLP_ref × 0.25 (which are also smaller than those of the deposition parameters sensitivity run) are larger than those of
NLP_ref × 4, indicating fairly good accuracy convergence of
the LM model.
B2

Appendix B: Statistical accuracy of the current
simulation setting

Econ (x, y) =

B1

13167

Areal emission case

NLP_ref was 16 h−1 grid−1 for the areal emission case simulating re-suspension from soil and forests in 2013. The lower
half of Table B1 is the same as the upper half except for
the areal emission case (re-suspension from forest). Both
Econ and Edep of NLP_ref × 4 were much lower than thosefor the deposition parameters sensitivity run, indicating that16 h−1 grid−1 supports a statistically significant simulation
for the purpose of this study, as shown in Figs. 7–12. Usually, Lagrangian-type models are not appropriate for solving
areal emission problems, because they require many more
LPs for areal emission cases compared to point source cases
and thus become computationally too expensive to obtain statistical accuracy (i.e., to set large number of LPs). In the caseof this simulation, especially for the concentration, sensitivity to deposition parameters was much more significant thansensitivity to model uncertainty in NLP using the sufficiently
small number of NLP_ref = 16 h−1 grid−1 . Econ and Edep of
NLP_ref × 4 are smaller than those of NLP_ref × 0.25, indicating fair accuracy convergence of the LM model.

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13168

M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium

Table B1. Statistical measures of temporal mean 137 Cs surface concentration (Econ ) and cumulative deposition (Edep ) of the sensitivity runs
against the reference run for (top) the point source case and (bottom) the areal emission case.
Number emission
rate of LP, NLP(h−1 grid−1 )

Eca
(–)

vdb
−1
(cm s )Median(25th–75th percentile)c (%)
of EconMedian(25th–75th percentile)d (%)
of Edep

32 000

0.04

0.01

–

–

32 000128 000
8000

0.03–0.05
0.04
0.04

0.05–0.01
0.01
0.01

2.7 (0.37–5.0)
0.51 (0.21–0.96)
0.95 (0.39–1.8)

7.5 (3.6–13)
0.72 (0.32–1.5)
1.6 (0.73–2.9)

16

0.04

0.10

–

–

16
64
4

0.03–0.05
0.04
0.04

0.05–0.01
0.10
0.10

7.3 (2.6–13)
0.39 (0.17–0.78)
0.78 (0.34–1.5)

7.7 (3.6–17)
2.0 (1.3–2.6)
2.3 (1.5–3.1)

Point source case, March 2011
Reference run
Sensitivity runs
Deposition parameters
NLP_ref × 4
NLP_ref × 0.25
Areal emission case, 2013
(re-suspension from forest)
Reference run
Sensitivity runs
Deposition parameters
NLP_ref × 4
NLP_ref × 0.25

a Collection efficiency; see Eq. (A1). b Dry deposition velocity over land; see Eq. (A3). c Relative errors of temporal mean surface concentration at each grid cell of thesensitivity run to that of the reference run; see Eq. (B1). d Same as Econ but for cumulative deposition; see Eq. (B2).

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M. Kajino et al.: Long-term assessment of re-suspended airborne radiocesium
Acknowledgements. This research was mainly supported by
Grants-in-Aid for Scientific Research on Innovative Areas(JP24110002 and JP20110003) from the Ministry of Education,
Culture, Sports, Science and Technology (MEXT), and partly
supported by the Japan Society for the Promotion of Science(JSPS), Ministère des Affaires Étrangères et du Développement
International (MAEDI) under the Japan-France Integrated Action
Program (SAKURA), as well as by KAKENHI grant numbers
JP15K16121, JP24340115, and JP26310201 from MEXT. Partial
expenses were covered by the Japanese Radioactivity Survey fromthe NRA, Japan. The authors are grateful for useful discussions
with and comments from Yuichi Moriguchi of University of Tokyo,
Hiromi Yamazawa of Nagoya University, and Haruo Tsuruta of
Remote Sensing Technology Center of Japan (RESTEC). The
authors are thankful to Masayuki Takigawa of JAMSTEC for the
maintenance of the website of LM. The LM was developed by
M. Inatsu, C. Yoshikawa, M. Takigawa, and M. Kajino.
Edited by: Y. Balkanski
Reviewed by: three anonymous referees

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Journal
Decemberof the2017
Meteorological Society of Japan, Vol.
T. T.95,
SEKIYAMA
No. 6, pp. 447−454,
et al.
2017
DOI:10.2151/jmsj.2017-025

447

NOTES AND CORRESPONDENCE
The Impact of Surface Wind Data Assimilation on the Predictabilityof Near-Surface Plume Advection in the Case of the Fukushima Nuclear Accident
Tsuyoshi Thomas SEKIYAMA, Mizuo KAJINO, and Masaru KUNII
Meteorological Research Institute, Tsukuba, Japan(Manuscript received 24 March 2017, in final form 4 August 2017)

Abstract
We investigated the predictability of plume advection in the lower troposphere and the impact of AMeDAS
surface wind data assimilation by using radioactive cesium emitted by the Fukushima nuclear accident in March2011 as an atmospheric tracer. We conducted two experiments of radioactive plume predictions over easternJapan for March 15, 2011 with a 3-km horizontal resolution using the Japan Meteorological Agency non-hydrostatic weather forecast model and local ensemble transform Kalman filter (JMANHM-LETKF) data assimilation
system. The assimilated meteorological data were obtained from the standard archives collected for the Japan
Meteorological Agency operational numerical weather prediction and the AMeDAS surface wind observations.
The standard archives do not contain land-surface wind observations. The modeled radioactive cesium concentrations were examined for plume arrival times at 40 observatories. The mean error of the plume arrival times
for the standard experiment (assimilating only the standard archives) was 82.0 min with a 13-h lead-time on anaverage. In contrast, the mean error of the AMeDAS experiment (assimilating both the standard archives andAMeDAS surface wind observations) was 72.8 min, which was 9.2 min (11 %) better than that of the standard
experiment. This result indicates that the plume prediction has a reasonable accuracy for the environmental emergency response and the prediction can be significantly improved by the surface wind data assimilation.
Keywords surface wind data assimilation; plume advection predictability; Fukushima nuclear accident; environmental emergency response1.

Introduction

The advection of minor constituents in the troposphere (e.g., water vapor, oxidants, and aerosols) is
one of the key processes for numerical weather orenvironment prediction. Advection, emission, deposition, and chemical/physical changes define constituent
Corresponding author: Tsuyoshi Thomas Sekiyama,
Atmospheric Environment and Applied Meteorology
Research Department, Meteorological Research Institute,
1-1 Nagamine, Tsukuba, Ibaraki 305-0052, Japan
E-mail: tsekiyam@mri-jma.go.jp
J-stage Advance Published Date: 28 August 2017©2017, Meteorological Society of Japan

distributions. For weather prediction, the distributionof water vapor influences the coverage and strengthof precipitation. Moreover, the distributions of oxidants, aerosols, and their precursors directly impact
human health and indirectly modify the weather andclimate. However, it is difficult to simulate advectionin the lower troposphere due to poor reproducibility ofnear-surface wind velocities in numerical weather simulations. Numerical weather forecast models cannot
explicitly resolve fine-scale structures of real surface
wind velocities due to the topographical heterogeneity
and the instability (or nonlinearity) of the atmosphericboundary layer (ABL). These sub-grid scale featurescause discrepancy between the measured and forecast-448

Journal of the Meteorological Society of Japaned variables. Therefore, the observational datasets ofsurface wind velocities are usually very limited or are
avoided for operational weather prediction (e.g., Japan
Meteorological Agency (JMA) 2015, in Japanese)to prevent the degradation of initial conditions. To
improve near-surface wind prediction, recent studies
tried to assimilate surface wind observations over land(e.g., Hacker and Snyder 2005; Benjamin et al. 2010;
Hacker and Rostkier-Edelstein 2007; Hacker et al.
2007; Rostkier-Edelstein and Hacker 2010; Ancell
et al. 2011, 2015; Ingleby 2015; Bédard et al. 2015,2017). In these studies, however, the surface wind observations have an influence only on extremely shortterm and local forecasts (less than 6 h) even if they
have a positive impact. Besides, these past studies
have not particularly investigated the predictability of
constituent advection.
In this context, validating constituent advection
predictions and understanding the impact of surface
wind data assimilation have garnered increasing attention. Especially, the predictability of near-surface
plume advection is a crucial factor for the environmental emergency response (EER), in which we haveto prepare for the dispersion of hazardous materials
emitted from a point source on the ground to the ABL.
If the plume prediction has a reasonable accuracy likeweather forecast, the area in danger will have to be
evacuated in advance (World Meteorological Organization 2006). As mentioned above, the surface winddata assimilation is often very limited or avoided forweather analyses and predictions, although nudging
of surface wind observations has been performed forlower tropospheric advection simulations (e.g., theWRF-CMAQ ozone simulation conducted by Li et al.
2016). However, the impact of surface wind data
assimilation has not been well investigated for theseadvection simulations, especially for predictions incorporating high-performance data assimilation.
The investigation of plume advection requires
widespread tracer observations with a fine time resolution. Therefore, we used radioactive cesium emitted
by the Fukushima nuclear accident in March 2011 as
the atmospheric tracer. The radioactive cesium plumewas dispersed from a single source (the Fukushima
Daiichi Nuclear Power Plant). The surface concentration of the plume was measured at many locations
at a high frequency and with high accuracy (Tsuruta
et al. 2014; Oura et al. 2015). This point source plumeis advantageous due to its simple tracer isolation andvalidation without cross-dispersion. This paper shows
the accuracy of the 137Cs plume advection prediction
and the impact of surface wind data assimilation on

Vol. 95, No. 6

the predictability of the 137Cs plume advection in thefollowing sections.
2.

Methodology

In general, atmospheric advection models requirea gridded point value (GPV) dataset that contains
a meteorological analysis or prediction that is precalculated via a high-performance data assimilation
system. For example, the JMA operational meso-GPV
dataset was used by all regional models that participated in the multimodel intercomparison of Fuku­
shima nuclear pollution predictions (Science Councilof Japan (SCJ) 2014) except for two models. One
was our model (Sekiyama et al. 2015), which used
an original meteorological dataset. The other model
used the European Centre for Medium-Range Weather
Forecasts (ECMWF) operational global GPV dataset.
All models except Sekiyama et al. (2015) utilized
the GPV datasets that were prepared by others (i.e.,
JMA or ECMWF). However, for the purpose of ourinvestigation, the GPV datasets have to be preparedby our own data assimilation system because we needto arrange a comparison of meteorological analyses
performed in the presence or absence of surface winddata assimilation. Hence, we utilized the meteorological data assimilation system of Sekiyama et al. (2015)
in this study to prepare the GPV datasets with andwithout land-surface wind data assimilation.
The meteorological data assimilation system of
Sekiyama et al. (2015) was developed by Kunii (2014)
and is composed of the JMA’s non-hydrostatic regional weather prediction model (JMANHM, cf., Saito
et al. 2006, 2007) and the local ensemble transform
Kalman filter (LETKF, cf., Miyoshi and Aranami2006). In this study, this system was driven by 20
ensemble members and a 3-km horizontal resolution
within the model domain of eastern Japan (215 × 259
grids, cf., Fig. 2b of Sekiyama et al. 2015). The covariance localization parameters were set to 50 km in thehorizontal, 0.1 natural-logarithm-pressure coordinate
in the vertical, and 3 hours in the time dimension for
all observations. The boundary conditions for themodel domain were provided by the JMA operational
global analysis and prediction. Incidentally, JMANHM
has been used to produce the JMA operational mesoGPV dataset at a 5-km horizontal resolution since
before March 2011 with a four-dimensional variational (4D-Var) data assimilation method. This weather
forecast system is known as JNoVA (Honda et al.
2005).
Using the abovementioned JMANHM-LETKF
data assimilation system, two types of meteorologicalDecember 2017

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T. T. SEKIYAMA et al.

initial conditions were prepared from March 11 to 31,
2011 at 3-h intervals. One was the standard analysis(STD), which was produced by assimilating only
the observations archived for the JMA operational
analysis of JNoVA. These observations were collectedby land-surface observatories (pressure measurementsonly), satellites (including sea surface wind), radio­
sondes, pilot balloons, wind profilers, aircrafts, ships,
and buoys. The satellite radiances and radar precipitation analyses were excluded from this study. This
JNoVA dataset does not contain land-surface windvelocity observations.
The other analysis was produced by assimilating
the JNoVA dataset and the land-surface wind-velocity
observations collected by AMeDAS. AMeDAS is
the acronym of the automated meteorological dataacquisition system managed by JMA, which is a landsurface observation network that comprises approximately 1,300 stations throughout Japan with an average interval of 17 km. In this study, we assimilated
the land-surface wind velocities from more than 200
AMeDAS stations within the model domain (shown in

Fig. 1. Locations of the AMeDAS surface windobserving stations used for the AMeDAS experiment from 00:00 UTC on March 11 to
12:00 UTC on March 14, 2011. The letter “P”
indicates the location of the Fukushima Daiichi
Nuclear Power Plant.

Fig. 1).
For the surface wind data assimilation of theAMeDAS experiment, the height of the surface windvelocity was intended to be fixed at 10 m, because
almost every AMeDAS anemometers are installed at aheight of 10 m according to the World MeteorologicalOrganization guideline. All observed velocities Uobs
were transformed into U10 by using the following
formula:

( )
( )

ln 10
z0
U10 =
U ,
zobs obslnz0
where zobs is the height of the anemometer installationand z0 is a constant roughness length (1 m). However,
the transformation was negligible in the experiment.
The observation errors of the AMeDAS surface windvelocity were set to 2 m s−1, which is a comparable
level to the root mean square errors (RMSEs) of the
STD experiment shown in Fig. 2. The RMSEs andcorrelation coefficients (r) in Fig. 2 were calculated by
1-h forecasts of the STD and AMeDAS experimentsthat were performed every 3 h from 00:00 UTC on
March 11 to 12:00 UTC on March 15, 2011 at theAMeDAS stations used in this study.
Using these meteorological analyses as the initialconditions, two forecast runs (STD and AMeDAS experiments) were performed with the same JMANHM
configuration for 48 h. The boundary conditions were
provided by the JMA operational global prediction
with the same initial time as JMANHM. The forecast
initial time was set to 12:00 UTC on March 14 toreproduce the radioactive cesium plume behavior
over the inland area of eastern Japan. According to
Nakajima et al. (2017), the radioactive plumes moved
primarily over the ocean during the Fukushima nuclear accident in March 2011. However, on March 15 – 16
and 20 – 21, the plumes moved deeper inland. However, because there was a widespread precipitation
over the Tohoku and Kanto regions on March 20 – 21,
we investigated only the case of March 15 – 16 in thisstudy. When the precipitation is strong and widespread, the influence of wet deposition on the plumeconcentration becomes extremely large. The model
simulations of precipitation strength and wet deposition are complicated and challenging. Therefore, weplan to investigate the March 20 – 21 event in future.
Plume advection simulations were then driven
by the 48-h meteorological forecast runs by using
the Eulerian regional air quality model version 2(RAQM2). RAQM2 was developed by Kajino et al.

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Vol. 95, No. 6

Fig. 2. Scatter diagrams with RMSEs and correlation coefficients (r) between the AMeDAS surface wind observations used in this study and 1-h forecasts (U10 and V10: zonal and meridional winds at a height of 10 m) of
the STD/AMeDAS experiments. The 1-h forecasts were performed every 3 h from 00:00 UTC on March 11 to
12:00 UTC on March 15, 2011.

(2012) and was used for the Fukushima nuclear accident simulation (Adachi et al. 2013; Sekiyama et al.
2015). RAQM2 and JMANHM share the same model
domain (eastern Japan) and horizontal resolution (3km), although the vertical resolution was converted
from JMANHM’s 60 layers (from the surface to 22
km asl) to RAQM2’s 20 layers (from the surface to 10
km asl). All radioactive cesium (137Cs) was expected
to be contained in sulfate aerosol particles mixed
with organic compounds as shown by Sekiyama et al.
(2015).

The emission of radioactive cesium was fixed at
a constant value (1 Bq h−1), because we intended tovalidate only the plume arrival time at each station torigorously examine the plume advection predictability.
The Fukushima radioactive cesium plume was emitted
from the point source; the background concentration
was nearly zero before March 2011. The modeled concentration contrast between the background and theedge of arriving plume was larger than ten orders ofmagnitude. Therefore, we were easily able to identify
the plume arrival time in simulations. We tried many

December 2017

T. T. SEKIYAMA et al.

settings of the Fukushima nuclear pollutant simulation(e.g., Adachi et al. 2013; SCJ 2014; Sekiyama et al.
2015) by using realistic emission rate datasets (e.g.,
Chino et al. 2011) and confirmed that we need only
relative concentration values if we want to define
the edge of radioactive plumes. This is because thebackground concentration of anthropogenic nuclear
products is zero in simulation models; thus, theplume edge has a jump of values by more than ten (orsometimes twenty) orders of magnitude. Furthermore,
wet deposition in the model was turned off to avoid aplume disappearance caused by the erroneous precipitation in the model.
The modeled plume arrival times were validated bycomparing them with the hourly averaged radioactive
cesium concentrations measured by Tsuruta et al.
(2014) and Oura et al. (2015). Tsuruta et al. (2014)
developed a method to retrieve the hourly averaged
concentrations of radioactive cesium in the lower
atmosphere using suspended particulate matter (SPM)
sampling tapes with a detection limit of less than 0.6
Bq m−3 during the Fukushima nuclear accident. The
SPM tapes were collected from the air pollution monitoring network managed by the national government
and maintained by prefectural governments. Tsuruta
et al. (2014) and Oura et al. (2015) reported theconcentration data at 99 SPM-tape sampling stations.
However, we screened these stations to clearly detect
the radioactive plume arrival, which is discussed
in the next section. Consequently, the data from 40
stations of the 99 stations were used in this study, as
shown in Fig. 3.
3.

Results and discussion

The plume arrival times at the SPM-tape samplingstations were determined with a threshold of 1 Bq m−3
for the observations and 1 × 10−15 Bq m−3 for the modelexperiments. In general, when the plume arrived, theconcentration rose sharply because the backgroundwas nearly zero (e.g., Fig. 4a). The observed concentrations changed rapidly from less than the detection
limit (approximately 0.6 Bq m−3) to more than 1 Bq
m−3. The modeled concentrations increased by morethan 10 orders of magnitude. Note that the highest
modeled concentrations at the SPM tape sampling stations were on the order of 10−13–10−11 Bq m−3. Hence,
the threshold for the modeled plumes was set to 1 ×
10−15 Bq m−3. For example, the observed plume arrival
time and the STD forecasted plume arrival time areshown in Fig. 4a at Station #49 in Kuki City, Saitama
Prefecture, based on Oura et al. (2015).
When a concentration surge was not clearly ob-

451

Fig. 3. Closed squares indicate the SPM-tape“radioactive cesium observing” stations of Ouraet al. (2015) used in this study. Open squares
indicate the SPM-tape stations shown in Oura etal. (2015) but not used in this study. The letter “P”
indicates the location of the Fukushima Daiichi
Nuclear Power Plant.

served, e.g., as in Fig. 4b (Station #59 in Chiba City),
arrival time data were not used. In addition, more than
one increase in concentration over a short period oftime was also observed in a few cases, e.g., as in Fig.
4c (Station #81 in Ota Ward, Tokyo). In this case,the edge of the observed plume could not be clearly
identified and compared with modeled plumes. Thus,
we omitted these data from the statistical calculations.
After the screening, 40 out of the 99 stations of Ouraet al. (2015) remained, as shown in Fig. 3. Fortunately,
most of the 40 stations were located inland; therefore,
it was expected that we could distinctly observe theinfluence of land-surface wind observations on theplume simulation.
The STD experiment did not assimilate the landsurface wind observations; namely, it assimilated theobservation dataset regularly used in the operational
JNoVA system, except for the satellite radiance andradar precipitation data. We have confirmed that thedifference between the STD experiment analysis and452

Journal of the Meteorological Society of Japan

Vol. 95, No. 6

Table 1. Mean plume arrival time errors at the 40 SPMtape sampling stations.

mean error (min)standard deviation (min)significance level (p-value)

Fig. 4. Examples of time series of the radioactive
cesium (137Cs) concentrations derived from the
SPM-tape sampling observations and two modelexperiments (STD and AMeDAS) from 12:00
UTC on March 14 to 12:00 UTC on March 15,
2011. The advection of the plumes was examined using the data from (a) Kuki City, Sai­tama
Prefecture, and not using the data from (b)Chiba City, Chiba Prefecture, or (c) Ota Ward,
Tokyo.

AMeDAS
experiment

82.0
83.4

72.8
79.6
0.008

the realistic ability of state-of-the-art operational
weather forecast models and data assimilation systems
to predict plume advection for emergency evacuations.
Conversely, the mean error of the AMeDAS experiment was 72.8 min, which was 9.2 min (11 %) smaller
than that of the STD experiment, with a statisticalsignificance level of p-value of 0.008 (Table 1).
This result indicates that the land-surface wind data
assimilation significantly improved the predictabilityof near-surface plume advection even after a halfday forecast. Although the AMeDAS surface windobservations had only a small positive impact on thesurface wind reproducibility as shown in Fig. 2, theimpact on the plume predictability was large with ahigh statistical significance. The accuracy of a halfday plume prediction depends on not only the accuracy of 12-h-forecasted wind velocities in the vicinityof the plume but also very short-term wind-velocity
forecasts near the emission source. Therefore, there ispossibility that the improvement of the plume arrival
time predictability is caused by only the improvementof very short-term forecasts near the emission source.
However, at least from the viewpoint of advection
prediction, the improvement of the plume predictability was surely maintained for longer than 6 h by the
AMeDAS surface wind data assimilation. Besides, the72.8-min error will be acceptable for the EER evacuation if the prediction is available with a half-day lead
time.
4.

the operational JNoVA analysis is small. The influenceof the satellite radiance and radar precipitation data
was negligible for surface wind predictability in thisstudy. The averaged difference (mean error) in theplume arrival times at the 40 SPM-tape samplingstations between the observational data and the STD
experiment was 82.0 min (Table 1). Here, the average
forecast length was 13 h. As indicated by the standard
deviation (83.4 min) compared to this average, theforecasted plume arrival time was often very close to
the observations. However, there were a few instancesof errors exceeding 3 h. This is thought to represent

STD
experimentSummary

The assimilation of AMeDAS surface wind data
has a positive impact on the predictability of plumeadvection in the lower troposphere, at least in cases of
wintertime air pollution over complex terrain, e.g., theFukushima nuclear accident. The plume arrival predic­
tion has a 72.8-min error with a half-day lead timefor Tohoku and Kanto regions for March 15, 2011 byusing the AMeDAS surface wind data assimilation. If
the plume arrival prediction was obtained for the EER
with an accuracy of the AMeDAS experiment in thisstudy, the information could be used for the evacuation or sheltering. Furthermore, a similar strategy forDecember 2017

T. T. SEKIYAMA et al.

improving advection predictability would be applicable to near-surface water vapor, oxidant, and aerosol
predictions by using surface wind data assimilation.
Acknowledgments
We thank Dr. Tsuruta and his colleagues for providing us with their SPM-tape radioactive cesium
data. The JMA headquarters provided the weather
analysis, prediction, JNoVA, and AMeDAS datasets
used in this study. This work was supported by JSPS
KAKENHI Grant Number JP24340115, JP16H02946,
JP16H04051, JP26800247, and JP17K00533; MEXT
KAKENHI Grant Number JP24110003; and the JapanFrance Integrated Action Program (SAKURA) of
JSPS/MAEDI.
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Tellus B: Chemical and Physical Meteorology

ISSN: (Print) 1600-0889 (Online) Journal homepage: http://www.tandfonline.com/loi/zelb20

137

Mass flux analysis of Cs plumes emitted fromthe Fukushima Daiichi nuclear power plant
Tsuyoshi Thomas Sekiyama & Toshiki Iwasaki
To cite this article: Tsuyoshi Thomas Sekiyama & Toshiki Iwasaki (2018) Mass flux analysis of137

Cs plumes emitted from the Fukushima Daiichi nuclear power plant, Tellus B: Chemical andPhysical Meteorology, 70:1, 1-11, DOI: 10.1080/16000889.2018.1507390
To link to this article: https://doi.org/10.1080/16000889.2018.1507390© 2018 The Author(s). Published by Informa
UK Limited, trading as Taylor & Francis
Group.
Published online: 08 Oct 2018.

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TellusSERIES B
CHEMICAL
AND PHYSICALMETEOROLOGYPUBLISHED BY THE INTERNATIONAL METEOROLOGICAL INSTITUTE IN STOCKHOLM

Mass flux analysis of 137Cs plumes emitted from the
Fukushima Daiichi nuclear power plant
By TSUYOSHI THOMAS SEKIYAMA1, and TOSHIKI IWASAKI2
1

Meteorological Research Institute, Tsukuba, Japan; 2Graduate School of Science, Tohoku University,
Sendai, Japan(Manuscript received 6 March 2018; in final form 24 July 2018)

ABSTRACTThe flow vectors of radioactive cesium-137 (137Cs) plume emitted from the Fukushima Daiichi nuclear power
plant in March 2011 were quantitatively depicted by a mass flux analysis in this study. 137Cs plumes were
calculated by an Eulerian dispersion model with a 3-km horizontal resolution. The vertically columnintegrated mass flux was consistent with the flow approximation based on ground surface 137Cs observations,even though there were some discrepancies that were caused by differences in the wind direction between the
ground surface and the dominant plume layer. These discrepancies were explained by combining the use of
the ground surface horizontal mass flux with the column-integrated mass flux. The mass flux analysis clearlyprovided an illustration of 137Cs dominant stream locations, directions, and depositions by reducing highdimensional model outputs into a lower-dimensional plot. Mass flux (i.e. the product of the mass density andwind velocity) has often been used in dynamic meteorology but has not been used as frequently inatmospheric chemistry or pollutant dispersion studies. However, the concept of mass flux is a robust
alternative for conventional validation approaches that only utilize a time series of pollutant concentrations.
Mass flux analyses can be used further in atmospheric chemistry as a quantitative visualization tool to trackthe emission, advection, dispersion, and deposition of atmospheric constituents.
KEYWORDS: plume dispersion, numerical simulation, mass flux analysis, radioactive cesium-137, Fukushima
nuclear accident

1. Introduction
The dispersion modeling of atmospheric constituents is essential for numerical simulation (or prediction) of precipitation,
climate change, and air pollution. To improve dispersion
modeling, it is crucial to quantitatively depict the modeleddispersion of pollutants and validate those results with observations. An ideal validation could be accomplished by using
atmospheric constituents that are (1) chemically inert to avoid
errors in the reaction rate estimation, (2) emitted from a single
source to isolate plumes without cross-dispersion, and (3)observed at many locations with a high accuracy. The radioactive cesium-137 (137Cs) from the Fukushima nuclear accident considerably matches these criteria. The Fukushima
nuclear accident was triggered by the 2011 Tohoku earthquake and tsunami and was the largest nuclear disaster since
Chernobyl, which resulted in the extensive dispersion of a
Corresponding author. email: tsekiyam@mri-jma.go.jp

large amount of radionuclides in Japan (e.g. 131I, 133Xe,
134
Cs, and 137Cs). Among these radionuclides, 137Cs has arelatively long half-life (30 years) and characteristics such as
chemical inertness and high detectability. It was confirmedthat 137Cs was emitted from a single location: the Fukushima
Daiichi nuclear power plant (FDNPP). The ground surface
concentration of 137Cs has been retrieved at many locationsat an hourly frequency with a high accuracy (e.g. Tsuruta
et al. 2014; Oura et al. 2015).
After the accident, many numerical simulations were
performed using atmospheric dispersion models or chemistry transport models (e.g. Science Council of Japan, 2014;
Draxler et al. 2015). However, most of the simulations
were validated by comparing the results with total groundsurface 137Cs deposition or surface dose rates. Depositionand dose rates are strongly affected by precipitation anddeposition processes. However, the time variability of pollutant concentrations is rarely detectable in these types of

Tellus B: 2018. # 2018 The Author(s). Published by Informa UK Limited, trading as Taylor & Francis Group.
This is an Open Access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/4.0/),
which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.
Citation: Tellus B: 2018, 70, 1507390, https://doi.org/10.1080/16000889.2018.1507390

12

T. T. SEKIYAMA AND T. IWASAKI

Fig. 1. Fukushima 137C plumes identified by Tsuruta et al. (2014). Horizontal bars show periods with 137Cs concentrations greaterthan 10 Bq m3. Closed (open) circles indicate areas where the highest concentrations were larger (smaller) than 100 Bq m3. This figure
was reprinted from fig. 3 in Nakajima et al. (2017).

observations. Therefore, in the majority of previous modelstudies, deposition model errors and dispersion model
errors were barely distinguished between. In contrast,
Nakajima et al. (2017) isolated the individual 137Cs plumes
simulated by their chemistry transport models comparing
with hourly concentration observations from Tsuruta et al.
(2014) and Oura et al. (2015) and successfully examined thedispersion model performance separately from the deposition model performance. It should be noted that Tsuruta
et al. (2014) and Nakajima et al. (2017) classified 137Cs
propagations over land in March 2011 into nine plumes
based on their time, location, and direction, as shown in
Figs. 1 and 2. These plume numbers (P1–P9) are used without any change in this study.
Nakajima et al. (2017) clearly identified these nine
plumes, as shown in Fig. 2, and exhibited their model performance with hourly 137Cs concentrations observed at 99
sample stations. However, we realized that several of theplume directions in Fig. 2 were inconsistent with the otherforward trajectory model simulations. For example,
Nakajima et al. (2017) claimed that plume P3 was transported toward the Nakadori region and crossed the
Abukuma Mountains from south to north in the afternoonof 15 March 2011, Japanese Standard Time (JST); then, apart of plume P3 exhibited a northern detour from northto south in the late afternoon, as shown in Fig. 2a (geographical names are shown in Fig. 3). On the other hand,
Kajino et al. (2016) calculated forward trajectories beginning at 12:00 JST, 15:00 JST, 18:00 JST, and 21:00 JST on
15 March 2011, from the FDNPP, which emitted the pollutant at 25-m height (Fig. 4). The forward trajectories
described in Kajino et al. (2016) indicated that plume P3
rotated invariably clockwise and moved from south tonorth over the Abukuma Mountains and the Nakadori
region in the late afternoon on 15 March 2011, JST. Thereis a discrepancy between observation-based analyses andforward trajectory analyses.

This discrepancy might be caused by dispersion model
errors, meteorological analysis errors, or the vertical wind
shear in the presence of the height difference between observations and the dominant stream of pollutant flows. In addition, neither the pollutant plumes in Fig. 2 nor thepollutant trajectories in Fig. 4 are quantitatively depicted
regarding pollutant mass flows. We cannot determine whichflow depiction is more plausible using only these simple figures. First, it is desirable that plumes of atmospheric pollutants should be identified and quantitatively determinedwhen considering pollutant mass flows using dispersion
models. This quantitative plume identification allows us toobjectively validate the modeled dispersion performanceand distinguish between dispersion and deposition errors.
In this study, we quantitatively depicted all of theplume streams (P1–P9) from Tsuruta et al. (2014) andNakajima et al. (2017) using our 137Cs dispersion modeland attempted to determine the reasons for the abovementioned discrepancy. In this context, we took advantage of a mass flux analysis to quantify the pollutant
flows. Mass flux is often used in fluid dynamics ordynamic meteorology (e.g. Iwasaki et al. 2014; Yano
2014); however, it has not been used as much in atmospheric chemistry or dispersion modeling, except formeridional transport analyses (e.g. Stohl et al. 2003;
Belikov et al. 2013) or regional pollutant outflows (e.g.
Bey et al. 2001). The concept of mass flux could be arobust alternative to traditional concepts using time-variable concentrations to assess horizontal pollutant flows.
We attempt to exhibit the mass flux analysis as a modelvisualization method or a validation metric through the
depiction of Fukushima 137Cs plumes. In this analysis, wedefined not only the single-layer mass flux but also thecolumn-integrated mass flux to track pollutant massflows that are vertically distributed. Details for the massflux equations are described in Section 2, followed by amodel description in Section 3. Then, a comparison of

MASS FLUX ANALYSIS OF 137CS PLUMES

3

Fig. 2. Schematic diagrams of the transport routes analyzed in Nakajima et al. (2017) for plumes P1–P9 isolated in Tsuruta et al.
(2014). Black arrows indicate the general movement trend for each plume. This figure is reprinted from fig. 15 in Nakajima et al. (2017).
The letters, B, C, E, H, J, K1, K2, K3, 9, 12, and 15, indicate 137Cs sample station locations identified in Nakajima et al. (2017) but
these indications are not used in this study.

Fukushima 137Cs plumes between our analysis results andthose of Nakajima et al. (2017) is shown in Section 4, followed by concluding remarks in Section 5.

2. Mass flux and continuity equation
The mass flux j is defined by the product of the mass concentration q and the wind velocity v  ðu; v; wÞ. j is represented as a mass per unit area per unit of time (e.g. kgm2 s1) or as the rate of mass flow per unit area (e.g. [kgs1] m2), which perfectly corresponds with the momentum
density or momentum per unit volume (e.g. [kg m s1]m3). The mass flux indicates the extent that a pollutant isflowing over one’s head or how much a pollutant collides
into one’s body per unit of time. Even if the mass concentration is known, the amount of total exposure cannot be
estimated without the mass flux information. Furthermore,mass flux is a vector not a scalar (e.g. concentration); therefore, we can recognize the streams and directions of pollutant transport using the mass flux distribution.
In the atmosphere, the mass concentration and massflux are governed by the continuity equation, which conserves the total mass of a pollutant as shown belowoqoq
oq oqu oqv oqwþrj¼þ r  qv ¼
þ
þ
þ
ot
ot
ot
ox
oy
oz
oq
oqw
þ r  qvh þ
¼ r;
¼
ot
oz

where t represents time, vh  ðu; vÞ represents the horizontal wind, and r represents the generation of the pollutant per unit of volume and time. When r > 0, r isreferred to as a source term. Conversely, when r < 0, r isreferred to as a sink term. To achieve a vertically comprehensive assessment of mass flux in the atmosphere, we
integrate Eq. (1) over a column from the surface to the
top of the atmosphere;
1

ð
oq
oqw
þ r  qvh þ
dz
ot
oz
0

1
ð

¼
0

¼

1
1
ð
ð
oq
oqw
dz þ r  qvh dz þ
dz
ot
oz
0

0

1
ð

1
ð

0

0

o
qdz þ r  qvh dz
otðx; y; z; and t are mutually independentÞ
þ ½qwz¼1
z¼0
1
ð


o
qdz þ r  jcolumn ∵qð1Þ ¼ wð0Þ ¼ 0
¼
ot
0

1
ð

¼ rdz ¼ rs ;

(2)

0

(1)

where rs represents deposition or emission at the surface,assuming that the pollutant is inert and is added/removed

4

T. T. SEKIYAMA AND T. IWASAKI

Fig. 3. Map of the area around the FDNPP. Closed diamonds
indicate the 137Cs sample stations in Tsuruta et al. (2014) andOura et al. (2015).

only at the surface. Here, we introduce the column inteÐ1gral of the horizontal mass flux, 0 qvh dz  jcolumn , whichindicates the total mass flow per unit width integrated
vertically from the surface to the top of the atmosphere(e.g. [kg s1] m1 ¼ kg m1 s1). Conveniently, the timeintegral of jcolumn illustrates the dominant stream of thepollutant two-dimensionally. Because the column integral
Ð1z¼1
of the vertical mass flux 0 oqw
oz dz ¼ ½qwz¼0 ¼ 0 is negligible, the dominant stream can be calculated using only
horizontal wind components and neglecting the verticalwind component, which is rarely obtained with highaccuracy. This illustration is helpful when tracking pollutant plumes with a vertical structure.
Next, we integrate Eq. (2) over time for a typicalplume situation, where the mass concentration is zero
before a plume arrives and after it leaves. In thissituation,
2
3t¼1
1
1
ð 1
ð
ð


o
7
qdzdt ¼ 6
¼ 0 ∵qðt ¼ 0Þ  qðt ¼ 1Þ  0 ;4 qdz5
ot
0

0

0

t¼0

therefore, the integration of Eq. (2) over time is represented as
1
ð

1
ð

r  jcolumn dt ¼ rs dt;
0

(3)

0

which indicates that the time-integral for the divergence ofthe column integral in the horizontal mass flux equals thetotal emission or deposition per unit area (e.g. kg m2).

Fig. 4. Forward trajectories that began at 12:00 JST (dashed
two-dot line), 15:00 JST (dashed-dotted line), 18:00 JST (dashedline), and 21:00 JST (solid line) on 15 March 2011 from the
FDNPP at a height of 25 m (calculated by Kajino et al. 2016).

Actually, negative divergence (¼ convergence) and deposition are often very similar in 137Cs plume model simulations. However, they are not exactly consistent becausemodel simulations have numerical errors (mainly due totruncation processes) and assumption errors (e.g. non-zerobackground concentrations and finite limits of the model
top height), as shown in Fig. 5. This figure was compiledusing the 137Cs dispersion model results detailed in the next
section. The time-integrated period (21:00 JST on 20
March to 21:00 JST on 22 March 2011) was chosen toobserve a noticeable example of wet deposition induced bywidespread precipitation.
In addition, we can calculate a vertically mass-weighted
mean wind, vweighted , by dividing the column integral of
the horizontal mass flux by the total column mass
jcolumnvweighted ¼
Total column mass

1
ð

1
ð

qvh dz=
0

qdz;

(4)

0

which describes the wind velocity that forces the columnintegrated mass from the ground surface to the top of
the atmosphere.

MASS FLUX ANALYSIS OF 137CS PLUMES

5

Fig. 5. Convergence (¼ negative divergence) of the column-integrated 137Cs mass flux and the total deposition of 137Cs values
temporally integrated from 21:00 JST on March 20 to 21:00 JST on 22 March 2011 for the model simulation used in this study.

3. Model description
In this study, the simulation of the Fukushima 137Cs plumedispersion was performed by an offline Eulerian regional air
quality model, which was originally developed by Kajino
et al. (2012) for non-radioactive aerosol simulations. Thisdispersion model has been used for Fukushima nuclear pollutant simulation by Adachi et al. (2013), the Science
Council of Japan (2014), Sekiyama et al. (2015, 2017), andKajino et al. (2016). Radioactive 137Cs nuclides wereassumed to be contained in sulfate aerosol particles thatwere mixed with organic compounds, as described in detailby Sekiyama et al. (2015). We applied the 137Cs emission
scenario estimated by the Japan Atomic Energy Agency(Katata et al. 2015) to this simulation. Based on the time series of this emission scenario, 137Cs-containing aerosol particles were injected into a grid cell above the FDNPP. The
injection height was assigned based on the emission scenario, which varied temporally from 20 to 150 m.
This offline dispersion model was driven by meteorological grid point value (GPV) data calculated by the data
assimilation system in Kunii (2014) and Sekiyama et al.
(2015, 2017). The data assimilation system was composed ofa non-hydrostatic regional weather prediction model(referred to as NHM; cf. Saito et al. 2006, 2007) and a local
ensemble transform Kalman filter (LETKF; cf. Miyoshi andAranami 2006). The NHM has been operationally used by
the Japan Meteorological Agency (JMA) for daily national
weather forecasts with a four-dimensional variation method;
this weather forecast system is called JNoVA (cf. Honda
et al. 2005). The LETKF was driven by 20 ensemble members at a 3-km horizontal resolution within a model domain
over eastern Japan (215  259 grids; cf. fig. 2b of Sekiyama
et al. 2015) and 60 vertical layers from the surface to 22 km

above the surface. Details of the LETKF settings were
described in Sekiyama et al. (2015, 2017). The boundary
conditions for the model domain were provided by the JMA
operational global analysis. Using the NHM and LETKF
system (referred to as NHM-LETKF), we simultaneously
assimilated the observations archived by the JNoVA systemand the land surface wind observations collected by the
JMA automated meteorological data acquisition system(AMeDAS), as described in Sekiyama et al. (2017). The
AMeDAS is a land surface observation network that comprises 1300 stations throughout Japan, with an averageinterval of 17 km. The operational JNoVA dataset contains
land surface pressure, satellite-observed sea surface winds,
and observations from radiosondes, pilot balloons, windprofilers, aircrafts, ships and buoys.
The dispersion model shared the same model domain andhorizontal resolution as those of the NHM-LETKF calculated meteorological GPV data, although the vertical resolution was converted from the original 60 layers (from thesurface to 22 km) to 20 layers (from the surface to 10 km) to
reduce the computational burden of calculations within thestratosphere. Meteorological GPV data were input into thedispersion model at 10-min intervals. In the dispersion
model, the time step was set to 24 s using a linear interpolation of the meteorological GPV data. The model simulations were performed from 10 March to 31 March 2011.

4. Mass flux of Fukushima 137Cs plumes
4.1. Insight into plume P3
First, we focus on Plume P3, which has a discrepancy
between the observation-based analysis and the forward6

T. T. SEKIYAMA AND T. IWASAKI

trajectory analysis, as mentioned in the Introduction. To
track the dominant stream of the plume, the time-integral
Ðtof the column-integrated mass flux ( t12 jcolumn dt; hereafter called the TC mass flux) of plume P3 is illustrated
in Fig. 6a. The time-integration period was chosen to follow the classification of Tsuruta et al. (2014), shown in
Fig. 1, and Nakajima et al. (2017), shown in Fig. 2, asmuch as possible. In this period (15:00 to 21:00 JST on
15 March 2011), the dominant stream (red or dark-red
shading) flowed northwestward directly from the FDNPP
and crossed the Abukuma Mountains toward the
Nakadori region (Fig. 6a). Meanwhile, the widespread
mass flux (which likely includes the residuals of plume P2
as shown in the panel P2 of Fig. 7) generally flowed
northward or northeastward in the vicinity of the dominant stream of plume P3. This is consistent with the forward trajectory results of Kajino et al. (2016). The
widespread and northeastward-directed TC mass flux
occurred at an order of magnitude lower than that of thedominant stream directly emitted from the FDNPP.
However, the arrow of plume P3 at 18:00 JST on 15
March from Nakajima et al. (2017), shown in Fig. 2 (also
depicted as a large gray arrow in Fig. 6a), turns southwestward over the Nakadori region.
Here, we pay attention not only to the TC mass flux
but also to the horizontal mass flux at the ground surface(20 m) for plume P3 (Fig. 6b). The horizontal mass flux
at the ground surface turned southwestward over the
Nakadori region. This flow direction is consistent withthat of Nakajima et al. (2017). The horizontal southwestward mass flux at the surface over the Nakadori region
was a non-dominant stream that was approximately twoorders of magnitude lower than that in the immediate
vicinity of the FDNPP. The plume P3 direction shown by
Nakajima et al. (2017) was analyzed only by means of
the surface observations from Tsuruta et al. (2014).
Therefore, these studies derived only the southwestward
stream during this time slot. The two opposite stream
directions are clearly depicted by the vertical cross-sectionof the pollutant concentration and the meridional wind
(Fig. 6c). The maximum concentration core of plume P3
was located at a height of 900 m above the ground level(AGL), where southerly winds were occurring. In contrast, the surface wind direction was northerly beneaththe plume core, although it was very weak. This indicatesthat a dominant stream of atmospheric pollutants cannot
be determined using only ground surface observations.

4.2. Other plumes
Next, we surveyed each plume (P1–P2 and P4–P9) toanalyze the TC mass flux direction and magnitude.
Figure 7 shows the TC mass flux of Plumes P1–P2 andFig. 6. Time-integrals from 15:00 to 21:00 JST on 15 March2011 of (a) the column-integrated mass flux of plume P3 and (b)the horizontal mass flux of plume P3 at the ground surface. Colorshading indicates the magnitude of the mass flux. Small
arrowheads indicate the direction of the mass flux. A large grayarrow illustrates the depiction in Nakajima et al. (2017) for plumeP3 at 18:00 JST on March 15, shown in Fig. 2. (c) Latitude-altitude
cross-sections at 140.5 E for 137Cs concentrations (gray shading),
meridional winds (arrows), and potential temperature (h; contours)
at 17:00 JST on 15 March 2011. The cross-section is located at thegreen line in panel (b).

MASS FLUX ANALYSIS OF 137CS PLUMES

7

Fig. 7. Time-integrals of the column-integrated mass fluxes of plumes P1–P2 and P4–P9. Color shading indicates the magnitude of themass flux. Small arrowheads indicate the direction of the mass flux. Gray arrows illustrate the depictions from Nakajima et al. (2017)
for plumes P1–P2 and P4–P9, shown in Fig. 2. In figure P2, only part of the plume P3 arrows from Nakajima et al. (2017) in thevicinity of the FDNPP is illustrated.

8

T. T. SEKIYAMA AND T. IWASAKI

Fig. 8. (a) Same as Fig. 7 for plume 2, but the plottedarrowheads include the direction of small mass fluxes <1 GBq m1.
Gray diamonds indicate the 137Cs sample stations from Tsuruta
et al. (2014) and Oura et al. (2015). (b) Convergence of the columnintegrated 137Cs mass flux time-integrated during the period ofplume 2 in the model simulation. (c) Total deposition measured byairplanes, which were recorded by the Japanese government (Torii
et al. 2012). Note that this deposition represents all Cs-137
emissions and plumes until autumn of 2011.

Fig. 9. (a) Time-integral of the horizontal mass flux at the
ground surface for plume P8. The time-integrated period is
consistent with that of plume P8 in Fig. 7. Color shading indicatesthe magnitude of the mass flux. Small arrowheads indicate thedirection of the mass flux. Latitude-altitude cross-sections at
140.4 E for 137Cs concentrations (gray shading), meridional wind(arrows), and potential temperature (h; contours) are shown at (b)15:00 JST on 20 March 2011 and (c) 17:00 JST on 20 March 2011.
The cross-sections are located at the green line in panel (a).

MASS FLUX ANALYSIS OF 137CS PLUMES

P4–P9 with plume pathway arrows from Nakajima et al.
(2017) that are the same as those in Fig. 6a. The corresponding time-integration periods were chosen to follow
the classification of Tsuruta et al. (2014), shown in Fig.
1, and Nakajima et al. (2017), shown in Fig. 2 (hereinafter called T&N), as much as possible.
The model-calculated TC mass flux of plume P1 flowed
mainly northeastward over the Pacific Ocean and partly
crossed the Oshika Peninsula located 120 km north-northeast of the FDNPP. The mass flux slightly flanked acoastal area north of the FDNPP, which was consistent
with the plume depiction in T&N. The mainstream of
Plume P1 was detected by dose rate monitoring posts at
the Onagawa nuclear power plant in the Oshika Peninsula(the Tohoku Electric Power Company, http://www.tohokuepco.co.jp/ICSFiles/afieldfile/2011/03/14/11031401_t1.pdf, inJapanese) during the P1 time slot (18:00 JST on 12 March to
06:00 JST on 13 March 2011). The monitoring post data
indicated that the dose rates at the Oshika Peninsula slowly
increased after 19:00 JST on 12 March, suddenly increased
at 00:00 JST on 13 March, peaked at 02:00 JST on 13
March, and plateaued after 05:00 JST on 13 March; all ofthese stages were occurred during the P1 time slot.
The model-calculated TC mass flux of plume P2 mainly
flowed south-southwestward from the FDNPP, turned
west, and spread over a large area in the Kanto region.
This pathway was very consistent with those depicted in
T&N. In addition, a northwestward flow was distinguishedin the vicinity of the FDNPP, which was likely classified aspart of plume P3 in T&N. Tsuruta et al. (2014) andNakajima et al. (2017) proposed three terminal directions
for plume P2: northeastward, northwestward, and southward over the Kanto region. The northeastward and northwestward flows were completely consistent with those
illustrated by the TC mass flux. The dominant southwardflows in T&N did not seem to be reproduced by the TC
mass flux in Fig. 7. However, a southward (or southwestward) flow can be seen by plotting arrowheads for verysmall mass fluxes (Fig. 8a), although this flow was not
dominant. Unfortunately, the observations in Tsuruta
et al. (2014) and Oura et al. (2015) were not derived fromthe northern area of the Kanto region (shown in Fig. 8a).
This biased distribution caused a bias in the plume analysis
from T&N when the southward flow was supposed to be
dominant. In practice, the northeastward/northwestward
flows were presumably dominant and converged over themountainous area of the northern Kanto region during
this time period (shown by a gray rounded square in Fig.
8b). It is noted that the convergence (¼ deposition) was
inconsistent or unnatural near the FDNPP because thetime-integral length (¼ 6 h) was too short for the sourcearea. The convergence area (shown by the gray roundedsquare) resembles the 137Cs-polluted land surface over the9mountainous area of the northern Kanto region in regard
to shape and depth (Fig. 8c), which indicates that the convergence (¼ deposition) of Plume P2 is likely responsible
for this land pollution. During this time period, it was drizzling (approximately <0.1 mm per 6 h) over this mountainous area in the model, which caused wet deposition.
Tsuruta et al. (2014) and Nakajima et al. (2017)
inferred that plume P4 was transported by northerly
winds over the ocean, which eventually reached the
Choshi Peninsula. This depiction is clearly consistent with
one of the two streams from the model-calculated TC
mass flux during the P4 time slot. According to the TC
mass flux depiction, the two streams were southward andeastward; the edge of the southward stream passed over
the Choshi Peninsula, which is the easternmost area ofthe Kanto region.
Tsuruta et al. (2014) and Nakajima et al. (2017) reported
that plumes P5 and P6 were short-duration events that
occurred on 18 and 19 March, respectively, and both wereobserved over a limited area just north of the FDNPP.
During these time slots, the plumes were mainly transported towards the Pacific Ocean, as shown by the modelcalculated TC mass flux in Fig. 7. However, the peripheriesof the plumes passed slightly over a coastal area north of
the FDNPP. The ground surface observations in Tsuruta
et al. (2014) likely detected these peripheries.
The model-calculated TC mass flux for plume P7 depicts
a pathway characteristic of pollutant flows. The plume wastransported over a long distance; first, it flowed southeastward with the offshore winds; second, it sharply turned
clockwise toward the west over the ocean; then, it flowedonshore and landed over the Kanto region, as shown in
Fig. 7. This pathway is obviously consistent with that referenced in T&N. Meanwhile, another northwestward flow
was also depicted from the FDNPP, which was likely partially attributed to plume P8 because the P7 time slot overlapped with that of P8 (Fig. 1).
The depiction of plume P8 was somewhat similar tothat of plume P3. The model-calculated TC mass flux
indicated that a dominant stream flowed northwestward
from the FDNPP and turned toward the northeast at
38 N. However, Tsuruta et al. (2014) and Nakajima
et al. (2017) reported that plume P8 was transported
northwestward from the FDNPP and redirected toward
the south near the Nakadori region at the northern edgeof the Abukuma Mountains (shown by gray arrows in
Fig. 7). In the TC mass flux and T&N depictions, theplume rotates in the opposite direction. This inconsistency is discussed in the next section in detail.
Plume P9 followed a southern route and reached the
Kanto region, which had sporadic rainfall during thistime slot. This rainfall caused 137Cs wet deposition in awide range of areas in the Kanto region, which dispersed

10

T. T. SEKIYAMA AND T. IWASAKI

the plume (cf. Science Council of Japan 2014; Tsuruta
et al. 2014; Sekiyama et al. 2015; Nakajima et al. 2017).
The TC mass flux depicted the plume in a tongue shapethat flowed southward from the FDNPP; then, it flowed
to the northern part of the Kanto region and spread oversouthern parts of the Kanto region, as shown in Fig. 7.
This pathway is precisely consistent with that of T&N,
which is based on ground surface observations.

4.3. Inconsistency of plume P8
In the previous section, we encountered another inconsistent depiction between the TC mass flux and the T&N illustration for plume P8. The dominant stream of plume P8
flowed with the southerly winds throughout the entire
period for the model-calculated TC mass flux; however, thestream exhibited an opposite rotation at the Abukuma
Mountains for the T&N illustration. The discrepancy islikely to be caused by the difference in wind direction
between the ground surface and the dominant layer of themass flux, similar to that for plume P3. The horizontal
mass flux at the ground surface (20 m), shown in Fig. 9a,clearly explains the two situations. The surface mass flux
turned southwestward and propagated over the Nakadori
region. Here, this flow direction was completely consistent
with that of T&N based on ground surface observations.
As shown in Fig. 9b, the high concentration core of plumeP8 was distributed from the ground surface to 1000 m
AGL at the beginning of the P8 time slot. At this time, thewind direction was northerly at <500 m AGL, and it was
southerly at higher than 500 m AGL. This vertical wind
shear caused an inconsistent depiction between the TC
mass flux analysis and the surface observation analysis.
Then, the lower part of plume P8 was transported
southward along the ground surface, while the upper partwas transported northward along approximately isentropic surfaces. Consequently, the vertical cross-section
shape of plume P8 arched over the ground in 2 h, asshown in Fig. 9c. Although the column-integrated massflux is a useful tool to track atmospheric pollutants, across-check between the column-integrated and theground surface mass fluxes is essential when the verticalwind shear is expected to be large.

using together with the TC mass flux because the inconsistencies were only caused by the vertical wind shear
between the ground surface and the dominant plumelayer. The convergence of the TC mass flux was also able
to explain the 137Cs deposition distribution over northern
Kanto region measured by airplanes.
The mass flux analysis clearly provided an overall illustration of dominant plume locations, directions, and depositions without developing movie/animation files.
Unfortunately, such a useful mass flux analysis has rarely
been used for dispersion simulations in atmosphericchemistry. The concept of the mass flux analysis can befurther used as diagnostic or quantitative visualization
tools because mass flux illustrations are a robust way for
researchers to reduce high-dimensional model outputs
into a lower-dimensional plot. It is inadequate to examine
only a time series of concentrations at each observation
point because such a conventional approach cannot validate the reproducibility of model-calculated mass flow balance. In contrast, the mass flux analysis can track theemission, advection, dispersion, and deposition of atmospheric constituents.

Acknowledgments
We thank Dr. Mizuo Kajino from the MeteorologicalResearch Institute for providing us with his Lagrangian
dispersion model. We also thank Prof. Teruyuki
Nakajima from the Japan Aerospace Exploration Agency
for kindly providing us with the figures in his paper.

Funding
This work was supported by JSPS KAKENHI (underGrant JP16H02946, JP16H04051 and JP17K00533) andtheEnvironment
ResearchandTechnology
Development Fund (5-1605) at the Environmental
Restoration and Conservation Agency (ERCA).

ORCID
Toshiki Iwasaki

http://orcid.org/0000-0003-2110-0687

5. Conclusions

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137
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performance. The model-calculated TC mass flux was
almost entirely consistent with the surface observations
from Tsuruta et al. (2014) and Oura et al. (2015),
although with some exceptions. However, these exceptions can be explained by the ground surface mass flux

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Sakaguchi, A., Kondo, H., Yamamoto, M., Miyata, Y., and Igarashi, Y., Temporal changes indissolved 137Cs concentrations in groundwater and stream water in Fukushima after the
Fukushima Dai-ichi Nuclear Power Plant accident, Journal of Environmental Radioactivity,166,
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Hirose, K., Kikawada, Y., Igarashi, Y., Fujiwara, H., Jugder, D., Matsumoto, Y., Oi, T., Nomura,M., Plutonium, 137Cs and uranium isotopes in Mongolian surface soils, Journal of Environmental
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2016
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Igarashi, Y., Kajino, M., Zaizen, Y., Adachi, K., and Mikami, M., Atmospheric radioactivity
over Tsukuba, Japan: a summary of three years of observations after the FDNPP accident,
Progress in Earth and Planetary Science 2:44, 2015. doi: 10.1186/s40645-015-0066-1.
Kajino, M., G-3. NHM-Chem: Sensitivity of Cs deposition to the size and hygroscopicity of Csbearing aerosols, Technical Reports of the Meteorological Research Institute, 76, 132–142,
2015.
Sekiyama, T. T., Kunii, M., Kajino, M., and Shinbori, T., Horizontal Resolution Dependenceof Atmospheric Simulations of the Fukushima Nuclear Accident Using 15-km, 3-km, and 500m Grid Models, J. Meteorol. Soc. Jpn., 93B, 49–64, 2015.

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2014
Abe, Y., Iizawa, Y., Terada, Y., Adachi, K., Igarashi, Y., and Nakai, I., Detection of uranium
and chemical state analysis of individual radioactive microparticles emitted from the Fukushima
nuclear accident using multiple synchrotron radiation X-ray analyses, Analytical Chemistry, 86,
8521-8525, 2014.
2013
Aoyama, M., Uematsu, M., Tsumune, D., and Hamajima, Y., Surface pathway of radioactive
plume of TEPCO Fukushima NPP1 released 134Cs and 137Cs, Biogeosciences, 10, 30673078, 2013. doi: 10.5194/bg-10-3067-2013.
Tsumune, D., Tsubono, T., Aoyama, M., Uematsu, M., Misumi, K., Maeda, Y., Yoshida, Y.,
and Hayami, H., One-year, regional-scale simulation of 137Cs radioactivity in the ocean
following the Fukushima Daiichi Nuclear Power Plant accident, Biogeosciences, 10, 5601-5617,
2013.
Povinec, P.P., Aoyama, M., Biddulph, D., Breier, R., Buesseler, K., Chang, C., Golser, R., Hou,X., Ješkovský, M., and Jull, A., Cesium, iodine and tritium in NW Pacific waters-a comparisonof the Fukushima impact with global fallout. Biogeosciences, 10, 5481-5496. 2013.
Kumamoto, Y., Murata, A., Kawano, T., and Aoyama, M., Fukushima-derived radiocesium inthe northwestern Pacific Ocean in February 2012, Applied Radiation and Isotopes, 81, 335-339,
2013. doi: 10.1016/j.apradiso.2013.03.085.
Toyama, C., Muramatsu, Y., Igarashi, Y., and Aoyama, M., Atmospheric fallout of 129I in Japan
before the Fukushima accident: Regional and global contributions, 1963-2005, Environmental
Science & Technology, 47(15), 8383-8390, 2013. doi: 10.1021/es401596z.
Adachi, K., Kajino, M., Zaizen, Y., and Igarashi, Y., Emission of spherical cesium-bearing
particles from an early stage of the Fukushima nuclear accident. Scientific Reports, 3, Article
number 2554, 2013. doi:10.1038/srep02554.
2012
Toyama, C., Muramatsu, Y., Uchida, Y., Igarashi, Y., Aoyama, M., and Matsuzaki, H.,
Variations of 129I in the atmospheric fallout of Tokyo, Japan 1963-2003, Journal ofEnvironmental Radioactivity, 113, 116-122, 2012. doi: 10.1016/j.jenvrad.2012.04.014.
Inomata, Y., Aoyama, M., Tsumune, D., Motoi, T., and Nakano, H., Optimum interpolation
analysis of basin-scale 137Cs transport in surface seawater in the North Pacific Ocean, Journalof Environmental Monitoring, 14, 3146-3155, 2012. doi: 10.1039/C2EM30707C.
Aoyama, M., Tsumune, D., Uematsu, M., Kondo, F., and Hamajima, Y., Temporal variation of
134Cs and 137Cs activities in surface water at stations along the coastline near the Fukushima Daiichi Nuclear Power Plant accident site, Japan. Geochem. J, 46, 321-325, 2012.
Aoyama, M., Tsumune, D., and Hamajima, Y., Distribution of 137Cs and 134Cs in the North
Pacific Ocean: impacts of the TEPCO Fukushima-Daiichi NPP accident, Journal ofRadioanalytical and Nuclear Chemistry, 1-5, 2012. doi: 10.1007/s10967-012-2033-2.
Honda, M. C., Aono, T., Aoyama, M., Hamajima, Y., Kawakami, H., Kitamura, M., Masumoto,Y., Miyazawa, Y., Takigawa, M., and Saino, T., Dispersion of artificial caesium-134 and-137
in the western North Pacific one month after the Fukushima accident, Geochem. J, 46, e1-e9,
2012.
Povinec, P. P., Hirose, K., and Aoyama, M., Radiostrontium in the Western North Pacific:
characteristics, behavior, and the Fukushima impact, Environmental Science & Technology, 46,10356-10363, 2012. doi: 10.1021/es301997c.
Tsumune, D., Tsubono, T., Aoyama, M., and Hirose, K., Distribution of oceanic 137Cs from the
Fukushima Dai-ichi Nuclear Power Plant simulated numerically by a regional ocean model,
Journal of Environmental Radioactivity, 111, 100-108, 2012.

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2011
Aoyama, M., Fukasawa, M., Hirose, K., Hamajima, Y., Kawano, T., Povinec, P.P., and SanchezCabeza, J. A., Cross Equator transport of 137Cs from North Pacific Ocean to South Pacific Ocean,
BEAGLE2003
cruises,
Progressin
Oceanography,
89,
7-16,
2011.
doi:10.1016/j.pocean.2010.12.003.
Buesseler, K., Aoyama, M., and Fukasawa, M., Impacts of the Fukushima nuclear power plants
on marine radioactivity. Environmental Science & Technology, 45(23), 9931-9935, 2011. doi:
10.1021/es202816c.
Povinec, P. P., Aoyama, M., Fukasawa, M., Hirose, K., Komura, K., Sanchez-Cabeza, J.A.,
Gastaud, J., Jeskovsky, M., Levy, I., and Sykora, I., 137Cs water profiles in the South Indian
Ocean - an evidence for accumulation of pollutants in the subtropical gyre. Progress in
Oceanography, 89 , 17-30, 2011. doi:10.1016/j.pocean.2010.12.004.
Gastaud, J., Povinec, P.P., Aoyama, M., Hirose, K., Sanchez-Cabeza, J.A., Levy, I., Roos, P.,
Eriksson, M., Bosc, E. and Rezzoug, S., Transport and scavenging of Pu in surface waters of theSouthern Hemisphere Oceans, Progress in Oceanography, 89, 92-100, 2011.
doi:10.1016/j.pocean.2010.12.009.
Hirose, K., Kim, C.S., Yim, S.A., Aoyama, M., Fukasawa, M., Komura, K., Povinec, P.P., andSanchez-Cabeza, J.A., Vertical profiles of plutonium in the central South Pacific, Progress in
Oceanography, 89, 101-107, 2011. doi:10.1016/j.pocean.2010.12.010.
Tsumune, D., Aoyama, M., Hirose, K., Bryan, F., Lindsay, K. and Danabasoglu, G., Transportof 137Cs to the Southern Hemisphere in an Ocean General Circulation Model, Progress in
Oceanography, 89, 38-48, 2011. doi:10.1016/j.pocean.2010.12.006.
Sanchez-Cabeza, J.A., Levy, I., Gastaud, J., Eriksson, M., Osvath, I., Aoyama, M., Povinec, P.P.
and Komura, K., Transport of North Pacific 137Cs labeled waters to the south-eastern Atlantic
Ocean. Progress in Oceanography, 89, 31-37, 2011. doi:10.1016/j.pocean.2010.12.005.
Levy, I., Povinec, P.P., Aoyama, M., Hirose, K., Sanchez-Cabeza, J.A., Comanducci, J-F.,
Gastaud, J., Eriksson, M., Hamajima, Y., Kim, C.S., Komura, K., Osvath, I., Roos, P. and Yim,
S.A., Marine anthropogenic radiotracers in the Southern Hemisphere: new sampling andanalyticalstrategies,
Progressin
Oceanography,
89,
120-133,
2011.
doi:10.1016/j.pocean.2010.12.012.
2010
Inomata,Y., Igarashi, Y., Yoshioka, K., Tanaka, T. Y., and M, Chiba., Temporal variation of
222Rn at the summit of Mt. Fuji associated with the Asian continental outflow, Atmospheric
Environment, 44,3856-3865, 2010.
Nakano, H., Motoi, T., Hirose, K., and Aoyama, M., Analysis of 137Cs concentration in the
Pacific using a Lagrangian approach. Journal of Geophysical Research, 115, C06015, 2010.
doi:10.1029/2009JC005640.
Hirose, K., Igarashi, Y., Aoyama, M., and Inomata, Y., Depositional behaviors of plutonium
and thorium isotopes at Tsukuba and Mt. Haruna in Japan indicate the sources of atmosphericdust. Journal of Environmental Radioactivity, 101,106-112, 2010.
Zhang, Y., Zeng, J., Yamada, M., Wu, F., Igarashi, Y., and Hirose, K., Characterization of
Pu concentration and its isotopic composition in a reference fallout material. Science of the
Total Environment, 408(5), 1139-1144, 2010.
2009
Aoyama, M., Hamajima, Y., Fukasawa, M., Kawano, T., and Watanabe, S., Ultra low level deep
water 137Cs activity in the South Pacific Ocean, Journal of Radioanalytical and Nuclear
Chemistry, 2009. doi:10.1007/s10967-009-0253-x.
Hirose, K., Aoyama, M., and Povinec, P.P., 239, 240Pu /137Cs ratios in the water column of theNorth Pacific: a proxy of biogeochemical processes, Journal of Environmental Radioactivity,
100, 258-262, 2009.
Igarashi, Y., Inomata, Y., Aoyama, M., Hirose, K., Takahashi, H., Shinoda, Y., Sugimoto, N.,
Shimizu, A., and Chiba, M., Possible change in Asian dust source suggested by atmosphericanthropogenic radionuclides during the 2000s, Atmospheric Environment, 43, 2971-2980, 2009.
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Inomata, Y., Aoyama, M., and Hirose, K., Analysis of 50-y record of surface 137Csconcentrations in the global ocean using the HAM-global database, Journal of Environmental
Monitoring, 11(1), 116-125, 2009. doi: 10.1039/b811421h.
Hirose, K., Igarashi, Y., Aoyama, M., and Inomata, Y., Depositional behaviors of plutonium andthorium isotopes at Tsukuba and Mt.Haruna in Japan indicate the sources of atmospheric dust,
JournalofEnvironmental
Radioactivity,
101(2),
106-112,2009.
doi:10.1016/j.jenvrad.2009.09.003.
Igarashi Y., Kikawada, Y., Oda, K., Yamauchi, R., Nomura, M., Honda, T., Oi, T., and Hirose,K., Anomalous Uranium Isotope Ratio in Atmospheric Deposits in Japan, Journal of Nuclear
Science and Technology, 46(12), 1094-1098. 2009.

2008
Aoyama, M., Hirose, K., Nemoto, K., Takatsuki, Y., and Tsumune, D., Water masses labeledwith global fallout 137Cs formed by subduction in the North Pacific, Geophysical Research
Letters, 35, L01604, 2008. doi:10.1029/2007GL031964.
Hirose, K., Igarashi, Y., and Aoyama, M., Analysis of the 50-year records of the atmosphericdeposition of long-loved radionuclides in Japan, Applied Radiation and Isotopes, 66, 1675-1678,
2008.
Hirose, K., Aoyama, M., Igarashi, Y., and Komura, K., Improvement of 137Cs analysis in small
volume seawater samples using the Ogoya underground facility, Journal of Radioanalytical andNuclear Chemistry, 276,3,795-798, 2008.
Aoyama, M., Hirose, K., Nemoto, K., Takatsuki, Y., and Tsumune, D., Water masses labeledwith global fallout 137Cs formed by subduction in the North Pacific, Geophysical Research
Letters, 35, L01604, 2008. doi:10.1029/2007GL031964.
2007
Hirose, K., Aoyama, M., and Kim, C.S., Plutonium in Seawater of the Pacific Ocean, Journal ofRadioanalytical and Nuclear Chemistry, 274 No. 3, 635-638, 2007.
Hirose, K., Igarashi, Y., and Aoyama, M., 50 years records of atmospheric deposition of longlived anthropogenic radionuclides in Japan, International Conference on Environmental
Radioactivity: From Measurements and Assessments to Regulation, IAEA-CN-145, 95-96,
2007.
Hirose, K., Igarashi, Y., and Aoyama, M., Recent trends of plutonium fallout observed in Japan:
Comparison with natural lithogenic radionuclides, thorium isotopes, Journal of Radioanalyticaland Nuclear Chemistry, 273 No.1, 115-118, 2007.
Hirose, K., Aoyama, M., Fukasawa, M., Kim, C. S., Komura, K., Povinec, P. P., and SanchezCabeza, J. A., Plutonium and 137Cs in surface water of the South Pacific Ocean, Science of the
Total Environment, 381, 243-255, 2007.
2006
Yamamoto, M., Sakaguchi, A., Sasaki, K., Hirose, K., Igarashi, Y., and Kim, C.K., Seasonaland spatial variation of atmospheric 210Pb and 7Be deposition: features of the Japan Sea side ofJapan, Journal of Environmental Radioactivity, 86, 110-131, 2006.
Hirose, K., Aoyama, M., Kim, C.S., Kim, C.K., and Povinec, P.P., Plutonium isotopes inseawater of the North Pacific: effects of close-in fallout, INTERNATIONAL CONFERENCE
ON ISOTOPES AND ENVIRONMENTAL STUDIES. Eds. P. Povinec, J.A. Sanchez-Cabeza.
Vol. 8 of Radioactivity in the Environment, Elsevier, London, 2006.
Aoyama, M., Hirose, K., and Igarashi, Y., Re-construction and updating our understanding on
the global weapons tests 137Cs fallout, Journal of Environmental Monitoring, 8, 431-438, 2006.
Aoyama, M., Fukasawa, M., Hirose, K., Mantoura, R.F.C., Povinec, P.P., Kim, C.S., andKomura, K., Southern Hemisphere Ocean Tracer Study (SHOTS): An overview and preliminary
results, INTERNATIONAL CONFERENCE ON ISOTOPES AND ENVIRONMENTALSTUDIES. Eds. P. Povinec, J.A. Sanchez-Cabeza, Vol. 8 of Radioactivity in the Environment,
Elsevier, London, 2006.

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4.
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6.
7.

Inoue, H.Y., Matsueda, H., Igarashi, Y., Sawa, Y., Wada, A., Nemoto, K., Sartorius, H., andSchlosser, C., Seasonal and long-term variations in atmospheric CO2 and 85Kr in Tsukuba,
Central Japan, Journal of the Meteorological Society of Japan, 84/6, 959-968, 2006.

2005
Hirose, K., Aoyama, M., Igarashi, Y., and Komura, K., Extremely low backgroundmeasurements of 137Cs in seawater samples using an underground facility (Ogoya), Journal ofRadioanalytical and Nuclear Chemistry, 263, 349-353, 2005.
Povinec, P.P., Aarkrog, A., Buesseler, K.O., Delfanti, R., Hirose, K., Hong, G.H., Ito, T.,
Livingston, H.D., Nies, H., Noshkin, V.E., Shima, S., and Togawa, O., 90Sr, 137Cs and 239,
240
Pu concentration surface water time series in the Pacific and Indian Oceans- WOMARS
results, Journal of Environmental Radioactivity, 81, 63-87, 2005.
Igarashi, Y., Aoyama, M., Hirose, K., Povinec, P.P., and Yabuki, S., What anthropogenic
radionuclides (90Sr and 137Cs) in atmospheric deposition, surface soils and Aeoliandusts suggest
for dust transport over JAPAN, Water, Air, and Soil Pollution: Focus, 5, 51-69, 2005.
2004
Hirose, K., Kim, C.K., Kim, C.S., Chang, B.W., Igarashi, Y., and Aoyama, M., Wet and dry
depositions of plutonium in Daejeon, Korea, The Science of the Total Environment, 332, 243252, 2004.
Povinec, P.P., Hirose, K., Honda, T., Ito, T., MarianScott, E., and Togawa, O., Spatial
distribution of 3H, 90Sr, 137Cs and 239, 240Pu in surface waters of the Pacific and Indian Oceans GLOMARD database, Journal of Environmental Radioactivity, 76, 113-137, 2004.
Aoyama, M., and Hirose, K., Artificial Radionuclides database in the Pacific Ocean: Ham
database, TheScientificWorldJOURNAL, 4, 200-215, 2004.
Hirota, M., Nemoto, K., Wada, A., Igarashi, Y., Aoyama, M., Matsueda, H., Hirose, K.,
Sartorius, H., Schlosser, C., Schmid, S., Weiss, W., and Fujii, K., Spatial and Temporal
Variations of Atmospheric 85Kr Observed During 1995-2001 in Japan: Estimation of
Atmospheric 85Kr inventory in the Northern Hemisphere, Journal of Radiation Research,
45, 405-413, 2004.
Hirose, K., Chemical Speciation of Thorium in Marine Biogenic Particulate matter, The
Scientific World JOURNAL 4, 67-76, 2004.
2003
Hirose, K., and Aoyama, M., Analysis of 137Cs and 239, 240Pu concentrations in surface waters of
the Pacific Ocean, Deep-Sea Research II, 50, 2675-2700, 2003.
Hirose, K., Aoyama, and Povinec, P.P., Concentration of particulate plutonium in surface anddeep-water, samples collected during the IAEA’97 expedition, Deep-Sea Research II, 50, 26392647, 2003.
Povinec, P.P., Livingston, H.D., Shima, S., Aoyama, M., Gastaud, J., Goroncy, I., Hirose, K.,
Huynh-Ngoc, L., Ikeuchi, Y., Ito, T., Rosa, J.L., Kwong, L.L.W., Lee, S.H., Moriya, H., Mulsow,
S., Oregioni, B., Pettersson, H., and Togawa, O., IAEA’97 expedition to the NW Pacific OceanResults of oceanographic and radionuclide investigations of the water column, Deep-Sea
Research II, 50,2607-2637, 2003.
Ito, T., Povinec, P.P., Togawa, O., and Hirose, K., Temporal and Spatial variation ofanthropogenic radionuclides in seawater of the Japan Sea, Deep-Sea Research II, 50, 2701-2711,
2003.
Tsumune, D., Aoyama, M., and Hirose, K., Behavior of 137Cs and concentrations in the North
Pacific in an ocean general circulation model, Journal of Geophysical Research, 108, C8, 3262,
2003.
Hirose, K., and Aoyama, M., Present Background Levels of Surface 137Cs and 239, 240Pu
Concentrations in the Pacific, Journal of Environmental Radioactivity, 69(1-2), 53-60, 2003.
Tsumune, D., M. Aoyama, and K. Hirose, Numerical simulation of 137Cs and 239, 240Pu
concentrations by an ocean general circulation model, Journal of Environmental Radioactivity,69(1-2), 61-84, 2003.
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8.

Aoyama, M., and Hirose, K., Temporal variation of 137Cs water column inventory in the North
Pacific since the 1960s, Journal of Environmental Radioactivity, 69(1-2),107-117, 2003.
9. Hirose, K., Igarashi, Y., Aoyama, M., Kim, C.K., Kim, C.S., Chang, B.W., Recent trends ofplutonium fallout observed in Japan: plutonium as a proxy for desertification, J. Environ. Monit,
5, 302-307, 2003.
10. Igarashi, Y., Aoyama, M., Hirose, K., Miyao, T., Nemoto, K., Tomita, M., and Fujikawa, T.,
Resuspension: Decadal Monitoring Time Series of the Anthropogenic Radioactivity Depositionin Japan, Journal of Radiation Research, 44, 319-328, 2003.
11. Kim, C.K., Kim, C.S., Chang, B.U., Choi, S.W., Chung, C.S., Hong, G.H., Hirose, K., andIgarashi, Y., Plutonium isotopes in seas around the Korean Peninsula, The Science of the Total
Environment, 318, 197-209, 2003.
12. Kim, C.K., Kim, C.S., Chang, B.U., Choi, S.W., Chung, C.S., Hong, G.H., Hirose, K., andPettersson, H.B.L., 240Pu /239Pu atom ratios in the bottom sediments of the NW Pacific
Ocean, Journal of Radioanalytical and Nuclear Chemistry, 258(2), 265-268, 2003.
13. Lee, H.N., Tanaka, T., Chiba, M., and Igarashi, Y., Long range transport of Asian dust from dust
storms and its impact on Japan, Water, Air, and Soil Pollution, 3, 231-243, 2003
14. Hirose, K., Implication of POC/234Th ratios in oceanic particulate matter: an approach to particle
aggregation, Papers of Meteorology and Geophysics, 53, 4, 109-118, 2003.
1.
2.
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2.
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4.
5.
6.
7.

1.

2002
Hirose, K., Miyao, T., Aoyama, M., and Igarashi, Y., Plutonium isotopes in the Sea of Japan,
Journal of Radioanalytical and Nuclear Chemistry, 252, 293-299, 2002.
Hirose, K., and Aoyama, M., Chemical speciation of plutonium in seawater, Analytical andBioanalytical Chemistry, 372, 418-420, 2002.
Hong, G.H., Kim, Y.I., Lee, S.H., Cooper, L.W., Choe, S.M., Tkalin, A.V., Lee, T., Kim, S.H.,
Chung, C.S., and Hirose, K., 239+240Pu and 137Cs concentrations for zooplankton and nekton in
the Northwest Pacific and Antarctic Oceans 1993-1996, Marine Pollution Bulletin, 44, 660-665,
2002.
2001
Igarashi, Y., Aoyama, M., Nemoto, K., Hirose, K., Miyao, T., Fushimi, K., Suzuki, M., Yasui,
S., Asai, Y., Aoki, I., Fujii, K., Yamamoto, S., Sartorius, H., and Weiss, W., 85Kr measurement
system for continuous monitoring at Meteorological Research Institute, Japan, Journal ofEnvironmental Monitoring, 3, 688-696, 2001.
Igarashi, Y., Aoyama, M., Hirose, K., Miyao, T., and Yabuki, S., Is it Possible to use 90Sr and137Cs as tracers for the aeolian transport?, Water. Air. And Soil Pollution, 130, 349-354, 2001
Hirose, K., Igarashi, Y., Aoyama, M., and Miyao, T., Long-term trends of plutonium fallout
observed in Japan, In Plutonium in the Environment, 251-266, 2001.
Aoyama, M., Hirose, K., Miyao, T., Igarashi, Y., and Povinec, P.P., 137Cs activity in surface
water in the western North Pacific, J. Radioanal. Nucl. Chem., 248, 3, 789-793, 2001.
Aoyama, M., Hirose, K., Miyao, T., Igarashi, Y., and Povinec, P.P., Temporal variation of 137Cs
inventory in the western North Pacific, J Radioanal Nucl Chem, 248, 3, 785-787, 2001
Hirose, K., Aoyama, M., Miyao, T., and Igarashi, Y., Plutonium in seawaters on the western
North Pacific, J Radioanal Nucl Chem, 248, 3, 771-776, 2001.
Tsumune. D., Aoyama, M., Hirose, K., Maruyama, K., and Nakashiki, N., Caluculation of
Artificial Radionuclides in the Ocean by an Ocean General Circulation Model, J Radioanal Nucl
Chem, 248, 3, 777-783, 2001.
2000
Shiraishi, K., Kimura, S., Yonehara, H., Takada, J., Ishikawa, M., Igarashi, Y., Aoyama, M.,
Komura, K., and Nakazima, T., Survey of external dose around the JCO facility using sugar
samples and ESR method, Adv. ESR Appl, 16, 9-14, 2000.

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Komura, K, Yamamoto, M., Muroyama, T., Murata, Y., Nakanishi, T., Hoshi, M., Takada, J.,
Ishikawa, M., Kitagawa, K., Suga, S., Endo, A., Tozaki, N., Mitsugashira, T., Hara, M.,
Hashimoto, T., Takano, M., Yanagawa, Y., Tsuboi, T., Ichimasa, M., Ichimasa, Y., Imura, H.,
Sasajima, E., Seki, R., Saito, Y., Kondo, M., Kojima, S., Muramatsu, Y., Yoshida, S., Shibata,S., Yonehara, H., Watanabe, Y., Kimura, S., Shiraishi, K., Bannai, T., Sahoo, S.K., Igarashi, Y.,
Aoyama, M., Hirose, K., Uehiro, M., Doi, T., and Matsuzawa, T., The JCO criticality accident
at Tokai-mura, Japan: an overview of the sampling campaign and preliminary results, J. Environ.
Radioactivity, 50, 3-14, 2000.
Igarashi, Y., Miyao, T., Aoyama, M., Hirose, K., Sartorius, H., and Weiss, W., Radioactive noble
gases in the surface air monitored at MRI, Tsukuba, before and after the JCO accident, J.
Environ. Radioactivity, 50, 107-118, 2000.
Miyao, T., Hirose, K., Aoyama, M., and Igarashi, Y., Trace of the recent deep water formationin the Japan Sea deduced from historical 137Cs data, Geophys. Res. Lett., 27, 22, 3731-3734,
2000.
Aoyama, M., Hirose, K., Miyao, T., and Igarashi, Y., Low level 137Cs measurements in deep
seawater samples, Appl. Radiat. Isot, 53, 159-162, 2000.
Igarashi, Y., Sarutorius, H., Miyao, T., Weiss, W., Fushimi, K., Aoyama, M., Hirose, K., andInoue, H.Y., 85Kr and 133Xe monitoring at MRI, Tsukuba and its importance, J. Environ.
Radioactivity, 48, 191-202, 2000.

1999
Hirose, K., Amano, H., Baxter, M.S., Chaykovskaya, E., Chumichev, V.B., Hong, G.H., Isogai,
K., Kim, C.K., Kim, S.H., Miyao, T., Morimoto. T, Nikitin, A., Oda, K., Pettersson, H.B.L.,
Povinec, P.P., Seto, Y., Tkalin, A., Togawa, O., and Veletova, N.K., Anthropogenic
radionuclides in seawater in the East Sea/Japan Sea: Results of the first-stage Japanese-KoreanRussian expedition, J. Environ. Radioactivity, 43, 1-13, 1999.
Igarashi, Y., Aoyama, M., Miyao, T., Hirose, K., Komura, K., and Yamamoto, M., Airconcentration of radiocaesium in Tsukuba, Japan following the release from the Tokai waste
treatment plant: comparisons of observations with predictions, Applied Radiation and Isotopes,
50, 1063-1073, 1999.
Igarashi, Y., Aoyama, M., Miyao, T., Hirose, K., and Tomita, M., Anomalous 90Sr deposition
during the fall, 1995 at MRI, Tsukuba, Japan, J. Radioanal. Nucl. Chem, 239(3), 539-542, 1999.
Ikeuchi, Y., Amano, H., Aoyama, M., Berezhnov, V.I., Chaykovskaya, E., Chumichev, V.B.,
Chung, C.S., Gastaud, J., Hirose, K., Hong, G.H., Kim, C.K., Kim, S.H., Miyao, T., Morimoto,T., Nikitin, A., Oda, K., Pettersson, H.B.L., Povinec, P.P., Tkalin, A., Togawa, O., and Veletova,
N.K., Anthropogenic radionuclides in seawater of the Far Eastern Seas, Sci. Total Environ.,
237/238, 203-212, 1999.
Pettersson, H.B.L., Amano, H., Berezhnov, V.I., Chaykovskaya, E., Chumichev, V.B., Chung,C.S., Gastaud, J., Hirose, K., Hong, G.H., Kim, C. K., Lee, S.H., Morimoto, T., Nikitin, A., Oda,
K., Povinec, P.P., Suzuki, E., Tkalin, A., Togawa, O., Veletova, N.K., Volkov, Y., and Yoshida,K., Anthropogenic radionuclides in sediments in the NW Pacific Ocean and its marginalseas:results of the 1994-1995 Japanese-Korean-Russian expeditions, Sci. Total Environ,
237/238, 213-224, 1999.
1998
Igarashi, Y., Hirose, K., and Otsuji-Hatori, M., Beryllium-7 Deposition and Its Relation to
Sulfate Deposition, J. Atmos. Chem, 29, 217-231, 1998.
Miyao, T., Hirose, K., Aoyama, M., and Igarashi, Y., Temporal variation of 137Cs and 239, 240Pu
in the sea of Japan, J. Environ. Radioactivity, 40, 239-250, 1998.
1997
Hirose, K., Complexation-scavenging of plutonium in the ocean, Radioprotection - Colloq, 32,
C2-225 - C2-230, 1997.

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1996
Igarashi, Y., Otsuji-Hattori, M., and Hirose, K., Recent deposition of 90Sr and 137Cs observed in
Tsukuba, J. Environ. Radioactivity, 31, 157-169, 1996.
Otsuji-Hatori. M., Igarashi, Y., and Hirose, K., Preparation of a Reference Fall out Material for
Activity Measurements, J. Environ Radioactivity, 31, 2, 143-155, 1966.
1995
Aoyama, M., and Hirose, K., The temporal and spatial variation of 137Cs concentration in thewestern North Pacific and its marginal seas during the period from 1979 to 1988, J. Environ.
Radioactivity, 29, 57-74, 1995.
Hirose, K., Geochemical studies on the Chernobyl radioactivity in environmental samples, J.
Radioanal. Nucl. Chem., Articles, 197, 315-335, 1995.
1994
Hirose, K., Takatani, S., and Aoyama, M., Deposition of 90Sr and plutonium isotopes derived
from the Chernobyl accident in Japan, J. Radioanal. Nucl. Chem., Articles, 182, 349-358, 1994.
1993
Hirose, K., and Sugimura, Y., Chemical speciation of particulate 238U, 239, 240Pu and Th isotopes
in seawater, Sci. Total Environ., 130/131, 517-524, 1993.
Hirose, K., Takatani, S., and Aoyama, M., Wet deposition of radionuclides derived from the
Chernobyl accident, J. Atmos. Chem., 17, 61-71, 1993.
Shiraishi, K., Igarashi, Y., Yamamoto, M., and Nakajima, T., Concentrations of thorium anduranium in freshwater samples collected in the former USSR, J. Radioanal. Nucl. Chem.,
185,157-165, 1993.
1992
Aoyama, M., Hirose, K., and Takatani, S., Particle size dependent dry deposition velocity of theChernobyl radioactivity, In Precipitation Scavenging and Atmosphere - Surface
Exchange.volume 3, The 22e Summers Volume: Application and Appraisals. Coordinated by
S.E. Schwartz and W.G.N. Slinn. Hemisphere Publishing Corporation, Washington, 1581-1593,
1992.
Hirose, K., Sugimura, Y., and Aoyama, M., Plutonium and 137Cs in the western North Pacific:
estimation of residence time of plutonium in surface water, Appl. Radiat. Isot, 43, 349-359,
1992.
1991
Aoyama, M., Hirose, K., and Sugimura, Y., The temporal variation of stratospheric fallout
derived from the Chernobyl accident, J. Environ. Radioactivity, 13, 103-115, 1991.
Hirose, K., and Sugimura, Y., Chemical speciation of particulate uranium in seawater, J.
Radioanal. Nucl. Chem., Articles, 149, 83-96, 1991.
1990
Hirose, K., and Sugimura, Y., Plutonium isotopes in the surface air in Japan: effect of Chernobyl
accident, J. Radioanal. Nucl. Chem., Articles, 138, 127-138, 1990.
Hirose, K., Aoyama, M., and Sugimura, Y., Plutonium and cesium isotopes in river waters inJapan, J. Radioanal. Nucl. Chem., Articles, 141, 191-202, 1990.
Hirose, K., Aoyama, M., and Sugimura, Y., Short and long term effects of Chernobyl
radioactivity on deposition and air concentrations in Japan, IAEA-SM 306/129, 141-149, 1990.
1988
Aoyama, M., Evidence of stratospheric fallout of caesium isotopes from Chernobyl accident,
Geophys. Res. Lett., 15, 327-330, 1988.
Miyake, Y., Saruhashi, K., Sugimura, Y., Kanazawa, T., and Hirose, K., Contents of 137Cs,
plutonium and americium isotopes in the southern ocean waters, Pap. Met. Geophys, 39, 95113, 1988.
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1987
Aoyama, M., Hirose, K., and Sugimura, Y., Deposition of gamma-emitting nuclides in Japanafter the reactor-IV accident at Chernobyl, J. Radioanal. Nucl. Chem., Articles, 116, 291-306,
1987.
Hirose, K., Aoyama, M., Katsuragi, Y., and Sugimura, Y., Annual deposition of Sr-90, Cs-137
and Pu-239, 240 from the 1961-1980 nuclear explosions: a simple model, J. Meteor. Soc. Japan,
65, 259-277, 1987.
1986
Aoyama, M., Hirose, K., Suzuki, Y., and Sugimura, Y., High level radioactive nuclides in Japanin May, Nature, 321, 819-820, 1986.
Hirose, K., Sugimura, Y., and Katsuragi, Y., 90Sr and 239+240Pu in the surface air in Japan: theirconcentrations and size distributions, Pap. Met. Geophys, 37, 255-269, 1986.
Katsuragi, Y., and Aoyama, M., Seasonal variation of Sr-90 fallout in Japan through the end of
1983, Pap. Met. Geophys., 37, 15-36, 1986.
Suzuki, Y., Inoue, H., Katsuragi, Y., and Sugimura, Y., The distribution of 85Krin the air over
the North and South Pacific Ocean Mem., Natl. Inst. Polar Res., Spec. Issue, 40, 462-466, 1986.
1985
Hirose, K., and Sugimura, Y., A new method of plutonium speciation in seawater, J. Radioanal.
Nucl. Chem, Articles, 92, 363-369, 1985.
1984
Hirose, K., and Sugimura, Y., Plutonium in the surface air in Japan, Health Phys., 46, 12811285, 1984.
Inoue, H., Katsuragi, Y., and Shigehara, K., Tritiated water vapor in the surface air at Tokyo,
Pap. Met. Geophys, 35, 11-20, 1984.
1983
Katsuragi, Y., A study of 90Sr fallout in Japan, Pap. Met. Geophys., 33, 277-291, 1983.
Katsuragi, Y., Kawamura, K., and Inoue, H., Tritium fallout in Tokyo, Pap. Met. Geophys, 34,
21-30, 1983.
1982
Inoue, H., and Katsuragi, Y., A study of tritium fallout in Japan, Pap. Met. Geophys, 32, 21-28,
1982.
Katsuragi, Y., Hirose, K., and Sugimura, Y., A study of plutonium fallout in Japan, Pap. Met.
Geophys, 33, 85-93, 1982.
1981
Hirose, K., and Sugimura, Y., Concentration of Uranium and the activity ratio of 234U /238U insurface air-effect of atmospheric burn-up of cosmos-954, Pap. Met. Geophys., 32, 317-322,
1981.
Hirose, K., and Sugimura, Y., Plutonium content of river water in Japan, Pap. Met. Geophys, 32,
301-305, 1981.
Sugimura, Y., and M. Mayeda, An improved method of analysis of uranium in sea water using
chelating resin, Pap. Met. Geophys, 32, 167-171, 1981.
1980
Sugimura, Y., The uranium content and the activity ratio 234U /238U in sea water in the Pacificocean, Isotope Marine Chemistry, 211-246, 1980.
1979
Sugiura, Y., Saruhashi, K., and Miyake, Y., Evaluation on the disposal of radioactive wastes
into the North Pacific. - the effect of steady flow and up - welling, Proc. The 3rd NEA Seminaron Marine Radioecology, 1979.
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1978
Miyake, Y., Nuclear weapons and radioactive pollution of the earth's environment, In, Proc.
Internal Symp. On the damage and after-effects of the atomic bombing of Hiroshima andNagasaki July 21-Aug., 9, 1977, Tokyo, 164-188, 1978.
1976
Miyake, Y., and Saruhashi, K., A critical study on the IAEA definition of high level radioactive
waste unsuitable for dumping at sea, Pap. Met. Geophys. 27, 75-80, 1976.
Miyake, Y., and Saruhashi, K., Disposal of radioactive waste into deep seas, J. Radiat. Res., 17,
42-43, 1976.
Miyake, Y., and Sugimura, Y., The plutonium content in the Pacific Ocean waters, Proc. of
Symp. on Transuranium Nuclides in the Environment, San Francisco, 17-21 Nov. 1975,
USERDA and IAEA, IAEA-SM-199/22, 91-105,1976.
Sugimura, Y., Symposium II plutonium in environment, II-1.Distribution and behavior ofplutonium in the global environments, J. Radiat. Res., 17, 4-5, 1976.
Sugiura, Y., Saruhashi, K., and Miyake, Y., Evaluation on the disposal of radioactive wastes
into the North Pacific, Pap. Met. Geophys, 27, 81-87, 1976.
1975
Miyake, Y., Shimada, T., Kawamura, K., Sugimura, Y., Shigehara, K., and Saruhashi, K.,
Distribution of tritium in the Pacific ocean, Rec. Oceanogr. Works Japan, 13, 17, 1975.
Miyake, Y., Katsuragi, Y., and Sugimura, Y., Plutonium fallout in Tokyo, Pap. Met. Geophys,
26, 1-8, 1975.
Miyake, Y., Sugimura, Y., and Hirao, Y., Uranium, thorium and potassium contents in granitic
and basaltic rocks in Japan, In, The Natural Radiation Environment II, II, Proc. 2nd Intnl. Symp.
On Natural Radiation Environment, Aug. 7-11, 1972, Houston, Texas, USA, ed. J. A. S. Adams,
W. M. Lowder, T. F. Gesell, 535-558,1975.
Saruhashi, K., I-1 The global radioactive contamination due to nuclear weapon testings, J. Radiat.
Res., 16, 47, 1975.
Saruhashi, K., Katsuragi, Y., Kanazawa, T., Sugimura, Y., and Miyake, Y., 90Sr and 137Cs in the
Pacific waters, Rec. Oceanogr. Works Japan, 13, 1-15, 1975.
1973
Miyake, Y., and Saruhashi, K., The radio-carbon model of the ocean, Pap. Met. Geophys, 24,
263-271, 1973.
Miyake, Y., Saruhashi, K., and Sugimura, Y., The isotopic ratio 234U /238U in sea water and its
bearings on the isotopic ratio in river waters, Oceanogr. Works Japan, 12, 23-25, 1973.
Miyake, Y., Sugimura, Y., and Saruhashi, K., Content of plutonium in river water in Japan, Pap.
Met. Geophys, 24, 75-78, 1973.
Miyake, Y., Sugimura, Y., and Yasujima, T., Thorium isotope content in river water in Japan,
Pap. Met. Geophys, 24, 67-73, 1973.
1972
Miyake, Y., Sugimura, Y., and Uchida, T., A new method of spectrophotometric determinationof uranium in sea water and uranium content with 234U /238U ratio in the Pacific water, Rec.
Oceanogr. Works Japan, 11, 53-63, 1972.
1970
Miyake, Y., Katsuragi, Y., and Sugimura, Y., A study on plutonium fallout, J. Geophys. Res.,75, 2329-2330, 1970.
Miyake, Y., Sugimura, Y., and Mayeda, M., The uranium content and the activity ratio 234U
/238U in marine organisms and sea water in the western North Pacific, J. Oceanogr. Soc. Japan,
26, 123-129, 1970.

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1969
Miyake, Y., Radioactive contamination of the ocean, Bull. Jpn. Soc. Fisheries Oceanogr., Nov.,
1969, 1969.
1968
Miyake, Y., and Sugimura, Y., Plutonium content in the western North Pacific waters, Pap. Met.
Geophys, 19, 481-485, 1968.
Miyake, Y., Katsuragi, Y., and Sugimura, Y., Deposition of plutonium in Tokyo through the endof 1966, Pap. Met. Geophys, 19, 267-276, 1968.
1967
Miyake, Y., and Kanazawa, T., Atmospheric ozone and radioactive fallout, Pap. Met. Geophys.,
18, 311-326, 1967.
Miyake, Y., Radioactive contamination of the ocean, J. Radiat. Res., 8, 1, 1967.
Miyake, Y., Sea, radioactivity in, International Dictionary of Geophysics, 1-7, 1967.
1966
Miyake, Y., and Saruhashi, K., On the radio-carbon age of the ocean waters, Pap. Met. Geophys.,
17, 218-223, 1966.
Miyake, Y., and Sugimura, Y., Ratio 234U /238U and the uranium concentration in seawater inthe western North Pacific, J. Geophys. Res., 71, 3083-3087, 1966.
1965
Kuroda, P. K., Miyake, Y., and Nemoto, J., Strontium isotopes Global circulation after the
Chinese nuclear explosion of 14 May 1965, Science, 150, 1289-1290, 1965.
Miyake, Y., Saruhashi, K., Katsuragi K., and Kanazawa, T., Radioactivity of dust and rain - theratio of Cs-137 and Sr-90 in the radioactive fallout, Proc. Internl. Conf. On Cloud Physics,Tokyo,
IAMAP, 395-399, 1965.
Park, K., George, M.J., Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Strontium90 and caesium-137 in Columbia river plume, July 1964, Nature, 208, 1084-1085, 1965.
1964
Miyake, Y., and Sugimura, Y., Uranium and radium in the western North Pacific waters, Studies
on Oceanography, Prof. K. Hidaka Sixtieth Birthday Commemorative Volume, Univ. of Tokyo
Press, Tokyo, 274-278, 1964.
Miyake, Y., and Ohtsuka, Y., Beryllium-7 in rain water, Pap. Met. Geophys, 15, 89-92, 1964.
Miyake, Y., A sequential procedure for the radiochemical analysis of marine material, Annex tosafety series 11."Method of surveying and monitoring marine radioactivity, IAEA, Vienna, 26,1964.
Miyake, Y., Atomic weapons and the pollution of the sea, The East, 3, 64-68, 1964.
Miyake, Y., Saruhashi, K., Katsuragi, Y., Kanazawa, T., and Sugimura, Y., Uranium, radium,
thorium, Ionium, strontium 90 and cesium 137 in coastal waters of Japan, In, Recent Researches
in the Fields of Hydrosphere, Atmosphere and Nuclear Geochemistry, published by Editorial
Committee of Sugawara Festival Volume, Maruzen Co. Ltd., 127-141, 1964.
Miyake, Y., Sugimura, Y., and Tsubota, H., Content of uranium, radium, and thorium in riverwaters in Japan, In, The Natural Radiation Environment, ed. By J. A. S. Adams and W. M.
Lowder, Univ. of Chicago Press, Chicago, Rice Univ. Semicentennial Series, 219-225, 1964.
Sugimura, Y., Torii, T., and Murata, S., Uranium distribution in drake passage waters, Nature,
204, 464-465, 1964.
1963
Folsom, T.R., and Saruhashi, K., A comparison of analytical techniques used for determinationof fallout cesium in sea water for oceanographic purpose, J. Radiat. Res., 4, 39-53, 1963.
Miyake, Y., and Tsubota, H., Estimation of the direct contribution of meteoric water to riverwaters by means of fall-out radio-cesium and radiostrontium, "Radioisotopes in Hydrology"
Proc. IAEA Symp., Tokyo. IAEA, Vienna, 425-431, 1963.
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Miyake, Y., Artificial radioactivity in the Pacific ocean, IUGG Monograph, 20 Radioactive
traces in oceanography, 1963.
Miyake, Y., Saruhashi, K., Katsuragi, K., and Kanazawa, T., Deposition of Cs-137 and Sr-90 in
Tokyo through the end of July 1963, Pap. Met. Geophys, 13, 180-181, 1963.
Miyake, Y., Saruhashi, K., Katsuragi, Y., Kanazawa, T., and Tsunogai, S., Deposition of 90Sr
and 137Cs in Tokyo through the end of July 1963, Pap. Met. Geophys, 14, 58-65, 1963.
Miyake, Y., Sugiura, Y., and Tsubota, H., II. Contents of uranium, radium, and thorium in riverwaters in Japan, The Natural Radiation Environment, RICE Univ. semicentennial Pub, 1963.

1962
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Seasonal variation of radioactive
fallout, J. Geophys. Res., 67, 189-193, 1962.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Penetration of 90Sr and 137Cs indeep layers of the Pacific and vertical diffusion rate of deep water, J. Radiat. Res., 3, 141-147,
1962.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., The peak in radioactive fallout inthe temperate zone of the northern hemisphere, J. Radiat. Res., 3,148-152, 1962.
Sugimura, Y., and Sugimura, T., Uranium in recent Japanese sediments, Nature, 194, 568-569,
1962.
1961
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Cesium 137 and strontium 90 insea water, J. Radiat. Res., 2 25-28, 1961.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Cesium 137 and strontium 90 insea water, Pap. Met. Geophys., 12, 85-88, 1961.
1960
Miyake, Y., and Saruhashi, K., Vertical and horizontal mixing rates of radioactive material in
the ocean, Disposal of Radioactive wastes IAEA Vienna, 167-173, 1960.
Miyake, Y., and Katsuragi, Y., Strontium 90 in western North Pacific surface waters, Pap. Met.
Geophys., 11, 188-190, 1960.
Miyake, Y., Saruhashi, K., Katsuragi, Y., and Kanazawa, T., Radioactive fallout in Japan andits bearings on meteorological conditions, Pap. Met. Geophys., 11, 151-158, 1960.
1959
Miyake, Y., Saruhashi, K., and Katsuragi, Y., The Sr-90 fallout and the air motion, Pap. Met.
Geophys., 9, 172-176, 1959.
Miyake, Y., Special committee on oceanic research, Working Group on Radioactivity in the
Ocean, 1959.
1958
Miyake, Y., and Saruhashi, K., Distribution of man-made radio-activity in the North Pacificthrough summer 1955, J. Mar. Res., 17, 383-389, 1958.
Miyake, Y., and Sugiura, Y., The method of measurement of radioactivity in sea water, Pap.
Met. Geophys., 9, 48-50, 1958.
Miyake, Y., Hazards to human health of radioactive dust, Radio Japan, 2, 3-5, 1958.
Miyake, Y., The distribution of artificial radioactivity in the equatorial region in the Pacific in
the summer of 1956, The Proc. 9th Pacific Science Congress, 16, 227, 1958.
1957
Miyake, Y., and Saruhashi, K., The world-wide strontium 90 deposition during the period from1951 to the fall of 1955, Pap. Met. Geophys., 8, 241-243, 1957.
Miyake, Y., The biological effects of nuclear tests - warning by Japanese scientists, Radio Japan,
1, 3-5, 1957.

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Miyake, Y., XI. Methods for the measurement of radioactivity in sea water, Annuals ofInternational Geophysical Year, 5, 1957.
Miyake, Y., XII. Radioactivity as a tracer of air motions in the atmosphere, Symp. On
radioactivity, IGY, Utrecht, 360, 1957.
Miyake, Y., Sugiura, Y., Saruhashi, K., and Kanazawa, T., The estimation of the amount of Sr90 deposition and the external infinite gamma does in Japan due to man-made radioactivity, Pap.
Met. Geophys., 8, 222-231, 1957.

1956
Miyake, Y., On the distribution of radioactivity in the North Pacific ocean in 1954-1955, Internl.
Conf. For Peaceful Uses of Atomic Energy in Geneva, August 1955, 13, pp.381-384, 1956.
Miyake, Y., Radioactivity in rain water and air, Proc. Internl. Conf. For Peaceful Uses of Atomic
Energy in Geneva, August 1955, 13, 345-349, 1956.
Sugiura, Y., and Kanazawa, T., On the radioactive fallout collected in Tokyo on 26th November,
1955, Pap. Met. Geophys., 7, 128-135, 1956.
1955
Miyake, Y., and Sugiura, Y., Radiochemical analysis of radio-nuclides in sea water collected
near BIKINI atoll, Pap. Met. Geophys., 6, 90-92, 1955.
Miyake, Y., Effects of atomic explosions on the atmosphere and sea, The research in the effects
and influences of the nuclear bomb text explosions, 1-6, 1955.
Miyake, Y., Kigoshi, K., and Saruhashi, K., Radiochemical analysis of fission products
contained in the soil collected at Tokyo, May, 1954, Pap. Met. Geophys., 6, 1, 93-94, 1955.
Miyake, Y., The artificial radioactivity in rain water observed in Japan, 1954-1955, Research in
the Effects and Influences of the Nuclear Bomb Test Explosions Science, 151-159, 1955.
Miyake, Y., The artificial radioactivity in rain water observed in Japan, from autumn 1954 tospring 1955, Pap. Met. Geophys., 6, 26-32, 1955.
Miyake, Y., Sugiura, Y., and Kameda, K., On the artificial radioactivity in the sea near Japan,
Pap. Met. Geophys., 6, 90-92, 1955.
Miyake, Y., Sugiura, Y., and Kameda, K., On the distribution of radioactivity in the sea around
Bikini atoll in June, 1954, Pap. Met. Geophys., 5, 253-262, 1955.
Miyake, Y., Sugiura, Y., and Kameda, K., Research in the effects and influence of the nuclear
bomb test explosions, Soc. For Promotion of Science, 415-417, 1955.
1954
Miyake, Y., Kigoshi, K., Sugiura, Y., and Saruhashi, K., A study on the productivity in coastal
waters by means of the radio-carbon, Pap. Met. Geophys., 5, 89-94, 1954.
Miyake, Y., The artificial radioactivity in rain water observed in Japan from May to August,
1954, Pap. Met. Geophys, 5, 173-177, 1954.

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6. Publication list (in Japanese) 1954-2018
論文リスト（和文誌）1954-2018- 43 -

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2017
足立光司, 福島原子力発電所事故で放出された不溶性放射性粒子の形態と組成, エアロゾル
研究, 32-4,255‒260, 2017.
五十嵐康人, 放射性物質の大気沈着・拡散および陸域からの再浮遊について, Proceedings ofthe 17th Workshop Environmental Radioactivity. KEK Proceedings, 2017-6, 50-55, 2017.
小野貴大, 飯澤勇信, 阿部善也, 中井 泉, 寺田康子, 佐藤志彦, 末木啓介, 足立光司, 五十嵐康人, 福島第一原子力発電所事故により１号機から放出された放射性粒子の放射光マイク
ロビーム X 線分析を用いる化学性状の解明, 分析化学 66（4）, 2017.
梶野瑞王, 大気エアロゾルの環境動態シミュレーション,安全工学, 57-6, 433-441, 2018
木村徹, 五十嵐康人, 財前祐二, ガンマ線スペクトル解析ソフト P-SCAN を用いた高精度な Cs134/Cs137 放射能比の導出, Proceedings of the 17th Workshop Environmental Radioactivity.
KEK Proceedings, 2017-6, 262-266, 2017.

1.

2016
木村徹, 五十嵐康人, 財前祐二, 2016: IAEA-TEL-2015-03 Proficiency Test 試料の測定,
Proceedings of the 17th Workshop Environmental Radioactivity. KEK Proceedings, 2016-8, 228-231.

1.
2.

2015
足立光司, 電子顕微鏡がとらえた放射性粒子, 地球化学, 49,185‒193, 2015.
木村徹, 五十嵐康人,「大気降下物試料の放射能分析の品質管理」，Proceedings of the 16th
Workshop Environmental Radioactivity. KEK Proceedings 2015-4 , 23-27, 2015.

1.

2014
青山道夫, 東電福島原発事故後の海洋での放射能汚染の推移 （With English Abstract）.
YAKUGAKU ZASSHI, Journal of the Pharmaceutical Society of Japan. 2014.

1.

2013
青山道夫, 極低レベル放射能測定：少量表層海水および極深層海水試料への応用.放射線,39(1), 17-20, 2013.

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4.

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2012
青山道夫, 福島第一原子力発電所から放出された放射性セシウム同位体の北太平洋における総量と分布, 日本原子力学会誌, 54, 780-783, 2012.
青山道夫, 五十嵐康人, 1957 年から始まる気象研究所における環境放射能研究の歴史, 放射線, 38(3), 123-128, 2012.
青山道夫, 海洋に放出された放射性物質の長期地球規模での挙動, 検証！福島第一原発事故-放射性物質の実際と科学者たちの活動の記録（化学 4 月号別冊）. 4, 33-38, 2012.
青山道夫, 五十嵐康人, 広瀬勝己, 月間降下物測定 660 ヶ月が教えること-90Sr, 137Cs および Pu 降下量 1957 年 4 月～2012 年 3 月-. 科学, 82(4), 0442-0457, 2012.
2011
青山道夫, 福島原子力発電所事故による海洋汚染. 理大科学フォーラム, 11(通 巻 329 号),
42-45, 2011.
津旨大輔, 坪野考樹, 青山道夫, 廣瀬勝巳, 福島第一原子力発電所から漏洩した 137Cs の海
洋拡散シミュレーション(With English Abstract). 電力中央研究所報告書, V11002, PP.18, 2011.
2009
五十嵐康人, 放射性同位体分析, In 岩坂泰信, 西川雅高, 山田丸, 洪天祥（編）, 黄砂, 古今書院, pp81-87, 東京, 2009.
Igarashi, Y., Anthropogenic Radioactivity in Aerosol -A Review Focusing on Studies during the
2000s-, 保健物理, 44(3), 2009.
- 44 -

3.

1.
2.

1.
2.

1.
2.

猪股弥生, 五十嵐康人, 千葉長, 青山道夫, 廣瀬勝己, 再現期間を用いた降水時における高ガンマ線線量率の異常値判定. Radioisotopes, 58, 37-42, 2009.

2007
三浦誓也, 貝森優希, 工藤英嗣, 野呂幸男, 時枝隆之, 廣瀬勝己, 青森県における大気中クリプトン-85 濃度調査, 青森県原子力センター所報, 1, 43-44, 2007.
廣瀬勝己, 青山道夫, Kim、S.A. Kim, 太平洋の海水中のプルトニウムの挙動について, 日本
放射化学会誌 別冊, 8 supplement, 103, 2007.
2006
青山道夫, 大気と海洋深層における核実験起源放射性降下物の超低レベル放射能測定,
ADIOISOTOPES, 55, 429-438, 2006.
小村和久, 青山道夫, 秋山正和, 氏平祐輔, 大田裕二, 大塚良仁, 大西勝基, 櫻井敬久, 篠原
伸夫, Ge 検出器の使用状況に関するアンケート報告, RADIOISOTOPES, 55, 699-707,2006.
2004
廣瀬勝己, 化学トレーサーを利用した海洋変動の研究, 測候時報第 71 巻特別号, 71, S81S86, 2004.
五十嵐康人, 廣瀬勝己, 放射性降下物標準試料の調製を通じた 90Sr, 137Cs,プルトニウムなどの人工放射性核種分析の相互比較と参照値の決定, 放射化分析, No. 17, 39-42, 2004.
2002

1.

五十嵐康人, 90Sr

1.

1986
鈴木款, 井上久幸, 杉村行勇, 大気中の放射性クリプトンの分布, 地球化学, 20, 106-108,
1986.

1.
2.
3.

1.
2.
1.

2002.

と 137Cs を用いたダスト輸送過程と再飛散,

エアロゾル研究, 12, 17(4), 252,1978
葛城幸雄, 人工放射性物質, 天気, 25, 第 11 号, 20-23, 1978.
猿橋勝子, 海洋中の放射性物質を指標とした物質の移動・拡散, 水汚染の機構と解析－環境
科学持論－, 日本地球化学編, 東京, 産業図書, 271-287, 1978.
三宅泰雄, 猿橋勝子, 杉村行勇, 葛城幸雄, 金沢照子, 核兵器と地球環境の放射能汚染, 被爆の実相と被爆者の実情 －1977 年 NGO 被爆問題シンポジウム報告書, ISDA－ JNPC 編集,
朝日イブニングニュース社発行, 182-212, 1978.
1976
葛城幸雄, 日本における放射性降下物, 天気, 23, No. 7, 333-345, 1976.
杉村行雄, 環境のなかのプルトニウム, 自然, 3 月号, 1976.
1975
三宅泰雄, 島田利夫, 川村清, 重原好次, 地上大気のトリチウム含量と気象との関係, 日本
地球化学会 “環境問題特別号”, 131-135, 1975.

1.

1974
三宅泰雄, 葛城幸雄, 最近の 90Sr 降下物の起源, 天気, 21, 639-644, 1974.

1.

1971
葛城幸雄, 人工放射能による大気汚染, 気象研究ノート, 107, 258-282, 1971

- 45 -

1.
1.
2.

1970
三宅泰雄, 葛城幸雄, 最近の日本における放射性塵の降下, 天気, 17, 593-598, 1970.
1965
葛城幸雄, 日本における Cs-137 および Sr-90 降下について(I), 天気, 12, 323-328, 1965.
葛城幸雄, 日本における Cs-137 および Sr-90 降下について(II) "成層圏における人工放射性物質の滞留時間の推定", 天気, 12, 377-384, 1965.

1.

1964
三宅泰雄, 葛城幸雄, 金沢照子, 放射性降下物の現状と将来, 科学, 34, 142-148, 1964.

1.

1963
三宅泰雄, 放射性降下物の気象学の側面, 天気, 9, No. 1, 1962.

1962
1. 三宅泰雄, 放射性降下物の気象学的側面, 天気, 9, 1-6, 1962.
2. 三宅泰雄, 猿橋勝子, 放射能による水の汚染, 水利科学, 23, 1-14, 1962.
1.
1.

1.
2.
3.
4.
1.
2.
3.

1958
三宅泰雄, 猿橋勝子, 放射化学と海洋, 科学, 28, 510-513, 1958.
1956
三宅泰雄, 杉浦吉雄, 葛城幸雄, 1955 年 4 月旭川地方に降った放射性の落下塵,気象集誌, 第2 号, 34, 226-230, 1956.
1955
三宅泰雄, 放射能雨の性格, 測候時報, 22, 196-202, 1955.
三宅泰雄, 日本に降った人工放射性雨・雪[II]9 月～12 月, 天文と気象, 21, 1-8, 1955.
三宅泰雄, 杉浦吉雄, 亀田和久, 日本近海の海水放射能について, 昭和 29 年度海洋資源開発
調査報告（黒潮班）, 81-82, 1955.
三宅泰雄, 放射能雨の性格, 測候時報, 22, No. 7, 196-202, 1955.
1954
三宅泰雄, 日本に降った人工放射性雨(1954 年 5-7 月), 天文と気象, 20, 1-8, 1954.
三宅泰雄, 杉浦吉雄,亀田和久, ビキニ海域における人工放射能の分布とその海洋学的考察"
ビキニ灰とそれによる傷害 第三", 科学, 24, 601-605, 1954.
三宅泰雄, 有住直介, 磯野謙治, 田島英三, ビキニの灰とそれによる傷害（つづき）日本に降った放射能雨, 科学, 24, No. 8, 405-409, 1954.

- 46 -

年代・雑誌名不明
1.
2.
3.
4.

三宅泰雄, 放射能とその作用, 遺伝,（1975？）
三宅泰雄, 第 II 編 環境の放射能汚染と化学分析,（1970？）
三宅泰雄, 猿橋勝子, 杉村行勇, 海洋における放射性核種, 海洋化学講座, 第 6 巻, 海洋無機
化学, 第 4 章, 109-170,（1969？）
三宅泰雄, 田島英三, 日米放射能会議, 8-18,（1954？）- 47 -

7. History of the studies at the Meteorological Research Institute

表：研究の歴史- 48 -

研究の歴史
放射能調査研究費以前および、放射能調査研究費での研究
西暦

年度

1954

昭和29年

1955

S30

北太平洋観測（日・米・加）

1956

S31

北太平洋赤道海域観測（日・米・仏）

1957

S32

国際地球観測年事業費

1958

S33

国際地球観測年事業費

1959

予算項目

予算項目
課題名

課題名

課題名

課題名

課題名

放射化学分析(落下塵・降放射能調査研究費 水・海水中の放射性物質の研究)

深海水の循環に関する研究(以下深海水）S34

放射能調査研究費 放射化学分析

深海水

1960

S35

国際原子力機関委託研究 放射能調査研究費 放射化学分析

深海水

1961

S36

国際原子力機関委託研究 放射能調査研究費 放射化学分析

深海水

1962

S37

国際インド洋観測（日・米・放射能調査研究費 放射化学分析
ソ・英・仏）

深海水

1963

S38

国際インド洋観測（日・米・放射能調査研究費 放射化学分析
ソ・英・仏）

深海水

1964

S39

オレゴン州立大学との共同研究

放射能調査研究費 放射化学分析

深海水
海水中における放射廃棄物の化学的挙動の研究(以下廃棄物化学的挙動）1965

S40

放射能調査研究費 放射化学分析

海洋中の放射性廃棄物のモニタリングの測定法
深海水に関する研究（以下海洋廃棄物モニタリング）

1966

S41

放射能調査研究費 放射化学分析

海洋廃棄物モニタリング 深海水

廃棄物化学的挙動

1967

S42

放射能調査研究費 放射化学分析

海洋廃棄物モニタリング 深海水

廃棄物化学的挙動

1968

S43

放射能調査研究費 放射化学分析

海洋廃棄物モニタリング 深海水

廃棄物化学的挙動

1969

S44

放射能調査研究費 放射化学分析

深海水

廃棄物化学的挙動

1970

S45

放射能調査研究費 放射化学分析

深海水

廃棄物化学的挙動

1971

S46

放射能調査研究費 放射化学分析

深海水

廃棄物化学的挙動

課題名課題名

放射能調査研究費以前および、放射能調査研究費での研究
西暦

年度

予算項目

予算項目
課題名

課題名

課題名

課題名

課題名

1972

S47

放射能調査研究費 放射化学分析

放射性固体廃棄物の海洋処分に伴う鉛直拡散に関する研究(以下廃棄物鉛直
拡散）

1973

S48

放射能調査研究費 放射化学分析

廃棄物鉛直拡散課題名課題名

大気中におけるKr-85
およびH-3挙動と蓄積
廃棄物鉛直拡散に関する調査研究(以下Kr-85／H-3）
廃棄物鉛直拡散Kr-85／H-3
廃棄物鉛直拡散Kr-85／H-3数種の放射性核種 原子力施設に由来するの同時測定による深 放射性気体の広域分
海拡散の研究(以下 布に関する調査研究(以下放射性気体）
深海拡散）

1974

S49

放射能調査研究費 放射化学分析

1975
1976

S50
S51

放射能調査研究費 放射化学分析
放射能調査研究費 放射化学分析

1977

S52

放射能調査研究費 放射化学分析

1978

S53

放射能調査研究費 放射化学分析

深海拡散

放射性気体

環境における超ウラン元素の分布と挙動に関する研究（以下超ウラン元素）

1979
1980
1981
1982
1983
1984
1985
1986
1987
1988
1989
1990
1991

S54
S55
S56
S57
S58
S59
S60
S61
S62
S63
H1
H2
H3

放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費
放射能調査研究費

放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析
放射化学分析

深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散
深海拡散

放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体
放射性気体

超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素超ウラン元素放射性気体

超ウラン元素

放射性気体
放射性気体
放射性気体

超ウラン元素超ウラン元素超ウラン元素1992

H4

放射能調査研究費 放射化学分析

海洋における放射性核種の挙動に関する調査研究（以下海洋放射性
核種）

1993
1994
1995

H5
H6
H7

放射能調査研究費 放射化学分析
放射能調査研究費 放射化学分析
放射能調査研究費 放射化学分析

海洋放射性核種
海洋放射性核種
海洋放射性核種

放射能調査研究費以前および、放射能調査研究費での研究
西暦

年度

予算項目

予算項目
課題名

課題名

課題名

課題名

課題名

課題名

1996

H8

海洋環境における放射
大気圏の放射性核種の動性核種の挙動に関する放射能調査研究費 態に関する研究（以下大気研究（以下海洋環境放
圏放射性核種）
射性核種）

1997

H9

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

1998

H10

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

1999

H11

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

2000

H12

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

2001

H13

海洋環境における放射
大気圏の放射性核種の長
性核種の長期挙動に関
放射能調査研究費 期的動態に関する研究(以する研究(以下海洋環境
下大気圏放射性核種）
放射性核種）

大気中の放射性気体の実態把握に関する研究(以下放射性
気体)

2002

H14

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

放射性気体

2003

H15

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

放射性気体

2004

H16

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

放射性気体

2005

H17

放射能調査研究費 大気圏放射性核種

海洋環境放射性核種

放射性気体

2006

H18

海洋環境における放射放射性降下物の長期変動性核種の長期挙動に関
放射能調査研究費 と再浮遊に関する研究（以する研究(以下海洋環境
下放射性降下物）
放射性核種）

2007

H19

放射能調査研究費 放射性降下物

海洋環境放射性核種

2008

H20

放射能調査研究費 放射性降下物

海洋環境放射性核種

2009

H21

放射能調査研究費 放射性降下物

海洋環境放射性核種

2010

H22

放射能調査研究費 放射性降下物

海洋環境放射性核種課題名

放射能調査研究費以前および、放射能調査研究費での研究
西暦

年度

予算項目

予算項目
課題名

2011

H23(8月から）「大気を通じた人工放射性
核種の陸圏・水圏への沈放射能調査研究費 着およびその後の移行過程の解明研究（以下陸圏
水圏人工放射性核種）」

2012

H24

放射能調査研究費 陸圏水圏人工放射性核種

2013

H25

放射能調査研究費 陸圏水圏人工放射性核種

2014

H26

放射能調査研究費 陸圏水圏人工放射性核種

2015

H27(4月から）「人工放射性核種のバック
グラウンド大気監視と数値
放射能調査研究費
解析に関する研究（以下ＢＧ放射性核種）」

2016

H28

放射能調査研究費 ＢＧ放射性核種

2017

H29

放射能調査研究費 ＢＧ放射性核種

2018

H30

放射能調査研究費 ＢＧ放射性核種

課題名

課題名

課題名

課題名

課題名

課題名

表紙の図説明2018 年 3 月までの 90Sr および 137Cs 月間降下量
福島第一原発事故後の大気の放射能汚染の実態把握を継続しています。2011 年3 月の事故以降に確定できた90Sr および137Cs 月間降下量を対数軸で誤差と併せてプロットしています。
134Cs は事故直後には 137Cs とほぼ等量降下していますから、放射性セシウム全体ではこのプロットのほぼ倍量となります。ただし、90Sr、137Cs はどちらもおよそ 30 年の半減期で減衰しますが、134Cs はおよそ 2 年で半減します。図示した誤差は計測の統計誤差で、１シグマです。測定誤差は本来全ての測定値につき表示すべきですが、グラフが見づらくなるため、従来はあえて表示してきませんでした。また、数十年前のデータについては誤差データが伝えられていません。
福島第一原発事故以前に採取され、分析や前処理途上だった試料については、事故の汚染によって実験室環境および測定室環境、測定機器のバックグラウンドなどが大幅に上昇したため、蒸発濃縮工程においても試料の汚染（コンタミネーションと言います）の問題が発生します。そのため、観測値を求めること自体が困難になっていましたが、試料を汚染の水準の低い関西にて分析しこの問題の回避に努めました。しかし、細心の注意と努力にもかかわらず、2010 年の後半および 2011 年初のデータは欠測となってしまいました。
徐々に実験環境の除染や測定機器の入れ替えなどを実施して事故以前のきれいな実験環境を追求し、データを求めるように努めています。また、IAEA が実施した分析相互比較に参画するなどし、その精度の維持管理に努めていますので、現状のデータの信頼性は担保できていると考えています。

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8. A caption of the cover art
Monthly depositions of 90Sr and 137Cs before March 2018.
We have been monitoring the deposition amounts for the purpose of understanding the actual conditionof radioactive pollution after the accident of Fukushima Daiichi Nuclear Power Plant (FDNPP). The
monthly deposition amounts of both 90Sr and 137Cs, which were successfully fixed values after the accident
in March 2011, are shown in logarithmic axis with error widths. Since the deposition amounts of 134Cs were
nearly equal to those of 137Cs, the total amounts of radioactive cesium were approximately two times of thevalues in this figure. We would like to note that the half-lives of 90Sr and 137Cs are approximately 30 years,
however, 134Cs decays to half in about 2 years. The error widths are one sigma in statistical error ofmeasurements. Although the error widths of all measurement values were better to be shown, error values
used to be not shown on purpose to avoid busy figure in the former versions. In addition, error data before
several decades are not handed down.
For the analysis of samples collected before the accident of FDNPP and not measured then, significant
increase of background values were severe problem, because environments including laboratories andinstruments were polluted, and it was also problem that samples could have been polluted in the operationsof concentration or other preprocesses (contamination problem). We overcame the difficulty by carrying theanalysis out in western Japan (Kansai area) where the pollution level was low, however, the data during thelatter part of 2010 and the former part of 2011 were lacked in spite of our efforts.
We are making efforts to recover clean working environment as before the accident, by cleaning of theenvironments, exchanging instruments, etc. for the purpose of acquire the correct data. In addition, we
maintain the precision of the data, for example, by participating inter-comparison programs performed by
IAEA, so that, the credibility of the data are assured (please refer to the section: quality control of radioactive
analysis of atmospheric deposition samples ).

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