Environment International 85 (2015) 213–228

Contents lists available at ScienceDirect

Environment International
journal homepage: www.elsevier.com/locate/envint

Review article

An overview of current knowledge concerning the health andenvironmental consequences of the Fukushima Daiichi Nuclear Power
Plant (FDNPP) accident
Abubakar Sadiq Aliyu a,b,⁎, Nikolaos Evangeliou c, Timothy Alexander Mousseau d,e,Junwen Wu f, Ahmad Termizi Ramli ba

Department of Physics, Nasarawa State University Kefﬁ, P.M.B 1022 Kefﬁ, Nigeria
Department of Physics, Universiti Teknologi Malaysia, 81310 Johor Baru, MalaysiacNorwegian Institute for Air Research (NILU), Department of Atmospheric and Climate Research (ATMOS), Kjeller, NorwaydEnvironment and Sustainability Program, and Department of Biological Sciences, University of South Carolina, Columbia, SC 29208, United Statese
Faculty of Biotechnology, Chubu University, Kasugai, JapanfState Key Laboratory of Marine Environmental Science, Xiamen University, Xiamen 361005, Chinab

a r t i c l e

i n f oArticle history:
Received 7 July 2015
Received in revised form 15 September 2015
Accepted 17 September 2015
Available online xxxx
Keywords:
Biota
Human health
Fukushima nuclear accident
Radioactive contamination
Radioecology
Source terma b s t r a c t
Since 2011, the scientiﬁc community has worked to identify the exact transport and deposition patterns of radionuclides released from the accident at the Fukushima Daiichi Nuclear Power Plant (FDNPP) in Japan. Nevertheless, there still remain many unknowns concerning the health and environmental impacts of theseradionuclides. The present paper reviews the current understanding of the FDNPP accident with respect to interactions of the released radionuclides with the environment and impacts on human and non-human biota. Here,we scrutinize existing literature and combine and interpret observations and modeling assessments derived after
Fukushima. Finally, we discuss the behavior and applications of radionuclides that might be used as tracers of environmental processes. This review focuses on 137Cs and 131I releases derived from Fukushima. Published estimates suggest total release amounts of 12–36.7 PBq of 137Cs and 150–160 PBq of 131I. Maximum estimatedhuman mortality due to the Fukushima nuclear accident is 10,000 (due to all causes) and the maximum estimates for lifetime cancer mortality and morbidity are 1500 and 1800, respectively. Studies of plants and animals
in the forests of Fukushima have recorded a range of physiological, developmental, morphological, and behavioral consequences of exposure to radioactivity. Some of the effects observed in the exposed populations include thefollowing: hematological aberrations in Fukushima monkeys; genetic, developmental and morphological aberrations in a butterﬂy; declines in abundances of birds, butterﬂies and cicadas; aberrant growth forms in trees; andmorphological abnormalities in aphids. These ﬁndings are discussed from the perspective of conservation
biology.
© 2015 Elsevier Ltd. All rights reserved.

Contents
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Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Methods of literature search . . . . . . . . . . . . . . . . . . . . . .
Source term assessment . . . . . . . . . . . . . . . . . . . . . . . .
Atmospheric transport and deposition of Fukushima derived radionuclides .
Impacts of the FDNPP accident on humans . . . . . . . . . . . . . . .
Impacts of FDNPP accident on marine ecosystems . . . . . . . . . . . .
Impacts of FDNPP accident on terrestrial ecosystems . . . . . . . . . . .
Application of isotopes as tracers for environmental processes . . . . . .
Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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⁎ Corresponding author at: Department of Physics, Nasarawa State University Kefﬁ, P.M.B 1022 Kefﬁ, Nigeria.
E-mail addresses: abubakarsaliyu@nsuk.edu.ng, saabubakar2@live.utm.my (A.S. Aliyu).

http://dx.doi.org/10.1016/j.envint.2015.09.020
0160-4120/© 2015 Elsevier Ltd. All rights reserved.

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A.S. Aliyu et al. / Environment International 85 (2015) 213–228

Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction
The accident at the Fukushima Daiichi Nuclear Power Plant (FDNPP)
in Japan on March 11, 2011 resulted in a release of about 73 radionuclides (135 in total including radioactive progeny) (IRSN, 2012). It wasa direct consequence of a high-magnitude earthquake (~9.0), which occurred in the Paciﬁc Ocean near Japan's east coast, creating two massive
tsunamis that in turn struck Japan about 1 h following the earthquake(Akahane et al., 2012). The main electric power grid was disabled by
the earthquake and seawater intrusion to the backup power supply
caused power loss and the cooling systems of the four nuclear reactors
were disrupted, leading to increased pressure levels and the productionof hydrogen gas due to the extreme heating of the cooling water. During
the following days, hydrogen explosions released large amounts of radioactivity into the atmosphere. Fearing more severe damage and larger
releases, plant managers ordered seawater to be used as a cooling medium (IRSN, 2012). This highly radioactive water was discharged to the
Paciﬁc Ocean, with discharges (to both land and sea) ongoing morethan three years following the accident (Kumamoto et al., 2015).
Subsequently, several studies were conducted to document global
transport and deposition of the most important fallout radionuclides(e.g. Christoudias and Lelieveld, 2013; Evangeliou et al., 2013b;
Kristiansen et al., 2012) and the potential impacts to human populations(e.g. Evangeliou et al., 2014a; Ten Hoeve and Jacobson, 2012; WHO,2013) and to animals and plants (e.g. Aliyu et al., 2015a; Evrard et al.,2012; Garnier-Laplace et al., 2011; Hiyama et al., 2012; Møller et al.,2012; Møller and Mousseau, 2011b). Simultaneously, many researchgroups initiated experimental monitoring programs to assess potential
consequences in Europe, Asia, USA and Japan (e.g. Evrard et al., 2012;
Kim et al., 2012; Kinoshita et al., 2011; Kritidis et al., 2012; Truong
et al., 2012; MacMullin et al., 2012; Paatero et al., 2012; Pham et al.,2012; Povinec et al., 2012a).
Some of the major radioactive ﬁssion products released into the atmosphere and the ocean included 131I (t1/2 = 8.02 days), 134Cs (t1/2 =2.06 years), 137Cs (t1/2 = 30.07 years) and 90Sr (t1/2 = 28.78 years),
which were then dispersed globally by the prevailing winds (Masson
et al., 2011) resulting in the contamination of both terrestrial and marine ecosystems (Butler, 2011; Chino et al., 2011; NSCJ, 2011).
Many fundamental questions concerning the FDNPP accident stillneed to be resolved. For example, what was the exact amount of the released radioactivity? How is the environment being affected? What are
the predicted consequences for human populations? The present paper
aims to review the existing literature related to the FDNPP accident inorder to address these questions. Furthermore, we hope to stimulate
further discussion among the scientiﬁc and regulatory communities
concerned with nuclear safety and radiation protection (e.g., CTBTO,
IAEA, UNSCEAR, ICRP, etc.).
2. Methods of literature search
The discussion in this paper is based on keyword searches in Google
Scholar, Scopus and Web of Science. The keywords ‘source term’, ‘radioactive contamination’, ‘human health’ and ‘biota’ were combinedwith the primary keyword ‘Fukushima’ during literature searches.
The inclusion criteria for all publications were as follows: (i) they
discuss the release of radionuclides from FDNPP to the air or sea;(ii) they highlight the impacts of the FDNPP accident on either
human or non-human biota; (iii) they report the speciﬁc activities
of ﬁssion products at Fukushima prefecture and some parts ofJapan pre- and post- FDNPP accident. We conducted more than 200

224
224

literature searches to arrive at over 170 publications included inthis review. A few articles which considered the radioecological impacts of the Chernobyl nuclear accident and fallout from atomic
weapons tests were included for the purpose of comparisons withthe manifested impacts of the FDNPP accident.
3. Source term assessment
Following the FDNPP accident, there was much public concern
about the release of radioactivity into the atmosphere and ocean,and many independent groups attempted to estimate total releaseamounts. For the releases to the atmosphere, the Japanese government estimated the source terms for 131 I and 137Cs to be 160 PBq
and 15 PBq, respectively (RJG, 2011). The Nuclear Safety Commissionof Japan (NSCJ) estimated that the total amounts of 131I and 137Cs released into the atmosphere were 150 PBq and 12 PBq, respectively(NSCJ, 2011). Masson et al. (2011) estimated that 153 PBq 131I and
13 PBq 137 Cs were released, whereas other researchers reported
the total emission of 137 Cs to be as high as 13–35.8 PBq (Chinoet al., 2011; Stohl et al., 2012). In order to estimate the source termof 137Cs and 131I, Chino et al. (2011) adopted an estimation method
using an atmospheric dispersion model. Their results (~ 13 PBq)
were comparable to those of the Ministry of Education, Culture,
Sports, Science and Technology, Japan (MEXT, 2011) and have beenvalidated by subsequent analyses (Katata et al., 2012; Terada et al.,2012). In a more recent study, Katata et al. (2015a) modiﬁed thecomputational methodology to address some of the limitations ofprevious studies (Katata et al., 2012; Terada et al., 2012). They considered a new scheme of wet, dry, and fog-water parameterizationof both gaseous and particulate radionuclides in their simulation.
Observations from both land in different parts of Japan and the Pacific Ocean were coupled to this new model, leading to estimated total
release amounts of 151 PBq of 131I and 14.5 PBq of 137Cs. The reported
total release amount of 131I accords with the estimate of the Nuclear
Safety Commission of Japan (NSCJ, 2011) and lies within the range fortotal release amount of 131I by this review (150–160) PBq (Table 1).
However, Chino et al. (2011) Terada et al. (2012) and Katata et al.
(2015a) estimated the total release of 137Cs (13–15 PBq), based on an
inverse estimation of the source term by coupling Japanese environmental monitoring data with regional atmospheric dispersion simulations. In other words, they used data from Japanese stations only anda regional simulation domain and assumed constant radioactivity ratios
for the different radionuclides based on iodine and cesium concentrations in rain, snow and vegetation. Chino et al. (2011) noted that estimates of release amounts using this methodology are associated withan uncertainty of at least a factor of ﬁve.
Other research groups and national laboratories have also estimatedthe total release amounts of 131I and 137Cs from FDNPP, reporting releases that are higher than the Japanese estimates. For example, theFrench Institute of Radioprotection and Nuclear Safety (IRSN, 2012) reported releases of 137Cs of 20.6 PBq—25% higher than those of Katata
et al. (2015a). In addition, Stohl et al. (2012) used an inverse modelingtechnique employing data from the Comprehensive Nuclear Test-Ban
Treaty Organization (CTBTO) measurement network to address emission inventories of 137Cs and 133Xe. They reported 137Cs emissions of
36.7 PBq and 133Xe 15.3 EBq (×1018 Bq) based on a larger number of observations. The estimates of Stohl et al. (2012) found 2.5 times higher
emission rates compared to that of Chino et al. (2011); Terada et al.
(2012) and Katata et al. (2015a). However, the authors state that theirestimates are associated by an uncertainty of about 50%.

A.S. Aliyu et al. / Environment International 85 (2015) 213–228

215

Table 1
Source term of anthropogenic radionuclides in the atmosphere and the ocean (PBq).
Radionuclide

Global fallout
Atmosphere

131

I
132
I
137
Cs

Chernobyl
Ocean

Atmosphere

Fukushima
Ocean

–

1760b

–

950a

600a

85b

16b

600a
10.87a

380a
6.6a

1b
0.087b

–
–

–

133

Xe
All noble gases
90
Sr239,240
Pu

AtmosphereOcean

150–160c
35.8d
12–20e37 ± 18f
15,300 ± 7650f
12,134d0.01–0.14g(1–2.4) × 10–6g

–
4–27h

0.1–2.2g–a

UNSCEAR (2000).
NEA (2002).
Masson et al. (2011), NSCJ (2011), RJG (2011).
d
IRSN (2012).
e
IRSN (2012), Masson et al. (2011), NSCJ (2011), RJG (2011).
f
Stohl et al. (2012).
g
Povinec et al. (2012a).
hBailly du Bois et al. (2012), Kawamura et al. (2011), Tsumune et al. (2012).
bc

The Science Council of Japan (SCJ, 2014) compared various model
estimates for the total amounts of radioactive materials released
from Fukushima. They adopted the source term estimation of Stohl
et al. (2012) as a prior emission in order to compare the discrepancy
between ﬁeld observations and computed concentrations. The SCJ's
approach documented the total release of 137Cs between March 11
and 19 as 19.4 ± 3.6 PBq. This value lies between the estimates of
Terada et al. (2012) and Stohl et al. (2012), but they are closer tothose of the IRSN (2012). However, the SCJ noted some limitationsin their analysis. For example, only one Eulerian aerosol model(MASINGAR), developed by Tanaka and Chiba (2005), was used
and the bias of model transport could affect the source term estimate. In order to avoid such problems, various model simulations
with common experimental settings are usually used and the source
term estimates are compared. Another issue is the poor horizontal
resolution of the global Eulerian aerosol model; this can be eliminated using a regional chemical transport model and collecting hourly
observation data.
More recently, Saunier et al. (2013) used gamma dose rate observations to estimate the source term using inverse modeling. They
employed a Eulerian model developed by the IRSN and calculated
that 105.9 PBq of 131 I, 35.8 PBq of 132 I, 15.5 PBq of 137 Cs and12,134 PBq of noble gases were emitted. The details of the theoreticalformulation of the IRSN model are available in the literature (e.g.
Isnard, 2006). The authors compared their data with ﬁeld observations of activity concentration and surface deposition, and showedthat 137 Cs and 131 I emissions are likely realistic, 132 I are probably
underestimated, and noble gases are overestimated. Winiarek et al.
(2014) also used inverse modeling to reconstruct FDNPP releases.
Their technique can handle heterogeneous types of data in thesame inversion. They estimated that the released 137Cs ranged between 11.6 and 19.3 PBq, with an estimated standard deviation of
15–20%..
Finally, Chai et al. (2015) used an inverse emission estimation system based on a transfer coefﬁcient matrix (TCM) created using the
HYSPLIT Lagrangian dispersion model and a cost function that measuresthe differences between the model predictions and the actual air concentration measurements. The system was ﬁrst tested with identicaltwin experiments, in which pseudo observations are generated withthe same model used to estimate the transfer coefﬁcients. With thepseudo observations generated at the same location and time as the actual 137Cs observations to be assimilated later, the system is able to accurately recover the release rates and obtain better release estimates
than the use of the singular value decomposition (SVD) method. This inverse estimation approach is found to be robust and not overly sensitive

to the ﬁrst guess of the release. As for Fukushima, a penalty term isadded in the cost function to create smooth temporal changes. While
the temporal variations of the release rates at 6-h segments cannot be
fully retrieved using 24 h observations, the features at larger time scales
are well recovered. Given that only one ﬁrst guess estimation was taken
into account (Katata et al., 2015b), the results were close to those of
Katata et al. (2014).
Fission products were also released directly into the cooling water of
the FDNPP and were transported to the sea along the east coast of Japan.
Accurate estimation of the ﬁssion products directly discharged into thesea and their respective amounts is the basis for any impact assessment
on marine life. Therefore, since 2011, a number of groups have investigated the amounts of Fukushima-derived ﬁssion products thatwere directly discharged into the ocean (Buesseler, 2012; Buesseler
et al., 2012; Charette et al., 2013; Kumamoto et al., 2015; Tsumune
et al., 2012). The total amounts of 137Cs and 90Sr released into the Paciﬁc Ocean along the East Coast of Japan were estimated to be 1 to
~ 42 PBq (1 PBq = 1015 Bq) (Bailly du Bois et al., 2012; Kawamura
et al., 2011; Tsumune et al., 2012) and 0.1–2.2 PBq (Povinec et al.,2012a), respectively. Clearly, these estimates are associated withconsiderable uncertainty, which stems from the lack of available information from the FDNPP accident (Chino et al., 2011; Ohno andMuramatsu, 2014; Oza et al., 2013; Ten Hoeve and Jacobson, 2012).
Using a regional oceanic modeling system, ROMS, developed by
Shchepetkin and McWilliams (2005), Tsumune et al. (2012) estimated that the total 137Cs directly released was 3.5 ± 0.7 PBq from 26
March to the end of May, 2011. They noted that 137Cs moved southwards along the coast during the simulation period; then theeastward-ﬂowing Kuroshio current and its extension transported
137
Cs during May 2011, reducing 137Cs concentrations to less than
10 Bq L− 1 by the end of May 2011. Kawamura et al. (2011); Bailly
du Bois et al. (2012) and Tsumune et al. (2012) showed that thedischarged 137 Cs ranged between 1 and 42 PBq. Kawamura et al.
(2011) used two models driven by data obtained using the
WSPEEDI-II model. A limitation of this study was that it only considered deposited radionuclides from the aerosol, whereas releases via
direct discharge of the cooling waters were not taken into account.
Thus, it is likely that they underestimated the source term. The authors estimated the total amounts discharged into the ocean to be~ 11 PBq for 131I and ~ 4 PBq for 137Cs for the period from March 21
to April 30, 2011.
Table 1 shows the total amounts of major radionuclides released
from the FDNPP accident; these amounts are compared to global
fallout from Nuclear Weapons Testing and the Chernobyl accident.
Although uncertainty associated with total releases from the216

A.S. Aliyu et al. / Environment International 85 (2015) 213–228

FDNPP accident reaches 50% in certain cases, these data indicatethat the Chernobyl accident resulted in the largest (estimated) release of radionuclides, whereas the Fukushima accident resulted
in the largest direct release to the ocean. It is important to notethat FDNPP accident is argued to have resulted in largest release ofradionulclides into the ocean because the amount of radionuclides
released by the Dnjepr River into the Black Sea has not been estimated. Kumamoto et al. (2015) reported that the mean value forwater column inventories of the Fukushima-derived 134Cs and theatomic-bomb-testing-derived 137 Cs were nearly similar (1020 ±80 and 820 ± 120 Bq m − 2 , respectively) suggesting that the
FDNPP accident and nuclear bomb testing resulted in similar deposition amounts to the Paciﬁc Ocean.
A signiﬁcant fraction of the iodine released from the FDNPP accident remains as a gas (gaseous/total ratio: 77.2 ± 13.6% 23); hence,
it is difﬁcult to accurately measure total iodine. For this reason, theInternational Atomic Energy Agency (IAEA) has adopted the use of
the “iodine equivalent” calculated from measured activities of 134Cs
and 137Cs. Details of the methodology are presented in IAEA's International Nuclear Event Scale (INES, 2009). By using the IAEA/INES
method, the Nuclear and Industrial Safety Agency (NISA) of Japan estimated the release of “iodine-131 equivalent” due to the FDNPP

accident to be 770 PBq (NERHJ, 2011), which is considerably higher
than the 150 to 160 PBq reported in Table 1. There are still ongoing
efforts to computationally ascertain the source term of FDNPP nuclear accident with a great degree of accuracy. However, current estimates suggest total release amounts of 12–36.7 PBq of 137Cs and
150–160 PBq of 131I. These radioactive ﬁssion products have beentransported to and deposited at locations around the globe. The
two fundamental modes of transportation were the atmosphereand the sea.

4. Atmospheric transport and deposition of Fukushima
derived radionuclides
Once the radioactive fallout was injected into the atmosphere,
its trajectory was dispersed globally by the prevailing winds(Bolsunovsky and Dementyev, 2011; Huh et al., 2012; Leon et al.,2011; Manolopoulou et al., 2011; Pittauerová et al., 2011). Aerosol
species were removed from the atmosphere by wet and dry processes,
whereas noble gases were only lost due to radioactive decay.
Evangeliou et al. (2013b) used the atmospheric transport model
LMDZORINCA to simulate the global dispersion of radiocesium

Fig. 1. Total deposition of 137Cs (kBq m−2) during the ﬁrst 4 months after the FDNPP accident. The deposition is expressed in weekly intervals. The LMDZORINCA model was used in the
resolution of 2.50° × 1.27° over 19 sigma-p vertical layers.
Source: Evangeliou et al. (2013b)].

A.S. Aliyu et al. / Environment International 85 (2015) 213–228

217

Table 2
Fukushima-derived radionuclides detected in locations all over the world.
Location (reference)

Period(DD/MM/YY)

Maximum recorded activity concentrations in samples

National Institute of Radiological Sciences (NIRS) Chiba,
Japan (Ishikawa et al., 2014)
Regional scale ﬁeld work at Ibaraki, Fukushima, Chiba,
Tochigi, Saitama, and Tokyo Japan (Kinoshita et al.,2011)
Gunsan and Busan, South Korea (Kim et al., 2012)15/03/2011 to 1 year later

Air (Bq m−3): 132Te: 9.9, 131I: 16.2, 134Cs: 10.13, 137Cs:
10.27, 129mTe: 4.0, 136Cs: 123
Surface (Bq m−2): 137Cs: 9 × 105

Uljin and Ansan, South Korea (Hong et al., 2012)
Waste Isolation Pilot Plant, Carlsbad, New Mexico, USA(Thakur et al., 2012)
Krasnoyarsk (Russia) (Bolsunovsky and Dementyev,
2011)
Southwest USA (Wetherbee et al., 2012)
Alaska, USA (Wetherbee et al., 2012)
USA (Wetherbee et al., 2012)
La Spezia, Italy (Barsanti et al., 2012)
Athens, Greece (Kritidis et al., 2012)
Thessaloniki, Greece (Bolsunovsky and Dementyev, 2011;
Manolopoulou et al., 2011)
Nuclear Science Research Institute, Ibaraki, Japan (Saegusa
et al., 2013)
Hei Longjiang, Ji Lin, Qin Huangdao, Tian Jin and Beijingarea, China (Tuo et al., 2013)03/2011 to 05/2011

28/03/2011 to 31/05/2011

07/4/2011 to 28/04/2011
17/03/2011 to 04/04/2011
04/04/2011 to 03/05/2011
08/03/2011 to 05/04/2011
08/03/2011 to 05/04/2011
08/03/2011 to 05/04/2011
28/03/2011 to 12/04/2011
2011
24/03/2011 to 09/04/2011

01/03/2011 to 01/04/2011
02/04/2011 to 13/04/2011

University of Milano-Bicocca, Milan, Italy (Clemenza et al.,2012)

12/03/2011 to 05/04/2011

Ten air sampling sites, Czech Republic (Rulík et al., 2014)22/03/2011 to 17/05/2011

Vilnius, Lithuania (Lujanienė et al., 2013; Lujanienė et al.,2012a)
Pic du Midi observatory, France (Evrard et al., 2012)

23/03/2011 to 15/04/2011

Centre d'Etudes Nucléaires de Bordeaux Gradignan,
Université de Bordeaux, France (Perrot et al., 2012)
Comenius University Mlynská Dolina, Bratislava, Slovakia(Povinec et al., 2012b)

26/03/2011 to 14/05/2011

Spanish environmental radiological monitoring network
in Barcelona, Caceres and Seville, Spain (Baeza et al.,2012)
Stations in Novi Sad, Rimski Sancevi, Kisac, Melenci,
Krcedin, Subotica, Kumane, Zrenjanin, Bikovo, Nadalj,
Serbia (Bikit et al., 2012)
Hanoi, Dalat, and Ho Chi Minh City, Vietnam (Truong
et al., 2012)
Lancaster; Spark Bridge; Stirling, etc. (United Kingdom)(Beresford et al., 2012)
Cluj, Bihor, Bistriţa-Nŭsaŭd, Maramureş, and Arad, NW
Romania (Cosma et al., 2012)
Mt. Zeppelin global atmosphere watch station, NyÅlesund,Svalbard, Norway (Paatero et al., 2012)
Sacavém, Lisbon, Portugal (Carvalho et al., 2012)22/03/2011 to 10/04/2011

19/03/2011 to 25/05/2011

16/03/2011 to 17/04/2011

16/03/2011 to 20/04/2011

Surface (Bq m−2): 137Cs + 134Cs: ~3
Air (Bq m−3): 131I: 3.12 × 10−3, 134Cs: 1.19 × 10−3, 137Cs:.25 × 10−3
Surface (Bq m−2): 137Cs: 649, 134Cs: 802
Air (Bq m−3): 134Cs: 0.38 × 10−3, 137Cs: 0.45 × 10−3, 131I:
3.85 × 10−3
Melted snow (Bq L−1): 131I: 0.62, 137Cs: 0.075, 134Cs: 0.095
Surface (Bq m−2): 137Cs: 30–240, 134Cs: 2–46
Surface (Bq m−2): 137Cs: 16–27, 134Cs: ~55
Surface (Bq m−2): 137Cs: 1–45, 134Cs: 1–3
Surface (Bq m−2): 137Cs + 134Cs: ~0.3
Surface (Bq m−2): 137Cs: ~10
Air (Bq m−3): 131I: 497 × 10−6, 137Cs: 145 × 10−6, 134Cs:
126 × 10−6
Rain water (Bq L−1): 131I: 0.7
Evaporated water (Bq m−2): 137Cs: 1.4 × 104
Rain water (Bq L−1): 137Cs: 41
Air (Bq m−3): 131I: 1720 × 10−6, 134Cs: 247 × 10−6, 137Cs:
289 × 10−6, 136Cs: 23 × 10−6
(Tuo et al., 2013)
Rain water and vegetation (Bq kg−1): 131I: 2.68
Air (Bq m−3): 131I: ~4.2 × 10−4, 137Cs: ~9.4 × 10−5, 134Cs: 8
× 10−5Air (Bq m−3): 131I: 5.6 × 10−3, 137Cs: 0.72 × 10−3, 134Cs:
0.64 × 10−3
Air (Bq m−3): 131I: 3700 × 10−6, 137Cs: 1040 × 10−6
Surface (Bq m−2): 134Cs: ~2
Air (Bq m−3): 134Cs: 0.2 × 10−3, 137Cs: 0.7 × 10−3
Air (Bq m−3): 131I: 2.4 × 10−3,
Rain water (Bq L−1): 3.5
Surface (Bq m−2): 137Cs: 2–114, 134Cs: b100
Air (Bq m−3): 131I: 0.5 × 10−3,
137
Cs: 0.07 × 10−3, 134Cs: 0.0197 × 10−3
Rain water (mBq L−1): 131I: 1130, 137Cs: 12, 134Cs: 90
Air (Bq m−3): 131I: 3080 × 10−6, 137Cs: 690 × 10−6, 134Cs:
620 × 10−6, 132Te: 330 × 10−3
Air (Bq m−3): 131I: 2.7 × 10−3, 137Cs: 0.21 × 10−3
Rain water (Bq L−1): 1.8

27/03/2011 to 22/04/2011

Air (Bq m−3): 131I: 193 × 10−6, 134Cs: 33 × 10−6, 137Cs: 37 ×

01/04/2011 to 13/04/2011

10−6
Grass (Bq kg−1): dry matter for 131I: 55, dry matter for
137
Cs: 8
Rain water (Bq L−1): 131I: 1.69, 134Cs: 0.042, 137Cs: 0.067

28/03/2011 to 11/04/2011
12/03/2011 to 02/05/2011
17/03/2011 to 20/04/2011

Vancouver Island, Canada (Sinclair et al., 2011)
Countrywide, Finland (Leppänen et al., 2013)

20/03/2011 to 28/04/2011
19/03/2011 to 28/04/2011

Huelva Iberian Peninsula, Spain (Lozano et al., 2011)Çnaem, Istanbul, Turkey (Güngör et al., 2014)28/03/2011 to 07/04/2011
07/03/2011 to 12/09/2011

Mount Lulin, Taiwan (Huh et al., 2012)
Vienna Austria, Austria (Steinhauser et al., 2013a)

20/03/2011 to 29/04/2011
24/03/2011 to 13/04/2011

Budapest, Hungary (Homoki et al., 2013)
Monaco (Pham et al., 2012)08/03/2011 to 22/05/2011
23/03/2011 to 07/05/2011

Air (Bq m−3): 131I: 810 × 10−6, 134Cs: 659 × 10−6, 137Cs:
675 × 10−6
Surface (Bq m−2): 137Cs + 134Cs: ~1
Air (Bq m−3): 131I: 1.47 × 10−3, 137Cs: 134 × 10−3
Soil (Bq m−2): 131I: 1.03, 134Cs: 0.65, 137Cs: 0.74
Air (Bq m−3): 133Xe: 30–70
Air (mBq m−3): 131I: 10.6, 134Cs: 0.39, 137Cs: 0.41, 136Cs: 28,
129
Te: 129, 129mTe: 234, 132Te: 51
Higher concentrations were observed in Southern Finland
than in the North
Air (mBq m−3): 131I: 3.69, 132I: 0.162, 134Cs: 0.88, 137Cs: 0.95
Air (mBq m−3): 131I: 1.03, 134Cs: 0.25, 137Cs: 0.23
Rain water (mBq m−3): 134Cs: 4.1, 137Cs: 4.7
Air (mBq m−3): 131I: 0.06, 134Cs: 0.028, 137Cs: 0.030
Air (Bq m−3): 131I: 1.2 × 10−3
Rain water (Bq L−1): 131I: 5.2
Air (μBq m−3): 131I: 96, 137Cs: 102, 134Cs: 102
Surface (Bq m−2): 137Cs + 134Cs: 1–2
Air (μBq m−3): 131I: 354, 134Cs: 30, 137Cs: 37
Deposition rate (mBq m−2 d−1): 134Cs: 13.7, 137Cs: 19.1
(continued on next page)218

A.S. Aliyu et al. / Environment International 85 (2015) 213–228

Table 2 (continued)
Location (reference)

Period(DD/MM/YY)

Maximum recorded activity concentrations in samples

Bremen, Seefeld and Schifford, Germany (Pittauerová
et al., 2011)12/03/2011 to 14/05/2011

Darwin, Townsville, Melbourne, Perth and Cacos Island,
Australia (Carpenter and Tinker, 2012)01/40/2011 to 30/04/2011

Institute of Nuclear Physics Krakow, Poland (Mierelski
et al., 2014)21/03/2011 to 30/10/2011

River sediment (Bq kg−1): 131I: 0.113, 137Cs: 2.67, 134Cs:
0.08
Soil (Bq kg−1): 131I: 0.68, 137Cs: 3.00, 134Cs: 0.069
Rain water (Bq kg−1): 131I: 0.31, 137Cs: 0.02, 134Cs: 0.02
Air (Bq m−3): 133Xe: 12 × 10−3
No Fukushima-derived radio iodine and cesium was
detected in Australia, 133Xe was only detected in Darwin
Air (mBq m−3): 131I: 5.73, 134Cs: 0.461, 137Cs: 0.436

during the ﬁrst four months after the accident. A weekly intervals of
the total deposition of 137Cs is presented in Fig. 1.
Global scale simulations by Christoudias and Lelieveld (2013) andEvangeliou et al. (2014b) showed that the plume generated by the
FDNPP was initially affected by northerly and easterly winds, which
gradually reached the USA via the North Paciﬁc and Alaska.
Table 2 summarizes observations of Fukushima-derived ﬁssion
products all over the world, including air concentrations (in Bq m−3),
surface deposition densities (in kBq m− 2) and concentration in both
fresh and marine water (in Bq L− 1). The data in Table 2 can be used
for worldwide health assessments based on ﬁeld observations.
Based on observations from the CTBTO network (Christoudias andLelieveld, 2013), fallout from FDNPP reached Eastern Siberia after fourdays, Alaska and the West Coast of the USA after 8–10 days (Thakur
et al., 2012), the east coast after 11–12 days, the North Atlantic(i.e., Iceland) after about 13 days, Europe after 14–17 days, and Asia
after 18–20 days. This is much longer than the ~10-day period it took
for the Chernobyl fallout to spread across the globe (Harrison et al.,1993). This is mainly because in Chernobyl there was a tremendous explosion and nuclear ﬁre, which injected radionuclides directly into thetroposphere (Evangeliou et al., 2013a) resulting in faster transport,
whereas in Fukushima the releases occurred inside the planetary
boundary layer (PBL). The radioactive plume from Fukushima wastransported over the Paciﬁc Ocean; over the ocean, variation in theboundary layer is low and this may have resulted in the prolonged
time for travel of the Fukushima radioactive cloud around the world.
The variability in surface temperatures over the ocean are small compared to those over land, and this results in mixing of larger air masses
and hence longer transport times for aerosols that are transported overocean (Aliyu, 2014; Jilani, 2009; Sorbjan, 1989). In other words, the turbulence over the ocean is relatively low compared to terrestrial groundsurfaces where the presence of trees, mountains and other rough surfaces increases the exchange of mass and momentum between the
ground surface and the atmosphere (Jilani, 2009; Marshall and Plumb,1965).
Observations in Table 2 are very close to those of the CTBTO andagree with the estimation that the plume arrived in Europe more than
10 days after the FDNPP accident. Outside Japan, ﬁssion products weredetected in Korea, Vietnam, Taiwan, USA, Canada, and Europe (CTBTO,
2011; Huh et al., 2012; Kim et al., 2012; Leon et al., 2011; Masson
et al., 2011) (Table 2). In Europe, the ﬁrst traces of the radioactive fallout
were detected on 19 March 2011 (slightly more than one week after theaccident), while the ﬁrst peak of activity level was observed between 28
and 30 March. Measurements suggest that airborne activity levels are
currently of minimal concern for public health in Europe (Masson
et al., 2011). The activity concentrations of radionuclides in the plumedecreased exponentially with downwind distance (Hsu et al., 2012)
due to dry and wet scavenging and subsequent deposition along thepath of dispersion (Aliyu et al., 2015b).
In Japan, air concentrations of 134Cs and 137Cs in the city ofFukushima were still elevated (613 μBq m−3 and 829 μBq m−3, respectively) 7 months after the accident (Toyoshima et al., 2011). Intensive
monitoring was conducted on smaller Japanese Islands, which included

45 measuring stations (Hirose, 2012). These measurements suggested
that the FDNPP accident affected areas in the central and eastern partsof Honshu Island, although ﬁssion products were deposited all overJapan. Hirose (2012) also used daily deposition of radionuclides inrain water measured by Prefectural Governments in Japan to estimate
the atmospheric resident time of Fukushima-derived 137Cs within
300 km from the FDNPP site: ~10 days.
Earlier studies (e.g. Buck, 2011; Ten Hoeve and Jacobson, 2012;
Thakur et al., 2012) showed that radionuclides were transported bywinds over the Paciﬁc Ocean towards the West Coast of the USA. Fig. 2
depicts the atmospheric dispersion pattern of the ﬁssion products at different heights of the PBL during the ﬁrst 10 days of the FDNPP, using
single-particle trajectories employed by the HYSPLIT model (Draxler,
2004; Draxler and Hess, 1997). The start time of release of particles
from the FDNPP was 12 March 2011. The lower trajectories [100 and
500 m above ground level (AGL)] circulated along the East Coast ofJapan towards the northern part of the country. The particles at higher
altitudes (1 and 1.5 km AGL) were transported across the Paciﬁc
Ocean to the West Coast of North America. Thakur et al. (2012) showedthat 19% of the total fallout was deposited on Japan and only 2% reachedother land areas of Asia and North America, while the remainder was
deposited over the Paciﬁc Ocean and the Arctic. In a recent study,
Evangeliou et al. (2015) estimated that about 23% of the released 137Cs
was deposited over Japan and 76% of the radiocesium was depositedover the North Paciﬁc and North Atlantic oceans. From the remainder,
163 TBq was deposited in North America, mostly in the USA (95 TBq),
Canada (40 TBq) and Greenland (5 TBq), about 14 TBq in Europe (mostly in European Russia, Sweden and Norway), 47 TBq in Asia (mostly in
Asian Russia, Philippines and South Korea), and traces deposited overAfrica, Oceania and Antarctica. Considering that the plume followed anorthward direction before it arrived in the USA and then in Europe,
about 69 TBq was deposited over the Arctic (see Fig. 2).
Deposition of the Fukushima-derived radionuclides over land are directly linked to their atmospheric release. Surface deposition levels areassociated with the risk of internal radiation exposure via inhalation,
whereas aerosol species are important when considering ingestion
pathways. Soil around the FDNPP and neighboring prefectures wascontaminated with 137Cs and 134Cs at levels of more than 105 and
104 Bq m−2, respectively (Burns et al., 2012; Yasunari et al., 2011),
and other rare refractory elements might also have been important.
Kinoshita et al. (2011) reported that the surface deposition of 137Cs incentral and eastern Japan ranged from 0.4 to 900 kBq m− 2, which is
0.06–130 times the cumulative deposition (100–1000 Bq m−2) from atmospheric nuclear bomb testing (Kinoshita et al., 2011). It is estimatedthat a total of more than 5.6 PBq of 137Cs was deposited over Japan(Yasunari et al., 2011). For example, soil concentrations of 134Cs and137
Cs in the Iitate Village area located 25 km northwest of the FDNPP
were 62.4 and ~ 73 kBq kg−1, respectively, and the topsoil concentrations of 131I and 137Cs 23 km south of the Fukushima Reactor Unit 1
were reported to be 140 kBq m− 2 and 80 kBq m− 2, respectively(Garnier-Laplace et al., 2011).
Radiation dose measurements can be used to extrapolate the level ofsurface contamination due to the FDNPP accident in Japan. For instance,

A.S. Aliyu et al. / Environment International 85 (2015) 213–228

219

Fig. 2. Calculated trajectories for air masses released at the site of the Fukushima Daiichi reactor (Aliyu et al., 2015a).

the environmental external radiation dose rate before the nuclear accident in the Fukushima region was approximately 0.05 μSv h− 1
(Kinoshita et al., 2011). The external radiation dose rates in prefecturalareas due to the FDNPP accident have been estimated to be 1.14 μSv h−1
in Naka-Dori, 4.57 μSv h−1 in Iitate, 0.02 μSv h−1 in a region between
northern Ibaraki and eastern Saitama, and 0.23 μSv h−1 in southern
Ibaraki and northern Chiba prefectures (Kinoshita et al., 2011). Møller
et al. (in pressb) reported that dose rates at 400 locations around the
Fukushima region varied from 0.5 to 38 μSv h− 1. This indicates thatlarge areas in eastern and northern Japan were strongly contaminatedby 137Cs, whereas western regions were largely shielded by mountainous topography. However, similar to Chernobyl, deposition was highly
heterogeneous even within the most contaminated areas (Møller
et al., in pressb, in pressc).
The data presented in Table 2 can also be used to predict the doseand risk in Europe, North America and Asia using empirical environmental data rather than estimates based on mathematical models that
may result in over- or underestimation of the impacts of the accident.
The question of how ﬁssion products from FDNPP accident weretransported and deposited in locations within and outside Japan willcontinue to be of interest as it is the basis for any radioecological impacts
assessment involving Fukushima-derived long-lived radionuclides
which could, even at trace quantities, bioaccumulate in plant and animal
tissues, thereby increasing the risk of radiation induced effects in the exposed populations.
5. Impacts of the FDNPP accident on humans
Given that the potential health effects of the FDNPP accident will
likely be of great social and economic importance for decades to come,modeling of predicted doses can be a useful tool for a preliminary assessment of the health risks. Table 3 illustrates the predicted health effects of the FDNPP accident on local and global populations. In allcases, a Linear Non Threshold (LNT) model of human exposure wasused to calculate potential radiological health effects. This model assumes that the disintegration of each radionuclide has the same probability of causing cell transformation, and that each transformed cell has
the same probability of developing a cancer tumor. Although the LNTmodel has been employed extensively in radiation safety (NRC, 2006;
UNSCEAR, 2010), several arguments concerning its validity and response at low doses still remain unresolved (Cuttler, 2010; Tubiana
et al., 2009). Traditional epidemiological studies have only considereddoses above 100 mSv as causing statistically signiﬁcant increases in stochastic cancer risk, although there are a growing number of studies suggesting possible effects at doses well below 100 mSv (e.g. Brenner andSachs, 2006; Brenner et al., 2003; Eidemüller et al., 2010; Mathews etal. 2013). On the other hand, proponents of the LNT model claim thatthe difﬁculty in detecting and attributing a small number of cancers to
low doses does not indicate that there is an absence of risk (Hoffman
et al., 2011). Note that the NRC (Nuclear Regulatory Commission) andthe WHO (World Health Organization) accept the LNT hypothesis forsuch assessments.
Ten Hoeve and Jacobson (2012) estimated that there would be an
additional 130 (15–1100) and 180 (24–1800) lifetime cancer related
mortalities and morbidities related to the Fukushima radiological disaster, respectively, with over 90% of the projected impact occurring inJapan. However, Beyea et al. (2013) argued that the failure by Ten
Hoeve and Jacobson (2012) to consider the long-term dose estimate
from radiocesium in their studies may lead to an underestimation of
the global effects of the accident. By including a more realistic

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A.S. Aliyu et al. / Environment International 85 (2015) 213–228

Table 3Cancer related morbidities (cases) and mortalities (deaths) due to the FDNPP accident.
The table shows number of individuals that are expected to suffer from all solid cancers.
Reference

Morbidities(incidents)

Mortalities(deaths)

Collective dose inJapan man-Sv

IRSN (2012)UNSCEAR (2013)
IPPNW (2013)
Von Hippel (2011)
Ten Hoeve and Jacobson (2012)
Evangeliou et al. (2014b)

70–250–––
24–1800
160–880–
5000
10,000750–1500
15–1100
730–1260–
48,00095,000–––population perspective, Beyea et al. (2013) suggested that the estimate
for the number of future cancer mortalities due to the FDNPP accident isprobably closer to 1000 than 125. Other health estimates for
Fukushima-related cancers were made by Von Hippel (2011) who estimated the cancer mortalities to be 770–890 and the highest external
doses outside the 20-km evacuation area to be between 20 and 40 mSv.
Evangeliou et al. (2014a) estimated that the highest doses would
rise to between 50 and 100 mSv. Using the LNT model, in the northernmost area of the FDNPP (with about 300,000 inhabitants) they estimated the lifetime mortality risks to be 750–1500 individuals. The IRSN
(2012) estimated that a population of 21,100 received more than
16 mSv; a population of 3100 received more than 50 mSv, and anothergroup of 2200 individuals received 100–500 mSv near Fukushima. Assuming the LNT hypothesis, an exposure of 100 mSv will increase lifetime cancer risk by 1% (Cuttler, 2010), giving rise to 70–250 excess
cancers depending on the precise amount of radiation received byeach grouping.
On the other hand, several studies have reported much higher values
for projected cancer mortalities in Japan. For example, the most detailed
model, used in a report published by IPPNW (2013) in late March 2013,
estimated a collective effective dose of 95,000 mSv, whereas UNSCEAR(2013) estimated 48,000 mSv. In terms of fatal cancers (mortalities),
the UNSCEAR estimates imply (via the LNT theory) that ~5000 people
in Japan will die in the future from Fukushima-related cancers. This ﬁgure is obtained by applying a fatal cancer risk of 10% per Sv; the
UNSCEAR report, like the previous WHO reports, no longer applies a
DREFF (dose and dose-rate effectiveness factor) of 2 to risk estimates.
However, the methodologies and assumptions used in the new
UNSCEAR report will require scrutiny before a ﬁnal conclusion can be
drawn. It is noteworthy that the UNSCEAR report adds that “The collective effective dose to the population of Japan due to a lifetime exposure
following the FDNPS accident is approximately 10–15% of the corresponding value for European populations exposed to radiation following the Chernobyl accident. Correspondingly, the collective absorbed
dose to the thyroid was approximately 5% of that due to the Chernobyl
accident.”
In a recent study, Evangeliou et al. (2014b) considered numerous
ﬁssion products (131 I, 134 Cs, 136Cs, 137 Cs, 129mTe, 132 Te, 95Nb, 90Sr,
110m
Ag, 99Mo, 241Am, 238Pu, 239–240Pu, 241Pu, 242Cm and 243–244Cm)in their assessment. The authors also took into account several arguments reported after the Ten Hoeve and Jacobson (2012) estimates
were issued, pointing out shielding effects and the exponential decreaseof the deposition densities of the studied radionuclides as a result ofvertical migration in soil. For a discussion of vertical migration of deposited radionuclides from fallouts see, for example, Almgren and Isaksson(2006); Nakanishi et al. (2014), and Takahashi et al. (2015). Evangeliou
et al. (2014b) estimated that solid cancer incidences and mortalities
from Fukushima would be 160 to 880 and 110 to 640, respectively,
which are close to previous estimations. By adding thyroid cancers,
the total estimate rises from 230 to 850 for incidents and from 120 to
650 for mortalities (see Figs. 3 and 4).
Estimated fatalities due to workers' exposure and mandatory evacuation have been reported to be around 610, increasing total estimatedmortalities to 730–1260 (Evangeliou et al., 2014b). These estimatesare 2.8 times higher than those previously reported for radiocesium
and 131I, and 16% higher than those reported based on radiocesium only.
Prefectural Governments in Japan have adopted the use of highly
sensitive ultrasound systems to scan for radiation-induced thyroid cancer due to the Fukushima accident (Tronko et al., 2014). Tronko et al.
(2014) reported that, by the ﬁrst quarter of 2014, about 80% of the target population (36,000 of the residents aged up to 18 years) were examined. Out of the examined population, 75 cases of suspected malignancy
were identiﬁed. Thirty-four patients have received surgery; pathological diagnoses include one benign tumor, one suspected to be poorly differentiated thyroid carcinoma, and 32 papillary thyroid carcinomas.
This is arguably an alarming case of prevalence considering the examination period of less than 3 years following exposure. However, thestudy was inadequately controlled and failed to effectively assess thyroid cancers in preschool children in the broader unexposed population,a point that was used by Tronko et al. (2014) to argue that the highprevalence was not due to the FDNPP nuclear accident. Arguably, thehighly sensitive equipment that is in use in Fukushima might have resulted in this high prevalence of thyroid anomalies that would not normally have been detected. However, Tronko et al. emphasized that
additional analysis will need to address thyroid cancer cases that may
appear in the coming years, once the period of latency has passed.
They noted that attention should be paid to thyroid dose reconstruction,
age at exposure and diagnosis, and whether there will be a “harvesting
effect,” which is a spike in cases after introduction of intensive screening
programs.
Tokonami et al. (2012) and Hosoda et al. (2013) attempted tomeasure and estimate the thyroid doses in evacuees from the
Fukushima nuclear accident site, respectively. Tokonami et al.
(2012) conducted screening tests of 62 residents and evacuees
using a portable Na(TI) spectrometer to measure 131 I activity inthe thyroid. These measurements were conducted a month after
the FDNPP accident (12 to 16 April 2011). The age of examinees
ranged from b1 to 80 years; 17 of them were from Tsushima District of Namie Town, a region which is severely contaminatedwith radioactive materials released from the crippled FDNPP.
Forty-ﬁve of the examinees were evacuees from coastal areas, including Minami-soma City located to the north of FDNPP. The median (and maximum) thyroid equivalent doses were estimated as
4.2 (23) mSv and 3.5 (33) mSv for children and adults, respectively,
which are much smaller than the mean thyroid dose in the Chernobyl accident (490 mSv in evacuees). Hosoda et al. (2013) used themeasurement data for 131 I from the earlier study by Tokonami et al.
(2012) to estimate the external exposure of thyroid to 131I based on
the accumulated 134Cs among evacuees. The maximum values of
131 134
I/ Cs activity ratio corresponding to thyroid uptake factors of 0.3,
0.1 and 0.03 were evaluated to be 0.9, 2.6 and 8.7, respectively. The
maximum value of the 131I/134Cs activity ratio was used to obtain themost conservative thyroid equivalent dose estimation. The maximum
internal exposure of the thyroid to 131I on the basis of cumulative134
Cs measured by the whole-body counter was estimated to be
18 mSv. This value is higher than the maximum thyroid doses for children and adults reported in the earlier study.
Computer models are currently used to predict the human deaths
and health effects of the FDNPP nuclear accident based on the LNT
hypothesis. More ﬁeld research is needed in order to investigate
the effects of the nuclear accident in exposed human populations.
This project should involve continuous monitoring of the exposed
population as radiation effects on humans may take longer before
they are expressed especially for cancers which can have latency periods of decades. The study of non-human biota, which often havemuch shorter lifespans (compared to humans), provides opportunities to investigate the biological consequences of radiological contaminants on a short timespan and this approach has been fruitfulfor study of the consequences of both the Fukushima and Chernobyl
disasters.

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221

Fig. 3. Excess lifetime morbidity (incidence) and mortality (death) risk of all solid cancers from the uptake of 134Cs, 136Cs, 137Cs 133Xe, 129mTe, 132Te, 95Nb, 90Sr, 110mAg, 99Mo, 241Am, 238Pu,
239–240
Pu, 241Pu, 242Cm and 243–244Cm in Japan from all internal and external exposure pathways using the Linear Non Threshold models for a lifetime of 89 years. The calculations were
performed using the LMDZORINCA model for a spatial resolution of 0.45°–0.51°.
Source: Evangeliou et al. (2014b).

6. Impacts of FDNPP accident on marine ecosystems
Contamination of the marine environment following the FDNPP accident represents the most important anthropogenic radioactive release
into the sea ever recorded, on par with fallout from atomic bomb testing. The radioactive marine pollution came from atmospheric fallout
onto the ocean and direct release of contaminated water from the plant.
In the immediate vicinity of the plant, Cs levels of the ocean surface
at the point of discharge showed up to 103 times higher activities than
had been reported by Aoyama and Hirose (2004) before the Fukushima
accident (Aoyama and Hirose, 2004), with a peak of 68 Bq m−3 on 6
April 6, 2011 (Buesseler et al., 2011). Dilution of the contaminants occurs rapidly due to ocean mixing, particularly in the energetic coastal
waters off Japan where the Oyashio waters move south and interact
with the rapidly ﬂowing and offshore meandering of the Kuroshio Current. These currents, tidal forces and eddies mix the waters rapidly offshore. As a consequence, one month after the accident the 137Cs level
decreased by a factor of 103.

Honda et al. (2012) reported that the 137Cs activity in surface seawater ranged from 0.004 to 0.284 Bq kg− 1 in the western North Paciﬁc from 14 April to 5 May. Tsumune et al. (2012) argued that if
no additional releases occur in the future, the effect of the 137 Cs
may be smaller than that of global fallout in the North Paciﬁc. The
timescale of transport to other basins is several decades. Thus, the effect of the 137Cs release on other oceanic basins might be negligible.
With respect to dose effects on humans, Buesseler et al. (2011) calculated that the dose due to direct exposure during human indirectcontact with the water was 0.1 μSv day− 1 when the 137Cs level approaches 102 kBq m− 3, which is lower than the average dose fromall sources to the Japanese population of about 0.17 μSv day− 1.
Cesium-137 concentrations in suspended solids (SS) from surface
and subsurface waters off the western North Paciﬁc have been measured by Honda et al. (2012); it ranged from 5.92 to 32.42 Bq kg−1 ona dry weight basis (kg-dw−1). A similar pattern was observed for134
Cs. Thus, it is likely that SS were contaminated with radionuclides
from the FDNPP (Honda et al., 2012). The SS-seawater distribution

Fig. 4. Excess lifetime morbidity (incidence) and mortality (death) risk of thyroid cancer from the uptake of 131I from all internal and external exposure pathways using the Linear Non
Threshold model for a lifetime of 89 years. The calculations are performed using the LMDZORINCA model for a spatial resolution of 0.45°–0.51°.
Source: Evangeliou et al. (2014b).

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A.S. Aliyu et al. / Environment International 85 (2015) 213–228

coefﬁcient (Kds: Bq kg-dw−1 for SS/Bq kg−1 for seawater) was further
estimated to be 200–4400, with an average of 2100. This average Kds
is comparable to the suggested Kds for the open ocean: 2000 (IAEA,
2004).
In terms of potential biological impacts, radiation doses in marineorganisms are generally dominated by the naturally occurring radionuclides 210Po (an alpha emitter) and 40K, even when organisms are exposed to anthropogenic radioactivity discharged to coastal waters(Aarkrog et al., 1996). Fukushima-derived radiocesium was detectedin zooplankton and mesopelagic ﬁsh (Buesseler et al., 2012), with the137
Cs concentration in zooplankton off the western North Paciﬁc coast
at 13.46–71.46 Bq kg-dw−1, which is signiﬁcantly higher than that of
the southern Baltic Sea and the northern Adriatic Sea one month after
the Chernobyl accident: 0.4–5 Bq kg-ww− 1 (Knapinska-Skiba et al.,
2003; Marzano and Triulzi, 1994). In contrast, the 137Cs concentrationin zooplankton around Japan was less than 0.1 Bq kg-ww− 1 during
the past decade (Kaeriyama et al., 2008; Kasamatsu and Ishikawa,
1997). Thus, the observed concentrations were two orders of magnitude higher than before the accident. Meanwhile, the concentration factor [CF: i.e. the ratio of the Cs concentration of zooplankton to that ofambient seawater (Bq kg-ww-1/Bq kg−1)] was calculated to be 200–
840, which is an order of magnitude higher than the 10–100 of previous
observations (IAEA, 2004; Kaeriyama et al., 2008; Kasamatsu andIshikawa, 1997).
Buesseler et al. (2012) suggested that radiation risks of 137Cs derived
from the FDNPP to marine organisms and human consumers of seafood
were well below those from natural radionuclides. At the same time, thetotal cesium levels in demersal (bottom-dwelling) ﬁsh off Fukushima
was investigated (Buesseler, 2012), where the cesium levels in 40% ofﬁsh were above the new regulatory limit (100 Bq kg−1 wet) set by
the Japan Ministry of Agriculture (JMA, 2011). However, this is lower
than the new regulatory limit of cesium in ﬁsh products in four prefectures (Iwate, Miyagi, Ibaraki and Chiba).
The effect of the FDNNP accident on marine products in Fukushima
was also assessed by Wada et al. (2013). The authors concluded thatthe total activity concentrations of radiocesium in marine products
have decreased signiﬁcantly. However, the authors noted that higher
activity concentrations have been observed in shallower waters southof the FDNPP, and long-term monitoring of marine products fromaround the FDNPP is needed before restarting coastal ﬁsheries (Wadaet al., 2013). Chen (2013) reported the radioactivity level in ﬁsh
products from other prefectures adjacent to Fukushima, based on monitoring data of July 2013. Activity concentrations of 134Cs, and 137Cs varied from 0.1 to 338 Bq kg− 1 (average: 11 Bq kg− 1) and from 0.1 to699 Bq kg−1 (average: 18 Bq kg− 1), respectively. Three years later,
the dose rate to the most affected ﬁsh species near the FDNPP haveremained above baseline levels; a theoretical human consumer of
50 kg of ﬁsh caught a short distance (3 km) from the FDNPP in 2013
would have received a total committed effective dose of approximately
0.95 mSv year−1, with FDNPP contributing ~14% of total committed effective dose (Johansen et al., 2014).
Numerous studies have assessed the impacts of the Fukushima
accidents on the marine and terrestrial environments in both Japanand other countries (Aliyu et al., 2015a; Buck, 2011; Chen et al.,2014; Kim et al., 2012; Periáñez et al., 2013; Sin et al., 2014; Wu
et al., 2013). For instance, Aliyu and co-authors concluded that, within Japanese Coastal areas, the accident did heavier damage to marine
bionts compared to terrestrial ﬂora and fauna. Even though many of thestudies outside Japan have demonstrated that the levels of radiocesium
in seafood were not likely to have discernible health effects (Chen et al.,2014), investigations of the marine environment have continued inorder to address public concern about the impact of the accident.
Buck (2011) reported that the radiation in the ocean will be diluted
along the Japanese coast; however, exposed marine organisms and contaminated tsunami debris would likely reach the USA. Time-series measurements of 134Cs and 137Cs in seawater showed that the ﬁrst arrival of

the Fukushima derived radio-cesium in a location 1500 km west of British Columbia, Canada was in June 2012, which is 1.3 years after the accident (Smith et al., 2015). Wu et al. (2013) showed that the activity
concentration of 137Cs in the China Sea ranged from 0.75 ± 0.07 to
1.43 ± 0.08 Bq m−3, which is below the regulatory standard in China.
More recent reports have suggested that Fukushima-derived cesium arrived off the coast of British Columbia, Canada in February 2015, and off
the coast of Del Mar., CA, USA, in April 2015 (KO Buesseler, 2015 pers.
comm., 3 September).
Strontium-90 is of signiﬁcant health and environmental concern
given that it is a calcium analog and, thus, may bioaccumulate in
the bones and teeth of many organisms. The activity concentrationsof 90 Sr have been measured in selected “hotspots” around Japan(Steinhauser et al., 2013b). The authors reported that the 90Sr activity concentration in soil and vegetation samples from the hot spots
was lower than the Japanese regulatory limit. The activity concentrations of 90Sr were less than 10% of 137Cs. The contamination of themarine ecosystem by ﬁssion products from FDNPP had raised questions
about the safety of consuming seafood from the regions of Japan. Continuous monitoring programs were initiated by the Japanese authorities
to ensure that the levels of radioactivity in the Prefectural waters
remained low. This can only be achieved by ensuring that no further release of contaminated cooling water from the destroyed nuclear facility
occurs. Although few studies have explicitly examined the biologicalconsequences of radio-strontium in natural systems, perhaps in part because of the technical challenges and ﬁnancial costs related to its measurement, its presence in aquatic systems is of great concern becauseof the possibility of biomagniﬁcation in the food web.

7. Impacts of FDNPP accident on terrestrial ecosystems
Although it is perhaps fortunate for the people of Japan that due
to prevailing meteorological conditions at the time of the accident,
the bulk of radioactive emissions were transported to the ocean(Evangeliou et al., 2014b), there is a growing body of literature that
documents a range of physiological, developmental, genetic, morphological, and behavioral consequences of exposure to anthropogenic radioactivity derived from the FDNPP accident on the terrestrial ﬂora
and fauna of the Fukushima region of Japan. Hiyama et al. (2012) used
the pale blue grass butterﬂy, Zizeeria maha, as indicator species to evaluate the environmental impact of ionizing radiation. Some of the morphological abnormalities detected in some individuals of adultbutterﬂies collected in May 2011 increased in number and with time.
Similar results were also observed when the sampled adult butterﬂies
were compared to old museum specimens. Genetic studies indicatedthat certain abnormalities were inherited in the F2 generation of thegrass butterﬂy. These abnormalities may be explained by random mutation in important genes or by epigenetic mechanisms. By comparing
their data with those obtained from control populations, the authors
concluded that Fukushima-derived anthropogenic radionuclides likely
caused the morphological and genetic damage to the butterﬂy species.
The study by Hiyama et al. (2012) was signiﬁcantly strengthened bylaboratory experiments that used both internal and external radiation
sources, and these ﬁndings validated observations of the elevated mutation rates and phenotypic effects observed in the ﬁeld (Møller andMousseau, 2013). However, a recent article by Copplestone andBeresford (2014) has highlighted some drawbacks in the work of
Hiyama and colleagues. Copplestone and Beresford (2014) argued
that based on data reported by Hiyama et al. (2012) the lethal dose required to kill 50% of exposed individuals (LD50) was equivalent to theradiation levels that organisms receive from exposure to backgroundnatural radiation sources. This disconnect between empirical ﬁndings
and predictions based on theoretical models highlights the many uncertainties in accurately estimating doses in wild populations (Caffrey
et al., 2014); it is likely that wild animals are signiﬁcantly more sensitive

A.S. Aliyu et al. / Environment International 85 (2015) 213–228

to the effects of radiation than predicted by conventional approaches(Garnier-Laplace et al., 2013).
Low blood cell counts were reported in wild monkeys from theforests of Fukushima City, and compared to the counts in monkeys
from Aomori prefecture, 400 km away (Ochiai et al., 2014). The
total radiocesium concentration in the muscle of Fukushima wildmonkeys ranged from 78 to 1778 Bq kg− 1, while the cesium concentration in the control population was reported to be below detection
limits. A negative correlation was observed between activity concentrations of Cs and low blood counts in juvenile Fukushima monkeys. Ochiai
et al. (2014) concluded that the observed hematological aberration was
due to exposure to Fukushima-derived anthropogenic radionuclides.
This study was based on a rather limited ecological design. First, thepopulation sizes of the monkeys (61 from Fukushima and 31 from
Shimokita) used in the study were rather small and did not include individuals from the more contaminated areas of the region. A similar result showing hematological variations in the Fukushima monkeys had
been presented by the same group of authors in an earlier study(Hayama et al., 2013). The level of radiocesium in the Fukushima wildmonkeys was similar to that found in sheep in some parts of the UK following the Chernobyl accident—that is, extremely low in terms of damage to the animals themselves (Stallard, 2014). However, the ﬁndings ofthis study are greatly strengthened by comparisons to children living incontaminated areas of northern Ukraine near Chernobyl where similar
hematological effects were observed by Stepanova et al. (2008).
A recent study of bull sperm and testes from the Fukushima region
found no evidence of signiﬁcant histological changes (Yamashiro et al.,2013); however, this study was very preliminary with only two bulls
from a relatively uncontaminated part of Fukushima used for the analysis of sperm. However, further investigation of reproductive function ina diversity of organisms is warranted given ﬁndings of similar studies in
Chernobyl which have found evidence for signiﬁcant male reproductive
impairment stemming from exposure to radionuclides (e.g. Hermosell
et al., 2013; Møller et al., 2014).
Conservation biologists are concerned with the effects of environmental perturbation on population abundances and biodiversity
as these are indicators of overall ecosystem “health.” Initial studiesof the Fukushima region in July 2011 surveyed abundance and diversity of many species at 300 locations that varied from in exposures of
0.5 μSv/h to more than 30 μSv/h. Several bird species, as well as butterﬂies and cicadas, were signiﬁcantly reduced in numbers in the
more radioactive areas, even after correcting for other environmental factors known to inﬂuence population sizes (Møller et al., 2012,2013). An additional three years of biological inventories at 400 locations documented that bird abundance and diversity had continued
to fall at contaminated locations (Møller et al., in pressb; Møller
et al., in pressc). A more detailed analysis of barn swallow (Hirundo
rustica) populations found strong evidence of population declines,
although preliminary molecular analysis yielded no evidence for significant increases in genetic damage (Bonisoli-Alquati et al., 2015). A recent reanalysis of the Møller et al. (in pressa, in pressb) data uses
modern radioecological methods to reconstruct doses to individual
birds and provides evidence for dose response effects on bird abundances adding further support to the hypothesis that radioactive contaminants are the likely cause of the observed declines in Fukushima
birds (Garnier-Laplace et al., in press). Overall, these investigations of
population effects are similar to those found for Chernobyl-affected regions of Ukraine and Belarus, providing some additional support for the
hypothesis that radioactive contaminants are the likely causes for theobserved declines (Møller and Mousseau, 2009, 2006, 2007a, 2007b,
2011b, 2015; Mousseau and Møller, 2014).
The effects of the FDNPP accident on goshawk (Accipiter gentilis
fujiyamae) reproduction in North Kanto, Japan were investigated by
Murase et al. (2015). This is the only study comparing the situation before and after the accident using the same population. The authors of
this recent work compared the historical data concerning the four223

stages of goshawk reproduction namely occupancy (stage 1), incubating (stage 2), hatching (stage 3), and ﬂedging (stage 4). The study
spans more than two decades (22 years) of data collection. The results
suggest that radiation had a negative impact on reproductive performance with a progressive decline in the four stages of reproduction
over time being directly related to the air dose rate under nests. Anotherrecent paper by Matsui et al. (2015) has demonstrated that radiation exposure in avian populations depends on the quality (or weight) of
nesting material. The authors suggest that the effects of elevated levelsof radioactivity in nests could be of particular importance to offspringthat are hatched in highly contaminated nests or during the incubation
stage of reproduction given the known radio-sensitivity of early developmental stages.
A recent study of Japanese ﬁr trees (Abies ﬁrma) suggests plants havealso been affected by radioactive contaminants (Watanabe et al., 2015).
This study compared trees at four locations that varied from 0.13 μSv/h
to 33.9 μSv/h ambient dose rate and found a highly signiﬁcant positive
relationship between frequencies of morphological abnormalities andambient radiation levels. Although the mechanism underlying theobserved abnormalities is unknown, similar effects have been reported for pine trees in Chernobyl (e.g. Kozubov and Taskaev, 2006;
Mousseau et al., 2013) providing support for the suggestion that radionuclides are the cause of the observed developmental abnormalities.
The potential ecological consequences of the Fukushima accident
have been assessed using the ERICA assessment tool for different
types of biota in two studies (Aliyu et al., 2015a; Garnier-Laplace et al.,2011). The exposure proﬁle to 131I, 134Cs, and 137Cs one month after
the FDNNP was used by Garnier-Laplace et al. (2011) for estimateddose rate (mGy day− 1) calculations to a variety of organisms. Dose
rates to selected reference species were 1.5 mGy day− 1 for birds,
2.3 mGy day− 1 for soil invertebrates, and 3.9 mGy day− 1 for forest
trees. Although largely hypothetical, such dose rates could be biologically signiﬁcant, especially for longer-lived or radiosensitive organisms.
A long-term monitoring program of human health and ecosystemconsequences of chronic exposure to radioactive materials in the environment would contribute to a better understanding of effects exposure
to low-dose ionizing radiation. In studying the immediate and longterm effects of exposure to low dose radiation one also needs to understand a number of relevant issues such as the possible induction of longterm or multigenerational effects and the capacity of organisms to adapt
to stress. There is considerable evidence that many species are able toevolve resistance or tolerance to pollutants through the process of natural selection and a recent study of birds in Chernobyl indicates that adaptation to radiation is possible, at least for some species (Galván et al.,2014). On the other hand, there is considerable evidence for radiationinduced damage to DNA that likely reduces individual ﬁtness and thatof offspring when such genetic damage (i.e. mutations) is passed to subsequent generations. Chronically exposed populations are subject tomultigenerational mutational accumulation and are more likely to experience local extinction events (Higgins and Lynch, 2001), a subjectof great concern to conservation biologists (Møller and Mousseau,
2011a). At this stage, relatively little is known about the genetic (or epigenetic) mechanisms associated with population responses to stressful
contaminants. Much greater investment in basic research is thus
warranted.
8. Application of isotopes as tracers for environmental processes
Natural and anthropogenic radionuclides can be used as tracers forenvironmental processes, particularly oceanic transportation andwater mixing. In the oceans, the behavior of cesium is thought to be conservative; that is, it is soluble (with only a small fraction adsorbed tomarine particulates) and is carried primarily with ocean waters. As
such, cesium has been used as a tracer of water mass mixing and transport (Aoyama et al., 2011; Bowen et al., 1980; Buesseler et al., 1991).
Following the FDNPP disaster, cesium activity and the 134Cs/137Cs

224

A.S. Aliyu et al. / Environment International 85 (2015) 213–228

activity ratio have been measured to assess its environmental effects
and to study oceanographic processes. Studies based on samples of atmosphere, soil, seawater and marine biota off Fukushima have shownthat the 134Cs/137Cs activity ratio in the Fukushima derived radionuclides was nearly 1.0 (Buesseler et al., 2011; Hirose, 2012). This ratio ishigher than the 134Cs/137Cs activity ratio of 0.5 measured in the Chernobyl fallout (IAEA, 1986; UNSCEAR, 2000). This makes the tracking ofFukushima derived radionuclides in the ocean quite straightforward
since, given its relatively short half-life (2 years), the only source of134
Cs in the North Paciﬁc at this time would be the FDNPP. Hence, in addition to the elevated cesium activities, the presence of 134Cs is a uniqueisotopic signature for tracking these waters and provides a means tostudy the rates of vertical and horizontal mixing processes in the Paciﬁc
Ocean.
Pu has high particle afﬁnity compared with cesium, and thus provides a tool for studying a variety of processes in the marine environment such as particle ﬂuxes and scavenging, and for validating various
biogeochemical ocean models. However, the most useful aspect of Pu
as an environmental tracer is its well-deﬁned source terms (Lindahl
et al., 2010). This is because the composition of Pu isotopes from different sources in marine environments varies widely due to the characteristic isotopic composition corresponding to the means of production.
For example, reactor-grade Pu typically contains more than 35% 239Pu
with a 240Pu/239Pu atom ratio of 0.2–1.0 after fuel burn-up (Yamana
et al., 2001). In contrast, the 240Pu/239Pu ratio in weapons-grade Pu ismuch lower (0.02–0.06) because burn-up is kept low to minimize theproduction of higher Pu isotopes (Marka, 1993). This review showsthat the Fukushima derived 240Pu/239Pu ratio is 0.216–0.255, which ishigher than that of global fallout (0.18), but lower than that of Chernobyl. A summary of the composition of Cs and Pu isotopes for Fukushima,
Chernobyl and the global fallout is presented in Table 4.
9. Conclusions
Despite the large number of studies aimed at assessing the impactsof the FDNPP accident, considerable uncertainty remains concerning
all aspects of this disaster. Source term estimates vary from 12 to36.7 PBq for 137Cs and 150 to 160 PBq for 131I. Predictions of humanhealth impacts range to 10,000 estimated deaths, and 1500 and 1800
maximum cancer mortalities and morbidities. Studies of terrestrial systems, although few, have revealed possible signiﬁcant impacts on somespecies, notably birds and butterﬂies; the paucity of data may reﬂect thevery few funded research projects currently underway. Similarly, impacts on marine systems are largely unknown, with data limited to estimates of body burdens for commercially important species. It is perhaps

Table 4
Anthropogenic nuclide isotope ratio characteristics.
Source term

Fukushima

Chernobyl Global fallout

134

~1i
1.95 × 105–2.53 × 107j
1.2k
107.8m0.323–0.330m
0.128–0.135m

0.5n
–
0.5p
–
0.408q–Cs/137Cs (activity ratio)
137
Cs/239,240Pu (activity ratio)238
Pu/239,240Pu (activity ratio)241
Pu/239,240Pu (activity ratio)240
Pu/239Pu (atom ratio)241
Pu/239Pu (atom ratio)i

Buesseler et al. (2011).
Hirose (2012).
Lujanienė et al. (2012a, 2012b).
m
Zheng et al. (2012).
n
UNSCEAR (2000).
p
IAEA (1986).
q
Muramatsu et al. (2000).
r
Kelley et al. (1999), Krey et al. (1976).
t
Krey et al. (1976).
uPaatero et al. (2012).
v
Kelley et al. (1999), Krey et al. (1976).
w
Holm (1988).
j

k

–
38.5r
0.032t13–16u0.180v0.00194w

surprising that there have not been greater efforts to directly assess thehealth and environmental impacts of this disaster.
Our review of the current literature has shown that it took the radioactive plume resulting from the FDNPP accident a longer time to travel
around the world's atmosphere than that of the Chernobyl plume. The
explanation is that Fukushima radioactive clouds were released into
the PBL, whereas Chernobyl plumes were injected into the troposphere.
Of paramount need is a concise summary of ﬁeld measurement data for
FDNPP-derived radionuclides detected in locations around the world.
These ﬁeld measurement data could be used to drive computer models
for robust worldwide assessments of human and environmental impacts of the FDNPP accident. There is no substitute for direct measurements when attempting to assess the hazards and risks of any disaster.
Despite the environmental ramiﬁcations of ﬁssion products released
from nuclear accidents and atomic bomb testing, scientists have developed few methods for using radioactive ﬁssion products as tracers of environmental processes such as ocean mixing, atmospheric transport andsoil processes. In this review, we have provided a few examples for how
Fukushima-derived radionuclides have been used to study marine andatmospheric processes. There are undoubtedly many other potential applications of these approaches.
Overall, it is very clear that much remains to be learned concerning
the impacts of the FDNPP disaster and we hope that greater investments
will be made in the coming years into the basic scientiﬁc study of the
global health and environmental impacts of this accident.

Acknowledgments
A.S. Aliyu wishes to acknowledge the support of the Research
Management Center of the Universiti Teknologi Malaysia for its support through the Post Doc fellowship scheme under the project number (Q.J130000.21A2.01E98). Support for TAM has been provided by
the Samuel Freeman Charitable Trust and the University of South
Carolina College of Arts & Sciences. The authors are grateful to Elizabeth
Farnsworth (of Mount Holyoke College, Massachusetts) for help withmanuscript editing.

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