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Eight-year variations in atmospheric radiocesium in Fukushima City

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After the Fukushima nuclear accident, atmospheric 134Cs and 137Cs measurements were taken in Fukushima city for eight years, from March 2011 to March 2019. The surface air concentrations and deposition of radio-Cs were high in winter and low in summer; these trends are the opposite of those observed in a contaminated forest area. The half-lives of 137Cs in the concentrations and deposition before 2015 (275 d and 1.11 y) were significantly shorter than those after 2015 (756 d and 4.69 y). The dissolved fractions of precipitation were larger than the particulate fractions before 2015, but the particulate fractions were larger after 2016. The half-lives of 137Cs in the concentrations and deposition were shorter before 2015, probably because the dissolved radio-Cs was discharged from the local terrestrial ecosystems more rapidly than the particulate radio-Cs. X-ray fluorescence analysis suggested that biotite may have played a key role in the environmental behavior of particulate forms of radio-Cs after 2014. However, the causal relationship between the seasonal variations in particle size distributions and the possible sources of particles is not yet fully understood. The current study also proposes a method of evaluating the consistency of a numerical model for radio-Cs resuspension and suggests that improvements to the model are necessary.
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Eight-year variations in atmospheric radiocesium in Fukushima city
Akira Watanabe1,2, Mizuo Kajino3,4,5, Kazuhiko Ninomiya6,7, Yoshitaka Nagahashi1, and Atsushi
Shinohara6
5
1Faculty of Symbiotic Systems Science, Fukushima University, Fukushima, Fukushima 960-1296, Japan
2Institute for Climate Change, Fukushima, Fukushima 960-0231, Japan
3Meteorological Research Institute (MRI), Japan Meteorological Agency (JMA), Tsukuba, Ibaraki 305-0052, Japan
4Faculty of Life and Environmental Sciences, University of Tsukuba, Tsukuba, Ibaraki 305-8572, Japan 10
5Institute of Radiation Emergency Medicine (IREM), Hirosaki, Aomori 036-8564, Japan
6Graduate School of Science, Osaka University, Toyonaka, Osaka 560-0043, Japan
7Intitute for Radiation Sciences, Osaka University, Toyonaka, Osaka 560-0043, Japan
15
20
Correspondence to: A. Watanabe (watamay1948@yahoo.co.jp), M. Kajino (kajino@mri-jma.go.jp), and K. Ninomiya
(ninomiya@rirc.osaka-u.ac.jp)
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Abstract. After the Fukushima nuclear accident, atmospheric 134Cs and 137Cs measurements were taken in Fukushima city for
eight years, from March 2011 to March 2019. The surface air concentrations and deposition of radio-Cs were high in winter
and low in summer; these trends are the opposite of those observed in a contaminated forest area. The half-lives of 137Cs in the
concentrations and deposition before 2015 (275 d and 1.11 y) were significantly shorter than those after 2015 (756 d and 4.69 5
y). The dissolved fractions of precipitation were larger than the particulate fractions before 2015, but the particulate fractions
were larger after 2016. The half-lives of 137Cs in the concentrations and deposition were shorter before 2015, probably because
the dissolved radio-Cs was discharged from the local terrestrial ecosystems more rapidly than the particulate radio-Cs. X-ray
fluorescence analysis suggested that biotite may have played a key role in the environmental behavior of particulate forms of
radio-Cs after 2014. However, the causal relationship between the seasonal variations in particle size distributions and the 10
possible sources of particles is not yet fully understood. The current study also proposes a method of evaluating the consistency
of a numerical model for radio-Cs resuspension and suggests that improvements to the model are necessary.
Keywords: Fukushima nuclear accident, long-term observation, radiocesium, atmospheric radioactivity, precipitation
radioactivity 15
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1 Introduction
We conducted eight-year measurements of atmospheric 134Cs and 137Cs in Fukushima city after the Fukushima Daiichi Nuclear
Power Plant (FDNPP) accident that occurred in March 2011 to understand the time variations in and emission sources of 134Cs
and 137Cs and to propose effective ways to reduce atmospheric radioactivity. Among the various radionuclides released to the
environment, radio-Cs is particularly important due to its abundance in terrestrial ecosystems (the impacts of other nuclides 5
were negligibly small 100 days after the accident; Yoshimura et al., 2020), long half-lives (2.06 y for 134Cs and 30.17 y for
137Cs), and bioaccumulation (accumulation in muscle tissues, with biological half-lives of 30-150 d; WHO, 2011). Radio-Cs
forms aerosols in the air and is therefore efficiently deposited onto the ground surface via precipitation in addition to via dry
deposition. Approximately 30 % of the radio-Cs released in March 2011 was deposited onto the ground surface in Japan (the
aircraft-measured deposition on the ground was 2.7 PBq for 137Cs; NRA, 2012, and the most updated estimate of 137Cs 10
emissions by the Japan Atomic Energy Agency is 10 PBq; Terada et al., 2020). (The activity of 134Cs in the environment was
equivalent to that of 137Cs in March 2011). Once radio-Cs is deposited onto the ground surface, it circulates within local
terrestrial ecosystems, so the discharge from the local environment to downstream or downwind regions may not be substantial
(0.02 - 0.3 % y-1 to river; Iwagami et al., 2017, approximately 1 % y-1 to atmosphere1). Thus, long-term monitoring of
atmospheric radio-Cs at even one station may allow us to understand the mechanisms of its circulation in the local terrestrial 15
ecosystems, to estimate the external and inhalation exposure risks to the local residents, to propose efficient ways to reduce
health risks to the residents, and to assess the effectiveness of decontamination efforts.
To date, a great number of studies have focused on the circulation of radio-Cs in terrestrial ecosystems (Onda et al.,
2020). In terms of the long-term monitoring of atmospheric radio-Cs with a focus on resuspension from the ground surface to
the atmosphere, several papers have been published. Based on atmospheric measurements taken in the contaminated forest 20
area of the Abukuma Highlands (30 km northwest of the FDNPP) from October 2012 to December 2014, Ochiai et al. (2016)
reported that the surface activity concentrations of 137Cs were higher in summer and lower in winter and that the time variations
of the fine-mode (< 1.1 μm in diameter using an impactor) and coarse mode (> 1.1 μm) behaved differently. The coarse-mode
fractions were larger in summer, and the fine-mode fractions were larger in winter. Kinase et al. (2018) conducted surface
concentration measurements at four locations in the forest area of the Abukuma Highlands from July 2011 to March 2014 and 25
found that the concentrations of 134Cs and 137Cs were lower in winter and early spring and higher from late spring to autumn.
Their size-resolved measurements with a six-stage cascade impactor showed that the backup filter (< 0.39 μm) activity
concentrations were high in winter, consistent with Ochiai et al. (2016). However, Kinase et al. (2018) found through scanning
electron microscopy (SEM) that there were significant amounts of soil dust particles in the backup filter; these particles were
larger but bounced off the upper impactor stages. Therefore, they concluded that the sizes of radioactive particles were not 30
1 The annual resuspension rate to the atmosphere was estimated as 0.047 % y-1 by Kajino et al. (2016). However, the current
study found that the resuspension rate was likely substantially underestimated (see Sect. 3.5 and Fig. 9). A value of
approximately 1 % y-1 was obtained from improved simulations, but that manuscript is still in preparation.
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small but were actually large (coarse-mode particles). In late spring, the surface concentrations were positively correlated with
the wind speed, so they concluded that the wind-blown soil particles carried radio-Cs in this season. In the summer and autumn,
the concentrations were positively correlated with temperature but negatively correlated with wind speed, so they concluded
that the resuspension mechanisms were different in the winter and summer. The SEM analysis revealed that there were more
abundant bioaerosols in summer than in winter. Based on simulations, Kaijno et al. (2016) indicated that the summer peaks in 5
surface concentrations in the Abukuma Highlands could be accounted for by the bioaerosol emissions from forest ecosystems,
even though the emission mechanism remains unknown. Igarashi et al. (2019a) further investigated the mechanisms of
bioaerosol emissions in forests in summer by using fluorescent optical microscopic observation and high-throughput DNA
sequencing techniques. They suggested that the fungal spores that accumulate radio-Cs may be significantly involved in
resuspension in the forest in summer. Kita et al. (2020) suggested that rain induced the emission of radio-Cs associated with 10
fungal spores in the forest in summer. Minami et al. (2020) combined aerosol flux measurements and a multilayer atmosphere-
soil-vegetation model and estimated that the bioaerosol emission flux was on the order of 10-2 μg m-2 s-1, which could account
for the surface concentrations of 137Cs in the forests in summer (Kajino et al., 2016; Kinase et al., 2018; Igarashi et al., 2019a).
Kinase et al. (2018) also showed that there was no enhancement in the 137Cs concentration associated with forest fire events in
the region. The surface concentration of 137Cs was not correlated with that of levoglucosan, which is often used as a marker of 15
biomass burning. These results are distinct from those from Chernobyl, where wildfire plays a key role in the migration of
radio-Cs associated with the event (Ager et al., 2019; Igarashi et al., 2020). The contributions of additional 137Cs emissions
from the nuclear reactor buildings of FDNPP to the surface concentrations in Japan were negligibly small compared to the
resuspensions from the ground surface (Kajino et al., 2016). On the other hand, unintentional emissions in the premises of
FDNPP such as debris removal operations contributed to some observed sporadic peaks (Steinhouse et al., 2015; Kajino et al., 20
2016), although the impacts of such events might be small in terms of long-term averages and trends.
The current study is distinct from other studies, as it includes long-term comprehensive measurements (time-resolved
and size-resolved measurements of surface air activity concentrations together with measurements of dissolved and particulate
forms of activity in precipitation) at an urban/rural location in the Fukushima Basin in the vicinity of contaminated forests in
the Abukuma Highlands. The field observation and the simulation methods are described in Sect. 2. Sect. 3 presents the results 25
for the surface concentrations (Sect. 3.1), deposition amounts (Sect. 3.2), size distribution (Sect. 3.3), chemical compositions
(Sect. 3.4), comparison with simulations (Sect. 3.5), and comparison with measurements taken outside Fukushima Prefecture
(Sect. 3.6). The seasonal variations and possible emission sources are discussed in Sect. 4.1, the impacts of decontamination
and natural variations on the differences in trends before and after approximately 2015 are discussed in Sect. 4.2, the reasons
for the substantial deposition amount in January in Fukushima city are discussed in Sect. 4.3, and major findings and future 30
issues are summarized in Sect. 5. The observation data used in the study are provided as a Microsoft Excel file in the
Supplement.
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Figure 1: Map of Fukushima Prefecture and the surrounding prefectures. The locations mentioned in this study and terrestrial
elevations are depicted in the map.
2 Methods
2.1 Sampling site 5
The observation site, Fukushima University, is located in Fukushima city, located in the northernmost basin (Fukushima basin)
in the Nakadori Valley, surrounded by the Ou mountains to the west and the Abukuma Highlands to the east (Fig. 1). The
distance of the observation site from the FDNPP is approximately 60 km. The Nakadori Valley was formed by the Abukuma
River, which starts in the mountains in Fukushima Prefecture near the border of Tochigi Prefecture and flows northeast through
the central parts of Fukushima city to the Pacific Ocean in Miyagi Prefecture. The major radioactive plumes arrived twice in 10
Fukushima city, on March 15 and 20 (plume #3 and #8, as identified by Nakajima et al. (2017), respectively). These plumes
were transported over the Abukuma Highlands (where the peaks are mostly lower than 1,000 m) but were blocked by the
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higher Ou Mountains (peaks are 1,000 - 2,000 m) and thus transported along the Nakadori Valley (Nakajima et al., 2017). The
land surface of Fukushima city was contaminated mainly on the afternoon of March 15 with plume #3. The air dose rate in
Fukushima city started to increase at 17:00 local time (LT), associated with the weak rain that started at 13:00 LT, and peaked
at 19:30 LT at a value of 24.0 μSv h-1.
2.2 Surface air concentrations 5
2.2.1 High-volume air sampler, cascade impactor, and radioactivity measurement
The air samples were collected using high-volume air samplers (Kimoto electric Co., Ltd., Model-120SL) placed on the roof
of the building at Fukushima University (37.68°N, 140.45°E) at a height of 25 m from ground level. In this study, we carried
out two types of air sampling: time-resolved observations and aerosol size-resolved observations. In the former case, aerosol
samples were collected on a quartz fiber filter (Tisch Environmental, Inc., TE-QMA-100). The air suction rate of the sampler 10
was 700 L min-1. The typical duration of each sample collection was 24 hours, from May 8 to September 2, 2011. Then, we
switched to 72 hours of collection until December 27, 2017; after that, 1 week of continuous collection was performed until
March 28, 2019. For the latter observations, a cascade impactor system (Shibata Scientific Technology Ltd., HV-RW) was
placed into a high-volume air sampler. The air suction rate was 566 L min-1. The aerosols were collected separately by diameter
on six quartz filters (Kimoto, TE-236). The range of particle sizes in this system was 0.39-0.69, 0.69-1.3, 1.3-2.1, 2.1-4.2, 4.2-15
10.2, and >10.2 μm. (Note that the sizes in the manuscript indicate the 50 % cutoff aerodynamic diameters.) Fine particles
with a size of <0.39 μm were captured on a backup filter (Kimoto, TE-230-QZ). The typical sample collection time for the
size-resolved observations was three weeks. In types of both observations, activated carbon fiber filters (Toyobo Co., Ltd.,
KF-1700F 84 mmφ) were also placed at the exit of the high-volume air samplers to collect gas-state aerosols.
The collected aerosol samples were sealed into polyethylene bags at Fukushima University. After being shaped into 20
definite shapes, the gamma rays from the samples were measured by high-purity germanium detectors (coaxial with 15, 35
and 40 % relative efficiencies, SEIKO EG&G, ORTEC and coaxial with 40 and 60 % relative efficiencies, CANBERRA)
connected to a multichannel analyzer system (MCA7600, SEIKO EG&G) at the Radioisotope Research Center, Osaka
University. The radioactivities of 134Cs and 137Cs were identified at gamma-ray intensities of 605 keV and 662 keV,
respectively. The detection efficiencies of the respective detectors for each gamma ray were determined from the same-shape 25
filter samples from standard 134Cs and 137Cs solutions obtained from the Japan Radioisotope Association. The typical
measurement time of each sample was 1-3 days. Under these conditions, the detection limits of 134Cs and 137Cs were
approximately 5 × 10-3 Bq. The errors in the measured values are derived from the systematic error of geometrical configuration
and the standard sample itself in addition to statistical error. All radioactivities determined by our measurements were corrected
at mid sampling times. 30
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The radioactivities of both 134Cs and 137Cs were identified for most filter samples. The deviation in concentration
between 134Cs and 137Cs became larger over time due to the relatively short half-life of 134Cs. According to the radioactive
decay correction performed in March 2011, the activity ratios of 134Cs/137Cs were approximately 1. These ratios are consistent
with those in other reports related to FDNPP accident, so we concluded that the detected radiocesium originated from FDNPP
accident. During the measurement period, no radioactivity from 134Cs and 137Cs was detected from the carbon filters; that is, 5
the component of gaseous radioactive cesium was negligibly small.
2.2.2 Impactor/cyclone system
Since the filters for the high-volume air samples were quartz fiber filters, they could not be used for elemental analysis with
X-ray fluorescence spectrometry (XRF). For the XRF analysis, we used an impactor/cyclone system (Tokyo Dylec Corp., no-
number special order, 1100 L min-1) in which the aerosols were separated by size into < 2.5 μm and > 2.5 μm using an impactor; 10
those < 2.5 µm and >0.1 μm were sampled in glass bottles (As One corp., 2-4999-07) using a 0.1 µm cyclone with sampling
intervals of one month from September 2014 to January 2018. Aerosols larger than 2.5 µm were collected on quartz fiber
filters in the system. Aerosol samples in glass bottles (0.1 – 2.5 μm) were defined as fine-mode PM (PMf), and those on quartz
fiber filters (> 2.5 µm) were defined as coarse-mode PM (PMc). The radioactivities of 134Cs and 137Cs in the samples were also
measured in the same manner. 15
2.2.3 Possible artifacts of impactor measurements
Size separation by an impactor is associated with the artifacts caused by bouncing effects. In fact, in cascade impactor
measurements, Kinase et al. (2018) observed abundant coarse-mode particles such as mineral dust and bioaerosol particles in
the backup filters due to bouncing effects. In the impactor/cyclone system, the glass fiber filters used as an impaction surface
were immersed in silicone oil to prevent particles from bouncing (Okuda et al., 2015). In this study, silicone oil was not used 20
for the cascade impactor but was used in the impactor/cyclone system. However, the long-duration measurements (such as the
monthlong measurements) could be associated with the larger particles that rebounded at the impactor and were collected in
glass bottles (Okuda et al., 2015).
2.3 Deposition (dry plus wet deposition, dissolved and particulate fractions)
The total deposition (dry plus wet deposition or fallout) samples were collected with a precipitation sampler (Miyamoto Riken 25
Ind. Co., Ltd., RS-20) with a funnel diameter of 20 cm. Since a heating device was not installed on the sampler, any snow in
the funnel was manually melted in a water bath in winter. The accumulated snow in the funnel never reached the top of the
funnel during the whole observation period. A filtration device was installed in the sampler using membrane filters (Advantec,
4-880-03) with a pore size of 0.45 μm. The radioactivities of 134Cs and 137Cs in the filtered water stored in the polyethylene
bottle and those on the filters were both measured by high-purity germanium detectors at Osaka University and were defined 30
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as the dissolved and particulate fractions of the deposition, respectively. It should be noted here that this separation does not
perfectly differentiate water-soluble and insoluble radio-Cs. The clogging of the pores of the membrane filter can occur during
filtration. The measured total (dissolved plus particulate) deposition amounts were compared with those measured by the
official method at the Fukushima Prefecture Nuclear Power Center (Fig. 1), which is located 6.5 km north-northwest of
Fukushima University. Our method was found to be consistent with their official method: the correlation coefficient R was 5
0.81, with a slope of 1.158 (the values from Fukushima University were larger by 16 %). Differences in locations and sampling
intervals (daily at the Fukushima Prefecture Nuclear Power Center; monthly at Fukushima University) could also have
contributed to the differences in the measured values at the two sites.
2.4 X-ray fluorescence analysis (aerosols, deposition, and river sediments)
X-ray fluorescence (XRF) analysis was carried out by using a RIX1000 (Rigaku Corp.) at Fukushima University. The 10
measurement setup recommended by the manufacturer was used for the XRF. The major and trace element contents were
analyzed by the fundamental parameter method and calibration curve method, respectively (Takase and Nagahashi, 2007).
Measurements were conducted for PMf (see Sect. 2.2.2), the particulate fractions of precipitation (see Sect. 2.3), and the river
sediments. River sediments were collected at 15 sites upstream and downstream of Fukushima city in the Abukuma River and
its tributaries in 2010. Samples were taken from the gravel layer of the lower terrace at 5 sites, from alluvial fan deposits at 1 15
site, and from current riverbed sediments at 9 sites. The dried sediment samples were sieved and divided into two grain size
groups: particles smaller than 180 μm (defined as fine sediment particles) and particles 180 μm - 2 mm (coarse sediment
particles).
2.5 Numerical simulation and validation data
Kajino et al. (2016) used a Lagrangian model (LM) to simulate the atmospheric dispersion and deposition of 137Cs resuspended 20
from bare soil and forest ecosystems from January to December 2013. Since the resuspension fluxes and size distributions
were unknown, they adjusted the flux from bare soil (forest ecosystems) so that the simulated surface concentrations matched
those measured in Namie (Tsushima) (Namie High School Tsushima Campus, 37.56°N, 140.77°E, 30 km northwest of the
FDNPP) (Fig. 1) in the winter (summer) of 2013, and they adjusted the dry and wet deposition parameters (reflecting the size
distributions and hygroscopicity) so that the simulated total (dry plus wet) deposition over land in March 2011 matched those 25
measured by the aircraft measurements (NRA, 2012). Thus, note that the size distribution of the simulation was assumed to
have submicron size ranges that were consistent with those of the primary emissions (the direct emissions associated with the
FDNPP accident in March 2011) but that may not be applicable for resuspension events; the carrier aerosols are presumed to
be soil dust or bioaerosols, which are usually larger than the submicron size range. Kajino et al. (2016) concluded that their
simulations are likely reliable because the simulated differences between the surface concentrations in the contaminated area 30
(or emission source area) (i.e., Tsushima) and those in the downwind area (Meteorological Research Institute (MRI), Tsukuba
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city, 36.06°N, 140.13°E, 170 km southwest of the FDNPP) (Fig. 1) were consistent with the observed differences at the two
locations.
However, Kajino et al. (2016) used only surface concentration measurements to validate the simulations. The current
study also used concentration and deposition measurements from Fukushima University for model validation. The previous
study compared only the two locations in the contaminated forest areas and in the downwind urban/rural regions; the current 5
study includes an additional location in the urban/rural region near the contaminated forest of the Abukuma Highlands (60 km
northwest of the FDNPP).
3 Results
3.1 Surface air concentrations
Figure 2 shows the time variations in the atmospheric radioactivity concentrations of 137Cs from May 2011 to March 2019. 10
Just after the accident, the 137Cs concentrations were higher than 0.01 Bq/m3, and the maximum concentration of 0.0169 Bq/m3
was detected on May 23, 2011. The concentration quickly decreased to a level of 10-4 Bq/m3, and the minimum concentration
of 4.05 × 10-6 Bq/m3 was obtained on December 5, 2018. By taking the annual averaged value, the decreasing tendency in the
atmospheric concentration could be expressed as Y = 0.0418X-0.476, where Y indicates the annual mean 137Cs concentration and
X means the number of years elapsed. The coefficient of determination, R2, is 0.993. This demonstrates that the surface 15
concentration decreased exponentially and halved in approximately 4 years; thus, the decrease rate was higher than the rate of
radioactive decay of 137Cs.
It is remarkable that the decreasing trends in the earlier stage and the later stage were different. The regression lines
of the raw data time intervals for the whole period (red; May 2011 - March 2019), the earlier stage (blue; May 2011 - December
2014), and the later stage (green; January 2015 - March 2019) are shown in Fig. 2, with the half-life (Th) in days and the 20
decrease rate (Rd) in % y-1. The decreasing trend (Th = 275 d, Rd = 92.0 % y-1) of the earlier stage is approximately three times
faster than that of the later stage (Th = 756 d, Rd = 33.5 % y-1). It is shown later in Fig. 3 in Sect. 3.2 and discussed in Sect. 4.2,
but this could be related to the relative abundance of particulate and dissolved fractions of radio-Cs in the environment. The
dissolved fractions of radio-Cs may discharge faster than the particulate fractions from contaminated environments, such as
soils and plants. The relative abundance of the dissolved fractions was larger in the earlier stage than in the later stage such 25
that the decreasing trend in the surface air concentration was faster than that in the later stage. In addition to the natural
variability, decontamination work, which was completed by March 2018 in Fukushima city and the surrounding municipalities,
may also have contributed to the difference in the decrease rates; this possibility is also discussed in Sect. 4.2.
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Figure 2: Time series of surface air activity concentrations of 137Cs on the left axis. The red, blue, and green lines indicate the
regression lines of the whole period, before 2015, and after 2015, respectively. The half-lives (Th) and decay rates (Rd) are also
depicted. The gray line indicates the ratio of the running mean of 20 data points (an approximately monthly cycle) to the
running mean of 160 data points (an approximately annual cycle) on the right axis to show its seasonal variation. 5
It is also interesting that our data show different seasonal variations from those measured in Tsushima by Ochiai et
al. (2016) and Kinase et al. (2018). The levels in their studies were high in summer and low in winter, but as depicted in the
gray line in Fig. 2, the concentration rose starting in October, with maxima in the spring season around March and minima in
the summer. The maxima in the spring are approximately one order of magnitude larger than the minima in the summer. The 10
measurements of their studies were conducted in high-dose areas in the mountain forest (approximately 400 m above sea level
(a.s.l.)), and the high-volume samplers were set near the ground surface. In contrast, the current air sampling was conducted
in a relatively low-dose area (10 times lower than that in Tsushima) located in an urban/rural region on a hill (approximately
200 m a.s.l.) at the southern end of the Fukushima Basin. The heights of the samples were 25 m from the ground surface. Such
geographical and altitude differences could have caused these differences. 15
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3.2 Deposition amounts
Figure 3 shows the monthly cumulative deposition of 137Cs from March 2011 to March 2019. The monthly deposition amount
peaked in March 2011 as 202200 Bq m-2, decreased to 1 % of the initial amount after one year, and decreased to an order of 1
Bq m-2 after eight years. It also showed seasonal variation and was high from winter to spring. Nevertheless, the current level
is from two to three orders of magnitude larger than that before the Fukushima nuclear accident. The monthly trend is expressed 5
as Y = 48232X-1.944, where Y indicates the monthly cumulative 137Cs deposition and X means the number of months elapsed.
R2 is 0.697.
The decrease rates of deposition (Th = 1.11 – 4.69 y) were generally slower than those of the surface concentrations
(Th = 275 - 756 d). It is hard to identify the reason for this phenomenon. A perfect simulation could answer this question, but
high uncertainties in atmospheric deposition modeling and land surface modeling inhibit a perfect understanding of these long-10
term circulations of radio-Cs in the environment. It is safe to presume here that the decreasing trends in deposition and surface
concentrations are different because the contributions of major emission sources to deposition and surface concentrations are
different. If the dominant source of the surface concentration is near (far from) the observation site and that for of deposition
is far from (near) the site, the faster decrease rate in concentration is due to the faster (slower) reduction rate in the nearby
sources of emissions than in the far sources. 15
There is also a distinct difference in the decreasing trends before and after 2015. In addition to the effect of
decontamination work, as previously discussed in Sect. 3.1, the relative abundances of the dissolved and particulate fractions
of 137Cs could be a part of the reason. The particulate fraction made up 72.6 % of the deposition of March 2011, which is
presumed to be largely influenced by primary emissions. Here, it is interesting to note that most primary radio-Cs emissions
are thought to be composed of water-soluble submicron aerosol particles (e.g., Kaneyasu et al., 2012 and almost all numerical 20
simulations afterwards, such as Sato et al., 2020), while water-insoluble Cs-bearing microparticles (CsMP; Adachi et al., 2013,
Igarashi et al., 2019b) may contribute somewhat to primary emissions (Ikehara et al., 2020, Kajino et al., 2021). If the primary
radio-Cs in aerosols were 100 % in water-soluble forms, the particulate fraction should have made up 0% of the precipitation
in March 2011 (although, some of the water-soluble Cs could have converted to a water-insoluble form through adsorption to
soil particles accumulated on the membrane filter during filtration). After April 2011, as the contributions of resuspension 25
were thought to be dominant, the dissolved fractions became larger. The ratio varied, but the dissolved fractions were generally
higher before 2016, and the particulate fractions became dominant after 2016. There seemed to be a regime change in the
physicochemical properties of radio-Cs circulating in the environment in the area in approximately 2015 and 2016, which
could have changed the decreasing trends of both the surface concentrations and deposition before and after 2015. This result
is consistent with the finding of Manaka et al. (2016), who reported that the exchangeable proportions of radio-Cs rapidly 30
decreased in forest soils from two to four years after the accident, i.e., from 2013 - 2015.
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Figure 3: (Top) Time series of 137Cs activity deposition. The red, blue, and green lines indicate the regression lines of the
whole period, before 2015, and after 2015, respectively. The half-lives (Th) and decay rates (Rd) are also depicted. (Bottom)
Time series of particulate and dissolved forms of 137Cs deposition on the left axis and the ratio of particulate to dissolved 137Cs
on the right axis. 5
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The seasonal variations in particulate and dissolved 137Cs were slightly different from each other and different from
those of the surface concentration. The surface concentration peaked in March in almost all years, and the total deposition
peaked in January. The peaks of the total deposition in January coincided with those of the dissolved 137Cs before 2016, but
the peaks of the dissolved 137Cs became unclear afterwards. The peaks of particulate 137Cs occurred in March before 2016,
which coincided with those of the surface concentrations. After 2016, there were no clear seasonal variations in particulate 5
137Cs. There are clear and different seasonal variations in the surface concentration and deposition. However, at the current
stage, we have no knowledge of or numerical tools to reveal the hidden mechanisms underlying these variations.
3.3 Size distributions
Figure 4 shows the time series of the seasonal mean atmospheric radioactivity concentrations of 137Cs obtained from the
cascade impactor measurements. The sampling interval for the cascade impactor measurements was three weeks. The seasonal 10
means included a sampling period if any part of the sampling period was included in the season. For example, the raw data
from the sampling period from February to March contributed to the averages of both DJF (December, January, and February,
i.e., winter) and MAM (March, April, and May, i.e., spring). The seasonal mean total (all sizes) concentrations of cascade
impactor measurements during the sampling period agreed well with those of the time-resolved observations (Fig. 2), with R2
= 0.93. The same seasonal variation discussed for the time-resolved observations (Sect. 3.1) was also observed; the atmospheric 15
137Cs concentration was relatively high in DJF and MAM compared to that in JJA (June, July, and August, i.e., summer) and
SON (September, October, and November, i.e., autumn).
The most dominant size range in activity was the backup filter (< 0.39 μm, or rebounded particles such as soil dust
and bioaerosols; Kinase et al., 2018), and its seasonal variation agreed well with that of the total particle concentration (high
in DJF and MAM). On the other hand, the second largest contribution was made by the size range of 4.2-10.2 μm, which 20
showed the opposite seasonal variation and was relatively high in JJA and SON. The seasonal variations in the largest particle
fraction, larger than 10.2 μm, are interesting. The trend appears to be synchronized with that of the backup filter particles (high
in DJF and MAM), but the opposite trend was observed in 2016 and 2017 (high in JJA). The contributions of other fractions,
i.e., 0.49-4.2 μm, were small in the measured period. Even though the contributions were small, the seasonal trend of 0.39-
0.69 μm was similar to that of the backup filter particles, but that of 1.3-2.1 μm was similar to that of 4.2-10.2 μm. The current 25
measurement indicates that the dominant particles and their sizes may be distinct depending on the season. The decrease rates
of each size were different before and after approximately 2015, as discussed in Sects. 3.1 and 3.2, but the size distribution of
the surface activity did not change substantially before and after approximately 2015.
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Figure 4: (Top) Time series of seasonal mean size-resolved surface activity concentrations of 137Cs and (bottom) their relative
fractions.
5
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Figure 5: (Top) Time series of surface activity concentrations of 137Cs in PMf (0.1-2.5 µm) and PMc (>2.5 µm) collected by
the impactor/cyclone system and those of the backup filter of the cascade impactor. (Bottom) Correlation coefficients of
temporal variations among seasonal mean 137Cs activity concentrations of different sizes measured by the impactor/cyclone
and the cascade impactor. Correlation coefficients higher than approximately 0.4 and lower than approximately -0.4 are colored 5
blue and orange, respectively.
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Cascade impactor sampling is associated with the bouncing effect, whereas filters for the impactor/cyclone system
were immersed in silicone oil to prevent the bouncing effect. Thus, compared the cascade impactor and the impactor/cyclone
measurement data, as shown in Fig. 5. The top panel of Fig. 5 shows the data with the same measurement time intervals (three
weeks for the cascade impactor data and one month for the impactor/cyclone data). The surface activity concentrations of 137Cs
in the backup filters were well correlated with those of PMf. No remarkable seasonality was observed in PMc, but some 5
enhancements were observed in JJA in 2015 and SON in 2016.
The bottom panel of Fig. 5 shows the correlation coefficients among the seasonal mean size-resolved data from the
cascade impactor and impactor/cyclone measurements. If we assume that the bouncing effect on the impactor/cyclone
measurements was negligible, the cascade impactor data and the impactor/cyclone data were consistent. There was a positive
correlation between PMf and the backup filter data. There were also positive correlations between PMc and the 1.3-2.1 µm and 10
4.2-10.2 µm data. There was a negative correlation between PMc and PMf. We can assume that fine-mode particles are the
dominant carriers of 137Cs in winter and spring and that coarse-mode particles are the dominant carriers of 137Cs in summer
and autumn. However, there was also a contradiction in the data. There were low or negative correlation coefficients between
the backup filter data and the cascade impactor data at smaller size ranges, such as 0.39-0.69, 0.69-1.3, and 1.3-2.1 µm, but
the backup filter data were positively correlated with the impactor data for > 10.2 µm. It appears that bouncing effect occurred; 15
particles larger than 10.2 µm bounced in the latter stages and were captured in the backup filter. However, as previously
discussed, the behaviors of the >10.2 μm-particle data were not consistent in time, i.e., they were generally high in DJF and
MAM and were high in JJA in 2016 and 2017 (Fig. 4). Kinase et al. (2018) and Igarashi et al. (2019a) considered that the
dominant carriers of resuspended 137Cs were coarse-mode particles such as soil dust and bioaerosols. Ochiai et al. (2016)
conducted two-stage impactor sampling and measured the surface activity concentrations of 134Cs and 137Cs above and below 20
1.1 μm from 2012 to 2014. They showed that the contributions of coarse-mode particles (> 1.1 μm) were dominant, with
maxima in summer. The contributions of the fine-mode particles (< 1.1 μm) were much smaller, and no significant seasonal
variations were found. All of their measurement sites were surrounded by contaminated forests in the Abukuma Highlands
(Tsushima and the nearby sites), so the sampling sites were different from those in our study. Such larger particles may have
contributed to the backup filter data in the current measurements; however, based on the fact that the backup filter data were 25
positively correlated with PMf and not with PMc, fine-mode particles (< 2.5 μm) should also play a key role in determining the
surface air concentrations in Fukushima city.
On the other hand, if we assume that the bouncing effect is also significant in the impactor/cyclone system due to the
long sampling duration, as suggested by Okuda et al. (2015), the positive correlation between the backup filter particles and
PMf was simply due to the bouncing effects of the larger particles in both systems. 30
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Even though the emission sources of the dominant particles collected by the size-resolved measurements could not
be identified in this study, the possible aerosol sources that would explain the differences in size and seasonality of the two
locations are discussed later, in Sect. 4.1.
3.4 Chemical characterizations of particles in the air, rainfall, and river sediments
Figure 6 shows the relative abundance of the XRF-measured atomic number concentrations of elements in the PMf monthly 5
sample from September 2014 to January 2018. Among the 15 detected species, PMf was mainly composed of SiO2, Al2O3, and
SO3. The fractions of SiO2 show clear seasonal variations and were higher around May. The seasonal variations in Al2O3 and
SO3 are the opposite of that of SiO2. A positive temporal correlation was obtained between the 137Cs in PMf and SiO2 (R =
0.30). Negative correlations were obtained for Al2O3 and SO3, with correlation coefficients of -0.36 and -0.35, respectively.
Note that these results do not prove that the SiO2-bearing aerosols are the carriers of resuspended 137Cs, but we can safely 10
conclude that the origins of SiO2 and 137Cs may be close to each other (i.e., that both come from the same source or the same
area/direction).
Figure 7 shows comparisons of the relative abundance of the periodic mean XRF measured atomic number
concentrations in different samples, fine sediment particles, coarse sediment particles, PMf, and particulate fractions of
precipitation. The PMf and precipitation data over the same period, from October 2014 and December 2012, were averaged. 15
The sediment samples were collected in 2010. The 10 species that were common to all samples are shown in Fig. 7. The
correlation coefficients for the compositions among samples are above 0.9, showing that the samples have similar origins. The
features of the PMf composition were distinct from the others. PMf included SO3 (17.8 %) and Cl (2.65 %), while the others
did not.
Weathered biotite is abundant in the soil in Fukushima and absorbs radio-Cs efficiently (Kogure et al., 2019). The 20
compositional correlation coefficients between the weathered biotite (Takase, 2020) and the four samples were high, at 0.73
to 0.87. However, when the two major components SiO2 and Al2O3 were excluded, the compositional correlation coefficients
changed significantly. The eight compositional correlation coefficients between the fine and coarse sediment particles were
0.98, but those between the sediments and PMf were 0.01 and 0.19 for the fine and coarse sediment particles, respectively. The
eight compositional correlation coefficients for the particulate fractions of precipitation were moderate, at 0.36, 0.44, and 0.45 25
for fine sediment particles, coarse sediment particles, and PMf, respectively. The eight compositional correlation coefficients
for weathered biotite were 0.76, 0.71, 0.50, and -0.14 for fine sediment particles, coarse sediment particles, particulate fractions
of precipitation, and PMf, respectively.
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Figure 6: Temporal variations in the chemical composition of PMf as measured by XRF.
Figure 7: Periodic mean chemical compositions of fine particles and coarse particles in sediments, PMf, and the precipitation 5
filter (the particulate fraction of precipitation) measured by XRF.
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The findings from the current section are summarized as follows. The mean compositions of both fine and coarse
sediment particles are similar to those of biotite, which absorbs radio-Cs efficiently. The similar composition feature was
observed for the particulate fractions of precipitation. The composition of PMf was slightly different from those of the other
samples, but the 137Cs concentrations in PMf become larger when the relative fractions of SiO2, the major component of biotite,
increased. Thus, biotite may have played a key role in the environmental behavior of radio-Cs in Fukushima city since 5
September 2014. However, the major carriers of radio-Cs before September 2014 and those in the dissolved fractions in
precipitation are still unknown.
3.5 Comparison with the simulation results and climatological deposition velocity analysis
In Fig. 8, the surface concentrations of 137Cs in 2013 simulated by Kajino et al. (2016) are compared with the time-resolved
observation data (Fig. 2). Kajino et al. (2016) included 137Cs resuspended from bare soil, 137Cs resuspended from forest 10
ecosystems, and additional 137Cs emissions from the FDNPP. The additional 137Cs emissions were negligibly small, the
concentrations of which in East Japan were two to three orders of magnitude smaller than those from the two sources. Therefore,
they are not depicted in the figure. The simulation was successful in explaining the magnitude and seasonal variations in
surface concentrations at Tsushima and the MRI, but the simulation at Fukushima city disagreed with the observations. The
simulation showed an enhancement of 137Cs from forests in the summer, but that was not detected in the observations. The 15
observed magnitude and seasonal trends are rather similar to those simulated for 137Cs from soil dust.
Figure 8: Time series of (black) measured surface activity concentrations of 137Cs and those simulated (by Kajino et al., 2016;
K16) considering different emission sources, (red) mineral dust from bare soil and (green) aerosols emitted from forest
ecosystems. 20
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Kajino et al. (2016) used only the observed surface concentrations to estimate the regional budget of resuspended
137Cs in the air, but we used the observed deposition to evaluate the model, as shown in Fig. 9. Suppose there is a simple
nonlinear relationship between the deposition (D) and surface concentration (C):
𝐷𝑎𝐶
, (1)
where a represents a removal rate and b represents nonlinearity, such as spatial and temporal variabilities. If one can take a
long-term average of D and C, Eq. (1) may hold. Eq. (1) is reformulated as 5
𝑙𝑜𝑔𝐷  𝑏 𝑙𝑜𝑔𝐶  𝑙𝑜𝑔𝑎. (2)
The log-log scatter plot between the monthly mean surface concentrations and monthly cumulative deposition of observed
(purple) and simulated (orange) 137Cs are depicted in the left panel of Fig. 9. The coefficient of determination of the observation
was 0.678, with a risk factor for < 0.1 %. Eq. (1) holds for the monthly mean resuspended 137Cs at Fukushima University. As
seen in Eq. (2), the intercept of the Y-axis indicates the removal rate a. log(a) is dimensionless, but if b is close to one, the unit
of a can be m s-1. From Fig. 9, b of observation is close to one. Therefore, the ratios of the monthly deposition amounts to the 10
monthly mean surface concentrations are referred to as the climatological deposition velocity (m s-1). Time series of the
climatological deposition velocity are presented in the right panel of Fig. 9.
Figure 9: (Left) Scatter diagram of (purple circles) observed surface concentrations and deposition of 137Cs and those simulated
(by Kajino et al., 2016; K16) considering different emission sources, (orange open squares) mineral dust from bare soil and 15
(orange close squares) aerosols emitted from forest ecosystems. The purple and orange lines indicate the regression lines of
the observed data and the simulated (both dust and forest) data. (Right) Time series of (blue) climatological deposition velocity
on the left axis and (gray) precipitation amounts on the right axis.
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The left panel of Fig. 9 clearly shows that the removal rate (a) used in Kajino et al. (2016) can be underestimated by
one to two orders of magnitude. The deposition velocities used in Kajino et al. (2016) were estimated from the observation of
137
Cs in March 2011, which was supposed to be mainly composed of submicron water-soluble particles. However, the current
study and the series of previous studies regarding resuspended
137
Cs indicated that the host particles of
137
Cs could be
substantially larger (e.g., soil and bioaerosols). This may be the reason for the overestimation of simulated
137
Cs from forests 5
in the summer in Fukushima city. If the deposition velocities of the model increased by one to two orders of magnitude, the
transport of
137
Cs from the contaminated forest to Fukushima city in summer may decrease such that the simulated surface
concentration in Fukushima city agrees with the observation. Certainly, their simulated regional budget needs to be reassessed
using the realistic deposition velocities indicated in the current study.
The observed climatological deposition velocity varied by more than one order of magnitude over time. There are 10
two main deposition mechanisms: dry deposition and wet deposition. Wet deposition is associated with precipitation. The
variations in the climatological deposition rate seem to agree with the observed precipitation, but almost no correlation was
observed (R ~ 0.10). The mean climatological deposition velocity was 5.3×10
-1
m s
-1
, and the peak values occurred in January.
The maximum value was 4.9 m s
-1
in January 2013, when the monthly precipitation was not very high (81.2 mm). Possible
reasons for these peaks in January are discussed later, in Sect. 4.3. The typical order of the dry deposition velocity of 15
supermicron (1–10 µm in diameter) particles is approximately 10
-3
–10
-2
m s
-1
(e.g., Petroff and Zhang, 2010), which is
substantially lower than the values in our climatological deposition velocity analysis. Certainly, the magnitudes of the instant
deposition velocity and our climatological deposition velocity are not directly comparable, but it seems that wet deposition
plays an important role in the removal of resuspended
137
Cs-bearing atmospheric aerosols.
20
Figure 10: (Left) Scatter diagram of the observed surface deposition of
137
Cs at Fukushima University and the MRI from
March 2011 to March 2019, with a regression line. (Right) Time series of the ratio of deposition at Fukushima University to
deposition at the MRI, with a regression line.
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3.6 Comparison of deposition amounts at Fukushima University and the MRI
Figure 10 compares the deposition amounts of 137Cs at Fukushima University (60 km northwest of the FDNPP) and the MRI
(170 km southwest of the FDNPP) from March 2011 to March 2019. The deposition data at the MRI are available from
Environmental Radioactivity and Radiation in Japan (https://www.kankyo-hoshano.go.jp/data/database/, last accessed: June
14, 2021). There was a significant positive correlation between the deposition amounts of 137Cs at the two sites. The slope of 5
the regression indicates that the ratio of deposition at Fukushima University to that at the MRI did not change significantly
from the initial ratio during the eight years, which is approximately 8-9 times (202200 Bq m-2 at Fukushima University and
23100 Bq m-2 at the MRI). This indicates that deposition was influenced by emissions from nearby sources and was not
substantially influenced by long-range transport at either site. The right panel of Fig. 10 indicates that the deposition ratios at
the two sites were approximately 10, with a variation of more than one order of magnitude and peaks in winter (especially 10
January) that decreased slightly over time. The right panel of Fig. 9 shows that the January peak is a feature of Fukushima city
and was not observed at the MRI. The possible reasons for the January peak in Fukushima City are discussed later, in Sect.
4.3. The slight decreasing trend was probably due to decontamination, which was ongoing in Fukushima during the period
until 2018, as shown later in Table 1. Certainly, natural variations could also have contributed to the decreasing trend.
4 Discussion 15
Even eight years after the FDNPP accident, the surface air activity concentration of 137Cs had not fallen to the level before the
accident, which was at an order of magnitude of 10-6 Bq m-3. In difficult-to-return zones, the surface concentrations sometimes
still exceed 10-2 Bq m-3. Based on long-term measurements, this study tries to understand the characteristics of radio-Cs in the
air and its deposition and to reveal its origins in order to identify effective ways to reduce radioactivity in contaminated
terrestrial ecosystems. 20
4.1 Seasonal variation and possible sources
The current study clearly shows that the surface concentrations of 137Cs are high from winter to spring, with peaks in March
and lows from summer to autumn, in the urban/rural area of Fukushima city (60 km northwest of the FDNPP). It also shows
that the deposition amounts of 137Cs are high in the winter, especially in January, and low from summer to autumn. This
seasonal trend is the opposite of that observed in a forested area in the Abukuma Highlands (Tsushima, 30 km northwest of 25
the FDNPP), which was high in the summer (Ochiai et al., 2016; Kinase et al., 2017). From winter to spring, northwesterly
winds prevail over the region associated with migrating disturbances, while southeasterly winds prevail over the region
associated with the Pacific high. The three simulated monthly mean surface wind fields for January to March and June to
August are shown in Kajino et al. (2016).
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In summer, Fukushima city is downwind of Tsushima. The surface concentrations of 137Cs at Tsushima are
approximately ten times greater than those at Fukushima city, and Fukushima city is downwind of Tsushima, but there is no
enhancement of 137Cs in summer. The traveling distances of carrier aerosols depend on their aerodynamic diameters. The
distance between the two sites is approximately 30 km. The traveling distances of aerosols below < 10 μm are not very different
and are larger than 100 km because their gravitational deposition velocities are negligibly small. On the other hand, the 5
traveling distances rapidly decrease proportionally to a square of the diameter above 10 μm, and the traveling distance of an
aerosol with a diameter of 100 μm is an order of 1 km (Kajino et al., 2021). Aerosols below < 10 µm can travel longer distances,
but their dry deposition amounts increase significantly from 1 µm to 10 µm (Kajino et al., 2021). Igarashi et al. (2019a) reported
that the major proportions of bioaerosols in forests in summer are smaller than 5 μm in diameter and can travel a fairly long
distance. Pollen is much larger than 10 μm, but pollen emission is limited in summer (Igarashi et al., 2019a). Consequently, 10
there was a significant enhancement in surface concentrations in the forests in summer but no enhancement in the downwind
urban/rural areas, probably because the carrier aerosols were efficiently deposited onto the ground surface before significant
amounts of atmospheric 137Cs reached the downwind areas. Consistent with our findings presented in Fig. 5, 137Cs in PMc was
more abundant than that in PMf and the backup filter particles in summer. To obtain a quantitative understanding of the regional
cycle of atmospheric 137Cs in the northern part of Fukushima Prefecture, accurate simulations are required in the future. 15
In winter and spring, the surface concentrations of 137Cs are probably enhanced due to the local emissions from the
nearby sources because the location of the sampling site is upwind of the Abukuma Highlands and the ground surface in the
upwind areas of the sampling site in the season (northwest directions) is less contaminated than the site. In winter and spring,
137Cs in the backup filter particles and PMf are pronounced in Fukushima city. These characteristics are somewhat different
from those reported in previous studies. Miyamoto et al. (2014) measured the size distributions of radio-Cs with a cascade 20
impactor for two periods, from March 17 to April 1 and May 9 to 13, 2011, at a site 120 km southwest of the FDNPP. They
showed that the peak size ranges were 1.2 - 2.1 µm and 0.65 - 1.1 µm in the former and latter periods, respectively. Doi et al.
(2013) reported that the peak diameters of the 137Cs concentration from April 4 to 11 were 1.0 µm and 1.5 µm at Tsukuba, 170
km southwest of the FDNPP. Kaneyasu et al. (2017) measured the size distributions of 137Cs and other chemical components
six times at Tsukuba from April to September 2011. The peak diameter ranges were 0.49 - 0.7 μm in the earlier stages (before 25
June 9), but the contributions of coarse-mode particles (> 1 μm) increased after June 9, and the second modes appeared in the
ranges of 3.5 - 5.2 μm and 7.8 - 11 μm in July and September, respectively. Judging from their measured mass size distributions
of Ca, which is assumed to originate from mineral soil, Kaneyasu et al. (2017) concluded that soil particles could be the major
carrier of resuspended radio-Cs in Tsukuba. Our XRF analysis indicated that radio-Cs is carried mainly by soil particles in
Fukushima city, but the size distributions are greater in PMf (<2.5 μm) and in the backup filter particles. If radio-Cs is carried 30
by soil particles, it is natural to presume that the fractions of radio-Cs in PMc would be large (e.g., Fig. 3 of Kaneyasu et al.,
2012 or Fig. 4 of Kaneyasu et al., 2017). One could argue that the bounced coarse mode soil particles are observed in the
backup filters, but in fact, the seasonal mean Cs concentrations in the backup filter are positively correlated with PMf and
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negatively correlated with PMc (Fig. 5). One could further argue that bounced large particles are also collected in PMf despite
the special procedures employed to prevent rebound in the impactor/cyclone system.
There are four possible explanations for these results: (1) If the bouncing effect did not occur in either system, the
major sources of radio-Cs in Fukushima city are probably related to combustion (a mass peak below 0.39 μm means that the
number peak is approximately 100 nm). If the bouncing effect occurred only in the cascade impactor, (2) if the origin of radio-5
Cs is soil particles, the size distributions of soil particles in Fukushima city are smaller, or radio-Cs in the soil exists more
within finer particles; (3) if the origin of radio-Cs is soil particles, the coarse-mode fraction deposits to the ground surface
faster than the fine-mode fraction, such that the proportion of Cs in PMf is larger in Fukushima city. (4) The bouncing effect
occurs in both systems, and the origin of radio-Cs is coarse-mode soil particles. (1) is less likely because there is little chance
of the artificial combustion of contaminated biomass. In fact, there were no temporal correlations between the 137Cs and 10
levoglucosan (a biomass burning marker) surface air concentrations at Tsushima during the forest fire event in the Abukuma
Highlands that occurred in March 2013 (Kinase et al., 2018). (3) is also less likely because long-range transport (at least 100
km) is required for the major proportions of coarse-mode particles to deposit to the ground surface, whereas Fukushima city
is characterized as the emission source region in that season. In terms of (2), the latter sentence, “radio-Cs in the soil exists
more with finer particles”, contradicts Kaneyasu et al. (2017), suggesting that the radio-Cs is uniformly distributed on the 15
surface of soil particles. (4) is possible, but rebound is prevented in the impactor/cyclone system, and there is also no evidence
that rebound occurred in the impactor/cyclone system (certainly, there is also no evidence that rebound did not occur). Further
experiments are required to determine whether (2) or (4) is more likely and whether some sources are missing. As Kaneyasu
et al. (2012) and (2017) reported, comparing the size distributions of 137Cs with those of other chemical components in
Fukushima city would be an effective way to investigate the origin of resuspended radio-Cs from winter to spring. Alternatively, 20
a PM2.5 cyclone or virtual impactor could be used to separate the fine-mode and coarse-mode particles, that can completely
exclude the bouncing effect.
4.2 Differences in trends before and after approximately 2015 (natural variation and decontamination)
As described in Sects. 3.1 and 3.2, distinct decrease rates were observed before and after approximately 2015 in both the
surface concentrations and the deposition. There may be two main reasons for this: natural variation and decontamination. The 25
natural variation (the dissolved fractions of precipitation or the exchangeable proportions of forest soils discharging faster than
other forms from the local ecosystems (Manaka et al., 2016)) was previously described, and the effect of decontamination is
presented in some detail here.
30
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Table 1: Decontamination achievement ratios in Fukushima City and the surrounding municipalities (Nihonmatsu City,
Kawamata Towna, Date City, and Koori Town)
March 2014 March 2015 March 2016 March 2017 March 2018
Fukushima Cit
y
Residential area (Number of houses)
Public facility (Number of facilities)
Road (km)
Agricultural field (ha)
Forest (livin
g
area)b (ha)
50.2 %
89.3 %
9.1 %
94.0 %
5.0 %
62.3 %
92.3 %
16.1 %
94.4 %
6.3 %
100.0 %
97.5 %
39.6 %
95.2 %
37.3 %
100.0 %
100.0 %
50.2 %
96.0 %
80.8 %
100.0 %
100.0 %
100.0 %
100.0 %
100.0 %
The surroundin
g
municipalities
Residential area (Number of houses)
Public facility (Number of facilities)
Road (km)
Agricultural field (ha)
Forest (livin
g
area)b (ha)
87.2 %
34.3 %
48.7 %
99.0 %
23.9 %
97.4 %
55.3 %
56.8 %
99.0 %
36.7 %
99.6 %
80.2 %
67.5 %
99.6 %
64.1 %
100.0 %
94.9 %
82.0 %
100.0 %
88.6 %
100.0 %
100.0 %
100.0 %
100.0 %
100.0 %
aOnly the western part of Kawamata town. The decontamination of areas with an annual cumulative dose exceeding 20 mmSv
was assigned to the central government, and that in areas with a dose below 20 mmSv was assigned to municipal governments.
The decontamination of the eastern part of Kawamata town was conducted by the central government. 5
bRemoval of the litter layer in forests within 20 meters of the forest edge.
Table 1 summarizes the achievement ratios of the scheduled decontamination of different land use types in Fukushima
city and the surrounding municipalities (available at https://www.pref.fukushima.lg.jp/site/portal/progress.html, last accessed:
June 14, 2021). The municipalities are in the northern part of Fukushima Prefecture, which comprises 55 % forest area, 15 % 10
farmland area, 6 % residential area, and 23 % other areas (https://www.pref.fukushima.lg.jp/uploaded/attachment/42042.pdf,
last accessed: June 14, 2021). More than 94 % decontamination was achieved for the farmland area by March 2014. For the
residential and public facility areas, some parts were decontaminated by March 2014, but some others were not fully
decontaminated until March 2018. For the road and forest areas, decontamination was not completed in most areas by March
2014, but extensive decontamination was conducted from 2014 to 2018. Note that only a part of the forest (20 m from the 15
forest edges) was decontaminated, which accounts for approximately 1 % of the whole forest area of the northern part of
Fukushima Prefecture. Additionally, only the litter layer of the forest was removed, and the soil layer remained.
Suppose that if contamination occurred independently of the land use type, approximately 30 % (farmland + half of
residential and other) of northern Fukushima was decontaminated by 2014, and an additional 15 % (half of residential and
other) was continuously decontaminated by 2018. The difference between the decrease rate from May 2011 to December 2014 20
(93.1 % y-1) and that after (30.7 %y-1) was higher than the decontamination rate (30 - 45 % per three to seven years). If the
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surface concentration at Fukushima University was affected mainly by the emissions from nearby sources (i.e., within the
northern part of Fukushima Prefecture), decontamination would not be the sole reason for the change in the decrease rates
before approximately 2015 and after. Natural variation (i.e., regime changes in the chemical forms of radio-Cs) would likely
occur during that period. As previously discussed, biotite may have played a key role in the environmental behaviors of radio-
Cs in Fukushima city after approximately 2015, but the current study could not identify the key aerosol particles that carried 5
dissolved (or exchangeable) radio-Cs and were abundant in Fukushima city before approximately 2015.
4.3 Substantial deposition amounts in January in Fukushima city
The climatological deposition velocities (or the ratios of the deposition rate to the mean surface concentration) in
Fukushima City were remarkably high in January 2013, 2014, 2015, and 2017 (Fig. 9). They were approximately one order of
magnitude larger than those in the other months. The ratio of the deposition in Fukushima and to that at the MRI was 10
approximately 10 on average, but those in January of those years exceeded 100 (Fig. 10). On the other hand, no peaks were
observed in January of 2012, 2016, or 2018.
There are two possible explanations for these results: vertical distribution and the existence of superlarge particles.
In terms of the former, the substantial proportions of 137Cs in the upper air may have caused lower surface concentrations but
higher deposition due to the wet removal of 137Cs aloft. However, due to the northwesterly winter monsoon, the upper air over 15
Fukushima city is also upwind of the Abukuma Highlands; thus, this possibility is less likely. In terms of the latter, superlarge
particles (~100 µm or larger in diameter), have settling velocities that are too high (as high as those of drizzle droplets) to enter
the high-volume air sampler but that allow them to settle efficiently in a deposition sampler. A similar feature has been
observed in the relationship of the deposition and surface concentration of sodium at observation sites near coastal areas (e.g.,
particles denoted as large sea salt particles (LSPs) in Kajino et al., 2012). The travel distance of such large particles is 20
approximately 1 km (e.g., Kajino et al., 2021), and Fukushima University is surrounded by major roads, such as Route 4 and
National Highway E4, within 1 km. January is the month when the highest snow depth occurs in the Fukushima Basin, and
the road surface may be wet and muddy due to snow removal work using deicing agents and daytime snow melt on pavements;
therefore, road dust emissions from busy transportation activities may be enhanced. The muddy surface conditions may
produce even larger road dust particles. Although there is no evidence of the existence of such superlarge particles, they may 25
be a possible reason for the substantial deposition amounts in January in Fukushima city. In fact, substantial amounts of road
salt from deicing agents could contribute to roadside PM10 samples in winter (Denby et al., 2016), indicating that there could
be emissions of particles even larger than 10 µm in diameter. In addition to the direct deposition of 137Cs to the rain sampler
at Fukushima University, the immediate resuspension of deposited 137Cs associated with road dust or road salt from nearby
roads around the university could contribute additional deposition to the rain sampler. Unfortunately, analyses of the surface 30
meteorological observational data for Fukushima City from the JMA, such as temperature, precipitation, snow cover, and wind
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speed data, did not reveal the differentiating features between the years with (2013, 2014, 2015, and 2017) and without (2012,
2016, and 2018) high deposition peaks.
5 Conclusions
Eight-year measurements of atmospheric 134Cs and 137Cs conducted at Fukushima University from March 2011 to March 2019
are summarized in this study. A high-volume sampler, a cascade impactor, and an impactor/cyclone system were used to collect 5
aerosol samples, and the activity concentrations of radio-Cs were detected by high-purity germanium detectors. A precipitation
sampler was used to collect deposition samples, and the dissolved and particulate fractions of radio-Cs in the samples were
measured. X-ray fluorescence (XRF) analysis was carried out to measure the elemental contents of the aerosol and precipitation
samples. The concentration and deposition data measured at Fukushima University were compared with numerical simulation
results. 10
The major findings are itemized as follows:
(1) The observed surface concentrations and deposition at Fukushima University (an urban/rural area of Fukushima city, 60
km northwest of the FDNPP) were high in winter and low in summer; these seasonal trends are the opposite of those observed
in a contaminated forest area (Ochiai et al., 2016; Kinase et al., 2018) (30 km northwest of the FDNPP, in the Abukuma
Highlands). Resuspension due to bioaerosol emissions (Kinase et al., 2018; Igarashi et al., 2019a) may be substantial in forests 15
but may not be in urban/rural areas. The half-life (Th) and decrease rate (Rd) for the eight years were 456 d and 55.6 % y-1 for
the concentrations and 2.35 y and 29.5 % y-1 for the deposition, respectively.
(2) The decreasing trends changed in approximately 2015 and were associated with changes in the dissolved/particulate
fractions of 137Cs in precipitation. The Th and Rd for concentrations before 2015 were 272 d and 93.1 % y-1, whereas they were
825 d and 30.7 % y-1 after 2015. The Th and Rd for deposition before 2015 were 1.10 y and 63.2 % y-1, whereas they were 5.39 20
y and 12.9 % y-1 after 2015. The dissolved fractions were higher before 2015, whereas the particulate fractions were higher
after 2016. This may have been because the dissolved proportion of radio-Cs discharged faster than its particulate forms from
the local terrestrial ecosystems. This is consistent with the findings of Manaka et al. (2016). Decontamination likely also
contributed to the difference because the decontamination of some land use types, such as agricultural fields, was completed
before 2014, and 100 % of the planned decontamination was completed by March 2018. The contribution of decontamination 25
was estimated in this study to be 30 - 45 % for the three to seven years, which is significantly smaller than the differences in
the Rd of the concentrations (93.1 % y-1 before 2015 and 30.7 % y-1 after 2015). Therefore, decontamination may play a partial
role in explaining the differences in Th and Rd before and after 2015, but changes in the chemical forms of radio-Cs likely play
a major role.
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(3) The size-resolved measurements showed that the dominant size range in activity in the cascade impactor data was the
backup filter (<0.39 μm in diameter, or particles rebounded from larger stages), followed by the 4.2-10.2 μm and >10.2 μm
sizes. The backup filter particles were abundant in winter. The seasonal mean 137Cs concentrations in the backup filter of the
cascade impactor were positively correlated with those in the fine-mode aerosols collected by the impactor/cyclone system
(PMf) and negatively correlated with those in the coarse-mode aerosols (PMc). PMc was high in summer. The impactor/cyclone 5
system prevented the bouncing effect, but bouncing may still have occurred during long-duration samplings. The XRF analysis
showed that biotite may have played a key role in the environmental circulation of particulate forms of resuspended radio-Cs
in Fukushima city after September 2014.
(4) The deposition amounts of 137Cs in January were remarkably high compared to the surface concentrations of 137Cs and the
deposition amounts of 137Cs at the MRI. Although we have no observational evidence, we hypothesize that the existence of 10
superlarge particles (~100 µm or larger, with a distance of ~ 1 km or less) associated with snow removal operations on major
roads near Fukushima University may be one of the reasons for the remarkable high deposition amounts in January.
Certain issues remained unresolved, and topics for future study are summarized as follows:
(1) The Abukuma Highlands are upwind of Fukushima city in summer. The enhancement of 137Cs in PMc in summer is
consistent with the fact that most bioaerosols exist in coarse mode. However, if radiocesium is carried mainly by biotite (i.e., 15
soil particles) in winter, there should be an enhancement of 137Cs in PMc because major proportions of soil particles exist in
coarse mode (e.g., Kaneyasu et al., 2017). On the other hand, sources of Cs-bearing fine-mode particles such as combustion
emissions may be less likely. Thus, the main carrier of radio-Cs may be biotite in winter, but this is still not fully confirmed.
XRF measurements were conducted for PMf from September 2014 to January 2018, when the particulate proportions were
dominant in the precipitation. Thus, the carrier aerosols of dissolved radio-Cs in Fukushima city are still unknown. As 20
Kaneyasu et al. (2012) and (2017) reported, comparisons of the size distributions of 137Cs with those of other chemical
components in Fukushima city would be an effective way to investigate the origin of resuspended radio-Cs from winter to
spring. Alternatively, a PM2.5 cyclone or virtual impactor could be applied to separate the fine-mode and coarse-mode particles,
which can completely exclude the bouncing effect.
(2) The simulation used in this study was made to be consistent with the surface concentrations in a contaminated forest 25
(Tsushima) and those in a downwind area (the MRI, 170 km southwest of the FDNPP). However, the current study found that
the simulated seasonal variation in Fukushima city was totally opposite to the observations. The current study indicated that
the deposition velocities applied in the simulation were significantly underestimated. Numerical simulation is a powerful tool
for quantitative assessment, but the current simulation requires further improvement. The reasons for the seasonal variations
in concentrations and deposition in the different locations need to be investigated with an improved model. 30
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Data availability
The observation data used in the study are provided as a Microsoft Excel file in the supplement.
Author contribution
AW conducted the long-term measurements together with KN and AS. YN performed the XRF analysis, and MK performed
the numerical simulation. AW, MK, and KN designed the manuscript structure and completed the draft together with all 5
authors.
Competing interests
The authors declare that they have no conflicts of interest.
Acknowledgments
This work was supported by Grants-in-Aid for Scientific Research (KAKENHI), grant numbers JP16H01777 and JP24110009. 10
This work was also supported by the Sumitomo Foundation’s Environmental Studies Grant and partly supported by the
Environment Research and Technology Development Fund (JPMEERF20215003) of the Environmental Restoration and
Conservation Agency of Japan. The authors thank Dr. Teruyuki NAKAJIMA of the National Institute for Environmental
Studies and Dr. Haruo TSURUTA of the Remote Sensing Technology Center for proving and installing a high-volume sampler
and Prof. Naohiro YOSHIDA of Tokyo Institute of Technology for the installation of an impactor observation system. The 15
authors thank Dr. Shogo TOGO of the Isotope Research Center of the University of Tokyo for the radioactivity measurements
of the initial samples and Dr. Yoshiaki YAMAGUCHI, Dr. Zijian ZHANG, Dr. Makoto INAGAKI, Mr. Shunsuke KAKITANI,
Mr. Kazuya FUJIHARA and Mr. Nobufumi FUJITA of Osaka University for their support with the radioactivity measurements
of the air samples. The authors also thank Mr. Kakeru KONNAI of the University of Tsukuba for data visualization, Dr. Akane
SAYA of the MRI for a useful discussion of aerosol deposition processes, and Prof. Tomoaki OKUDA of Keio University for 20
a useful discussion of the possible artifacts in the impactor measurements.
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... On the other hand, Adachi et al. (2013) isolated hydrophobic supermicron Cs-bearing particles (referred to as Cs-bearing microparticles; CsMPs) from aerosol filters collected in March 2011; the atmospheric behaviors of these CsMPs could be quite different from those of hydrophilic submicron particles. Detailed analyses of CsMPs are helpful for understanding emission events and mechanisms (Igarashi et al., 2019a;Kajino et al., 2021) and deposition 80 processes (Dépée et al., 2019). Vertical measurements obtained on mountains (Hososhima and Kaneyasu, 2014;Sanada et al., 2018) have revealed the importance of cloud deposition over mountainous forests in East Japan. ...
... The numerical simulations conducted by K16 were consistent with the findings described above: the surface concentrations in the mountainous forest area are low in winter and high in summer, the contributions of mineral dust are high in winter, and those of bioaerosols are high in summer. However, Watanabe et al. (2021) found that the simulations of K16 underestimated the observed deposition amounts by approximately two orders of magnitude. The major reason for this large 110 discrepancy in deposition is the incorrect assumption of the physical properties of resuspended 137 Cs by K16. ...
... The concentration and deposition measurements are conducted at Fukushima University (140.45°E, 37.68°N) (Watanabe et al., 2021). Fukushima University is located on a small hill at the southern edge of the Fukushima Basin 145 and is surrounded by major roads. ...
Preprint
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Resuspension of 137Cs from the contaminated ground surface to the atmosphere is essential for understanding the environmental behaviors of 137Cs and estimating external and inhalation exposure of residents. Kajino et al. (2016) assessed the 137Cs resuspension flux from bare soil and forest ecosystems in East Japan in 2013 using a numerical simulation constrained by surface air concentration measurements. However, the simulation was found to underestimate the observed deposition amounts by two orders of magnitude. The reason for this underestimation is that the simulation assumed that resuspended 137Cs is carried by submicron aerosols, which have low deposition rates. Based on the observational indications that soil dust and bioaerosols are the major carriers of resuspended 137Cs, a new simulation is performed with higher deposition rates constrained by both surface concentrations and deposition amounts. In the new estimation, the areal total annual resuspension of 137Cs in 2013 is 25.7 TBq, which is equivalent to 0.96 % of the initial deposition (2.68 PBq). Due to the rapid deposition rates, the annual redeposition amount is also large at 10.6 TBq, approximately 40 % of the resuspended 137Cs. The resuspension rate through the atmosphere (0.96 % y−1) seems slow, but it (2.6 × 10−5 d−1) may not be negligibly small compared to the actual decreasing trend of the ambient gamma dose rate obtained in Fukushima Prefecture after the radioactive decay of 137Cs plus 134Cs in 2013 is subtracted (1.0–7.9 × 10−4 d−1): Resuspension can account for 1–10 % of the decreasing rate due to decontamination and natural decay through land surface processes. The current simulation underestimated the 137Cs deposition in Fukushima city in winter by more than an order of magnitude, indicating the presence of additional resuspension sources. The site of Fukushima city is surrounded by major roads. Heavy traffic on wet and muddy roads after snow removal operations could generate superlarge (approximately 100 µm in diameter) road dust or road salt particles, which is not included in the model but might contribute to the observed 137Cs at the site.
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The dispersion and deposition of radio-cesium (¹³⁷Cs) carried by two types (type A and type B) of water-insoluble Cs-bearing solid microparticles (CsMPs) released due to the Fukushima nuclear accident were simulated for the first time. The presence of type B CsMPs (70–400 μm found in soil and 1–5 μm found in air), associated with the hydrogen explosion of Unit 1 in the afternoon of March 12, could partly explain the simulated underestimation of total deposition over land by assuming that 100% of the Cs carriers were water-soluble submicron particles (WSPs). Type A CsMPs (0.1–10 μm), released from Units 2 or 3 in the midnight between March 14 and 15, traveled over the Kanto Plain, the most populated plain in Japan. Differences in the size distribution of type A CsMPs altered the surface air concentration over Kanto substantially, by up to more than one order of magnitude. The major deposition mechanisms varied among dry, wet, and fog (and/or cloud) depositions depending on the size distribution and locations. The simulated activity fractions due to the CsMPs in the total deposition were compared to those observed in surface soil for the first time. The observations could be explained by the simulations for the locations under the influence of type B CsMPs. However, the simulations were substantially underestimated for the locations influenced by type A CsMPs. There could be more fractions of type A CsMPs emission in the source term and/or the simulated deposition rates of type A CsMPs were underestimated.
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We have measured artificial radionuclides, such as 90 Sr and 137 Cs, in atmospheric depositions since 1957 in Japan. We observed the variations in 90 Sr and 137 Cs, which were emitted from atmospheric nuclear tests and nuclear power plant accidents, due to their diffusion, deposition, and resuspension. In March 2011, the Fukushima Daiichi Nuclear Power Plant accident occurred in Japan, and significant increases in 90 Sr and 137 Cs were detected at our main site in Tsukuba, Ibaraki. Our continual observations revealed that the 137 Cs monthly deposition rate in 2018 declined to ~ 1/8100 of the peak level, but it remained more than ~ 400 times higher than that before the accident. Chemical analysis suggested that dust particles were the major carriers of 90 Sr and 137 Cs during the resuspension period at our main site. Presently, the effective half-life for 137 Cs deposition due to radioactive decay and other environmental factors is 4.7 years. The estimation suggests that approximately 42 years from 2011 are required to reduce the atmospheric 137 Cs deposition to a state similar to that before the accident. The current 90 Sr deposition, on the other hand, shows the preaccident seasonal variation, and it has returned to the same radioactive level as that before the accident.
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