Journal of Oceanography, Vol. 62, pp. 767 to 775, 2006
Mass Balance and Sources of Mercury in Tokyo Bay
M ASAHIRO S AKATA1*, KOHJI MARUMOTO 2, M ASAHIRO NARUKAWA 3 and KAZUO ASAKURA 4
1
Institute for Environmental Sciences, University of Shizuoka, Yada, Shizuoka 422-8526, Japan
National Institute for Minamata Disease, Hama, Minamata, Kumamoto 867-0008, Japan
3
Solar-Terrestrial Environment Laboratory, Nagoya University,
Honohara, Toyokawa, Aichi 442-8507, Japan
4
Environmental Science Research Laboratory, Central Research Institute of Electric Power Industry,
Abiko, Abiko, Chiba 270-1194, Japan
2
(Received 23 February 2006; in revised form 15 May 2006; accepted 8 June 2006)
The mass balance and sources of mercury in Tokyo Bay were investigated on the
basis of observations from December 2003 to January 2005. Estimated input terms
included river discharge (70 kg yr–1) and atmospheric deposition (37 kg yr–1), and
output terms were evasion (49 kg yr–1), export (13 kg yr –1) and sedimentation (495
kg yr –1). Thus, the outputs (557 kg yr –1) considerably exceeded the inputs (107
kg yr –1). In addition, the imbalance between the inputs and outputs of mercury was
much larger than that of other trace metals (Cd, Cr, Cu, Pb and Zn), which suggests
that there are other major inputs of mercury to Tokyo Bay. The mercury concentrations in rivers correlated significantly with the concentrations of Al and Fe, major
components of soil. In Japan, large amounts of organomercurous fungicides (about
2500 tons as Hg) were used extensively in fields in the past, and most of the mercury
was retained in the soil. In this study, the mercury concentration in rivers was measured primarily in ordinary runoff. These observations lead to the hypothesis that
field soil discharged into stormwater runoff is a major source of mercury in Tokyo
Bay. As a preliminary approach to validating this hypothesis, we measured the concentrations of mercury and other trace metals in river water during a typhoon. The
mercury concentrations in stormwater runoff increased to 16–50 times the mean value
in ordinary runoff, which is much higher than the increases for other metals. This
tends to support the hypothesis.
Keywords:
⋅ Mercury,
⋅ mass balance,
⋅ Tokyo Bay,
⋅ stormwater runoff,
⋅ organomercurous
fungicides.
in the supply of mercury to the lake, followed by tributary inputs (280 kg yr–1) and particulate remineralization
(250 kg yr–1). For Lake Michigan, Landis and Keeler
(2002) reported that atmospheric deposition contributed
approximately 84% to the total annual input of mercury
(1403 kg). This remarkably large contribution of atmospheric input may be due to the fact that both lakes have
small ratios of watershed area/lake area.
On the other hand, the mass balance of mercury in
coastal regions such as bays and estuaries has hardly been
studied. These regions have relatively large ratios of watershed area/water area. This may reveal the greater importance of runoff over direct atmospheric deposition of
mercury input to these regions (Kang et al., 2000; Balcom
et al., 2004a). Humans are exposed to a significant level
of methylmercury due to their consumption of fish and
fish products from marine systems. In addition, it is known
that significant in situ methylation of mercury occurs in
coastal marine systems (Mason et al., 1999; Balcom et
1. Introduction
In recent years, there have been increased human
health and environmental concerns associated with the
toxicity of methylmercury bioaccumulated in fish through
the food chain. Atmospheric deposition has been identified as the most likely source of mercury in aquatic environments, even in regions (e.g., pristine inland lakes in
North America and Scandinavia) far from direct discharges of mercury (Sorensen et al., 1990; Lindqvist et
al., 1991). A significant amount of methylmercury is produced by the in situ methylation of inorganic mercury
deposited into aquatic environments (Pirrone, 2001;
Rolfhus et al., 2003). From a mass balance study of mercury in Lake Superior, Rolfhus et al. (2003) concluded
that atmospheric deposition (740 kg yr–1) predominated
* Corresponding author. E-mail: [email protected]
Copyright©The Oceanographic Society of Japan/TERRAPUB/Springer
767
Chiba
Tokyo
Edo River
Narashino
Ara River
Komae
2
1
3
4
Tama River
Aqua line
5
Kanagawa
6
7
Futtsu
Kannonzaki
Umihotaru
Atmospheric deposition
River water
Seawater and sediment
1970. A similar pollution history was observed for other
trace metals such as Cd, Cr, Cu, Pb, and Zn (Matsumoto,
1983).
In this study, the mass balance and sources of mercury in Tokyo Bay were investigated on the basis of analysis of mercury in atmospheric deposition, river water,
seawater and sediment samples collected from December 2003 to January 2005. We considered river discharge
and atmospheric deposition as input terms, and evasion,
export and sedimentation as output terms. The terms effluent discharges (e.g., sewage treatment and manufacturing) and particulate remineralization were not included
in this mass balance, because no such data were available.
2. Methods
Fig. 1. Location of sampling sites.
al., 2004a). Hence, there is an urgent need for an increased
understanding of the marine biogeochemical cycling of
mercury, especially in coastal regions that are affected
by various anthropogenic sources (Balcom et al., 2004b).
Tokyo Bay, which refers to the inner part enclosed
by the line connecting Futtsu with Kannonzaki (Fig. 1)
in this study, has an area of 960 km2 and a mean depth of
15 m (Kasuya et al., 2004). The bay is surrounded by
densely populated (~26 million), highly industrialized
areas, including many steel mills, petrochemical plants
and thermal power plants. The bay bottom is covered with
a black silty clay containing sulfides. The sediment accumulation rate determined by 210Pb dating and the mercury pollution level in Tokyo Bay sediments have been
reported by Matsumoto (1983) and Matsumoto et al.
(1983), respectively. These results show that the average
sedimentation rate in the entire bay is 0.18 g cm–2yr–1,
and that the mercury pollution level in Tokyo Bay increased abruptly after 1950, reaching a maximum around
768
M. Sakata et al.
2.1 Wet and dry depositions
Precipitation and dry deposition (gaseous and
particulate) samples were collected every half-month for
one year from December 2003 to November 2004 at three
sites in the Tokyo Bay area (Fig. 1). The Komae and
Narashino sites are located about 20 km and 6 km inland
of the bay, respectively. The Umihotaru site is on an artificial island used as a parking lot on the “Aqua line” highway with a total length of 15.1 km (tunnel part of ~10 km
plus bridge part of ~5 km), which crosses the bay.
Sampling was conducted using an automatic sampler, shown in Fig. 2. The sampler is composed of a
precipitation sampler (177 cm2 collection area) and a
water surface sampler (1017 cm2 collection area) for dry
deposition (gases and particles). Details of the sampler
were described previously by Sakata and Marumoto
(2004, 2005). The precipitation sampler is equipped with
two sampling trains. The sample bottle for mercury measurement initially contains approximately 50 mL of 5
mol L–1 HCl solution for stabilizing mercury without adsorption and volatilization after collection. The water
surface sampler has a water surface plate with an airfoil
leading edge to minimize airflow disruptions caused by
the collector geometry. A water surface can act as an infinite sink for both atmospheric gases and particles. This is
advantageous to mercury, which is deposited in gaseous
and particulate forms. Five liters of 0.25 mol L–1 HCl
solution was used for the collection of deposited materials. The bulk of the trace elements associated with atmospheric particles were expected to be soluble in this
solution (Sakata and Marumoto, 2004).
Circulated water on the water surface plate was maintained at a temperature of roughly 5–10°C lower than the
air temperature by refrigerating the reservoir system to
reduce water evaporation during sampling. At the
Umihotaru site, which has high humidity during summer,
however, atmospheric water vapor often conversely con-
Moisture sensor
Automatically
covered
Anemometer
Water surface
plate
Air foil leading edge
Water surface (36 cm diameter)
Outlet
Water surface
holder
Reservoir
system
Refrigerating
(5 − 10 °C lower than
air temperature)
Tubing
pump
Precipitation
(for Hg)
Precipitation
(for others)
Fig. 2. Diagram of automatic wet and dry deposition sampler.
densed on the water surface due to refrigeration. As a
countermeasure, the degree of refrigeration of the reservoir system was reduced during this period and the amount
of sampling solution was also reduced from 5 L to 4 L.
To prevent water splashing on the surface plate owing to
the extremely strong wind on the sea, the water surface
plate was kept covered with a Teflon net with a mesh size
of 6.5 mm. In addition, the water surface plate was placed
on the lid when the wind speed monitored using an anemometer exceeded a predetermined limit of 20 m s–1.
However, these periods were very short relative to the
entire sampling periods. The measured amounts (weighed
values) of the residual sampling solution for dry deposition sampling corresponded roughly to the values calculated from the mass balance of Sr (~0.4 mg L–1) previously added to the solution, i.e., C 1W1 = C 2W2, where C1
and C 2 are the Sr concentrations in solution at the start
and the end of dry deposition sampling, respectively and
W1 and W2 are solution weight at the above times. This
supports the fact that a splash of water on the surface
plate owing to a strong wind does not induce a significant loss of sampling solution.
The sampling was started by pumping the sampling
solution to the water surface plate at a rate of 400 mL
min–1. The water surface sampler was allowed to run for
approximately 20 min, after which 250-mL samples
(blanks) were obtained from the water surface using a
50-mL syringe to determine the background concentration of mercury. At the end of the sampling period (~2
weeks), the pump was shut off, and the sampling solution on the water surface plate was poured into the bottle.
The samples collected at each site were weighed immediately in the laboratory of the Central Research Institute
of Electric Power Industry in Komae City (Fig. 1), after
which BrCl (0.002 mol L–1) was added to both the pre-
cipitation and dry deposition (i.e., sampling solution) samples for mercury measurement to oxidize all Hg compounds to Hg2+. All the samples were stored in a refrigerator until analysis.
The precipitation and dry deposition samples were
filtered through 0.4 µm filters. The mercury concentration in each filtrate was measured by cold-vapor atomic
fluorescence spectrometry (CVAFS; Tekran, model 2600),
following Hg0 generation with SnCl2 as the reducing agent
(Fitzgerald and Gill, 1979). The method’s detection limit
(MDL) for water samples was approximately 0.1 ng L–1.
The MDL was defined as 3 times the standard deviation
of the replicate measurements of a blank solution. The
mercury concentrations in the samples were >50 times
higher than the MDL. The blanks obtained at the start of
dry deposition sampling had Hg levels ranging approximately from 0 to 1 ng L–1. For almost all (92%) of the
samples, these levels corresponded to less than 10% of
the concentrations in the samples. The relative standard
deviation (RSD) of replicate measurements using a standard solution (3 ng L–1) was less than 3%. The annual wet
and dry deposition fluxes (mg m–2yr –1) were calculated
on the basis of the sum of the deposition amounts (=Hg
concentration × sample volume) for each sampling. Dry
deposition flux was corrected for the background blanks.
2.2 Water
Three rivers, Edo, Ara and Tama, account for most
of the water flowing into Tokyo Bay (Hattori, 1983). Water
samples were collected from them using a Teflon sampler on bridges near estuaries (Fig. 1). In principle, sampling was conducted during low tide on the same date as
the collection of atmospheric deposition samples (i.e.,
twice per month). The levels of seawater intrusion into
the samples were estimated to be less than 2% on the baMass Balance and Sources of Mercury in Tokyo Bay
769
sis of the Na concentration. On the other hand, the
seawater samples were collected at 3 or 4 depth layers at
7 stations in Tokyo Bay (Fig. 1) in June, August and October 2004 and January 2005. A Van Dorn sampler was
used for sampling. Prior to sampling, the sampler was
cleaned carefully by HCl solution. Both the river and
seawater samples were transferred to Teflon bottles for
storage. In the laboratory, BrCl (0.002 mol L–1) was added
into the samples and stored in a refrigerator until analysis. The mercury concentration in the samples was measured by the method described in the preceding section.
The bulk of mercury associated with suspended particles
was dissolved by adding BrCl. This was confirmed on
the basis of analysis of mercury in suspended particles
using a HNO 3 - HClO4 - H2SO4 digestion (refer to the
“sediments” section). The RSD of mercury measurements
of water samples was approximately 10%.
For evasion flux measurements, in addition, the concentrations of dissolved gaseous mercury (DGM) in surface seawater and total gaseous mercury (TGM) above
the sea surface were determined at 4–7 stations (Fig. 1)
in December 2003, October 2004 and January 2005. The
method of Wängberg et al. (2001) was used for the DGM
measurement. The surface seawater samples were collected directly from the sea surface using a peristaltic
pump. Two liters of the sample was poured into a Teflon
impinger, which consisted of a tube of about 1.8 m length
and 4 cm inside diameter. The sample was sparged by
introducing a stream of prepurified nitrogen at a flow rate
of 0.5 L min–1 for 2 hours. The gaseous mercury separated was collected on a Au trap. All DGM samples (in
Au traps) were sealed in containers and stored at room
temperature until analysis. The procedural blanks were
approximately 3 ng m –3. This level corresponds generally to less than 10% of the concentrations in the samples. The RSD of replicate measurements was about 10%.
Mercury in the air passing through Teflon filters on
the boat was measured hourly as TGM (primarily Hg0),
using a continuous mercury vapor analyzer employing Au
trap amalgamation and atomic absorption spectrometry
(AAS; Nippon Instruments, AM-2). TGM was measured
during transport as well as at anchor stations. The sampling inlet was placed into the wind, thus avoiding the
risk of contamination. The RSD of replicate measurements
of standard mercury vapor was less than 3%.
2.3 Sediments
A sediment core was collected by a diver inserting a
plastic tube (diameter 20 cm) into the sediment at station
1 in Tokyo Bay (Fig. 1) in December 2003. The core was
cut vertically into 2-cm sections. Sedimentation rate was
measured by Metocean Environment Inc. using the 210Pb
method (Matsumoto and Wong, 1977). 210Pb in the sediment was determined by counting the beta activity of its
770
M. Sakata et al.
daughter 210Bi, after acid digestion of the sediment. For
the age determination of the sediment, sediment porosity
was obtained from the measured density of the solid phase
and sediment water content.
Surface sediments (approximately the top 5 cm) were
collected by a diver inserting a plastic container into the
sediment at 7 stations (Fig. 1) in August and October 2004.
The samples were sealed in containers and stored in a
refrigerator until analysis. Wet samples (0.5–1.0 g) were
digested using HNO3 - HClO4 - H 2SO4 (1+1+5) for 30
min above 200°C (Ministry of the Environment, Japan,
2004). The mercury concentration in this solution was
measured by cold-vapor AAS (Nippon Instruments, AM2), following Hg0 generation with SnCl2. Mercury concentration in the sediments was normalized to dry-sample weight. The RSD of replicate measurements was less
than 4%.
3. Results
3.1 Wet and dry depositions
The annual wet and dry deposition fluxes of mercury at the Komae, Umihotaru and Narashino sites are
shown in Fig. 3. The mean (±standard deviation: SD) wet
and dry deposition fluxes at 10 sites in Japan during the
same period (Sakata et al., 2006) are also shown in the
figure. These sites include urban and industrial (3 sites),
remote (6 sites) and background (1 site) sites located
across the nation. There was no marked difference in the
annual wet deposition fluxes of mercury between the three
sites in the Tokyo Bay area. For the annual dry deposition fluxes, the highest flux (30.0 µg m–2yr –1) was observed at the Narashino site. Next, the mean atmospheric
deposition fluxes (wet plus dry: 38.7 µg m–2yr –1) of mercury at the Komae, Umihotaru and Narashino sites are
compared with those of Lakes Superior and Michigan.
Rolfhus et al. (2003) estimated a deposition flux of 9
µg m–2yr–1 in Lake Superior, on the basis of the work of
Fitzgerald et al. (1991) at a site in northern Wisconsin.
Moreover, Landis and Keeler (2002) calculated a deposition flux of 20.3 µg m–2yr –1 in Lake Michigan, using a
hybrid receptor modeling framework. Thus, the mercury
deposition fluxes in Tokyo Bay were 2–4 fold higher than
those in both lakes.
Sakata et al. (2006) found that the annual wet deposition fluxes of mercury at 10 sites in Japan are correlated significantly with the annual precipitation amount
(r2 = 0.80, P < 0.001). Thus, about 80% of the variance in
the mercury wet deposition flux is explained by the precipitation amount. This is probably because mercury wet
deposition is dominated by the precipitation scavenging
of reactive gaseous mercury (RGM) via Hg0 oxidation
by O3 and other oxidants in the gas and aqueous phases
(Pai et al., 1997; Schroeder and Munth, 1998; Shia et al.,
0
Hg concentration (ng g-1)
200 400 600 800 1000 1200
0
2000
10
1990
1980
1970
1960
30
1950
1940
1930
1920
1910
1900
40
50
Fig. 3. Annual wet and dry deposition fluxes of mercury at
Komae, Umihotaru and Narashino sites. For comparison,
the average (±SD) deposition fluxes at 10 sites in Japan are
given.
60
70
0
1999). The annual average (±SD) concentration of TGM
(primarily Hg 0 ) at 211 sites in Japan was 2.2 ± 0.5
ng m–3 in the fiscal year of 2004 (Ministry of the Environment, Japan, 2005), showing the relatively small difference in TGM concentration among sites. It is assumed
that there is no marked difference in atmospheric RGM
concentration throughout the entire country, although such
data are not available. As shown in Fig 3, the annual wet
deposition fluxes of mercury at the three sites in the Tokyo Bay area exceeded their mean at 10 sites in Japan,
which suggests that mercury wet deposition at the three
sites is affected by the precipitation scavenging of RGM
and particulate mercury (Hg(p)) from local sources.
The local emissions dominate the dry deposition of
trace elements in industrial and urban areas (Sakata et
al., 2006), which is clear from the significantly higher
dry deposition fluxes of mercury in the Tokyo Bay area
than the Japanese mean dry deposition flux (Fig. 3). At
the Komae site, the Hg(p) level is closely related to the
level of mercury emission from municipal solid waste
(MSW) incinerators near the site (Sakata and Marumoto,
2002). It appears that at the Umihotaru and Narashino
sites, MSW incinerators and industrial sources along the
bay contribute to mercury emission.
3.2 Water
The annual average (±SD) mercury concentration for
all samples (n = 70) from the three rivers was 6.3 ± 7.1
ng L–1. Furthermore, the averages (±SD) for each river
were 3.7 ± 1.6 ng L –1 for the Tama River, 9.7 ± 9.9
ng L–1 for the Ara River and 5.4 ± 5.6 ng L–1 for the Edo
River. Thus, the mercury concentration in river water
showed exceedingly large variations. On the other hand,
the average (±SD) mercury concentration in all the
Year
Depth (cm)
20
500
1000
1500
2000
Amount of Hg used (tons yr-1)
Fig. 4. Vertical distribution of mercury concentration in sediment core as a function of time based on 210Pb dating. The
total amount of mercury used in Japan from 1965 to 1995
(Asami, 2001) is also given.
seawater samples (n = 94) collected at 7 stations in Tokyo Bay was 0.98 ± 0.61 ng L–1, which corresponds to
approximately one seventh of the average in river water.
The mercury concentration (1.19 ± 0.74 ng L–1) in water
near the bottom was higher than that (0.44 ± 0.41 ng L–1)
in surface water due to an increase in the amount of mercury in particulate forms. In addition, the average (±SD)
concentrations of DGM (n = 22) in surface seawater and
TGM (n = 22) above the sea surface in Tokyo Bay were
0.052 ± 0.026 ng L–1 and 1.9 ± 0.6 ng m–3, respectively.
The DGM concentrations accounted for about 5% of the
total mercury concentrations in surface seawater.
3.3 Sediments
A sedimentation rate of 0.16 g cm–2yr–1 was obtained
for the sediment core collected at station 1 in Tokyo Bay
(Fig. 1). The vertical profile of 210Pb in the core shows
that the sedimentary strata are hardly disturbed by physical or biological mixing. The vertical distribution of mercury concentration in the sediment core is shown in Fig.
4 as a function of time based on 210Pb dating. The total
amount of mercury used in Japan from 1965 to 1995
(Asami, 2001) is also given in the figure. The mercury
concentration increased abruptly after 1950 and then
reached a maximum around 1970, which is consistent with
results reported by Matsumoto et al. (1983) for cores collected in 1980–1982. In addition, it is clear that there is
no marked difference in the mercury concentration in the
Mass Balance and Sources of Mercury in Tokyo Bay
771
sediment after the early 1980s, which is sufficiently high
compared with the background level of 50 ng g –1
(Matsumoto et al., 1983). On the other hand, the total
amount of mercury used in Japan decreased abruptly during the 1965–1975 period, and the amount (36 tons) of
mercury used in 1995 corresponded to only about 2% of
that (1683 tons) in 1965 (Asami, 2001).
An average (±SD) mercury concentration of 430 ±
89 ng g –1 was obtained for surface sediments (n = 14).
This value was slightly lower than the average (±SD)
mercury concentration of 538 ± 89 ng g –1 in the surface
sediments (n = 33) collected by Matsumoto et al. (1983)
in 1981–1982. This coincides with the fact that there is
no marked difference in the mercury concentration in the
sediment core after the early 1980s (Fig. 4).
4. Discussion
4.1 Mass balance of mercury in Tokyo Bay
The mass balance of mercury in Tokyo Bay was constructed on the basis of the analysis of mercury in atmospheric deposition, river water, seawater and sediment samples. We considered river discharge and atmospheric deposition as input terms, and evasion, export and sedimentation as output terms. The terms effluent discharges (e.g.,
sewage treatment and manufacturing) and particulate
remineralization were not included in this mass balance,
because no such data were available. In Japan, however,
there are currently few industrial uses of mercury, and
mercury is being eliminated from paint products and electrical apparatus. As mentioned earlier, the total amount
of mercury used in 1995 corresponded to only about 2%
of that in 1965 (Fig. 4). Balcom et al. (2004a) reported
that direct effluent/sewage inputs are only 5% of total
mercury inputs to Long Island Sound. Thus, it is assumed
that there is a very small contribution of mercury from
effluent discharge. Moreover, remineralization is regarded
to occur at a lower level (~2 kg yr–1) based on the value
used in the mass balance study for Lake Superior (Rolfhus
et al., 2003).
4.1.1 Atmospheric input
A relatively small difference was observed in the
annual wet-plus-dry deposition fluxes (38.7 ± 7.3
µg m –2yr–1) of mercury between the three sites studied
(Fig. 3). The atmospheric input of mercury to Tokyo Bay
was estimated using those fluxes and the bay area of 960
km2. The calculated annual input was 37 ± 7 kg yr –1.
4.1.2 River input
The river input of mercury to Tokyo Bay was estimated using the mercury concentration (6.3 ± 7.1 ng L–1)
in the three rivers studied and the best estimate (11.1 ±
2.8 km3yr–1) from the literature (Matsumura and Ishimaru,
2004) for the total annual water discharge from rivers
flowing into the bay. This provided an annual input of
772
M. Sakata et al.
70 ± 96 kg yr–1. The variance term was calculated by
propagation of errors based on the variability (SD) associated with the mercury concentration in river water and
the total annual water discharge. The extremely large variance in the river input of mercury is mainly due to the
large variations in the mercury concentrations in river
waters. These values were obtained primarily in ordinary
runoff. They are different from those obtained in
stormwater runoff during rainfall, as will be described
later. Thus, a considerable uncertainty exists about the
estimate of river input.
4.1.3 Evasion
The evasional flux of mercury from surface seawater
in Tokyo Bay was calculated using a gas-exchange model
(Wängberg et al., 2001) based on the measured DGM,
TGM, water temperature and wind speed. Details of calculations were described by Narukawa et al. (2006). The
mean (±SD) flux at 7 stations (n = 22) was 140 ± 120
ng m–2d–1. Assuming that this value corresponds to the
annual average in Tokyo Bay, the annual evasional flux
of mercury from the entire bay is estimated to be 49 ± 42
kg yr–1. There is a great variability in this estimate, because a considerable uncertainty exists about the model
and parameters. Nevertheless, the value (49 ± 42 kg yr–1)
employed was close to the atmospheric input of 37 ± 7
kg yr–1, which is similar to the results for Lakes Superior
and Michigan (Landis and Keeler, 2002; Rolfhus et al.,
2003).
4.1.4 Export
On the basis of the estimate of the water balance in
Tokyo Bay, the export flux of water from the bay was
assumed to be equivalent to the net water discharge into
the bay. This value is obtained from the difference between the total discharge (=rivers + effluent discharges +
precipitation) and evaporation. The export amount of
mercury from Tokyo Bay was then estimated by multiplying the mean mercury concentration (0.98 ± 0.61
ng L–1) in seawater by the net water discharge (13.7 ± 2.8
km3yr–1). The variability (SD) associated with the net
water discharge was assumed to equal to that of water
discharge from rivers, because there is much less variability associated with other terms (effluent discharges,
precipitation and evaporation) (Matsumura and Ishimaru,
2004). The calculated annual export flux was 13 ± 11
kg yr–1.
4.1.5 Sedimentation
The sedimentation flux of mercury to the bay bottom was estimated using the sedimentation rate (0.18 ±
0.09 g cm–2yr –1) reported by Matsumoto (1983), the measured mercury concentration (430 ± 89 ng g–1) in surface
sediments, and the area (640 km2) where sedimentation
occurs (Matsumoto et al., 1983). The variability (SD)
associated with the sedimentation rate was assumed to
be 50% of the mean value based on measurement of sedi-
Atmospheric input
37 ± 7 kg yr-1
River input
70 ± 96 kg yr-1
Evasion
49 ± 42 kg yr-1
Export
13 ± 11 kg yr-1
Table 1. Inputs and outputs of mercury and selected trace metals to Tokyo Bay.
Storage 14 ± 9 kg
Missing input
450 ± 506 kg yr-1
Sedimentation
495 ± 350 kg yr-1
Fig. 5. Mass balance of mercury in Tokyo Bay.
Hg
Cd
Cr
Cu
Pb
Zn
Input (tons yr– 1 )(a)
Output (tons yr– 1 )(b )
Input/Output
0.11
3.0
38
150
26
410
0.56
5.6
71
70
36
278
0.19
0.54
0.54
2.1
0.72
1.5
(a)
mentation rate (Matsumoto, 1983) as data were not available. The annual sedimentation flux of mercury obtained
was 495 ± 350 kg yr–1.
Figure 5 shows the results obtained for the mass balance of mercury in Tokyo Bay. The amount (14 ± 9 kg) of
mercury stored in the bay water was estimated using the
concentration of 0.98 ± 0.61 ng L–1 in seawater and the
bay volume of 14 km3. Figure 5 indicates that the sum of
river and atmospheric inputs is 107 ± 103 kg yr–1, while
that of sedimentation, evasion and export is 557 ± 403
kg yr–1. Thus, the removal flux considerably exceeds the
input. Sedimentation (495 ± 350 kg yr–1) contributed predominantly to the removal flux of mercury in the bay.
Cd, Cr, Cu, Pb and Zn have a similar pollution history to mercury (Matsumoto, 1983). As for mercury, we
are conducting mass balance estimates for those metals.
The procedures used for metal analyses were described
by Sakata et al. (2006). The results will be reported elsewhere. However, since their concentrations in the bay
water are not available at present, the export fluxes remain unknown. It is assumed that the export fluxes are
very small relative to the sedimentation fluxes, as inferred
from the mass balance of mercury in Tokyo Bay (Fig. 5).
For metals other than mercury, hence, the sum of river
and atmospheric inputs was compared with sedimentation as the sole output term. The result indicates that the
inputs of those metals are within the range of 0.54–2.1 of
their respective outputs (Table 1). Thus, the imbalances
between the inputs and outputs for Cd, Cr, Cu, Pb and Zn
are much smaller than that for mercury (input/output =
0.19, Table 1). The extremely large deficiency in mercury input suggests that there are other major inputs of
mercury to Tokyo Bay, corresponding to approximately
450 kg yr–1 (Fig. 5).
4.2 Sources of mercury in Tokyo Bay
There were exceedingly large variations in the mercury concentration between waters from the three rivers,
as described earlier. It was found that mercury concentration in each river correlates significantly with the concentrations of Al (r2 = 0.33–0.75, P < 0.001 except for
the Edo River (P < 0.01)) and Fe (r2 = 0.60–0.88, P <
Input terms: river discharge and atmospheric deposition.
Output terms: sedimentation, evasion and export for
mercury and sedimentation for other trace metals.
(b)
0.001), as shown in Fig. 6. This suggests that mercury in
rivers may originate mainly from soil, because Al and Fe
are major components of soil.
In Japan, large amounts (about 2500 tons as Hg) of
organomercurous fungicides, primarily phenylmercury
acetate, were used extensively in fields in 1952–1974 (The
Chemical Society of Japan, 1977), and most of the mercury was retained in the soil. In fact, the mercury levels
(290 ± 460 ng g–1, n = 469) in Japanese field soils are
significantly higher than those (<100 ng g–1) in background soils (Asami, 2001). Masunaga et al. (2001) reported that pentachlorophenol (PCP) and chloronitrophen
(CNP), which were used extensively as paddy field herbicides in the past, contribute to about 50% (in terms of
TEQ) of the dioxins, even in recent Tokyo Bay sediment,
based on congener composition (Masunaga et al., 2001).
This means that dioxins scattered as herbicide impurities
are still in soil and are gradually running off. Similarly,
mercury originating from organomercurous fungicides in
field soil may contribute significantly to the discharge to
Tokyo Bay.
During rainfall, a large amount of suspended sediment from terrestrial soil and river bottom sediment is
discharged into estuaries through stormwater runoff. In
this study, the mercury concentration in rivers was measured primarily in ordinary runoff. In addition, fields,
particularly paddy fields, which are potentially contaminated with mercury, spread generally along rivers. This
leads to the hypothesis that field soil discharged into
stormwater runoff is a major source of mercury in Tokyo
Bay. River bottom sediment that originated from fields
also contributes to the mercury discharge due to its suspension in stormwater runoff. High concentrations (>400
ng g–1) of mercury have been observed in some river bottom sediments downstream in the Tokyo area (Geological Survey of Japan, 2006). As a preliminary approach to
validating this hypothesis, we measured the concentrations of mercury and selected trace metals (Cd, Cr, Cu,
Mass Balance and Sources of Mercury in Tokyo Bay
773
50
50
Ara River
r2 = 0.75
40
Tama River
r2 = 0.59
30
Hg (ng L-1)
Hg (ng L-1)
40
20
Edo River
r2 = 0.33
10
1
2
3
Al (mg L-1)
4
Tama River
r2 = 0.60
30
20
Edo River
r2 = 0.61
10
0
0
Ara River
r2 = 0.88
5
0
0
1
2
3
Fe (mg L-1)
4
5
Fig. 6. Relationship between mercury concentration and concentrations of aluminum and iron in each river.
Table 2. Concentrations of mercury and selected trace metals in Tama River during typhoon storm (July 27 and August 26, 2005).
Storm water runoff (A)
–1
Hg (ng L )
Cd (µg L– 1 )
Cr (µg L– 1 )
Cu (µg L– 1 )
Pb (µg L– 1 )
Zn (µg L– 1 )
(a)
July 27
August 26
60
0.47
2.8
13
7.0
28
185
0.84
7.3
35
27
80
Ordinary runoff (B)(a)
July 27
August 26
16
2.6
1.8
1.2
8.0
1.5
50
4.7
4.6
3.2
31
4.2
Average concentration (±SD) of each metal primarily in ordinary runoff from December 2003 to November 2004.
Pb and Zn) in the Tama River on July 27 and August 26,
2005 during a typhoon. The samples were collected on a
bridge near the Komae site (Fig. 1). The concentration of
each metal was compared with the concentration primarily in ordinary runoff from December 2003 to November 2004. The result is indicated in Table 2. The mercury
concentrations in stormwater runoff on July 27 and August 26 increased to 16 and 50 times the mean value in
ordinary runoff, respectively, which is much higher than
the increases for other metals. This does not contradict
the fact that the imbalances between the inputs and outputs for Cd, Cr, Cu, Pb and Zn in Tokyo Bay are much
smaller than that for mercury (Table 1).
Thus, the limited result tends to support the above
hypothesis, viz., field soil discharged into stormwater
runoff is a major source of mercury in Tokyo Bay. This
implies that the present pollution level of mercury has
been maintained in Tokyo Bay over a very long time,
which may be a reason why there is no marked difference
in mercury concentration in the sediment core after the
early 1980s (Fig. 4). It is likely that discharge of mercury
774
3.7 ± 1.6
0.18 ± 0.15
1.6 ± 0.8
11 ± 5
0.88 ± 0.65
19 ± 11
A/B
M. Sakata et al.
through stormwater runoff explains the imbalance between the inputs and outputs of mercury in Tokyo Bay
(Fig. 5). However, further research is required to quantify the relative importance of mercury discharge due to
stormwater runoff and to consequently construct an accurate mass balance of mercury in Tokyo Bay. In addition, whether this stormwater runoff is a principal source
of methylmercury should be confirmed on the basis of a
mass balance study.
Acknowledgements
Experiments in this study were carried out at the
Central Research Institute of Electric Power Industry,
where three of the authors (M.S., K.M. and M.N.) were
research scientists. We wish to thank T. Okabe, K.
Fukumori and H. Narutaki (Electric Power Engineering
System Co., Ltd.) for assistance in the sampling and analyses, and M. Arai (Metocean Environment Inc.) for help
with the sampling in Tokyo Bay. We are grateful to Dr. P.
H. Balcom and an anonymous reviewer for their helpful
comments and suggestions.
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