35

Long-term Mercury Behavior in Sulfurized/Solidified Mercury
Wastes by a Simulated Landfill Experiment Using Lysimeters
Akira SANO1*, Keizo KAWASE2, Ryuji YANASE2, Masaki TAKAOKA3,
Akito MATSUYAMA4, Fumitake TAKAHASHI5 and Takafumi KATO1
Department of Chemical Engineering, Fukuoka University
8-19-1, Nanakuma, Jonan-ku, Fukuoka 814-0180, Japan
2
Environmental Protection Center, Fukuoka University
8-19-1, Nanakuma, Jonan-ku, Fukuoka 814-0180, Japan3
Department of Environmental Engineering, Graduate School of Engineering, Kyoto University
C-1-3-461, Katsura, Nishikyo-ku, Kyoto 615-8540, Japan
4
National Institute for Minamata Disease4058-18, Hama, Minamata, Kumamoto 867-0008, Japan
5
Department of Environmental Science and Technology, Tokyo Institute of Technology
4259, Nagatsuta, Midori-ku, Yokohama, Kanagawa 226-8502, Japan*E-mail: sanoakira@fukuoka-u.ac.jp*1

Abstract
Mercury (Hg) released from landfill Hg waste poses serious health and environmental risks. The
Minamata Convention on Mercury (2017) emphasizes preventative management: landfills are to be stabilized
and Hg waste controlled. When developing stabilization/solidification technologies to ensure the safety oflandfilled Hg waste, leaching and headspace tests are currently used to assess dissolution and diffusion.
However, it remains unclear whether Hg wastes remain stable in real landfill environments because there arefew findings from long-term landfill experiments on Hg wastes. We used lysimeters and followed previous
long-term semi-aerobic/anaerobic simulated-landfill experiments on dry-cell batteries to investigate Hgdissolution and diffusion of sulfurized/solidified Hg wastes in simulated landfill environments. Total Hg (THg) release from the lysimeters, containing sulfurized/solidified Hg waste, to the atmosphere was extremely
small. Examination of Hg behavior in different landfill types, however, found semi-aerobic types to be
associated with lower Hg dissolution and diffusion risks than anaerobic types. Regarding Hg behavior of
sulfurized/solidified Hg wastes, Hg sulfide (HgS) waste solidified by cement was found to be more stable
than non-solidified HgS waste and dry-cell batteries. Methyl Hg (M-Hg) monitoring in leachates found M-Hg
concentration in leachates to fall below 0.5 µg/L under almost all landfilling conditions, with time change
trends moving in tandem with changes in leachate T-Hg. Volatilization and leaching tests involving sulfurized/
solidified Hg wastes were performed while simulating landfill changes to judge important factors in
landfilling conditions regarding Hg dissolution and diffusion through comparison with lysimeter experimentresults. Laboratory tests showed temperature and pH (alkaline) to affect the stability of sulfurized/solidified
Hg wastes and elucidated effects assumed to enhance Hg dissolution and diffusion in the lysimeter
experiments.
Key words :	anaerobic, cement solidification, mercury sulfide, semi-aerobic, stabilization

1. Introduction
Mercury (Hg) is emitted into the atmosphere fromboth anthropogenic and natural sources. Because it ishighly volatile, it subsequently enters oceans, lakes andrivers either directly from the atmosphere or from
deposits in surrounding basins, even when no specificsource of Hg exists. Hg handling and disposal are
addressed by the 2017 Minamata Convention on Mercury
Global Environmental Research
24/2020: 035-043
printed in Japansano.indd

35(UNEP, 2017). Hg release from landfills poses serious
health and environmental risks (Tao et al., 2017), andbiogeochemical transformations caused by landfilling of
unstable Hg waste, which is highly susceptible todissolution and diffusion, increase these risks.
Many researchers have developed stabilization/solidification technologies applicable to Hg waste
treatment methods, including sulfurization of Hg to Hg
sulfide (HgS) (López et al., 2010), sulfur (S) polymer©2020 AIRES

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A. SANO et al.

encapsulation (López et al., 2015), and treatments using
magnesium potassium phosphate (Cho et al., 2014) or
activated carbon and cement (Zhang & Bishop, 2002).
Japan adopted the policy on environmentally soundmanagement of mercury wastes recommended by theCentral Environment Council in 2015 (Central
Environment Council, 2015). To ensure environmentallysound management of Hg wastes, wastes consisting of
Hg or Hg compounds are designated as “industrial waste
requiring special management” due to their toxicity. In
the Japanese guideline for Hg wastes (JGL-HgW), wasteconsisting of Hg must be stabilized to HgS by
sulfurization using a ball-mill prior to landfill disposal toprevent Hg dissolution and diffusion. Then, solidification
using S polymer is undertaken to enhance the stability ofthe sulfurized Hg (Ministry of the Environment, 2017).
The stability of the sulfurized/solidified Hg waste isassessed by observing its dissolution and volatilization.
Although fundamental stability of sulfurized/solidified
Hg wastes can be obtained from leaching and headspacetests, it is unclear whether sulfurized/ solidified Hg
wastes remain stable in real landfill environments overthe long term. The latest research was undertaken fromthe viewpoint of long-term management of Hg wastes,
assessing the stabilization/ solidification technologies
developed through multiple laboratory tests (Kusakabe &Takaoka, 2020). Hg was positioned as a hazardous metal
and its distribution and emission in and from landfills
were surveyed on the basis of data derived from real
landfills containing industrial and domestic wastes (Chaiet al., 2013; Yang et al., 2018). However, there have beenfew studies involving long-term monitoring of Hg in
specialized landfills for Hg wastes. Yanase provides arare insight into Hg behavior in a continuous simulatedlandfill lysimeter experiment using dry-cell batteries(Yanase, 2005). Long-term monitoring data would bevaluable for Hg waste management in real landfills.
We identified the Hg behavior of sulfurized/
solidified Hg wastes in lysimeters following Yanase’s
long-term simulated-landfill lysimeter experiment toinvestigate how conditions in the landfill interior affectHg behavior. Eight lysimeters were operated concurrently
in different conditions of landfill type (semi-aerobic or
anaerobic type) and landfilling Hg waste variety(sulfurized or solidified) over the long term (five years).
On the other hand, there have been few comparative
studies of methyl Hg (M-Hg) formation in landfilled
sulfurized/solidified Hg wastes over the long term. In thisstudy, Hg behavior in leachate was examined separately,
divided into total Hg (T-Hg) and M-Hg for valuable dataacquisition.
The sulfurized/solidified Hg wastes used in our
simulated-landfill experiment are subject to multiple
influences from landfilling conditions such as changes in
the gaseous environment in terms of being semi-aerobic
or anaerobic, and changes in the liquid environment insano.indd

36

terms of dissolution and adsorption states. Importantfactors in Hg dissolution and diffusion under landfilling
conditions are hard to judge solely from simulatedlandfill experiment results. We performed laboratory tests
using the same samples used in the simulated-landfill
experiment to determine landfilling conditions that affectHg dissolution and diffusion. We explored the stability ofsulfurized/ solidified Hg wastes using laboratory tests ofisolated landfilling conditions including temperature, pH
and leachate materials. Here, we present our results on
the solubility and volatility of sulfurized/solidified Hg
wastes under different conditions of leaching andvolatilization tests. Finally, we investigate similaritybased findings between the simulated-landfill lysimeter
experiment and the laboratory tests.

2. Materials and Methods
2.1 Simulated-Landfill Lysimeter Experiment2.1.1 Mercury Waste Samples
Hg recovered from waste was ball-milled withpowdered S at an S/Hg molar ratio of 1.05 to yield a basic
sulfurized Hg waste (SMW) (Fukuda et al., 2014). The
SMW powder was also solidified using cement to further
enhance stability; the cement/Hg weight ratio was 1. This
100-mm cubic monolith was termed C-SMW. We usedsulfurization and solidification methods specified in JGLHgW (Ministry of the Environment, 2017) and a wellknown solidification technology used to prevent heavy
metal dissolution (Yu et al., 2005), respectively. The
technologies were selected in a phased manner under aplan to compare with a previous study targeting nontreated Hg waste (Yanase, 2005) and a future study
targeting new solidification technologies such as an S
polymer recommended by JGL-HgW.
2.1.2 Landfill Simulation Lysimeters
Landfills were simulated using lysimeters (length
250 × width 250 × depth 600 mm) operated under semiaerobic or anaerobic conditions by sidewall and leachatecollection pipes as shown in Fig. 1 (a: semi-aerobic, b:
anaerobic). Atmospheric air could penetrate into thelysimeter’s layers through a side wall pipe opened only
for the semi-aerobic type (Zhang et al., 2019) and could
not gain entrance to the lysimeter at height of the leachatecollection pipe only in the anaerobic type (Hirata et al.,2011). This lysimeter aerobic control using piping wasused in previous studies and the obtained findings (e.g.,leachate quality and gas composition) confirmed theauthenticity of simulated semi-aerobic/anaerobic landfill
conditions (Threedeach et al., 2012; Yanase, 2005). The
waste layer (depth 450 mm) in the lysimeters wascomposed of solid incineration residue and sewage sludge
compost. The SMW was either center-spotted onto thelayer as shown in the spotted pattern (Fig. 1b), or
thoroughly admixed with the layer as shown in the mixed
pattern (Fig. 1a). The C-SMW monolith was center-2020/07/30

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37

Fig. 1 Schematic diagram of simulated-landfill lysimeters (a: semi-aerobic type with mixed pattern, b: anaerobic type with spotted pattern).

Table 1 Landfill conditions for simulated landfill experiment.

Name of lysimeter

LS-1

Landfill type

LS-2

LS-4

LA-1

LA-2

Semi-aerobic

LA-3

LA-4

Anaerobic

Mercury waste samples

SMW

SMW

C-SMW

none

SMW

SMW

C-SMW

none

Landfill patternmixedspottedspotted-mixed

spottedspotted-Mercury content in landfilled wastes

335 g

335 g

421 g

none

335 g

335 g

421 g

none

Weight of solid wastes
Composition of solid wastes

34 kg-wet (1.2 kg/L of layer)
Incineration residue (water content: 85%)
Sewage sludge compost (water content: 85%)

spotted. Table 1 lists the weight proportions of the Hg
wastes, incineration residue and sewage sludge compostin each of the layers. The Hg load was approximately
100-fold more than the usual level of an industrial solid
waste landfill. Eight lysimeters were prepared simulating
different landfill types (semi-aerobic [LS] and anaerobic[LA]) and examples of either landfill pattern were
numbered 1 (mixed SMW), 2 (spotted SMW), 3 (spotted
C-SMW), and 4 control (no Hg waste), as shown in
Table 1.
The lysimeter experiment was performed over a spanof five years in a ventilated room. The upper layer of
each lysimeter underwent semimonthly water sprinkling(6 L each time). The leachates were drained from thecollection pipes after sprinkling. The T-Hg concentrationin the leachate samples was measured via a direct thermal

sano.indd

LS-3

37

: 80%
: 20%

decomposition Hg analyzer (MA-3000; Nippon
Instruments Co., Ltd., Japan) and the M-Hg concentrationin the leachate samples was determined via gas
chromatography with electron capture detection (GCECD, G-3800; Yanaco Technical Science Co., Ltd.,
Japan) using a procedure that involves the precondensation of the Hg species by the extraction of Hg
dithizonates (Akagi & Nishimura, 1991; Matsuyama etal., 2011). The pre-condensation of the NIMD method(Akagi method) provides a lower limit detection of
0.02 ng/L at a higher accuracy of measurement than the
more commonly used method (0.5 µg/L).
The gas used to fill each lysimeter after eachprevious gas sampling, was vacuum-aspirated for 20 min.
at a rate of 0.5 L/min. and passed through tubes packed
with gold amalgam to capture the Hg, which was detected

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A. SANO et al.

via a direct thermal decomposition Hg analyzer equippedwith an atomizer (RH-MA3; Nippon Instruments Co.,
Ltd.). The amount of Hg diffused from the lysi-meters
was calculated by dividing the amount of Hg detected by
the number of days elapsed from the previous samplingday. For the gas sampling method we referred to a GHG
emission assessment for simulated landfills using
lysimeters (Hirata et al., 2011) and applied it to mercury
measurement (Yanase, 2005).
2.2	Sulfurized/solidified Mercury Waste Stability
Assessment
2.2.1 Mercury Waste Samples and Pretreatment
SMW and C-SMW were subjected to volatilizationand leaching tests. Because C-SMW ages during cement
curing, the surface and inner part of the C-SMW monolith
may differ in terms of Hg volatility and dissolution.
Therefore, the tests were performed using C-SMW
monolith samples that had or had not been washed withwater. The C-SMW surface was rinsed away with 500 mL
water and then scrubbed with a plastic brush on a tray
filled with 1.5 L of water.
The washed and unwashed C-SMW monolithsamples were used intact for volatilization testing;another pair of monolith samples was crushed into
particles <0.5 mm in diameter prior to the leaching test;
while SMW in powder form was neither crushed nor
screened before the volatilization and leaching tests.
2.2.2	Mercury Volatilization Tests at Different
Temperatures
To explore the effects of temperature on Hg volatility
from SMW, C-SMW monoliths, and washed C-SMW
monoliths, we prepared a laboratory-scale device
featuring a desiccator with a petri dish. SMW (50 g) or a
C-SMW monolith (1.2 kg) was placed on the dish andheld at different temperatures (10 °C, 30 °C and 50 °C)
for seven days. Desiccator gas (21 L) was vacuumaspirated for 42 min. at 0.5 L/min. by a vacuum pump for
one, two, four and seven days. The gaseous Hg wasmeasured via the tube capture method described above.
The amount of Hg detected was added cumulatively at
every temporal sampling for each volatilization test andthe Hg volatilization rate was calculated by dividing the
cumulative Hg amount by the test time and effective
surface area of the Hg waste samples. Here, the values of2
the effective surface area were 60.8 cm for the SMW and2
365 cm for the C-SMW monoliths.
2.2.3 Leaching Tests Using Various Solvents
We explored the effects of pH and ions on
dissolution of Hg from SMW, C-SMW powder andwashed C-SMW powder using various solvents. The
sample and solvent were placed in sealed vials at a liquid/
solid ratio of 10. Each vial was mixed for 6 hrs. on a tube
rotator operating at 30 rpm at room temperature, eachsolvent was filtered through a 0.45-µm-pore-sized
membrane filter, and the pH and Hg concentrations were
measured using a pH meter and a thermal-decomposition
Hg analyzer (MA-3000), respectively. We varied the pH++
and the cation and anion levels: cations (Na , K , and2+Ca ): 100 –10,000 mg/L; anions (Cl , NO3 , and SO42-):
50 – 2,000 mg/L; pH (adjusted using HNO3 or NaOH):
3–11. These settings encompassed all known real landfill
environments (J. FIWMRA, 2010).

3. Results and Discussion
3.1 Mercury Behavior in Lysimeters3.1.1 Mercury Diffusion into the Atmosphere
Figure 2 shows the changes over time of the amountof Hg diffused from the lysimeters. The amounts of Hg

Fig. 2 Changes in amount of Hg diffused to the atmosphere for five years in lysimeter experiment.

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Long-term Mercury Behavior in Sulfurized/Solidified Mercury Wastes
from the mixed pattern (LS-1 and LA-1) were large
because SMW particles were present at the surface of the
top layer. The amounts of Hg diffused from all of thelysimeters fluctuated seasonally and the peaks of theamounts of Hg diffused, observed uniformly at 7, 19, 31
and 43 months, corresponded to summer season. Figure 3
presents a plot of the amounts of Hg diffused from LS-1
and LA-1 against room temperature. We obtained a
relationship between the amounts of Hg diffused from thelysimeters and the air temperature. Matsuto et al. foundthat temperature differences in the layer were the driving
force behind gas flow through the layer (Matsuto et al.,2015) and the driving force occurred due to seasonalvariation in temperature affecting Hg diffusion from thelysimeters.
Over time, the amounts of Hg diffused from thesemi-aerobic type (LS-2, LS-3, and LS-4) remained at
extremely low levels and gradually fell with time (Fig.
2a). This indicates that Hg diffuses less to the atmosphere

Fig. 3 Plot of amounts of Hg diffused in LS-1 and LA-1 againstroom temperature.

39

from SMW and C-SMW through semi-aerobic-type
landfills over the long term. LS-3 was similar to LS-4
(the control) in terms of amounts of Hg diffused. This
indicates that cement solidification effectively inhibits Hg
diffusion in semi-aerobic type landfills.
When comparing the landfill types (LS and LA
series), the amounts of Hg diffused were slightly larger
from the anaerobic lysimeters than from the semi-aerobic
lysimeters for both SMW and C-SMW. The amounts of
Hg diffused were represented by spikes at 34 and 44
months in LA-2 or LA-3 (Fig. 2b), the spikes may indicatethat gas clogged within the layer was released all at oncefrom the anaerobic type in which there was little air flow
through the layer.
3.1.2 Mercury in Leachates
The changes over time of T-Hg concentration in theleachates are shown in Fig. 4a (semi-aerobic lysimeters).
The T-Hg concentration of LS-1 was about 10 mg/L after
one month, decreasing gradually to about 0.01 mg/L overthe last four years. However, even this concentration
exceeded the effluent standard for wastewater in Japan(0.005 mg/L). The LS-2 leachate T-Hg concentrations
ranged from 0.01 to 0.1 mg/L initially and fell below theenvironmental standard for water pollution in Japan(0.0005 mg/L) over the final two years. This is assumed
to be due to small amounts of center-spotted SMW
powder escaping into the bottom layer. The LS-3 leachate
T-Hg concentrations ranged from 0.005 to 0.01 mg/L
initially and became stable below the environmental
standard from the second year. The Hg dissolution risk
under real landfill environments is thus lower withcement solidification than with ball-milling sulfurization.
When the two landfill types were compared, leachate
T-Hg concentrations were higher from the anaerobic type(Fig. 4b) than the semi-aerobic type. An aquifer attaining
half the depth of an anaerobic type landfill was found tonegate Hg adsorption in the layer, although the leachate
Hg was adsorbed by incineration ash (Yanase, 2005). The
times taken for the T-Hg concentrations in the anaerobic
type to fall below the effluent standard were longer thanFig. 4 Changes in T-Hg concentration in leachate during the five years of the lysimeter experiment.

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A. SANO et al.

Fig. 5 Changes in M-Hg concentration in leachate during the five years of the lysimeter experiment.

Fig. 6 Changes in pH of leachate during the five years of the lysimeter experiment.

in the semi-aerobic type, but the leachate T-Hg
concentrations stabilized below the environmental
standard by the fourth year. From the viewpoint of longterm management, we found both ball-milling
sulfurization and cement solidification to have low Hgdissolution risk under real landfill environments
regardless of landfill type.
Changes over time in the M-Hg concentration in theleachates are shown in Fig. 5. The M-Hg concentrationsof the LS series (semi-aerobic types) fell below thecommon lower limit detection of 0.5 µg/L (Fig. 5a),because Hg methylation, caused by anaerobic bacteria,
hardly occurs under aerobic conditions. Anaerobic
conditions promote Hg methylation in landfills as is wellknown (Chai et al., 2015). The leachate M-Hg in LA-3
was at almost the same level as in LA-4 (the control),while, in LA-1 and -2 M-Hg was detected transiently atcommon levels (Fig. 5b). When the two landfill types

sano.indd40

were compared, leachate M-Hg concentrations werehigher from the anaerobic type than from the semiaerobic type. The changes in leachate M-Hg
concentration over time in the leachates tended to movein tandem with leachate T-Hg changes in each lysimeter.
Hg methylation in real landfills is assumed to be
associated with Hg dissolution in Hg waste, and thestability of sulfurization and cement solidification islinked with a reduced Hg methylation risk.
We checked the water quality of the leachates to seek
factors that could cause increased Hg dissolution under
landfill environments. The changes over time in the pHof the leachates are shown in Fig. 6. The pH in the firsthalf period, in which Hg in the leachates was detected,
hovered around 8.6 in the LS series and around 9.1 in the
LA series. Pure HgS reagent (metacinnabar) has theproperty of increasing Hg solubility as pH becomes
alkaline (Clever et al., 1985). The pH effect was

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Long-term Mercury Behavior in Sulfurized/Solidified Mercury Wastes

41

Fig. 7 Results of volatilization and leaching tests (a: plot of Hg volatility against temperature, b: effects of solvent pH on Hg solubility).

consistent with the result of the lysimeter experiment,
where the anaerobic type, which showed a higher pH ofleachates, had higher Hg concentrations detected in theleachates.
3.1.3 Mercury Remaining in Lysimeters
Yanase examined the Hg behavior of dry-cell battery
waste by using a lysimeter landfill simulation (spotpattern) similar to our study (Yanase, 2005). For
comparing Hg stability in lysimeter experiments between
sulfurized/solidified Hg wastes and dry-cell battery
waste, the amount of Hg remaining in each lysimeter was
calculated by subtracting the cumulative amount of Hgemissions over the five years from the initial amount of
Hg. The percentage of Hg remaining was 99.9583% in
LS-1 and 99.9087% in LA-1, even for the mixed pattern,
for which higher Hg emissions were detected in our
experiment. Yanase reported that the percentage of Hg
remaining after seven years was 99.337% in the semiaerobic type and 99.41% in the anaerobic type (Yanase,2005). The comparison showing that the percentage of
Hg remaining was higher in our study than in Yanase’s
study proves that Hg sulfurized by ball-mill following
JGL-HgW has significantly higher stability than Hg
contained in dry-cell batteries. When comparing thepercentage of Hg remaining between semi-aerobic andanaerobic types, Yanase’s results contradict our results
showing the percentage for spotted SMW to be 99.9997%
in the semi-aerobic type (LS-2) and 99.9989% in theanaerobic type (LA-2). This conflicting finding issurmised to indicate Hg dissolution from dry-cell battery
to depend greatly on whether the exterior of the dry-cellbatteries is corroded due to oxygen in the layer (Yanase,2005); while Hg dissolution from SMW influenced bycontact with an aquifer occurred only in anaerobic
landfilling in our lysimeter design. Moreover, theremaining Hg percentage was highest in LS-3 (oversano.indd

41

99.9999%) and LA-3 (99.9999%). This result indicatesthat cement solidification maintains stability almost
completely when treated Hg waste is exposed to real
landfill environments over the long term.
3.2	
Mercury Volatility of Sulfurized/solidified Hg
WastesFigure 7a plots Hg volatility against temperature in
the volatilization tests. SMW and washed C-SMW
exhibited similar Hg volatility but the temperaturedependence of such volatility was more evident for
SMW. Real landfill temperatures are high due tobiodegradation of organic matters and seasonaltemperature changes (J. FIWMRA, 2010). We describedseasonal temperature variation affecting diffused Hgconcentrations, finding Hg diffusion of SMW to be higher
than of C-SMW in lysimeters simulating landfills. The
results obtained from the lysimeter experiment arethought to be related to differences in temperaturedependence obtained from the volatilization tests.
The Hg volatility from unwashed C-SMW was 25fold greater than that from washed C-SMW. Volatility of
C-SMW differs between the surface and inner portions
because only the surface of C-SMW is exposed to air
during cement curing. Thus, it is thought that most of the
diffused Hg in lysimeters LS-3 and LA-3 (Fig. 2) was
attributable to the C-SMW surface. The unstable Hg on
the surface was removed by washing with water. Thus,
washing C-SMW reduces the risk of Hg diffusion in
landfills.
3.3	
Mercury Dissolution from Sulfurized/solidified
Hg Wastes
JLT-13 for industrial wastes stipulates assessment of
whether Hg concentrations in solvents exceed a criterionof 0.005 mg/L (Environment Agency, 1973). We used the2020/07/30

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A. SANO et al.

criterion of 0.05 mg/g in the present study because the Hg
content varied between SMW and C-SMW. Figure 7b
presents the effects of pH on dissolved Hg content; thedashed line is the “converted” JLT-13 criterion. The
amount of HG dissolved from SMW increased underalkaline conditions, as found in a previous study(Mizutani, 2014). Svensson et al. found that someresidual sulfur in SMW is ionized to HgS22- under strongly
alkaline conditions (Svensson et al., 2006). While,
C-SMW (surface-washed or not) met the JLT-13 criterion
over the entire pH range. HgS was weakly soluble underalkaline conditions; the use of C-SMW resolves thisproblem. The alkaline effect obtained from the leaching
test results is assumed to play a role in increasing thedissolution risk in anaerobic-type landfills in thesimulated-landfill lysimeter experiment. The washed
SMW leaching test results implied that surface washingsolidified Hg waste marginally reduces Hg solubility.
The amounts of HG dissolved from all samples met
the JLT-13 criterion at all cation and anion levels tested.
++
Both SMW and C-SMW were stable to cations (Na , K ,and Ca2+) and anions (Cl−, NO3−, and SO42−) at
concentrations encompassing all known real landfill
environments.

4. Conclusions
We operated eight lysimeters to clarify the behaviorof Hg in sulfurized/solidified Hg wastes in lysimeterssimulating landfills and explored the effects of controlled
landfill conditions. Cement-solidified Hg waste was morestable than ball-milling-sulfurized Hg waste. The semiaerobic landfill type reduced the solubility and volatility
risks of sulfurized/solidified Hg wastes.
We performed volatilization and leaching tests ofsulfurized/solidified Hg wastes under conditions
simulating landfill environments, and defined the factors
rendering such waste unstable. Sulfurized Hg waste was
weakly soluble under alkaline conditions and weakly
volatile with temperature; cement solidification was
effective. Alkaline and temperature effects are assumed tobe related to Hg dissolution and diffusion behaviors inthe simulated-landfill lysimeter experiment.
The similarity-based findings between the simulatedlandfill lysimeter experiment and laboratory tests
suggested that changing operating conditions and using
solvents for volatilization and leaching tests enableassessment of Hg stability, which was affected by landfill
conditions such as temperature and pH.
Acknowledgements
This research was financially supported by JSPS
KAKENHI Grant Number JP17K06623 and Environment
Research and Technology Development Funds(3K143002, 3-1701) of the Environmental Restoration
and Conservation Agency of Japan.

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42

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Akira SANO
Akira Sano is an assistant professor in the
Department of Chemical Engineering at Fukuoka
University. He studies the synthesis technology
and assessment methodologies for functional
porous materials, contributing to biomass
recycling, bioenergy, bioreactors and mercury
waste management. He formerly worked at Kyoto
University as a researcher participating in a project on mercury control(2014 – 2015), at the National Institute for Environmental Studies (NIES) asa researcher participating in projects on developing and assessing biomass
recycling system (2009 – 2013), and at Yokohama National University as a
post-doctoral entrepreneur for a biohydrogen production systems (2006–2008). He holds a Ph.D. in Engineering from Nagoya University, Japan.

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Keizo KAWASE
Keizo Kawase is a technical staff member at theEnvironmental Protection Center, Fukuoka
University. His job involves analytical work forenvironmental management at the university. He
received his Master of Engineering from the
Graduate School of Environmental Engineering
at Kitakushu University. His research fields
include industrial health and safety, and environmental chemistry.

Ryuji YANASE
Ryuji Yanase directs the Environmental Protection
Center in Fukuoka University. He is certified as a
Professional Engineer in Japan and serves as amember in formal groups for management andassessment of landfill sites and wastes. His
primary subject of study is stability and control of
toxic solid waste landfills and his other researchinterests are long-term landfilling and mercury waste landfilling.

Masaki TAKAOKA
Masaki Takaoka is a professor in the Departmentof Environmental Engineering, Graduate Schoolof Engineering, Kyoto University, Japan. He is amember of the Minamata Convention Group ofTechnical Experts on Mercury Waste Thresholds
and a resource person under the UNEP Global
Mercury Partnership ― Waste Management
Partnership Area. His main research subject is management and treatmentof solid waste. He is particularly interested in controlling toxic substancessuch as heavy metals including mercury, persistent organic pollutants andradioactive nuclides in waste treatment processes and waste-to-energy
projects.

Akito MATSUYAMA
Akito Matsuyama is currently chief director for
research planning at the National Institute for
Minamata Disease and director of the Department
of International Affairs and Research. He
previously studied the behavior of mercury in
Minamata Bay, and is now focusing particularly
on mercury behaviors in sediment and seawatercolumns. He worked at Taisei Corporation from 1988 to 2003, entering the
Ministry of the Environment (National Institute for Minamata Disease) in
2003, where he continues to this study.

Fumitake TAKAHASHI
Fumitake Takahashi is an associate professor in
the Department of Transdisciplinary Science andTechnology, School of Environment and Society,
Tokyo Institute of Technology. He received his
Ph. D. in Urban Engineering at the University ofTokyo. His current research fields include waste
management, waste recycling, environmental risk
assessment, socio-psychological analysis of waste recycling behaviors andrecycling tools.

Takafumi KATO
Takafumi Kato is an associate professor at theLaboratory of Industrial Inorganic Chemistry,
Department of Chemical Engineering, Faculty ofEngineering, Fukuoka University. He received his
Ph. D. in Engineering from Kyushu University in
1994. His research interests are primarily in thefield of materials science, specifically, synthesis
and applications of porous carbon materials from sustainable precursors.
(Received 13 January, Accepted 26 June 2020)

sano.indd

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