3Emission Control of Mercury from Stabilized and Solidified
Products under Monofill Conditions
Masahiro SATO1*, Tomonori ISHIGAKI2, Kazuto ENDO2 and Masato YAMADA2
Faculty of Engineering, Hokkaido University
Nishii 8, Kita 13, Kita-ku, Sapporo, Hokkaido 060-8628, Japan
2
Center for Material Cycles and Waste Management Research, National Institute for Environmental Studies (NIES)
16-2, Onogawa, Tsukuba, Ibaraki 305-8506, Japan*E-mail: satomasahiro@eng.hokudai.ac.jp*1

Abstract
A column leaching test was conducted for 258 days to demonstrate Hg emissions in an engineered
landfill where stabilized and solidified mercury by using sulfur and a binder such as cement was disposed ofunder monofill conditions with a granite soil layer. Thermodynamical analysis considering leachate quality
showed that the mercury in the leachate could be chlorides, indicating that the mercury in metacinnabar instabilized/solidified products (below, s/s products) dissolved in percolation water. However, the gaseous andliquid emissions of soluble mercury were extremely low because of adsorption in the granite soil. Nearly all
mercury released from the stabilized/solidified product into the soil stayed within 1.5 cm of the product.
Mercury contents were greater in deeper soil than in upper soil layers. It seems that migration of Hg out of thecolumns was delayed due to absorption of mercury in the soil and moisture. Sulfur polymer effectively
immobilized mercury than low-alkali cement. However, both s/s products could achieve effective containmentof mercury-containing waste under monofill conditions with a mercury-adsorbing soil layer. It is important
not only to achieve stabilization/solidification of mercury but also to install a mercury-adsorption layer
surrounding the treated waste to minimize mercury emissions from landfills.
Key words :	column leaching test, landfill, mercury emissions, solidified/stabilized mercury

1. Introduction
The Minamata Convention on Mercury was agreed
on by nations in 2013. After the convention enters into
force, the utilization of metallic mercury will be restricted
and appropriate and safe storage and disposal will be
required to minimize mercury emissions into theenvironment. Fossil fuel (primarily coal) combustion andcement production processes are estimated to be the main
sources of global anthropogenic mercury emissions to theatmosphere (Kim & Zoh, 2012; Zhao et al., 2019). In
Japan, mercury has mainly been emitted from waste
incineration facilities; however, the amount of emissions
was 5.44 tons/year in 2014 (Takiguchi & Tamura, 2018),
much less than those of other countries around the world.
According to a review by Tao et al. (2017), emissions
from working faces in 10 landfills located in China,
Korea and the USA are on the order of one to a thousand
g/year, which may depend on differences in Hg content
of received waste (0.0328 –46.2 mg/kg). Immediately
covering them with soil has been described as a strategy
for emission control. Hg discharge with leachate fromsites for landfilling municipal solid waste was also
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investigated by Chai et al. (2011) and Ilgen et al. (2008).
Yanase et al. (2009) described long-term Hg behavior ina lysimeter filled with Hg-containing waste, i.e., usedbatteries and municipal solid waste. On the other hand,waste consisting of elemental mercury is never landfilled
because elemental mercury can be recovered for resource
use from the waste and its landfilling is discrouaged.
Hence, the long-term behaviour of elemental mercury
and its compounds when wastes with high concentrationsof immobilized mercury are disposed of in landfills
remains unknown.
International organizations and countries aredeveloping regulations or guidelines for environmentallysound management of mercury waste. With regard to
final disposal, elemental mercury waste in the EU can bepermanently stored in geological repositories such as
abandoned salt mines or hard rock (European Council,
2008). The U.S. Department of Energy has alreadyconstructed an above-ground facility for storage ofelemental mercury containers (Lee et al., 2012). In Japan,
there has been discussion of the safe disposal of
immobilized metal mercury in an engineered landfill if itsatisfies certain criteria, e.g., regulation of leachability,©2020 AIRES

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M. SATO et al.

and is separated from other wastes.
It is essential to control the mercury emissions fromlandfill bodies to the surroundings almost eternally, andappropriate treatment techniques are needed toimmobilize mercury in a landfill. There are several
mercury waste treatments which include mercury sulfide
formation and a solidification method (UNEP, 2012;
Donatello et al., 2012; López et al., 2015). As forstabilization/solidification processes, elemental mercury,which has high solubility and volatility, is reacted withsulfur powder to produce metacinnabar to reduce themobility of the mercury by changing its chemical
properties (UNEP, 2012). In addition, the stabilized
mercury is mixed with a binder such as sulfur polymer toencapsulate and physically isolate it from the surrounding
environment (UNEP, 2012; Rodríguez et al., 2012). A
treatment technique for elemental mercury has beendeveloped, with which leached and vaporized Hgconcentrations from sulfur-polymer-treated waste
containing mercury have been revealed to be at allowablelevels and extremely low (Fuhrmann et al., 2002).
However, these results were obtained under experimental
conditions for assessing treatability. It is important tounderstand the behaviour of mercury in real landfilling
situations in which both liquid and gas come in contact
with the treated waste, i.e, under unsaturated percolationand landfill gas flow, from the point of view ofidentifying long-term mercury emissions, especially inpost-aftercare.
In this study, a column leaching test under simulated
landfill conditions was conducted to clarify the release of
mercury into the atmosphere and leachate, and to suggest
appropriate landfilling of treated metallic mercury forsafe management of its final storage and disposal.

2. Materials and Methods
2.1.	
Stabilized/Solidified Products and Soil for the
Column Leaching Test
As a pre-treatment of elemental mercury,
metacinnabar was formed by reacting mercury (purity
99.9%) and sulfur following the method shown by
Fukuda et al. (2014). The product material was groundand mixed with a binder for solidification. Four types ofstabilized/ solidified product (hereinafter called s/s
products), were produced in the shape of cubes of about
7.8 cm and named as SP, CA, CB and CB+ according totype of binder and amounts of components as shown in
Table 1. Sulfur polymer was produced by mixing thecomponents under decreased pressure in N2 gas at 130°C
for 1 hour. Low-alkali cement cubes were produced witha curing period of 28 days. The mercury content in the s/s
products was calculated by the amount of elemental
mercury component during the sulfurization (shown in
Table 1). The sulfur polymer was a binder whichcontained ~95% elemental sulfur (Rodríguez et al.,2012). The low-alkali cement was composed of Portland
cement, silica fume and fly ash (Codina et al., 2008;
Mihara et al., 2008). The pH of the cement paste waslower than the equilibrium pH of portlandite. The two
types of low-alkali cements (A and B) were commercial
products and produced from different raw materials.
The mercury content of the SP (shown in Table 1)
was slightly higher than that of the others because of thehigh metacinnabar content. The compounded amount of
the components differed for each s/s product to facilitatecomparison of the ability to bind mercury at the sameratio by weight of the binder and metacinnabar.
Decomposed granite soil was selected for the
mercury-absorbing material due to its high permeability−5(2.3×10 m/s). The weight percentages of soil withparticle sizes of less than 0.005, 0.005 – 0.075, 0.075 – 2
and 2–75 mm were 3.8, 11.8, 51.2, and 33.2%,
respectively. The maximum size of the soil particles was
19 mm. The optimum moisture content was 13.4%.
2.2. Column Leaching Test
Considering safer landfilling of s/s products at anexisting sanitary engineered landfill in Japan, we assumed
the following landfilling conditions: The s/s product wasfilled above an old-waste layer that included organic
waste, and was not mixed with other wastes; rainwater
infiltrated under unsaturated conditions; the soiladsorption layer was placed around the s/s product;
landfill gas migrated to the mercury-storage layer fromthe existing waste layer so that the conditions in the latter

Table 1 Types of binders and components of the s/s products.

s/s
product

Type of binderComponent ratio by weight (%)

Wp*

MC**

Binder

Water

Me

Agant

(g)

% w/w

a

b

SP

Sulfur polymer

50

0

50

0

1509

43

CA

Low-alkali cement “A”

40

20

40

0

1202

34

CB

Low-alkali cement “B”

40

20

40

0

1145

34

CB+

Low-alkali cement “B”

40.8

18.3

40.8

0.2

1147

35

a: Metacinnabar (ground samples), b: Water-reducing agent for cement* Weight of s/s product, ** Mercury content in the s/s product佐藤.indd

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Emission Control of Mercury from Stabilized and Solidified Products under Monofill Conditions
layer became anaerobic. Figure 1 shows a schematic of
one experimental column. The columns consisted of threemain parts: a sprinkling device, column space for filling,
and drainage. Each part was tightly sealed with a flange
connection. Drops of water fell from the needles of thesprinkling device. A gas bag was connected to the venthole of the sprinkling device to allow the gas pressure in
the column to adjust to that of atmosphere. A Hgadsorption tube containing a gold trap and air pump wasconnected to the vent hole to constantly circulate thegases and trap gaseous mercury in the top space of thecolumn. The columns were set in a room with ambientotemperature kept around 25 C.
Decomposed granite soil was filled at a density of
1,692 kg-dry/m3 above glass beads on the punchedstainless-steel plate. The s/s product was placed at thecenter of the soil layer with a maximum thickness of
15 cm. The top of the filling layer was covered with glass
beads. The moisture content of the soil was adjusted to13.4% (OMC) before filling.
Approximately 320 mL of degassed water was added
once a day from the top of the column, so that the amountof infiltration per cross-sectional area of the column was
about 4.5 mm/d, approximately corresponding to theaverage annual precipitation in Japan. Precipitation in alandfill cannot all percolate into the waste layer becauseof evaporation and run-off from the cover (Nakajima etal., 1991; Baccini et al., 1987). This experiment,
however, assumed more Hg-leachable conditions. The
cumulative amount of infiltration water per crosssectional area was 1,027–1,074 mm, depending on thecolumn, at the end of the column leaching test (258 days).
Aiming to create anaerobic circumstances in the columns

5

and adjust the pH of the supplied water, L-cysteine andsodium hydroxide were added to the degassed water.
Hence, the average quality of the supplied water was as−follows: pH 7.5; Cl concentration of 10.4 mg/L; andoxidation-reduction potential (Eh) of 28.4 mV. The
leachate was kept constantly at a particular water level inthe drainage system and smoothly collected, then storedin polypropylene bottles without exposure to air. The
leachate weights were measured and their qualities wereanalyzed; the pH and Eh were measured by dedicatedelectrodes; concentrations of Ca2+, Cl−, SO42− wereanalyzed by ion chromatography and inductively coupledplasma atomic emission spectroscopy; and total mercury
amounts in the leachate were measured by a direct
thermal decomposition mercury analyzer (MA-3000,
Nippon Instruments Corp., Tokyo, Japan).
As for simulating landfill gas exposure to the s/s
products, 20 L of standard gas mixture was sent into thecolumns through the down vent hole. This ventilation
was conducted 21 times during days 22 –50 and 155–258
of the column leaching test. The gas mixture wascomposed of 40% carbon dioxide, approx. 20 ppm
hydrogen sulfide, and nitrogen (balance gas), chosen withreference to landfill gas concentration in anaerobic
conditions (Nagamori et al., 2016; Kim et al., 2005;
Gebert et al., 2011), except for CH4 gas which wasconsidered not to affect the behavior of Hg. Gas inside
the soil layer was sampled via the sampling pipe and thecomposition of CO2, N2 and O2 by GC-TCD and H2S was
measured by GC-FID. The gas in the upper part of thecolumns was constantly circulated by air pump so thatgaseous mercury could be collected by the Hg adsorption
tube (gold trap) installed in the upper vent hole of eachcolumn. The release rate was calculated by dividing thereleased Hg via leachate or gas by the initial amount ofmetallic Hg in each s/s product and the interval ofsampling leachate or cumulative gaseous Hg.
After 258 days of the column leaching test, different
zones of the granite soil were sampled from the columns
at different depths. The zone surrounding the s/s productwas less than 2 cm in thickness with an area of about
2
100 cm . The upper zone was at a depth of 0 to 2.5 cm,
the lower zone was below the surrounding zone, at depthof 11.5 to 15.0 cm, and the horizontal zone was otherbodies of soil at depths of 2.5 to 11.5 cm. The portions of
the soil sampled from the columns, approx. 10 g each,were extracted using a solution containing nitric acid,sulfuric acid and potassium permanganate (30 g/L). The
amount of Hg in the extract was analyzed using the
mercury analyzer. Distributions of Hg in the granite soil
were calculated by dividing amount of Hg in each part ofthe soil by the total Hg in all of the granite soil.

Fig. 1 Outline of the experimental column.

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M. SATO et al.

3. Results and Discussionthat of O2 gas. It was thought that the CO2 and H2S gases
were dissolving in the pore water. During times with noventilation, the gas composition in the soil layer was thesame as the atmospheric condition.
The ranges of leachate quality in each column are
shown in Table 3. The pH of the leachates was almost
stable and neutral. Figure 2 shows the Eh andconcentrations of SO42−, Cl−, and Hg in the leachates. The
Eh did not greatly change and ranged mostly from 400 to
500 mV in each column because of the high concentrationof O2 gas and insufficiency of reducing agents(L-Cysteine hydrochloride monohydrate and sodium

3.1.	Gas Composition in the Columns and Leachate
Quality
Table 2 shows the range of gas compositions in thesoil layer during gas ventilation. The CO2 concentrations
were higher than the atmospheric composition, and O2
gas also existed in the soil layer. After the gas
replacement, the CO2 concentration in columns gradually
declined. In contrast, the O2 concentration increased. No
H2S gas was detected even just after the replacement. The
solubility of CO2 and H2S gas in water are higher than

Table 2 Gas composition in the soil layer during gas ventilation.

CO2 composition [%]O2 composition [%]

SP

CA

CB

CB+

SP

CA

CB

CB+

Minimum

5

2

4

2

4

5

3

6

Maximum

26

24

27

21

16

19

17

15

Average

14

9

13

11

10

12

10

10

Table 3 Leachate quality in the column leaching test (“Ave” is average value during the test).
–

Cl [mg/L]

SO42– [mg/L]

Ca2+ [mg/L]

pH

Max

Ave

Min

Max

Ave

Min

Max

Ave

Min

Max

Ave

11

54

15

31

194

62

11

71

20

5.6

6.5

6.1

CA

7.8

15

13

33

182

133

30

66

53

5.6

6.7

6.2

CB

7.8

13

12

25

106

63

14

28

26

5.6

6.7

6.2

CB+

9.3

13

12

41

129

56

13

35

23

5.5

7.3

6.2

600

250

500

200

400
300

Gas ventilation

200

SP

100

CA

50

0

0

60

100 150
Days

200

250

Gas ventilation

40

SP

20

CA

0

CB
0

50

100 150
Days

200

250

CB
CB+
0

CB+

Hg [µg/L]

50

Gas ventilation

150

100
0

Cl-[mg/L]

SO42-[mg/L]

Eh [mV]

Min
SP

50

100 150
Days

0.1
0.08
0.06
0.04
0.02
0

200

250

SP
CA
Gas ventilation

050

100 150
Days

200

250

CB
CB+

Fig. 2 Eh and SO42–, Cl –, Hg concentrations in the leachate during the leaching test.

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Emission Control of Mercury from Stabilized and Solidified Products under Monofill Conditionshydroxide) in the supplied water. SO42− concentrations in
the leachate from each column, except the SP column,
increased during the periods of gas ventilation andgradually decreased in the other periods. Highconcentrations of SO42− in the leachate were detected in
the CA column. The changes in Ca2+ concentrations weresimilar to those of SO42−. The concentration of Cl− in theleachate from the SP column initially rose to a maximum
and then declined to the range of 12–15 mg/L. In othercolumns, the Cl− concentration increased from aminimum to a maximum as time went by.
The change in Ca2+ and SO42− concentrations in theleachates of the CA, CB, and CB+ columns indicateddissolution of cement hydrates such as ettringite,
including calcium and sulfate. Hydration of the low-alkali
cement “A” was said the main cause of ettringite
formation. The low-alkali cement “B” contained calcium
silicate hydrates and a little ettringite (Hoshino et al.,2014). Hence, the different SO42− concentrations in theleachates resulted from leaching of ettringite that seemed
to decompose by carbonation during gas ventilation(Nishikawa et al., 1992).
Hg concentrations in the leachate were higher thanthe limit for detection (0.03 μg/L) of Hg several times,
never exceeding that for quantification (0.1 μg/L) in theanalytical method used in this experiment. It is widely
known that the leaching behavior of metacinnabar isaffected by pH (Ravichandran et al., 1999). In addition,according to the pH – Eh diagram (shown by Gavis &Ferguson, 1972; Hem, 1970), the chemical form of Hg
and its solubility in water depend on the presence of Cl
and S and on the pH and Eh values. A thermodynamic
analysis using software (FactSage) indicated that Hg in

7leachate could form chlorides (HgCl2, [HgCl4] ) underthese leaching test conditions ([Cl−] = 10−3.53 mol/L,[SO42−] = 10−2.82 mol/L, [Ca2+] = 10−2.7 mol/L). Hence, Hgleached from the s/s products and changed to chloride,
which has higher solubility than metacinnabar.
The cumulative amounts of Hg in the leachates andgases are shown in Fig. 3. The amount of Hg released in
the leachate increased prominently in the initial quarterterm in which cumulative percolation was about 300 mm,
except for the SP column. Hg was constantly released via
gases thoroughout the days of the experiment.
Table 4 shows the cumulative amount of Hg releasedin leachate or gas and adsorbed in the soil at the end of
the column leaching test per the initial amount of metallic
mercury in the s/s products. Emissions from the columns
were extremely low, at a few ng/g, and almost the same
via gas and leachate in any of the columns. On the other
hand, the amount of Hg adsorbed by the soil was higher
than that of the emissions and differed between the HgSbinding sulfur polymer and low-alkali cement.
Metacinnabar has low solubility (Drott et al., 2013), but
Hg in leachate may have formed chloride, which has
higher solubility, under the conditions in this column test,
as mentioned above. It has been well observed that not
only dissolution of soluble matter on the surface of asolid, especially a porous one, but also diffusion frominside to the surface of it contribute to leaching behavior.
Gavis and Ferguson (1972) suggested that Hg forms
hydroxide, which also has a higher solubility thanmetacinnabar, under the high pH and Eh of a system with
S, Cl and Hg coexistence. The pH of the low-alkali
cement is maintained at over 10, as Codina et al. (2008)
showed, so more Hg may be diffused and dissolved from2−

Fig. 3 Cumulative amounts of Hg released from the column b via leachate (left) and evaporation (right).

Table 4 Mercury release (ng/g) from s/s products under landfill conditions.

Gas
SolubleAdsorption in granite soils
Release from s/s products

Emission from column

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SP
0.85
2.7
5,660
5,664

CA
2.4
2.1
122,000
122,004CB
2.2
1.5
39,20039,204CB+
4.5
1.9
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the products of low-alkali cement than those of the sulfur
polymer. Although the amounts of dissolved Hg frommetacinnabar differed between the s/s products, the Hgemissions from the columns under the simulated landfill
conditions were at the same level. This could beconsidered the result of adsorption in the soil.
Regarding the distribution of Hg in the granite soil atthe end of the column leaching test (Table 5), 64% of
mercury released from s/s product CB to the granite soil
existed in the surrounding part within approx. 1.5 cm of
the product, suggesting delayed migration of Hg byadsorption into the soil. The other 25.7% of the Hg wasmobilized to the horizontal bulk parts. This indicatesmigration by diffusion. The upper and lower zones of s/s
product CB over 1.5 cm from the products contained
0.4 % and 9.9 % of Hg released to the soil from the s/s
products, respectively. Similar results were also obtained
from the column leaching test for s/s products SP, CA,
and CB+. In terms of Hg content in each part of the soil
(Table 5), the Hg content values were maximized in thesurrounding soil within 1.5 cm of the s/s products. In
addition, the Hg content tended to increase in the deeper
zone of the columns. These results indicate that the soil atleast 1.5 cm away from the s/s products was nonsaturated with Hg, although the strictly local soil might
not be. Therefore, Hg leached from the s/s products
seems to be slowly emitted from the columns due torepeated adsorption/desorption of soluble and gaseous Hg
to/from the soils, and dissolution/evaporation of gaseous
Hg to/from moist soils under conditions of unsaturated
water flow. On the other hand, this Hg distribution among
the three phases might change spatially and temporally,
and the distributional variation more likely affects the
temporal variation of Hg concentration than variations in
pH and ion species in the leachates, as shown in Fig. 2.
Moreover this variation results in increased release via
leachates in the SP column, as shown in Fig. 3, but the
Hg content of the soil was lower yet than in the othercolumns, as shown in Table 5: Thus soil has the potential
for Hg adsorption. Therefore, it was thought that therelease of Hg via the leachates could not constantly
increase unless the metacinnabar in the SP leacheddramatically.

Although discussion is needed on how much theamount of anthropogenic emissions affects theenvironment, selection of a stabilization/solidification
method for waste mercury should also consider treatment
costs and the containable amount per solidified volume.
A higher Hg content per volume could lead to less spacerequired and easier waste management in landfills.
For greater minimization of long-term mercury
emissions from landfilling of s/s products, control ofamounts of penetration should be discussed in terms ofwhether or not it would be effective at delaying migrationof Hg via both leachates and gases.

4. Conclusions
In this study, a column leaching test for the solidified
product of metacinnabar, which was obtained bystabilizing metallic mercury, was conducted in monofill
conditions with a granite soil layer. With unsaturatedwater flow of 4.5 mm/day, the column entered condtionsof oxidation with a natural pH. The results ofthermodynamical analysis considering leachate quality
show that the Hg in the leachate could be chlorides,
indicating that the metacinnaber was soluble in porewater. However, the emissions of soluble Hg via leachates
and gaseous Hg were extremely low because of Hg
adsorption in the granite soil. Almost all Hg released
from the s/s products existed in the soil within 1.5 cm
away from it. Moreover, the deeper layers of soil
contained more Hg than the upper ones. It seems that
migration of Hg out of the columns was delayed due toadsorption/desorption of soluble Hg and dissolution/
evaporation of gaseous Hg between the soil, moisture,and pores in unsaturated water flow. The sulfur polymer
immobilized Hg in the s/s product. However, both thesulfur polymer and low-alkali cement s/s products could
achieve effective containment of mercury-containing
waste under monofill conditions with a Hg-adsorbing soil
layer. It is important not only to stabilize/solidify mercury
but also to install a Hg-adsorption layer surrounding thetreated waste to minimize Hg emissions from landfills.

Table 5 Distribution and content of Hg in soil in each zone at the end of the column leaching test.

Distribution of Hg (%)

Zone in column

Hg Content (µg/g-wet)

SP

CA

CB

CB+

SP

CA

CB

CB+

Surrounding part

58

70

64

57

1.7

19.9

5.7

7.8

Upper part

0.5

0.1

0.4

0.6

0.008–0.011

0.018–0.033

0.016–0.054

0.023–0.232

17

19

26

31

0.059, 0.066

0.53,1.1

0.32, 0.36

0.49, 0.81

24

10

9.3

9.3

0.12–0.31

0.97–1.8

0.57–0.58

0.51–1.9*1*2

Horizontal part
Lower part*4

*3

*1 within 1.5 cm of the s/s product, *2 at depths of 0 – 2.5 cm, *3 over 1.5 cm horizontally from the s/s product at depths of 2.5 – 11.5 cm, *4 depths of
11.5–15.0 cm

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Aknowledgement
This research was partially supported by theEnvironment Research and Technology Development
Fund (3K153004 and 3-1802) of the Ministry of theEnvironment, Japan.
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M. SATO et al.

Masahiro SATO
Masahiro Sato is an assistant professor at theLaboratory of Sustainable Material Cycle
Systems, Hokkaido University, Japan. He was a
postdoctoral fellow at the National Institute forEnvironmental Studies. Recently his main fields
of interest have been landfill gas emissions,storage of recycling incineration residues and theuse of anaerobic digestate in microalgae cultivation.

Tomonori ISHIGAKI
Tomonori Ishigaki is a senior researcher at the
National Institute for Environmental Studies. He
serves as an editorial board member for theemission factor database and lead author of the
Guidelines for National Greenhouse Gas
Inventories for the Intergovernmental Panel on
Climate Change. He is also serving as a memberof the Technical Committee (Solid Recovered Fuels) of International
Organization for Standardization. His published works cover subjects suchas the mechanism of waste conversion in landfills, technology development
for emission reduction from landfills, assessment of the fate of hazardous
substances in the environment, and appropriate management of solid waste
in urban areas of Asia to prevent urban floods.

Kazuto ENDO
Kazuto Endo heads the Fukushima branch of the
National Institute for Environmental Studies,Japan. His research fields are environmental
geotechnics and waste management. At present,
his team is researching the management ofradioactively contaminated waste and removed
soil, aiming at early revitalization of Fukushima.
He has researched the construction and management of seashore landfill
sites through computer simulation and field surveys. He is also interested in
the geotechnical use of recycled materials. He has published a guideline on
geotechnical use for quality control of recycled waste gypsum powder.

Masato YAMADA
Masato Yamada has been head of the International
Waste Management Engineering Section at the
Research Center for Material Cycles and Waste
Management, National Institute of Environmental
Studies, Tsukuba, Japan since 2000. His main
fields of interest are landfill engineering andlogistics of industrial waste. He is studying
appropriate waste management in developing countries. He has also been in
charge of several studies on disasters and radioactive contaminated waste
management in the aftermath of the Great East Japan Earthquake and the
Fukushima Daiichi nuclear accident.
(Received 10 December 2019, Accepted 16 May 2020)

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