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Radioactive fallout cesium in sewage sludge ashproduced after the Fukushima Daiichi nuclearaccident

Naofumi Kozai*, Shinichi Suzuki, Noboru Aoyagi, Fuminori Sakamoto,Toshihiko Ohnuki

Japan Atomic Energy Agency, 2-4 Shirane, Shirakata, Tokai-mura, Naka-gun, Ibaraki 319-1195, Japan

a r t i c l e i n f o

Article history:

Received 22 August 2014

Received in revised form

17 October 2014

Accepted 18 October 2014

Available online 29 October 2014

Keywords:

Sewage sludge ash

Cesium-137

Fukushima

Dissolution

Iron

Silicate

* Corresponding author.E-mail address: kozai.naofumi@jaea.go.jp

http://dx.doi.org/10.1016/j.watres.2014.10.0380043-1354/© 2014 Elsevier Ltd. All rights rese

a b s t r a c t

The radioactive fallout cesium (137Cs) in the sewage sludge ashes (SSAs) produced in Japan

after the Fukushima Daiichi Nuclear Accident was tested. Five samples of SSAs produced in

2011 and 2012 were tested. Two of the samples contained 137Cs (23 and 9.6 kBq/kg,

respectively) above the radioactivity criterion (8 kBq of radioactive Cs/kg of solid) for

controlled landfill disposal in Japan. The mineral components of SSA are roughly divided

into two groups: an HCl-soluble phase mainly composed of phosphates and oxides; and

silicates, including quartz, feldspar, and clay. Both phases contained 137Cs. The majority

(up to 90%) of 137Cs was contained in the HCl-soluble phase. Among the HCl-soluble sub-

phases, Fe-bearing phases that were probably iron oxides were mainly responsible for 137Cs

retention. No positive evidence was obtained that showed that phosphate-bearing phases,

which were included most in SSAs along with the silicate phase, retained 137Cs. Pre-

pulverizing SSAs and heating them at 95 �C in a 6 M or a concentrated aqueous HCl was

the most effective method of dissolving the HCl-soluble phase. The radioactivity concen-

trations of 137Cs in all the HCl-treatment residues were below the radioactivity criterion.

This residue was mostly composed of silicates. After static leaching tests of the residue at

60 �C for 28 days, no 137Cs was detected in simulated environmental water leachates (pure

water and synthetic seawater), demonstrating that 137Cs in the residue is very stably

immobilized in the silicates.

© 2014 Elsevier Ltd. All rights reserved.

1. Introduction

In March 2011, huge quantities of radionuclides were emitted

into the atmosphere by the consecutive explosions in the re-

actors and reactor buildings at the Fukushima Daiichi Nuclear

Power Plant (FDNPP). The emission of radionuclides began on

(N. Kozai).

rved.

March 12 and the total emission of 137Cs into the atmosphere

was estimated to be in the range of 10e36 PBq (Chino et al.,

2011; Morino et al., 2011; Stohl et al., 2012). The radioactive

Cs fallout was dispersed from the FDNPP into the ocean and

across large areas of Japan, including the Tohoku and north

Kanto regions (Fig. S1 in the Supplementary content) (Chino

et al., 2011; Kawamura et al., 2011). The soil, animals, plants,

wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6 617

architectural structures, and natural water in these regions

were contaminated with radioactive Cs (Koarashi et al., 2012;

Kozai et al., 2012; Shibata et al., 2012; Nakanishi et al., 2013;

Tanaka et al., 2013; Saegusa et al., 2014; Sakai et al., 2014).

Subsequently, rainwater has dissolved some of the radio-

active Cs from contaminated objects or washed the contam-

inated objects containing radioactive Cs away. Some of the

radioactive Cs has been discharged into combined sewerage

systems, where rainwater and sewage are collected together

and conveyed to sewage treatment plants. The wastewater is

purified by physical screening or sedimentation to remove

solids, and then by removing soluble and fine suspended

pollutants through biological processes. These treatments

produce sewage sludge. In Japan, sewage sludge is dried and

incinerated at 800e850 �C, and the end product is called

sewage sludge ash (SSA). These treatment processes reduce

the sewage sludge to SSA that is several hundredths of the

original volume of the sludge, thereby enriching the radioac-

tive Cs in SSA. Since the Fukushima Daiichi accident, highly

concentrated radioactive Cs has been detected in the SSAs

produced at many sewage treatment plants in Japan. The

concentration of radioactive Cs varies considerably among

different plants and production dates. The highest reported

concentration is 4 � 105 Bq/kg (in dehydrated sludge) as of

October 28, 2011 (the Ministry of Land, Infrastructure,

Transport and Tourism of Japan, 2011).

SSA, which is produced daily in large quantities, had been

used before the Fukushima Daiichi accident mainly as raw

material for cement and was not disposed of in landfills.

However, due to the accident, reuse of contaminated SSA in

the cement industry is no longer possible due to the high

radioactivity. The contaminated SSA is now being stored in

sewage treatment plants.

Currently, it is intended that the contaminated SSAswill be

buried in controlled or strictly controlled landfill sites based

on a radioactivity criterion (8 kBq of radioactive cesium per kg

of solid) (Ministry of the Environment, 2011). It may be also

possible to reduce the radioactivity of SSAs to achieve safe,

readily controlled landfill disposal. However, for the imple-

mentation of any of these options, it is essential to elucidate

the physical and chemical properties of the radioactive Cs in

SSAs, because currently, very little is known this.

To contribute to establishing themethodology for reducing

the radioactivity of SSAs and their subsequent safe landfill

disposal, we investigated SSAs contaminated with radioactive

Cs from the FDNPP. The maximum radioactivity concentra-

tion of 137Cs in the SSA tested in this study is 23 kBq/kg. This

Table 1 e Sampling site, date, and radioactivity concentration

Sewage treatment plant Date ofoccurrence

Rad

Iwaki city central sewage treatment plant,

Iwaki, Fukushima

May, 2011

Iwaki city central sewage treatment plant,

Iwaki, Fukushima

May, 2012

Teganuma sewage treatment plant, Abiko, Chiba June, 2011

Teganuma sewage treatment plant, Abiko, Chiba May, 2012

Nakakuji sewage purification center,

Hitachinaka, Ibaraki

May, 2012

value is high for radioactivity, although themolality of 137Cs is

only 0.052 nmol/kg. Therefore, it is impossible to analyze the

chemical states of 137Cs directly in the SSAs. The properties of137Cs in the SSAs were analyzed in terms of the static leaching

and acid-dissolution behaviors of 137Cs and the major

component elements of the SSAs. The effect of pulverizing

SSAs on acid-dissolution of 137Cs was also examined to

develop a method to reduce radioactivity effectively.

2. Materials and methods

2.1. Sewage sludge ash

Five samples of SSA generated in 2011 and 2012 were collected

from three sewage treatment plants (Table 1 and Fig. S1 in the

Supplementary content). The SSA samples were fine brown or

reddish-brown powders. We tested as-incinerated SSAs and

pre-pulverized ones. As-incinerated SSA samples were used

with no pretreatment except being dried at 30 �C to remove

the moisture that had been added to SSAs for storage at

sewage treatment plants. Pre-pulverized SSA samples were

prepared by an automatic mortar grinder with an agate

mortar or a planetary ball mill (PM200, Retsch GmbH, Ger-

many) with a zirconia grinding jar and 3-mm zirconia balls.

2.2. Chemicals

All reagent solutions used in the present study were prepared

using ultrapure water (18.2 MU cm�1) and reagent-grade

chemicals.

2.3. Static leaching and dissolution experiments

The leaching and dissolution properties of 137Cs from the SSA

samples were examined in the following two experiments.

2.3.1. Static leachingStatic leaching behaviors of 137Cs were investigated in pure

waterandsyntheticseawater. For thisexperiment, Iwaki’11and

Teganuma’11 samples, which had the highest and the second

highest radioactivity concentrations of 137Cs, respectively,were

tested. The SSA sample (2.00 g) was added to purified water

(100mL) or synthetic seawater (NaCl 25 g/L, MgCl2$6H2O 11 g/L,

CaCl2$2H2O 1.4 g/L) in a plastic bottle, and the bottle was left to

stand at 25 or 60 �C without agitation. The static leaching ex-

periments were performed for up to 28 days. After the

of SSAs.

ioactivity concentrationof 137Cs kBq/kg

Specific surfacearea m2/g

Abbreviation

23 1.3 � 102 Iwaki’11

6.5 e Iwaki’12

9.6 1.3 � 102 Teganuma’11

2.8 e Teganuma’12

6.8 e Nakakuji’12

wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6618

experiment, the liquid phase was collected by centrifugation,

passed through a filter with pore size 0.45 mm, and the radio-

activity of 137Cs in the liquid phase was determined.

2.3.2. Dissolution in various reagentsDissolution behaviors of 137Cs and the major component ele-

ments (Na, K, Mg, Ca, Al, Fe, Si, and P) of SSAs in various re-

agent solutions were investigated. The dissolution of 137Cs

was tested in a larger scale experiment (initial weight of SSA,

1.00 g) and elemental analysis was carried out a smaller scale

experiment (initial weight of SSA, 0.100 g). The SSA sample

was soaked in a reagent solution in a plastic bottle at a solid/

liquid ratio of 1/200 g/mL. The bottle was left to stand at 25, 60,

or 95 �C for up to 24 h without agitation. The solid phase was

separated from the liquid phase by centrifugation (10,000 rpm,

5 min), washed with purified water three times, and dried at

60 �C. In the larger scale experiment, the radioactivity of 137Cs

in the solid phase and the dry weight of the solid phase were

determined before and after the experiment. In the small-

scale experiment, the liquid phase and all of the rinsing so-

lutions were collected, combined, diluted with pure water,

and passed through a filter with a pore size of 0.45 mm to

determine the concentrations of themain dissolved elements.

Aqueous 0.1 and 1 M citric acid, 0.1 and 1 M sodium citrate,

0.1MEDTA$2Na, 0.5Mand saturated (ca. 1M) oxalic acid, nitric

acid at various concentrations, hydrochloric acid at various

concentrations, andconcentratedhydrofluoric acidwereused.

2.4. Analysis

The radioactivity of 137Cs was determined by a g-ray spec-

trometer with a germanium detector (ORTEC GEM-15180,

Advanced Measurement Technology Inc., USA). The

Table 2 e Elemental compositions of SSAs.

Iwaki'11 Iwaki'12 Teganuma'11 Teganuma'12 Nak

SiO2 24.6 29.6 23.3 25.5

Al2O3 11.3 13.1 11.6 12.0

P2O5 24.7 20.3 32.3 28.5

MgO 2.8 2.4 3.1 2.7

CaO 13.9 11.5 9.7 11.9

Fe2O3 13.7 14.9 11.9 12.2

Na2O 1.2 1.1 0.7 0.4

K2O 2.3 2.2 2.6 2.8

TiO2 1.1 1.1 0.8 1.1

NiO 0.0 0.0 0.0 0.0

CuO 0.3 0.3 0.2 0.2

MnO 0.3 0.4 0.2 0.1

ZnO 0.6 0.6 0.8 0.5

SrO 0.1 0.1 0.1 0.1

ZrO2 0.0 0.0 0.0 0.0

BaO 1.3 0.8 1.6 0.7

PbO 0.0 0.1 0.0 0.0

As2O3 0.0 0.0 0.0 0.0

SO3 1.5 1.3 1.4 1.3

Cl 0.2 0.1 0.0 0.0

Total

weight %

100.0 99.9 100.2 100.0

a Treated at 60 �C for 24 h.

elemental compositions of SSA samples were determined by a

wavelength dispersive X-ray fluorescence (XRF) spectrometer

(Primus II, Rigaku Corp., Japan). The BET surface areas were

measured by a surface area analyzer (Nova 1000e, Quantach-

rome Instruments, USA). Concentrations of metals in the

liquid phase were determined with an inductively coupled

plasma optical emission spectrometer (ICP-OES; 720, Agilent

Technologies, Inc., USA), an inductively coupled plasmamass

spectrometer (NexION 300X, PerkinElmer, Inc., USA), and an

atomic absorption spectrometer (AA-6200; Shimadzu Corp.,

Japan). Powder X-ray diffraction (XRD) data were obtained

with an X-ray diffractometer (Ultima IV, Rigaku Corp.). The

surface morphology of the solid phases were observed by

scanning electron microscopy (SEM; Phenom ProX, Phenom-

World, Netherlands, and JSM-7800F, JEOL, Japan) with an en-

ergy dispersive X-ray (EDX) analysis system.

3. Results

3.1. Chemical and mineral compositions of as-incinerated SSAs

Table 2 shows the elemental compositions of as-incinerated

(original) SSA samples. Si and P were the most abundant ele-

ments contained in these SSAs, and Al, Ca, and Fe were the

second most abundant. The elemental compositions of the

samples varied according to the time and location of pro-

duction, and the Si and P oxide content showed maximum

differences between samples of 5% and 12%, respectively.

XRD patterns of the SSA samples are shown in Fig. 1.

Although the relationship between peak intensities were

different among the samples, most of the major diffraction

akuji'12 Iwaki '11 residue after6 M HCl treatmenta

Teganuma '11 residue after6 M HCl treatmenta

27.4 80.8 85.5

11.6 8.3 5.5

27.1 0.5 1.3

2.8 0.7 0.6

10.5 2.4 1.1

11.4 2.5 1.7

0.9 0.9 0.5

3.5 2.2 2.0

1.1 1.2 1.5

0.0 0.0 0.0

0.2 0.0 0.0

0.3 0.1 0.0

0.5 0.2 0.0

0.1 0.0 0.0

0.0 0.0 0.0

1.4 0.0 0.0

0.0 0.0 0.0

0.0 0.0 0.0

1.0 0.1 0.1

0.1 0.0 0.0

100.0 99.9 100.0

wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6 619

peakswereobserved inallof thesamples, showing that theSSAs

were composed of similar minerals. The mineral phases of the

samples were divided into two groups according to their solu-

bility in aqueous HCl. The major minerals soluble in aqueous

HCl were expected to be iron oxides (hematite, maghemite,

goethite), and various phosphates of aluminum, calcium, and

iron (Al(PO3)3, rodolicoite, brushite, whitlockite, wyllieite). No

XRD peaks of simple aluminum oxides (e.g., alumina, gibbsite)

were observed. No diffraction peaks for S-bearing phases were

present in the XRD patterns; however, the presence of fine pre-

cipitates of S-bearing phases containing Ca or Ba, probably cal-

cium sulfate and barium sulfate, was confirmed by SEM-EDX

analysis. The major minerals insoluble in aqueous HCl were

quartzandfeldspar. Inallof theXRDpatterns,aweakdiffraction

peak was observed at 9.4e9.5� (d¼ 0.93e0.94 nm) and this peak

may be the strongest peak for talc. This talc may have been

formedby the incinerationofsewagesludge,because talc canbe

synthesized at several hundred degrees Celsius from silicon

dioxide and magnesium (Dumas et al., 2013; Takahashi et al.,

1994), although the talc is more likely to originate from addi-

tives in paper flushed into sewers. XRD peaks of other clay

minerals were not observed.

3.2. Static leaching behavior of 137Cs from as-incinerated SSAs

Fig. 2A shows the time course of the leaching behavior of 137Cs

from the Iwaki’11 sample into pure water and synthetic

Fig. 1 e XRD patte

seawater. At 25 �C, the percentages of the leached 137Cs

reached constant values in 7 days in both pure water and

synthetic seawater. After 28 days, 0.5% and 3% of the 137Cswas

leached into pure water and synthetic seawater, respectively

(Table 3). At 60 �C, the percentages of the leached 137Cs in pure

water reached constant values (ca. 1%) in 1 day, whereas that

leached into the synthetic seawater increased with time and

did not reach a constant value (ca. 6% at 28 days). These re-

sults indicate the presence of loosely bound (water-soluble

and exchangeable) 137Cs in SSA. After 28 days in the static

leaching test for the Teganuma’11 sample at 25 �C, 2% and 3%

of 137Cs were leached in pure water and synthetic seawater,

respectively. These values are similar to those for the Iwaki’11

sample.

Fig. 2B and Figure S2AeC in the Supplementary content

show the time course of the concentrations of the main

metallic elements (Na, K, Mg, Al, P, Ca, and Fe) leached from

the Iwaki’11 sample into pure water and synthetic seawater.

The Na, Mg, and Ca that leached into synthetic seawater were

not measured because the synthetic seawater originally con-

tained these elements at high concentrations. The concen-

trations of Fe and Al leached into the synthetic seawater were

below the detection limits of the ICP-OES instrument

(10e20 ppb). Therefore, only the results for leached K and P are

shown in Figs. 2B and S2C, respectively. The characteristic

time course of the leaching behavior of 137Cs (Fig. 2A) was in

good agreement with that of K (Fig. 2B). Although the leaching

behaviors of Na and Ca in purified water (Fig. S2A) were

rns of SSAs.

Fig. 2 e Time course of leaching behaviors of the 137Cs and K of as-incinerated Iwaki '11 sample in purified water (PW) and

synthetic seawater (SSW). (A) Percent fraction of the amount of leached 137Cs with respect to the initial amount of 137Cs in

the as-incinerated sample, and (B) the concentration of K in leachates.

wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6620

similar to that of 137Cs (Fig. 2A), their relationship with 137Cs

could not be determined because of the lack of data for Na and

Ca in seawater. The leaching behaviors of Mg, Al, P, and Fe

(Fig. S2AeC) are different from those of 137Cs. These results

suggest that at least the potassium-bearing phases contained

the water-soluble and exchangeable 137Cs.

3.3. Dissolution behavior of 137Cs from as-incineratedSSAs in various reagents

Fig. 3 shows the fractions of the 137Cs and the elements dis-

solved from the Iwaki’11 sample in various reagent solutions

at 25 and 60 �C. The decrease in the percentage of 137Cs was

not proportional to the weight decrease of the ash and they

showed a sigmoidal relationship with large curvature below a

weight decrease of 50% (Fig. 3A). The greatest decrease in 137Cs

was achieved when the ash was dissolved in aqueous 6 M HCl

at 60 �C for 24 h (data labeled “K” in Fig. 3A). The decrease in137Cs was 73% and the weight decrease was 49%. The second

largest decrease in 137Cs was achieved in a saturated aqueous

oxalic acid solution at 60 �C for 24 h (data labeled “G” in

Fig. 3A). The decrease in 137Cs was 68% and the weight

Table 3 e Static leaching of 137Cs from as-incinerated SSAs, thball-milled SSAs. The leaching of 137Cs is expressed as the radiinitial radioactivity of 137Cs in the solid phase. The data shownof treatment.

Iwaki'11as-incinerated

Iwaki

Before 6 M HCltreatment

Residue after6 M HCl

treatment at60 �C for 24 h

Residue aft6 M HCl

treatment a60 �C for 24

Liquidphase

T

25 �C 60 �C 60 �C 60 �C

Pure water 0.6 ± 0.1 0.8 ± 0.3 0.24 ± 0.1 n.d

Synthetic

seawater

3.0 ± 0.1 5.6 ± 0.5 2.7 ± 0.4 5.2 ± 1.2

n.d ¼ 137Cs was not detectable in leachate.a The sample was treated by HCl twice.

decrease was 31%. The third greatest decrease in 137Cs was

obtained in a concentrated aqueous (12M) HCl at 60 �C for 24 h

(data labeled “J” in Fig. 3A). The decrease in 137Cs was 66% and

the weight decrease was 48%. The residues after exposure to

aqueous 6 M HCl, concentrated HCl, and concentrated HF at

60 �C for 24 hwere gray, indicating thatmost of the iron oxides

were dissolved, whereas the colors of the other residues were

pale brown or similar to that of the original ash.

Of the major component elements in the Iwaki’11 sample,

the relationship between the decrease in Fe and the weight

decrease (Fig. 3B) was sigmoidal with a large curvature, and

the relationship between the decrease in Al and the weight

(Fig. 3C) had a sigmoidal relationship with a small curvature.

The sigmoid curves of Fe and 137Cs were very similar. The

dissolution behavior of the other elements (P, Mg, Ca, and Si)

had no correlation with that of 137Cs (Figs. 3B, C, D). These

results strongly suggest thatmost of the HCl-soluble 137Cs was

retained by Fe-bearing phases (probably iron oxides). A weak

correlation can be seen between the dissolution behaviors of

Al and 137Cs (Fig. 3D), indicating that a small proportion of the

Al-bearing phases contained 137Cs. The Al-bearing phases that

contain 137Cs may be present in iron oxides.

eir HCl-treatment residues, and HCl-treatment residues ofoactivity percentage of the leached 137Cs with respect to theare the average values of duplicate experiments for 28 days

'11 ball-milledfor 8 h

Teganuma'11as-incinerated

Teganuma'11ball-milled for 8 h

er

th

Residue after6 M HCl

treatment at95 �C for 24 ha

Before6 M HCl

treatment

Residue after6 M HCl treatmentat 60 �C for 24 h

emperature

60 �C 25 �C 60 �C

n.d 2.2 ± 0.3 n.d

n.d 3.3 ± 0.6 6.7 ± 1.7

Fig. 3 e Dissolution of 137Cs and major component elements in the Iwaki’11 sample in various reagent solutions as a

function of the weight decrease of the SSA. The decrease of 137Cs was defined as the radioactivity percentage of the

dissolved 137Cs with respect to the initial radioactivity of 137Cs in the sample. The decrease of a given element was defined

as the weight percentage of the dissolved element with respect to the initial weight of the sample. Concentration of 137Cs

and those of the major component elements dissolved in reagent solutions were separately determined in separate

experiments.

wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6 621

Similar results were obtained for the Nakakuji’12 sample

and the dissolution of 137Cs was only related to the dissolution

of Fe (Fig. S3). We did not examine the dissolution behavior of

the Iwaki’12, Teganuma’11, and Teganuma’12 samples in

detail. Similar percentages of the 137Cs (71e78%) were dis-

solved from these three SSAs by 6MHCl treatment at 60 �C for

24 h.

3.4. Characteristics of residues after HCl treatment at60 �C for 24 h

The Iwaki’11 and Teganuma’11 samples had initial radioac-

tivity concentrations of 137Cs greater than the radioactivity

criterion (8 kBq/kg). In the Iwaki’11 and Teganuma’11 sam-

ples, about 73% and 72% of the 137Cs was dissolved, respec-

tively, by 6 M HCl treatment at 60 �C for 24 h. The 137Cs

radioactivity concentrations in the HCl treatment residues

were 15 kBq/kg for the Iwaki’11 sample and 8.6 kBq/kg for the

Teganuma’11 sample (Table 4). Therefore, HCl treatment did

not reduce the 137Cs radioactivity concentrations to less than

the radioactivity criterion. The decrease in 137Cs from HCl

treatment at 95 �C for 24 h (Iwaki’11, 72%; Teganuma’11, 78%)

was similar to that from the HCl treatment at 60 �C for 24 h.

After HCl treatment of the Iwaki’11 sample, the solid res-

idue was washed with pure water 5 times and dried to

investigate the static leaching behavior of 137Cs. After 28 days

at 60 �C, 0.2% and 3% of 137Cs in the residue were leached into

pure water and synthetic seawater, respectively (Table 3).

These values are less than those of the as-incinerated (orig-

inal) Iwaki’11 sample, although water-soluble and exchange-

able 137Cs still remained in the residue.

The residues of the Iwaki’11 and Teganuma’11 samples

still contained a small amount of Fe (Table 2). The mineral

phases of the residues were silicates. Faint XRD diffraction

peaks of iron oxides were observed in a few residues. Fig. 4

shows the SEM images of the Iwaki’11 sample before (4A)

and after the HCl treatment (4B). Although most of the iron

minerals and phosphate minerals were dissolved, the mor-

phologies of the particles were not changed greatly by the

treatment.

As-incinerated SSAs consisted of two main types particles

(Fig. S4A and B). The first type of particle had a smooth surface.

Most of these particles were quartz and the rest were feldspar.

Fine particles containing Al, P, S, or Fe were attached to these

silicates. The second type of particle was dense aggregates of

numerous fine particles (typical particles are indicated with

white arrows in Fig. S4A and B). These aggregates mainly

contained Si, Al, and P. The residues after the HCl-treatment

mostly consisted of Si (Table 2). The HCl-treatment residues

were composed of three main types of particles (Fig. S4C and

Table 4 e Decrease of 137Cs in as-incinerated and ball-milled SSAs. The Nakakuji’12 sample was treatedwith concentratedaqueous HCl and the others were treated with aqueous 6 M HCl.

SSA Condition beforeHCl treatment

Decrease of 137Csby HCl treatment at60 �C for 24 h (%)

Radioactivity concentrationof 137Cs in residue treated withHCl at 60 �C for 24 h (kBq/kg)

Decrease of 137Cs byHCl treatment at 95 �C

for 24 h (%)

Iwaki'11 As-incinerated 73 ± 1 14.8 ± 0.3 72 ± 1

Ball-milled 8h 85 ± 0 5.9 ± 0.1 90 ± 1

Iwaki'12 As-incinerated 78 ± 1 e e

Ball-milled 8h 85 ± 0 2.0 ± 0.1 90 ± 0

Teganuma'11 As-incinerated 72 ± 1 8.6 ± 0.1 78 ± 0

Ball-milled 8h 85 ± 0 4.1 ± 0.1 88 ± 0

Teganumai'12 As-incinerated 71 ± 1 e e

Ball-milled 8h 87 ± 0 1.9 ± 0.1 90 ± 0

Nakakuji'12 As-incinerated 71 ± 0 e 79 ± 0

Ball-milled 8h 73 ± 0 0.9 ± 0.0 83 ± 0

wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6622

D). The first type was quartz and feldspar particles with a

smooth surface. The second type was fibrous particles (typical

particles are indicated with a blue arrow in Fig. S4C). The third

type was particles that appear to be sparse aggregates of fine

particles (typical aggregates are indicated with red arrows in

Fig. S4D). The second and the third type of particles are

probably the silicate framework of the dense aggregates that

retain fine particles of phosphates and iron oxides. Some of

the third type of particles contained a small amount of Fe. In

addition, there was a further type of particles that contained

high concentrations of Fe. An example of one of such particle

is shown in Fig. 4C (particle 1). This particle appeared to be a

dense aggregate of numerous nano particles and had a high Fe

content of 8% byweight (Si 60% andAl 20%). Fewas detected in

the whole particle. The other particles in Fig. 4C contained

almost no Fe; Fe was not detected in small particle 2 that had

morphology similar to particle 1. These results suggest that

the internal Fe in the large aggregate was not dissolved

because the hydronium ions did not penetrate the particles. It

was predicted that pulverizing these aggregates would

enhance dissolution of radioactive Cs.

3.5. Effect of pulverization

To examine the effect of pulverization, preliminary experi-

ments were conducted using three pulverized powders from

the Iwaki’11 sample prepared by dry pulverization with an

automatic agate mortar for 1, 4, and 10 h, respectively. The

Fig. 4 e SEM images of (A) the as-incinerated Iwaki’11 sample an

The large particles with smooth surfaces were mainly quartz.

grain sizes of the pulverized samples decreased with

increasing pulverization time. However, the grain sizes were

not uniform and small aggregates were still present in the

sample pulverized for 10 h (Figure S5 in Supplementary

content). When these samples were immersed in aqueous

6 M HCl at 60 �C for 24 h, a larger amount of 137Cs was dis-

solved as the particle sizes decreased (Figure S6 in

Supplementary content). This demonstrates that pulverizing

SSA is a promising pretreatment method for effectively dis-

solving 137Cs.

Because the mortar pulverization did not produce an SSA

sample with a uniform grain size, wet pulverization with a

planetary ball mill was used to prepare homogeneous ultra-

fine powers. This method has another advantage; a larger

amount of material can be pulverized at once. Ball milling for

more than 4 h produced particles less than 1 mm in size with

uniform grain sizes, and no large aggregates (Figure S5D). No

further size reduction was achieved by ball milling for up to

24 h. The following experiments used SSA samples ball-milled

for 8 h.

The ball-milled SSA samples were treated with aqueous

HCl at 60 �C for 24 h. The ball-milled Nakakuji’12 sample was

treated with concentrated HCl and the other ball-milled SSAs

were treated with aqueous 6 M HCl. Greater amounts of 137Cs

were dissolved from the ball-milled samples than from the as-

incinerated (not pulverized) samples (Table 4). The radioac-

tivity concentrations of 137Cs in the HCl-treatment residues of

the ball-milled Iwaki’11 and Teganuma’11 samples were

d (B, C) its residue after 6 M HCl treatment at 60 �C for 24 h.

Table 5 e Elemental compositions of HCl-treatment residues determined by XRF.

Residues ofIwaki'11

Residues ofTeganuma'11

Residues ofNakakuji'12

Residues ofIwaki'12

Residues ofTeganuma'12

Ball-milled8 h

Ball-milled8 h

Ball-milled8 h

As-incinerated Ball-milled8 h

Ball-milled8 h

Ball-milled8 h

Ball-milled8 h

Ball-milled8 h

Ball-milled8 h

6 M HCl,60 �C, 24 h

6 M HCl,95 �C, 24 h

6 M HCl,95 �C,

24 h, twice

6 M HCl,95 �C,24 h

6 M HCl,60 �C,24 h

6 M HCl,95 �C,24 h

Conc. HCl,60 �C,24 h

Conc. HCl,95 �C,24 h

6 M HCl,95 �C, 24 h

6 M HCl,95 �C, 24 h

SiO2 88.0 91.6 94.1 82.6 90.2 92.3 88.8 91.1 91.6 92.3

Al2O3 4.8 3.1 2.4 7.7 3.4 2.3 4.3 3.6 2.8 2.3

P2O5 2.0 2.2 1.4 0.4 2.2 2.4 1.8 1.8 2.6 2.5

MgO 0.2 0.1 0.1 0.8 0.2 0.2 0.2 0.2 0.1 0.1

CaO 0.6 0.4 0.4 2.1 0.3 0.3 0.6 0.4 0.3 0.2

Fe2O3 0.9 0.4 0.2 2.1 0.7 0.4 0.7 0.3 0.4 0.3

Na2O 0.5 0.5 0.0 0.7 0.0 0.2 0.4 0.6 0.0 0.0

K2O 1.3 0.6 0.3 1.9 1.2 0.5 1.5 1.0 0.5 0.6

TiO2 1.2 0.9 0.8 1.2 1.7 1.3 1.3 0.9 1.5 1.4

NiO 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

CuO 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

MnO 0.0 0.0 0.0 0.1 0.0 0.0 0.0 0.0 0.0 0.0

ZnO 0.2 0.1 0.0 0.2 0.0 0.0 0.0 0.0 0.0 0.0

SrO 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

ZrO2a 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

BaO 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

PbO 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

As2O3 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0 0.0

SO3 0.1 0.0 0.0 0.1 0.0 0.0 0.0 0.0 0.0 0.0

Cl 0.2 0.1 0.1 0.1 0.1 0.1 0.2 0.1 0.1 0.1

Total

weight %

100.0 100.0 100.0 100.0 100.0 100.0 100.0 100.0 100.0 100.0

a Ball-milled SSAs contained 1% Zr in terms of ZrO2. Because as-incinerated SSAs contained no Zr (Table 2), elemental compositions of HCl-treatment residues were recalculated by setting ZrO2 to

zero.

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wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6624

reduced to less than the radioactivity criterion (8 kBq/kg).

From the ball-milled Iwaki’11 sample, slightly larger amounts

of Fe and Si (Fe 6.3%, Si 0.9%, weight percent of the dissolved

weight of the element with respect to the initial amount of

SSA) were dissolved than from the as-incinerated sample (Fe

6.0%, Si 0.3%). This result suggests that some of the 137Cs was

dissolved from the silicates, probably because of the increased

surface area created by pulverization.

The HCl-treatment residues of the ball-milled Iwaki’11 and

Teganuma’11 samples were subjected to the static leaching

test at 60 �C for 28 days. Before use, the HCl-treatment resi-

dues were washed with purified water five times, dried in air

at 60 �C, and lightly ground in an agate mortar. After the static

leaching test in purified water, 137Cs was not detectable in the

leachate (Table 3). However, the 137Cs leached into the syn-

thetic seawater (5.2e6.7%) was at similar levels to the 137Cs

leached from the as-incinerated samples into the synthetic

seawater (5.6%). These results mean that a fraction of the HCl-

soluble solid phases containing 137Cs remained undissolved

after the HCl treatment. A small amount of the Fe-bearing

phases remained in the HCl-treatment residues (Table 5).

Several minutes after adding the ball-milled samples to

aqueous HCl, the powders flocculated to form millimeter-

sized aggregates. It is likely that the flocculation impeded

the penetration of HCl into the aggregates.

3.6. Dissolution of 137Cs at 95 �C

The effect of raising the temperature on the dissolution of137Cs was investigated. As shown in Table 4, the 6 M HCl

treatment at 95 �C of the as-incinerated Iwaki’11 sample did

not increase the dissolution of 137Cs. For all the ball-milled

SSA samples, more 137Cs was dissolved at 95 �C than at 60 �C.When the ball-milled Iwaki’11 sample was treated with

aqueous 6 M HCl at 95 �C twice, the second treatment only

dissolved a small amount of 137Cs (0.4%) together with Fe

(0.02%) and Si (0.2%). Two 6 M HCl treatments at 95 �Cdecreased the radioactivity concentration of 137Cs in the ball-

milled Iwaki’11 sample residue from 5.9 (60 �C) to 4.8 kBq/kg.

After the two 6MHCl treatments at 95 �C, the Fe content in the

residue of the Iwaki’11 sample decreased to 0.2% Fe2O3 (Table

5). SiO2 and Al2O3 accounted for most of the residue in the

Iwaki’11 sample (96.5% in total) (Table 5), indicating that the

residue was mostly silicates. For this residue, XRD diffraction

peaks of quartz and feldspar were observed, although no

peaks for other minerals were observed (Fig. 5). This residue

was used for the static leaching test after washing with pure

water 5 times. After 28 days in pure water and synthetic

seawater at 60 �C, 137Cs was not detected in the leachates

(Table 3). This result demonstrates that the 137Cs in the res-

idue is strongly immobilized in the silicates.

Fig. 5 e XRD patterns of residues of ball-milled SSAs after

heating in aqueous HCl at 95 �C for 24 h. The ball-milled

Iwaki’11 and Teganuma’11 samples were heated in

aqueous 6 M HCl, and the Nakakuji’12 sample was heated

in concentrated aqueous HCl.

4. Discussion

Although there is no data available on the chemical states of137Cs in raw sewage sludge, the 137Cs was assumed to be

adsorbed on sewage sludge components, such as organic

substances and minerals. When dehydrated sewage sludge is

incinerated, organic compounds are decomposed and new

minerals form. The 137Cs in organic substances is thought to

have been redistributed to pre-existing and newly formed

mineral phases. Therefore, it is not possible to compare the

chemical states of 137Cs in SSA with those in other contami-

nants, such as contaminated soils.

We revealed that most of the 137Cs in SSAwas contained in

Fe-bearing phases, such as iron oxides (e.g. hematite), and

some of the 137Cs was strongly immobilized in silicates, such

as quartz, feldspar, and talc. However, iron oxides, quartz, and

feldspar have very low sorption capacities for Cs (Torstenfelt

et al., 1982; Akiba et al., 1989; Sasaki et al., 2007; Bellenger

wat e r r e s e a r c h 6 8 ( 2 0 1 5 ) 6 1 6e6 2 6 625

et al., 2008; Giannakopoulou et al., 2012; Ohnuki et al., 2013).

The sorption of Cs on talc is unknown. Talc (Mg3Si4O10(OH)2) is

a 2:1 clay mineral. Although its structure is similar to that of

mica, the sorption capacity of talc for Cs is thought to be very

low because talc has no negative charge in the structure. A

fraction of talc in the HCl treatment residues of the ball-milled

SSAs was probably lost, because planetary ball milling for 6 h

converts the structure of talc to an amorphous structure and

the Mg is leached by heating in aqueous HCl at 80 �C (Yang

et al., 2006). In fact, the XRD diffraction peak at 9.4�e9.5�

observed for as-incinerated SSAs was not observed for resi-

dues of the ball-milled SSAs after heating in aqueous HCl at

95 �C for 24 h (Fig. 5). Therefore, minerals with very low

sorption capacities retained 137Cs in SSAs.

There are two possible mechanisms for the retention of137Cs. The first is containment in aggregates, where dense

aggregates composed of fine mineral particles sorbing 137Cs

physically confined the 137Cs. The second is strong bonding

that becomes significant at extremely low Cs concentrations.

The sorption behavior of Cs on minerals depends on its

concentration; as the concentration of Cs decreases, the

fraction of the Cs tightly sorbed on minerals increases

(Ohnuki et al., 2013). This could be because many minerals

have strong affinity sites for Cs at very low site densities. In

the ball-milled Iwaki’11 sample, which had the highest con-

centration of 137Cs, the mole concentrations of 137Cs in the

iron oxides and the silicates are estimated to be 3 � 10�10 and

1 � 10�11 mol/kg, respectively. These concentrations are

extremely low. We should also consider the possibility that

when the iron oxides and silicates were heated during the

incineration of the sewage sludge, the sorbed 137Cs diffused

inside their crystals and this made it more difficult to desorb

the 137Cs.

Dissolution of 137Cs from SSAs was increased by heating at

95 �C. Kawamoto et al. developed a method for effectively

dissolving 137Cs from contaminated Fukushima soils by

heating in aqueous 0.5 M HNO3 at 95 �C (Kawamoto et al.,

2013). We tested dissolution behavior of the ball-milled

Iwaki’11, Teganuma’11, and Nakakuji’12 samples in aqueous

0.5 M HNO3 at 95 �C. For the ball-milled Iwaki’11 and Tega-

numa’11 samples, the decrease in 137Cs caused by aqueous

0.5MHNO3 treatmentwas similar to that achieved by aqueous

6 MHCl (Table S1). For the ball-milled Nakakuji’12 sample, the

decrease in 137Cs caused by aqueous 0.5 M HNO3 was

approximately 8% greater than that caused by concentrated

aqueous HCl. The residues after the 0.5 M HNO3 treatment

were brown, showing that a large amount of iron remained

undissolved. The decrease in Al and P caused by aqueous 0.5M

HNO3 treatment was similar to that caused by aqueous 6 M

HCl. A considerable amount of Si was dissolved by a aqueous

0.5 M HNO3, whereas Si was hardly dissolved in aqueous 6 M

HCl. The decrease in Fe by aqueous 0.5 M HNO3 was less than

that by aqueous 6 M HCl (Table S1). As shown in Table 5, the

content of Fe in the residue treated with aqueous 6 M HCl at

95 �C was very low. These results indicate that for the 0.5 M

HNO3 treatment, some of the 137Cs in the iron oxides remained

undissolved and some of the 137Cs in silicates were dissolved

together with Si. As discussed previously, the 137Cs in silicates

is non-leachable and that in HCl-soluble phases, including

iron oxides, is leachable. To develop SSA treatment methods,

it is necessary to consider the amounts of 137Cs and the min-

eral phases that dissolve.

5. Conclusions

We investigated the characteristics of 137Cs in radioactively

contaminated SSAs produced after the Fukushima Daiichi

nuclear accident. We can draw the following conclusions.

Most of the 137Cs in SSAs was contained in Fe-bearing

phases, which were probably iron oxides, and a minor frac-

tion of the 137Cs was contained in silicates. A tiny amount of

the 137Cs was contained in K-bearing phases, which were

probably alkali metal salts. There was no evidence to show

that the 137Cs was contained in the major mineral phases,

which were phosphates. At least some of the 137Cs in the Fe-

bearing phases and K-bearing phases were leachable to

environmental water.

Heating SSAs in aqueous HCl dissolved a large fraction of

Fe-bearing phases and the radioactivity concentration

decreased. However, in SSAs, there were many aggregates of

nanosize particles of Fe-bearing phases and silicates, and the

Fe in these aggregates was not easily dissolved in aqueous

HCl. Because of this, the radioactivity concentrations of 137Cs

in the SSAs did not decrease below the radioactivity criterion

for landfill disposal in Japan. Pulverization of SSAs was the

most effective pretreatment for dissolving Fe-bearing phases

in aqueous HCl in the following two respects. The method

reduces the radioactivity concentrations of 137Cs in SSAs

below the radioactivity criterion. The 137Cs was very strongly

immobilized on the HCl-treatment residues, which were

mostly silicates. Thus, 137Cs should not leach to environ-

mental water from these residues.

Appendix A. Supplementary data

Supplementary data related to this article can be found at

http://dx.doi.org/10.1016/j.watres.2014.10.038.

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