Glassy Wasteforms for Nuclear Waste Immobilization

Abstract

Glassy wasteforms currently being used for high-level radioactive waste (HLW) as well as for low- and intermediate-level radioactive
waste (LILW) immobilization are discussed and their most important parameters are examined, along with a brief description
of waste vitrification technology currently used worldwide. Recent developments in advanced nuclear wasteforms are described
such as polyphase glass composite materials (GCMs) with higher versatility and waste loading. Aqueous performance of glassy
materials is analyzed with a detailed analysis of the role of ion exchange and hydrolysis, and performance of irradiated glasses.

Full text

Glassy Wasteforms for Nuclear Waste Immobilization
MICHAEL I. OJOVAN and WILLIAM E. LEE
Glassy wasteforms currently being used for high-level radioactive waste (HLW) as well as for
low- and intermediate-level radioactive waste (LILW) immobilization are discussed and their
most important parameters are examined, along with a brief description of waste vitrification
technology currently used worldwide. Recent developments in advanced nuclear wasteforms are
described such as polyphase glass composite materials (GCMs) with higher versatility and waste
loading. Aqueous performance of glassy materials is analyzed with a detailed analysis of the role
of ion exchange and hydrolysis, and performance of irradiated glasses.
DOI: 10.1007/s11661-010-0525-7
The Minerals, Metals & Materials Society and ASM International 2010
I. INTRODUCTION
THE choice of wasteform to use for nuclear waste
immobilization is a difficult decision, and durability is
not the sole criterion. In any immobilization process
where radioactive materials are used, the process and
operational conditions can become complicated, partic-
ularly if operated remotely and equipment maintenance
is required. Therefore, priority is given to reliable,
simple, rugged technologies and equipment, which may
have advantages over complex or sensitive equipment. A
variety of matrix materials and techniques is available
for immobilization. The choice of the immobilization
technology depends on the physical and chemical nature
of the waste and the acceptance criteria for the long-
term storage and disposal facility to which the waste will
be consigned. A host of regulatory, process, and product
requirements has led to the investigation and adoption
of a variety of matrices and technologies for waste
immobilization. The main immobilization technologies
that are available commercially and have been demon-
strated to be viable are cementation, bituminization, and
vitrification.
[1]
Vitrification is a particularly attractive
immobilization route because of the high chemical
durability of the glassy product. This characteristic
was used by industry for centuries. The chemical
resistance of glass can allow it to remain in a corrosive
environment for many thousands and even millions of
years. In most countries, High-level radioactive waste
(HLW) has been incorporated into alkali borosilicate or
phosphate vitreous waste forms for many years, and
vitrification is an established technology. Large streams
of low- and intermediate-level radioactive waste (LILW)
are planned to be vitrified in the United States, South
Korea, and Russia. Vitreous waste forms represent one
end of the spectrum of the HLW waste forms shown in
Table I.
[2,3]
At the other end of the spectrum shown in Table I,
the use of predominantly crystalline ceramic wasteforms
(ceramication) has also been proposed including single-
phase ceramics such as zircon to accommodate a limited
range of active species such as Pu and multiphase
systems such as Synroc to accommodate a broader
range of active species.
[4]
To date, these systems have not
been used extensively to immobilize active waste.
Recently, however, there has been a trend to systems
intermediate between the ‘‘completely’’ glassy or crys-
talline materials.
[5]
An illustration of nuclear waste
forms used and developed for industrial application
is given in Figure 1.
[6]
Typically, nondurable crystals
are Na
2
SO
4
,Na
2
MoO
4
2H
2
O, CaMoO
4
, NaF, and
NaCl, whereas durable crystals include BaAl
2
Ti
6
O
7
,
CaZrTi
2
O
7
, CaTiO
3
, and TiO
2
.
GCMs include the following: (1) glass ceramics in
which a glassy waste form is crystallized in a separate
heat treatment
[7,78]
; (2) GCMs in which, e.g., a refrac-
tory waste is encapsulated in glass such as hot-pressed
lead silicate glass matrix encapsulating up to 30 vol pct
of La
2
Zr
2
O
7
pyrochlore crystals to immobilize minor
actinides
[9]
; (3) GCM formed by pressureless sintering
of spent clinoptiloite from aqueous waste processing
[10]
;
(4) some difficult wastes such as the French HLW
U/Mo-containing materials immobilized in a GCM
termed U-Mo glass formed by cold crucible melting
that partly crystallize on cooling
[11]
; (5) yellow phase
containing wastes are immobilized in Russia in a yellow
phase GCM containing up to 15 vol pct of sulfates,
chlorides, and molybdates
[12]
; and (6) GCM that immo-
bilizes ashes from incineration of solid radioactive
wastes.
[13]
Note that alkali-rich wastes at the Hanford
site are also immobilized in glassy wasteforms with high
crystal contents that characterize them as GCMs.
[14]
GCMs may be used to immobilize long-lived radio-
nuclides (such as actinide species) by incorporating them
into the more durable crystalline phases, whereas the
short-lived radionuclides may be accommodated in the
less durable vitreous phase. Historically, the crystalliza-
tion of vitreous waste forms has always been regarded
as undesirable, as it has the potential to alter the
MICHAEL I. OJOVAN, Assistant Professor, is with the Depart-
ment of Materials Science and Engineering, Immobilisation Science
Laboratory, University of Sheffield, Sheffield, S1 3JD, United King-
dom. Contact e-mail: M.Ojovan@sheffield.ac.uk WILLIAM E. LEE,
Professor, is with the Department of Materials, Imperial College
London, London SW7 2AZ, United Kingdom.
Manuscript submitted May 12, 2010.
Article published online November 17, 2010
METALLURGICAL AND MATERIALS TRANSACTIONS A VOLUME 42A, APRIL 2011—837

composition (and hence, durability) of the remaining
continuous glass phase, which would (eventually) come
into contact with water. However, there has been a
recent trend toward higher crystallinity in ostensibly
vitreous wasteforms so that they are more correctly
termed GCMs. This is particularly apparent in the
development of hosts for more difficult waste or where
acceptable durability can be demonstrated even where
significant quantities of crystals (arising from higher
waste loadings) are present, such as the high sodium
Hanford wastes. Acceptable durability will result if the
active species are locked into the crystal phases that are
encapsulated in a durable, low-activity glass matrix. The
GCM option is being considered in many countries
including Australia, France, Russia, South Korea, the
United Kingdom, and the United States. The process-
ing, compositions, phase assemblages, and microstruc-
tures of GCMs may be tailored to achieve the necessary
material properties.
II. STABILITY OF GLASSES
Glasses as amorphous materials are among the most
abundant materials on the earth. Moreover, glasses are
among the most ancient of all materials used by
humans. The geological glass obsidian was used first
by humans thousands of years ago to form objects
including knives, arrow tips, and jewelry. Human-made
glass objects were first reported in the Mesopotamian
region as early as 4500 BC, and glass objects dating as
old as 3000 BC have been found in Egypt. These glasses
have compositions similar to those of modern soda-
lime-silicate glass as soda ash from fires, limestone from
seashells, and silica sand from the beaches were readily
available. Current human-made glasses include a large
variety of materials from window panels and cookware
to aerospace windows and bulk metallic glasses, as well
as nuclear waste glassy materials.
[12,15–17]
Glasses have an internal structure made of a well-
developed, topologically disordered, three-dimensional
(3-D) network of interconnected microscopic structural
blocks. Glasses are formed typically on rapid cooling of
melts to avoid crystallization, because little time is
allowed for the ordering processes. Whether a crystalline
or amorphous solid forms on cooling depends also on
the ease with which a random atomic structure in the
liquid can transform to an ordered state. Most known
glassy materials are characterized by atomic or molec-
ular structures that are relatively complex and become
ordered only with some difficulty. Therefore, it has long
been assumed that the glassy state is characteristic of
special glass-forming or network materials such as
covalent substances that exhibit a high degree of
structure organization at length scales corresponding
to several atomic separations. However, after the
discovery of metallic glasses, it was realized that almost
any substance, if cooled sufficiently fast, could be
obtained in the glassy state.
[18,19]
Glasses are solid amorphous materials with a topo-
logically disordered structure of interconnected struc-
tural blocks, which in, e.g., silicate glasses are SiO
4
tetrahedra. After heating, glasses continuously change
most of their properties to those of a liquid-like state in
contrast to crystals where such changes occur abruptly
at a fixed temperature (the melting point). The solid-like
behavior of amorphous materials at low temperatures is
separated from liquid-like behavior at high temperatures
by the glass transition temperature T
g
. The glass
transition is a kinetically controlled, fairly sharp change
in derivative properties such as thermal expansion
coefficient or specific heat. T
g
depends on the rate of
cooling, although empirically it can be calculated
roughly from Kauzmann’s relation
Tgð2=3ÞTL½1
where T
L
is the liquidus temperature at which a phase
diagram shows a crystal-free melt. The liquid–glass
transition has been considered as a second-order phase
Fig. 1—Phase composition of nuclear waste forms.
Table I. Classification of Types of Glass/Ceramic Waste Forms
Glasses Glass Composite Materials (GCMs) Crystalline Ceramics
Magnox, UK Glass ceramics Single phase
Defense Waste Processing Facility and
West Valley Demonstration Project,
Savannah River, SC
Crystal waste encapsulated
in glass matrix
Multiphase (e.g., Synroc*)
Alumina phosphate, Russia
*Synroc consists of titanates hollandite (BaAl
2
Ti
6
O
16
), zirconolite (CaZrTi
2
O
7
), perovskite (CaTiO
3
), and TiO
2
. The hollandite mainly fixes the
fission products and some process chemicals, whereas actinides and rare earth elements are bound in zirconolite and perovskite.
838—VOLUME 42A, APRIL 2011 METALLURGICAL AND MATERIALS TRANSACTIONS A

transition in which a supercooled melt yields, on
cooling, a glassy structure and properties similar to
those of crystalline materials, e.g., of an isotropic solid
material.
[20]
Amorphous oxide materials have an internal structure
made of a 3-D network of interconnected structural
blocks where each broken bond is treated as an
elementary configurational excitation—configuron.
[21,22]
Whether a material is liquid or solid depends primarily
on the connectivity between its elementary building
blocks so that solids are characterized by a high degree
of connectivity, whereas structural blocks in fluids have
lower connectivity (Figure 2(a)). Melting of a material
can be considered as a percolation via broken bonds,
[23]
e.g., melting of an amorphous oxide material occurs
when the configurons form a percolation cluster.
[24,25]
T
g
depends on quasiequilibrium thermodynamic parame-
ters of the bonds, e.g., on the enthalpy (H
d
) and entropy
(S
d
) of configurons, which can be found from available
experimental data on viscosity
[24–26]
Tg¼Hd=SdþRln½ð1hcÞ=hc½½2
where R is the gas constant. For strong melts such as
SiO
2
, the percolation threshold in the previous equation
h
c
=0
c
, where 0
c
is the universal Scher-Zallen critical
density in the 3-D space 0
c
=0.15 ±0.01.
[24,25]
How-
ever, for fragile materials, the percolation thresholds are
material dependent, and h
c
<< 1.
[26]
The connectivity of
a bond lattice is characterized by its geometry and
Hausdorff dimension.
[27]
Configuron motion in the bond
network occurs in the form of thermally activated jumps
from site to site. If the temperature is much below T
g
,
the glass network is characterized as an ideal disordered
structure that is described by a Euclidean 3-D geometry.
The higher the temperature, the larger the clusters made
of configurons in the disordered bond network. Finally,
at T
g
they form a macroscopic so-called percolation
cluster, which penetrates the whole volume of the
disordered network.
[28]
Moreover, the formation of the
percolation cluster changes the topology of the bond
network from 3-D Euclidean below to fractal
D
f
=2.55 ±0.05-dimensional above the percolation
threshold.
[29]
Hence, the glass–liquid transition can be
considered as a percolation-type phase transition with
formation near the percolation threshold of dynamic
fractal structures
[30]
or twinkling fractal structures.
[31]
The glass–liquid transition is therefore associated with
the reduction of the Hausdorff dimension of bonds from
the 3-D Euclidean in the glassy state to the fractal
D
f
=2.55 ±0.05-dimensional in the liquid state.
[27,30]
Above T
g
, amorphous materials have a fractal geometry
of bonds and a liquid-like behavior, whereas glasses
are amorphous materials below T
g
, when they have
solid-like behavior and 3-D geometry of bonds alike
crystals (Figure 2(b)). Note that the fractal struc-
tures formed near the glass transition are dynamic
fractal structures.
[22,27,31]
The physical and chemical durability of glasses
combined with their high tolerance to compositional
changes makes glasses irreplaceable when highly toxic
wastes such as long-lived and highly radioactive wastes
need reliable immobilization for safe long-term storage,
transportation, and consequent disposal. Although,
compared with crystalline materials of the same com-
position, glasses are metastable, their relaxation to a
thermodynamically stable crystalline structure is im-
peded kinetically. Relaxation processes in amorphous
materials are controlled by viscosity. Maxwell’s relaxa-
tion time gives the characteristic relaxation time
required to attain stabilized parameters
sM¼g=G½3
where Gis the shear modulus and gis the viscosity. The
higher the viscosity, the longer the relaxation time.
Viscosity change is thermally activated, and glass-
forming amorphous oxides are characterized by high
viscosities under normal conditions,
[18]
e.g., fused silica
has an activation energy of viscosity at low temperature
Fig. 2—(a) State of crystal and amorphous materials as a function of connectivity. (b) Temperature behavior of density and geometry of bonds.
METALLURGICAL AND MATERIALS TRANSACTIONS A VOLUME 42A, APRIL 2011—839

of Q
H
=759 kJ/mol and a shear modulus 31 GPa,
which gives relaxation times as long as s
M
~10
98
years.
This time is incommensurably longer than the lifetime of
the universe (~1.5 910
10
years). Another example to
demonstrate that long-term stability of silicate glasses is
practical absent of stress relaxation at room tempera-
tures, e.g., a high permanent internal stress is preserved
in glass articles made more than 2000 years ago.
[32]
Moreover, extrapolation of measured data showed that
the time necessary for a 0.5 pct decrease in permanent
stress is equal to 2.7 910
11
years for toughened boro-
silicate glass.
[32]
An intriguing question for nuclear waste glasses is
whether the irradiation does or does not affect the
relaxation processes, e.g., crystallization of an oxide
glass. It was found recently that intensive electron
irradiation of silicate glasses can cause a significant
decrease of viscosity and spinodal decomposition.
[33]
However, the dose rates required to observe such effects
are much higher than those that occur in vitrified
nuclear waste.
[34,35]
Therefore, the metastability of
silicate glasses commonly used by various industries is
a theoretical rather than a practical question. Oxide
glasses are stable longer than any imaginable geological
timescale of our universe.
III. GLASSES FOR NUCLEAR WASTE
IMMOBILIZATION
Two main glass types are currently used for nuclear
waste immobilization: borosilicates and phosphates.
The exact compositions of nuclear waste glasses are
tailored for easy preparation and melting, avoidance of
phase separation and uncontrolled crystallization, and
acceptable chemical durability, e.g., leaching resistance.
Vitrification can be performed efficiently at tempera-
tures below 1500 K (1227 C) because of the volatility of
the fission products, notably Cs and Ru, so avoiding
excess radionuclide volatilization and maintaining vis-
cosities below 10 Pa second to ensure high throughput
and controlled pouring into canisters. A more fluid glass
is preferred to minimize blending problems. Phase
separation on melting is most important for waste
streams containing glass-immiscible constituents; how-
ever, these can be immobilized in form of isolated and
phase separated disperse phase (in GCMs). The leaching
resistance of nuclear waste glasses is a paramount
criterion as it ensures low release rates for radionuclides
on any potential contact with water.
Vitrification involves melting waste materials with
glass-forming additives so that the final vitreous product
incorporates the waste contaminants in its macrostruc-
ture and microstructure. Hazardous waste constituents
are immobilized either by direct chemical incorporation
into the glass structure or by physical encapsulation. In
the former, waste constituents are dissolved in the glass
melt; Si, B, and P are included in the glass network on
cooling, whereas others such as Cs, K, Na, Li, Ca, Pb,
and Mg act as modifiers. Several glass compositions
were designed for nuclear waste immobilization; how-
ever, few are used in practice.
[1,5,8,12]
Table II gives
compositions of several nuclear waste glasses.
High waste loadings and high chemical durability can
be achieved in both borosilicate and aluminophosphate
glasses. Moreover, such glasses immobilize well large
quantities of actinides; for example, borosilicate glasses
can accommodate up to 7.2 mass pct PuO
2.[37]
In
contrast to borosilicate melts, molten phosphate glasses
are highly corrosive to refractory linings; this behavior
has limited their application. Currently, this glass is used
only in Russia, which has immobilized HLW from
nuclear fuel reprocessing in alumina-phosphate glass
since 1987.
[16]
It should be emphasized that nuclear waste glasses are
never completely homogeneous vitreous materials but
contain significant amounts of bubbles, foreign inclu-
sions such as refractory oxides, and other immiscible
components. Figure 3shows an example of SEM
characterization of stimulant British Magnox Waste
glass, which reveals heterogeneities and phase separa-
tion characterized by small droplets from 10 to 20 lm
sizes, in which subsequent fine segregation on a scale of
about 100 nm was observed.
[38]
Encapsulation is applied to elements and compounds
that have low solubility in the glass melt and do not fit
into the glass microstructure either as network formers
or modifiers. Immiscible constituents that do not mix
easily into the molten glass are typically sulfates,
chlorides, and molybdates, as well as noble metals such
as Rh and Pd, refractory oxides with high liquidus
temperatures such as PuO
2
, noble metal oxides, and
spinels.
Encapsulation is carried out either by deliberate
dispersion of insoluble compounds into the glass melt,
immiscible phase separation on cooling, or by sintering
Table II. Compositions of Some Nuclear Waste Glasses, Mass Pct
Plant, Waste, Country SiO
2
P
2
O
5
B
2
O
3
Al
2
O
3
CaO MgO Na
2
O Miscellaneous Waste Loading
R7/T7, HLW, France 47.2 — 14.9 4.4 4.1 — 10.6 18.8 £28
DWPF, HLW, United States 49.8 — 8.0 4.0 1.0 1.4 8.7 27.1 £33
WVP, HLW, UK 47.2 — 16.9 4.8 — 5.3 8.4 17.4 £25*
PAMELA, HLW, Germany—Belgium 52.7 — 13.2 2.7 4.6 2.2 5.9 18.7 <30
Mayak, HLW, Russia — 52.0 — 19.0 — — 21.2 7.8 £33
Radon, LILW, Russia 43 — 6.6 3.0 13.7 — 23.9 9.8 <35
*It was demonstrated recently that waste loading can be increased up to 35 pct to 38 pct.
[36]
£10 for fission products and minor actinide oxides.
840—VOLUME 42A, APRIL 2011 METALLURGICAL AND MATERIALS TRANSACTIONS A

of glass and waste powders so that the waste form
produced is a GCM. However, this requires a more
complex melter supplied with a stirrer.
IV. GLASSES FOR HIGH-, LOW-,
AND INTERMEDIATE-LEVEL WASTES
Although developed initially for HLW, vitrification is
used currently for immobilization of LILW, such as
from operation and decommissioning of nuclear power
plants.
[39,40]
Vitrification is one technology that has been
chosen to solidify 18,000 tonnes of ore mining tailings at
the Fernald, OH plant.
[41]
Plans are in place to vitrify
vast volumes of waste; for example, the vitrification of
the low-level radioactive waste at Hanford, WA is
expected to produce more than 160,000 m
3
of glass.
[42]
The U.S. Department of Energy (DOE) plans to vitrify
54 million gallons of mixed radioactive waste stored at
its Hanford site in eastern Washington State, which
represents 60 pct of the United States’ volume of
radioactive waste.
[43]
The world’s largest waste vitrifica-
tion plant (Waste Treatment Project (WTP)) is now
under construction at Hanford. Borosilicate glass will be
used for immobilization of Hanford’s low-activity waste
(LAW). The vitrified LAW will be disposed of in a
shallow land-burial facility. The proposed disposal
system has been shown to retain the radionuclides
adequately and prevent contamination of the surround-
ing environment. Release of radionuclides from the
waste form via interaction with water is the prime threat
to the environment surrounding the disposal site; the
two major dose contributors in Hanford LAW glass that
must be retained are
99
Tc and
129
I.
[44]
Several glasses
were developed to immobilize Hanford low-activity
wastes with composition ranges that will meet the
performance expectations of the Hanford site burial
facility.
[44]
It is planned that the WTP will vitrify 99 pct
of Hanford’s waste by 2028. The WTP melter chosen to
vitrify HLW is a joule-heated ceramic melter (JHCM).
The JHCM has nickel–chromium alloy electrodes that
heat the waste and glass-forming additives to ~1450 K
(1150 C). The glass melt is stirred by convection and by
bubbler elements, and then poured into carbon steel
canisters to cool. Note that the carbon steel canisters are
not corrosion resistant and do not present a barrier in an
envisaged repository environment. Canisters with vitri-
fied HLW are sealed and decontaminated. It was
planned that the vitrified HLW would be disposed of
in the Yucca Mountain geological repository; although
because of the change in the U.S. government policy,
this HLW will need to be stored. Current plans also
provide for the vitrified LAW to be stored on site.
Moreover, at Hanford, it is planned to use a bulk
vitrification process in which liquid waste is mixed with
controlled-composition soil in a disposable melter.
[43]
The process of bulk vitrification involves mixing LAW
with Hanford’s silica-rich soil and surrounding it with
sand and insulation in a large steel box. Electrodes are
inserted to vitrify the mixture, and when cooled, the
melter, its contents, and the embedded electrodes will be
buried as low-level waste (LLW) in an on-site burial
ground.
The vitrification of LLW and intermediate-level waste
(ILW) was studied intensively in Russia in the middle of
1970s.
[12]
A number of glass compositions were devel-
oped for immobilization of liquid waste containing
mainly sodium nitrate. Various boron-containing min-
erals as well as sandstone were tested as glass-forming
additives. Datolite CaBSiO
4
(OH) was found to be most
suitable fluxing agent. Other systems such as Na
2
O
(LILW oxides)-2CaO B
2
O
3
-SiO
2
were studied and glass
forming regions, melt viscosity and resistivity, leach rate
of sodium (and
137
Cs for actual waste), density, radia-
tion stability, and compressive strength were measured.
Suitable glass composition areas were established.
[12]
The most important properties of these glasses are given
in Table III.
Loam and bentonite clays were also used as glass-
forming additives. Up to 50 pct of either loam clay or
bentonite in the batch was substituted for sandstone.
This substitution increases the chemical durability of
glass and, moreover, such batches containing 20 to
25 wt pct of water form homogeneous pastes, which are
stable for long times without segregation and are
transportable in pipes over long distances.
[12]
Sodium
nitrate is the major component of both institutional
liquid LILW and nuclear power plant (NPP) opera-
tional wastes from RBMK (channel type uranium-
graphite) reactors. NPP wastes from WWER reactors
contain boron, although the major components of this
waste are sodium nitrate and sodium tetrahydroxyl
borate NaB(OH)
4
. Thus, there is no need to add boron-
containing additives to vitrify WWER waste. Silica,
loam, or bentonite clay or their mixtures are suitable as
glass-forming additives. WWER waste glasses are in the
Na
2
O-(Al
2
O
3
)- B
2
O
3
-SiO
2
system for which glass-
forming regions are well known. Long-term tests of
vitrified LILW have been carried out in a shallow
ground experimental repository since 1987.
[46]
These
show a low and diminishing leaching rate of radionuc-
lides. Boron-free aluminosilicate glasses in the Na
2
O-
CaO-Al
2
O
3
-SiO
2
system for immobilization of institu-
tional and RBMK wastes were produced from waste,
sandstone, and loam clay (or bentonite).
Fig. 3—British Magnox-waste glass secondary electron image.
METALLURGICAL AND MATERIALS TRANSACTIONS A VOLUME 42A, APRIL 2011—841

Some liquid waste streams contain sulfate and
chloride ions, which limits the waste oxide content to
5-10 wt pct because of the low sulfate and chloride
solubility (~1 pct) in silicate and borosilicate melts.
Thus, LILW vitrification becomes inefficient. Excess
sulfate–chloride phases segregate as a separate phases
floating on the melt surface because of the immiscibility
of silicate and sulfate (chloride) melts. The same
phenomenon occurs for molybdate- and chromate-
containing waste vitrification, where the separate phase
is colored and named ‘‘yellow phase.’’
[45]
Vitrification
of this waste can be done by using vigorous melt
agitation followed by rapid cooling to the upper anneal-
ing temperature to fix the dispersed sulfate–chloride
phase into the host borosilicate glass. Sulfate–chloride-
containing GCM (see yellow phase GCM in Figure 1)
have only a slightly diminished chemical durability
compared with sulfate–chloride free aluminosilicate and
borosilicate glasses (Table III), which is sufficiently high
for them to be used for waste immobilization. GCM
produced using a thermochemical technique based on
exothermic self-sustaining reactions are also composed of
vitreous and crystalline phases, mainly silicates and
aluminosilicates.
[47]
V. NUCLEAR WASTE VITRIFICATION
Vitrification is most suitable for aqueous radioactive
wastes which should be solidified for safer storage and
disposal. Waste vitrification is attractive because of the
following reasons:
(a) High capability of glass to reliably immobilize a
range of elements
(b) Simple production technology adapted from glass
manufacture
(c) Small volume of the resulting wasteform
(d) High chemical durability of waste form glasses in
contact with natural waters
(e) High tolerance of these glasses to radiation
damage.
The high chemical resistance of glass allows it to
remain stable in corrosive environments for thousands
and even millions of years. Several glasses are found in
nature, such as obsidians (volcanic glasses), fulgarites
(formed by lightning strikes), tektites found on land in
Australasia and associated microtektites from the bot-
tom of the Indian Ocean, moldavites from central
Europe, and Libyan Desert glass from western Egypt.
Some of these glasses have been in the natural environ-
ment for approximately 300 million years with low
alteration rates of only tenths of a millimeter per million
years.
The excellent durability of vitrified radioactive waste
ensures a high degree of environment protection. Waste
vitrification gives high waste volume reduction along
with simple and cheap disposal facilities. Despite a high
initial investment and then operational costs, taking
account of transportation and disposal expenses, the
overall cost of vitrified radioactive waste is usually lower
than alternative options.
The drawbacks of vitrification are its high initial
investment cost, high operational cost, and complex
technology requiring well-qualified personnel. These
reasons made vitrification economically viable when
relatively large volumes of radioactive waste with rela-
tively stable composition are available such as HLW or
operational radioactive wastes from NPP. Self-sustaining
vitrification has no such limitations, in contrast to
conventional vitrification technologies; however, this
technology is limited to calcined waste streams.
[47]
Vitrification technology comprises several stages,
starting with evaporation of excess water from liquid
radioactive waste, followed by batch preparation,
Table III. Properties of Vitrified LILW
Properties
Borosilicate Glasses
GCM
High
Sodium Waste
Operational
WWER* Waste
Glass Immiscible
(High Sulfate) Waste
Waste oxide content, mass pct 30 to 35 35 to 45 30 to 35 and up to 15 vol pct
of immiscible waste*
Viscosity, Pa s, at 1500 K (1227 C) 3.5 to 5.0 2.5 to 4.5 3.0 to 6.0 (for vitreous phase)
Resistivity, Xm, at 1500 K (1227 C) 0.03 to 0.05 0.02 to 0.04 0.03 to 0.05
Density, g/cm
3
2.5 to 2.7 2.4 to 2.6 2.4 to 2.7
Compressive strength, MPa 80 to 100 70 to 85 50 to 70
Normalized Leach Rate, g/(cm
2
day),
(28-day IAEA test)
137
Cs 10
–5
to 10
–6
~10
–5
10
–4
to 10
–5
90
Sr 10
–6
to 10
–7
~10
–6
10
–6
to 10
–7
Cr, Mn, Fe, Co, Ni ~10
–7
to 10
–8
~10
–7
10
–7
to 10
–8
REE, An ~10
–8
~10
–8
~10
–8
Na 10
–5
to 10
–6
~10
–5
10
–4
to 10
–5
B<10
–8
<10
–8
£10
–8
SO
4
2
~10
–6
(when present) — 10
–4
to 10
–5
at content <15 vol pct
*WWER or VVER, water-water energetic reactor, Russian analog of Western PWR, pressurized water reactor.
For example, ‘‘yellow phase.’’
[45]
842—VOLUME 42A, APRIL 2011 METALLURGICAL AND MATERIALS TRANSACTIONS A

calcination, glass melting, and pouring and cooling of
vitrified waste blocks with potentially small amounts of
secondary waste (Figure 4).
In a two-stage process, the waste is calcined prior to
melting. In the one-stage process, both waste calcination
and melting occurs in the melter. Thin film evaporators
are typically used, and the remaining salt concentrate is
mixed with the necessary additives and, depending on
the type of vitrification process, is directed to one or
another process apparatus.
In the two-stage vitrification process with separate
calcinations, the waste concentrate is fed into the
calciner. After calcinations, the required glass-forming
additives (usually as a glass frit) together with the
calcine are fed into the melter. In both cases, two
streams come from the melter: the glass melt containing
most of radioactivity and the off gas flow, which
contains off gases and aerosols.
In the one-stage vitrification process, glass-forming
additives are mixed with concentrated liquid wastes, and
so a glass-forming batch is formed (often in the form of
a paste). This batch is then fed into the melter where
subsequent water evaporation occurs, followed by
calcination and glass melting, which occur directly in
the melter.
Two types of melters are currently used at waste
vitrification plants: JHCMs and induction-heated mel-
ters, which can either be hot (induction, hot crucible
[IHC]) or cold, e.g., cold crucible melters (CCMs)
(Figure 5).
The melt waste glass is poured into containers
(canisters) made of stainless steel when immobilizing
HLW or carbon steel for vitrified LILW. These may or
may not be cooled slowly in an annealing furnace to
avoid accumulation of mechanical stresses in the glass.
When annealing is not used, cracking occurs resulting in
a large surface area being potentially available for attack
by water in a repository environment. Despite the higher
final surface areas of nonannealed glasses, these are
sufficiently durable to ensure a suitable degree of
radionuclide retention. Hence, in many cases, annealing
is not used in vitrification facilities.
The second stream from the melter goes to the gas
purification system, which is usually a complex system
Fig. 4—(a) Simplified schematic of a vitrification process. (b) Technological flow-sheet diagram of LILW vitrification plant at SIA ‘‘Radon.’’
1—Interim storage tank; 2—concentrate collector; 3—rotary film evaporator; 4, 15—HEPA-filters; 5, 17, 21—heat-exchangers; 6, 19—reservoirs;
7—glass-forming additives hoppers; 8—screw feeder; 9—batch mixer; 10—mechanical activator; 11—peristaltic pump; 12—cold crucible;
13—annealing furnace; 14—sleeve (coarse) filter; 16—pumps; 18—absorption columns; 20—heater. HEPA-filter is high-efficiency particulate air
filter and LRW is liquid radioactive waste.
[5]
METALLURGICAL AND MATERIALS TRANSACTIONS A VOLUME 42A, APRIL 2011—843

that removes from the off gas not only radionuclide but
also chemical contaminants. Operation of this purifica-
tion system leads to generation of a small amount of
secondary waste. For example, the distribution of beta
gross activity at PAMELA waste vitrification plant was
>99.88 pct in waste glass and the rest in secondary
waste, e.g.,<0.1 pct in intermediate-level waste,
<0.01 pct in cold waste, and <0.01 pct in off gas.
[49]
Table IV summarizes data on radioactive waste vitrifi-
cation facilities.
VI. DURABILITY OF GLASSY WASTEFORMS
The reliability of radionuclide immobilization is
characterized by the rate at which radionuclides can
be released from the wasteform during long-term
storage. As the most plausible path for reintroduction
of radioactivity into the biosphere is via water, the
most important parameters that characterize the abil-
ity of glass to hold on to the active species are the
leach rates. The leaching behavior of wasteforms
containing different amounts of waste radionuclides
is compared using the normalized leaching rates NR
i
for each i
th
nuclide expressed in g/cm
2
day and the
normalized mass losses NL
i
, expressed in g/cm
2
. These
are determined measuring the concentrations c
i
(g/l)
of inactive constituents or activities a
i
(Bq/L) of
radionuclides in the water solution in contact with
the wasteform after a time interval Dtexpressed in
days. The mass fraction of a given nuclide iin a
wasteform is defined as
fi¼wi=w0½4
where w
i
is the mass of nuclide in the wasteform (g)
and w
0
is the mass of the wasteform (g). The specific
activity of a given radionuclide in a wasteform q
i
(Bq/g) is defined as
qi¼Ai=w0½5
where A
i
is the radioactivity of radionuclides in the
wasteform (Bq). The normalized leaching rate of non-
radioactive nuclides NR
i
(g/cm
2
day) is calculated using
the expression
NRi¼ciV=fiSDt½6
where Sis the surface area of the wasteform in contact
with the water (cm
2
), Vis the solution volume (l), and
Dtis the test duration in days. The normalized leach
rate of radioactive nuclides NR
i
(g/cm
2
day) is calcu-
lated using the expression
NRi¼aiV=qiSDt½7
where a
i
(Bq/l) is the specific radioactivity of the solu-
tion. Normalized mass losses NL
i
(g/cm
2
day) are
Fig. 5—(a) JHCM melter at Nuclear Research Centre (FZK), Karlsruhe, Germany for vitrification of highly active waste concentrate
(HAWC).
[48]
Courtesy S. Weisenburger, Forschungszentrum Karlsruhe. (b) A view of a CCM (1) and rectangular carbon steel containers
(2) used at Moscow SIA ‘‘Radon’’ for vitrification of LILW.
[13]
844—VOLUME 42A, APRIL 2011 METALLURGICAL AND MATERIALS TRANSACTIONS A

determined for nonradioactive and radioactive nuclides
respectively from
NLi¼ciV=fiS½8
and correspondingly for radioactive samples
NLi¼aiV=qiS½9
A set of standard tests to determine the water
durability of vitrified waste and other wasteforms was
developed at the Materials Characterization Centre
(MCC) of the Pacific Northwest National Laboratory,
Richland, WA. These MCC tests are now the interna-
tionally approved standards used worldwide. The most
important tests are given in Table V.
[50]
The glass-formulation methodology defines a range of
compositions around a reference formula to allow for
slight fluctuations in the composition of the waste feed
stream. Figure 6illustrates schematically the effect of
additives on leaching rates.
[5,15]
For example, the addi-
tion of Al
2
O
3
improves the leach resistance of borosil-
icate glasses, whereas higher content of alkalis (Li
2
O,
Na
2
O, K
2
O) reduces glass durability.
As the cost-saving incentive is to increase the waste
loading in a wasteform, the optimal glassy wasteform
compositions are tailored as a compromise between
waste loading and final glass durability accounting also
for processing parameters on vitrification.
[17]
Table VI
gives typical data on the parameters of HLW borosil-
icate and phosphate glasses.
[5,51]
Table IV. Operational Data of Vitrification Programs
Facility Waste Type Melting Process Operational Period Performance
R7/T7, La Hague, France HLW IHC Since 1989/1992 6811 tonnes (237.9 10
6
Ci in
17206 canisters) to 2009
AVM, Marcoule, France HLW IHC Since 1978 857.5 tonnes in 2412 canisters
R7, La Hague, France HLW CCM Since 2003 GCM: U-Mo glass
WVP, Sellafield, UK HLW IHC Since 1991 >5000 canisters to 2009
DWPF, Savannah River, SC HLW JHCM Since 1996 5000 tonnes in 2845 canisters to 2009
WVDP, West Valley, NY HLW JHCM Since 1996 ~500 tonnes in 275 canisters to 2002
EP-500, Mayak, Russia HLW JHCM Since 1987 ~8000 tonnes to 2009 (900 10
6
Ci)
CCM, Mayak, Russia HLW CCM Pilot plant 18 kg/h by phosphate glass
PAMELA, Mol, Belgium HLW JHCM 1985–1991 ~500 tonnes in 2200 canisters
VEK, Karlsruhe, Germany HLW JHCM 2010 ~60 m
3
of HLW (24 10
6
Ci),
to be completed in 2010
Tokai, Japan HLW JHCM Since 1995 >100 tonnes in 241 canisters
(110 l) to 2007
Radon, Russia LILW JHCM 1987 to 1998 10 tonnes
Radon, Russia LILW CCM Since 1999 >30 tonnes
Radon, Russia ILW SSV 2001 to 2002 10 kg/h, incinerator ash
VICHR, Bohunice, Slovakia HLW IHC 1997 to 2001, upgrading
work to restart operation
1.53 m
3
in 211 canisters
WIP, Trombay, India HLW IHPT Since 2002 18 tonnes to 2010 (110 10
3
Ci)
AVS, Tarapur, India HLW IHPT Since 1985
WIP, Kalpakkam, India HLW JHCM Under testing & commissioning
WTP, Hanford, WA LLW JHCM Pilot plant since 1998 ~1000 tonnes to 2000
Taejon, Korea LILW CCM Pilot plant, planned 2005 ?
Saluggia, Italy LILW CCM Planned ?
IHC, Induction, hot crucible; CCM, cold crucible induction melter, JHCM, Joule heated ceramic melter; IHPT, induction heated pot type melter;
SSV, self-sustaining vitrification.
Table V. Standard Tests on Immobilization Reliability
Test Conditions Use
ISO 6961, MCC-1 Deionized water. Static. Monolithic specimen.
Sample surface to water volume (S/V) usually 10 m
–1
.
Open to atmosphere. Temperature 298 K (25 C)
for ISO test, 313 K (40 C), 343 K (70 C),
and 363 K (90 C) for MCC-1 test
For comparison of waste forms.
MCC-2 Deionized water. Temperature 363 K (90 C). Closed. Same as MCC-1 but at high temperatures.
PCT (MCC-3) Product consistency test. Deionized water stirred
with glass powder. Various temperatures. Closed.
For durable waste forms to accelerate leaching.
SPFT (MCC-4) Single pass flow through test. Deionized water.
Open to atmosphere.
The most informative test.
VHT Vapor phase hydration. Monolithic specimen. Closed.
High temperatures.
Accelerates alteration product formation.
METALLURGICAL AND MATERIALS TRANSACTIONS A VOLUME 42A, APRIL 2011—845

Vitrified radioactive waste is chemically durable and
reliably retains radioactive species. Typical normalized
leaching rates NR of vitrified waste forms are below 10
–5
to 10
–6
g/cm
2
day. Moreover, as glasses and GCM are
highly corrosion resistant, their high nuclide retention is
expected to last for many millennia.
VII. LONG-TERM DURABILITY OF NUCLEAR
WASTE GLASSES (EFFECT OF TEMPERATURE,
pH, AND TIME)
Corrosion durability of vitrified waste is the most
important acceptance parameter for disposal.
[12,52]
Insight into the long-term behavior of nuclear waste
glasses is an important issue related to our ability to
assess the reliability of nuclear waste immobilization in
an envisaged repository environment. The release of
radioactive species, which in nuclear waste glasses are
invariably cations, can be caused by corrosion of the
glass in contact with groundwater. However, the
potential contact of water with glass is deferred in
actual disposal systems to times after the waste con-
tainer has been breached. The material selection of the
engineered barriers, e.g., canisters depends on each
particular country. Stainless steels are considered but
also carbon steel, nickel alloys, titanium alloys, and
copper may be used. For vitrified HLW containers,
which are made of stainless steel, these times are
expected to be of the order of many hundreds or even
thousands of years. High temperatures and radiation
dose rates are likely only for the first few hundred years
after HLW vitrification, so that container temperatures
will be close to those of the ambient rock by the expected
time of contact with groundwater. Moreover, the role of
bc-radiolysis will also become negligible because of low
radiation dose rates. Vitrified LILW is almost invariably
at the ambient temperature of a repository environment.
In addition, this type of waste is expected to be disposed
of in near-surface repositories, which are often charac-
terized by near-neutral groundwaters and relatively low
host-rock temperatures. Hence, the temperatures of
nuclear waste glasses at the times of expected contact
with groundwater are likely to be close to those of the
surrounding repository environment.
Aqueous corrosion of nuclear waste glasses is a
complex process that depends on many parameters
such as glass composition and radionuclide content,
time, temperature, groundwater chemical composition,
and pH. Corrosion of silicate glasses, including
nuclear waste borosilicate glasses, involves two major
processes: diffusion-controlled ion exchange and glass
network hydrolysis.
[5,12,18,53–57]
Diffusion-controlled ion
exchange reactions lead to selective leaching of alkalis
and protons entering the silicate structure to produce a
hydrated alkali-deficient layer on the glasses. Hydrolysis
being a near-surface reaction of hydroxyl ions with the
silicate network leads to its destruction, resulting in
congruent dissolution of glass constituents and subse-
quent precipitation of hydrous silica-gel layers as sec-
ondary alteration products.
The role of ion exchange in the overall corrosion
behavior is important because it is the principal release
mechanism when the glass network hydrolysis is sup-
pressed. In dilute near-neutral solutions, ion exchange
controls the initial cation release, and at low tempera-
ture and pH, it can dominate over hydrolysis for many
hundreds of years.
[58]
Ion exchange involves the inter-
diffusion and exchange of the cation in the glass with a
proton (probably as H
3
O
+
) from the water. The ion
exchange reaction of glass with water can be written as
ð SiOAÞglass þH2O$ðSiOHÞglass þAOH
½10
This reaction is controlled by the counter diffusion of
protons (probably as H
3
O
+
) from the water, which
replace cations in the glass structure, e.g., cations
bounded to nonbridging oxygens (NBO). The rate of
Fig. 6—Glass leach rates as a function of glass composition.
Table VI. Typical Properties of HLW Glasses
Glass
Density
(g/cm
3
)
Compressive
Strength (MPa)
NR, 28
th
day,
in 10
–6
g/cm
2
day
Thermal
Stability,* K (C)
Damaging
Dose,* Gy
Borosilicate 2.7 22 to 54 0.3 (Cs); 0.2 (Sr). ‡823 (550) >10
9
Phosphate 2.6 9 to 14 1.1 (Cs); 0.4 (Sr). ‡723 (450) >10
9
*The irradiation has a small impact on glasses, and the damaging dose is the absorbed dose above which the radionuclide NR’s increase several
times, whereas thermal stability is the temperature above which the radionuclide NR’s increase >10
2
times.
846—VOLUME 42A, APRIL 2011 METALLURGICAL AND MATERIALS TRANSACTIONS A

cation release into water via diffusion-controlled ion
exchange rx
A
is given by
[58]
rxA¼qfAaAD0H=ptðÞ
1=2100:5pH expðEdA=2RTÞ½11
where qis the glass density, f
A
is the mass fraction of
the cation Ain the glass (Eq. [4]), a
A
=j/C
A
(0), jis
a constant relating concentration of protons at the
glass surface and in the contacting water and C
A
(0) is
concentration of cations at the glass surface, D
0H
is
the preexponential coefficient in the diffusion coefficient
for protons in the glass DH¼D0HexpðEdH =RTÞ;
E
dH
is activation energy for diffusion of protons in the
glass, tis time and E
dA
is the activation energy of
effective diffusion (kJ/mol). The activation energy for
interdiffusion is the sum of the enthalpy of motion of
protons or H
3
O
+
cations, H
mH
, and the enthalpy of
formation of NBO, H
NBO
:E
dA
=H
mH
+H
NBO
. The
average normalized leaching rate caused by the ion
exchange processes of cation A,NRx
A
(g/cm
2
day), is
measured experimentally or it can be found theoreti-
cally by calculating the total normalized cation release
via ion exchange and dividing it by the leach test dura-
tion t(days) resulting in
[58]
NRxA¼2qDA=ptðÞ
1=2½12
where D
A
are the effective diffusion coefficients. The
rates of ion exchange with time diminish as the inverse
square root of time reflecting the fact that near surface
layers of glasses are depleted in cationic species, and
deeper and deeper layers are supplying cations for ion
exchange reaction. Equation [12] indicates that the ion
exchange rate is determined completely by D
A
, which
depends on the pH of the attacking solution and has
an Arrhenius temperature dependence of activation en-
ergy E
dA
. Table VII gives D
A
for several silicate glasses
tested in near-neutral waters.
[58]
The second mecha-
nism of silicate glass corrosion—hydrolysis—is repre-
sented schematically by the reaction
SiOSi þH2O$2ð SiOHÞ½13
Hydrolysis results in complete dissolution of the glass
network and formation of orthosilicic acid, H
4
SiO
4
.
This process leads to a congruent release of glass
constituents into the water. The rate of hydrolysis is
calculated using the transition state theory of silicate
mineral dissolution
[59]
rh ¼kag
Hþ1Q=KðÞ
r
½expðEa=RTÞ½14
where kis the intrinsic rate constant,aHþis the hydro-
gen ion activity, gis the pH power law coefficient, E
a
is the activation energy and Qthe ion-activity product
of the rate-controlling reaction, Kis the pseudoequilib-
rium constant of this reaction, and ris the net reac-
tion order. The rate of release of cations Aas a result
of hydrolysis is given by
rhA¼qfArh ½15
The average normalized leaching rate caused by hydro-
lysis NRh
A
(g/cm
2
day) is measured experimentally or
it can be found theoretically by calculating the total
normalized cation release via hydrolysis and dividing it
by the leach test duration t(days) resulting in
NRhA¼qrh ½16
The affinity term [1 (Q/K)
r
] characterizes the decrease
in solution aggressiveness with respect to the glass as it
becomes increasingly concentrated in dissolved elements
and as the ion activity product Qof the reactive species
approaches the material solubility product K,e.g.,rh fi0,
when QfiK. In dilute aqueous systems when KQ,
the affinity term is simply equal to unity [1 
(Q/K)
r
]=1. Note that g=0.5,
[60]
and the higher the
pH of the attacking water solution the higher the rate of
hydrolysis.
Both ion exchange and hydrolysis contribute to
aqueous glass corrosion. Because of rapid dissolution
of near surface layers, which are generally different from
the bulk, there is an additional contribution to leaching
termed instantaneous surface dissolution. This is
accounted for by an exponential term in the dissolution
rate
rsA¼nsAkAexp kAtðÞ ½17
where k
A
is the rate of instantaneous dissolution of
species iin water (day
1
), and n
sA
is the surface con-
centration of radionuclides (g Æcm
2
). The total rate
of species released from glass into the water is given
by the sum
rA¼rsAþrxAþrhA½18
Table VIII summarizes the three most important mech-
anisms of glass corrosion in nonsaturated aqueous
solutions.
[61]
Equations [11] and [14] reveal that ion exchange
occurs preferentially in acidic and neutral solutions but
diminishes quickly with the increase of pH, whereas
hydrolysis occurs preferentially in basic solutions but
diminishes quickly with the decrease of pH. It is
normally considered that for pH <9 to 10, ion
exchange dominates glass corrosion, whereas hydrolysis
reactions are significant when the pH exceeds 9.
[62,63]
In
acidic media below pH =6, the water concentration of
protons (or hydronium ions) is high, resulting in a high
rate of ion exchange. The role of glass network
Table VII. Effective Diffusion Coefficients in Some Silicates
Glass Cation A
Temperature,
K(C) D
A
(m
2
/s)
USA SM539 B 363 (90) 4 910
21
Li 7 910
21
Na 4 910
21
British
Magnox-waste
Li 313 (40) 1.9 910
–20
Na 4.4 910
–20
Russian O3O-6 Na 295 (22) 2.8 910
21
Russian K-26 Cs 277.5 (4.5) 5 910
21
Russian Bs-10 Cs 284 (11) 1.8 910
–20
Quartz artifacts H 279.6 to 297
(6.6 to 24)
~10
–25
Silica glass H 296 (23) 1.4 910
–21
METALLURGICAL AND MATERIALS TRANSACTIONS A VOLUME 42A, APRIL 2011—847

dissolution in this process is insignificant, and cation
leaching is ion-selective with different leaching curves
for different cations. Above pH =9, the role of ion
exchange becomes insignificant because of the high
water concentration of hydroxyl ions, and thus, the
glass network commences to dissolve rapidly via
reaction [13]. In such basic media, the release of cations
becomes congruent as destruction of the glass network
results in practically complete dissolution of all glass
constituents. Note that the hydrolysis reaction [13]
becomes impeded if solutions become silica saturated.
The pH dependence of corrosion rate has a U-form
curve with typical minimal changes in the near-neutral
water solutions. Figure 7illustrates the pH dependence
of corrosion rate for borosilicate glass K-26.
[61]
Figure 7indicates the mass release from glass follows
simple power laws only below pH =6 and above
pH =9. In the interval 6 <pH <9, the dependence is
a more complex function with a changing slope when
pH changes and with minimal corrosion rates achieved
close but not at pH =7. Because of the time dependence
of ion-exchange rates in corroding glasses, the minimum
rates drift with time to lower values of pH. Therefore,
attempts to model the pH dependences by simple power
laws separated at pH =7 will inevitably result in smaller
values of exponent terms mand g. For example, the
exponent terms for UK Magnox waste glass based on
data from the pH ranges 2 <pH <7 and 7 <pH <10
were m =0.39 for boron, m=0.43 for silicon, and
g=0.43,
[64]
which are somewhat smaller than the
theoretical value of m=0.5.
The ion-exchange reaction of glass with water leads to
a gradual diminution of cation content in the near-
surface glass layers. Because of this depletion in the glass
near-surface layers over time, the rate of ion exchange
diminishes. In contrast, the rate of glass hydrolysis,
although small in near-neutral conditions, remains
constant. Hence, hydrolysis will eventually dominate,
once the near-surface glass layers have become depleted
in cations. Depending on glass composition and the
conditions of aqueous corrosion, as well as on time, the
contribution of the basic mechanisms to the overall
corrosion rate can be different. For example, in dilute
solutions, ion exchange controls the initial corrosion
stage. Moreover, at expected disposal temperatures
(below 320 K or several tens of C), the corrosion of
glasses will occur via ion exchange for long periods of
time, even in contact with non-silica saturated ground-
water, although ion-exchange controls corrosion of
glasses over geological timescales when the contacting
groundwater is silica saturated and the hydrolytic
dissolution of the glass network is impeded.
The time required for silicate glasses to reach the
hydrolysis stage in near-neutral solutions depends
mainly on glass composition and temperature. Note
that we suppose unchanged water parameters, so there is
no coupling between water chemistry and the corroding
glass. More highly polymerized glasses are hydrolyti-
cally decomposed more slowly. Thus, glasses with higher
silica contents require longer times before hydrolysis
becomes dominant compared with high-sodium-content
glasses. The corrosion regimes of silicate glasses should
be characterized in terms of time-temperature parame-
ters as the higher the temperature the sooner hydrolysis
becomes dominant. This occurs because the activation
energy for hydrolysis E
a
is significantly higher than the
activation energies of diffusive processes E
dA
. It has been
shown that the diffusion-controlled ion exchange stage is
dominant up to a time, s(T), given by
[58,61]
sðTÞ¼s0exp½2EaEdA
ðÞ=RT½19
where s
0
is a preexponential term. The time s
0
increases
with silica concentration approaching the solubility limit
when QfiK, demonstrating that the only important
cation release mechanism remains ion exchange in
this case. For example, in near-neutral water solu-
tions, UK Magnox-waste glass undergoes incongruent
Table VIII. Main Characteristics of Corrosion Mechanisms
Mechanism, Rate Behavior Instantaneous Surface Dissolution Ion Exchange Hydrolysis
Time* Short-term effect exp(kt) Diminishes t
1/2
Independent*
Temperature Arrhenian Arrhenian, Universal
activation energy
Arrhenian, One high
activation energy
pH Dependent Decreasing 10
0.5pH
Increasing 10
0.5pH
Saturation effectsUnlikely Unlikely Impeded (1C
Si
/C
Si saturation
)
Selectivity Selective Selective Congruent
*Time behavior may be affected by saturation effects.
Changes in solution chemistry may affect solution pH.
Fig. 7—pH dependence of cesium normalized leaching rate of glass
K-26.
848—VOLUME 42A, APRIL 2011 METALLURGICAL AND MATERIALS TRANSACTIONS A

ion-exchange over a period of 28 days at temperatures
as high as 60 Cto90C and an estimate of s(60 C) 
28 days, E
a
60 kJ mol
–1
and E
di
36 kJ mol
–1
,
which enables the tentative identification of the most
likely scenarios for corrosion of UK Magnox-waste
glass as a function of temperature and time.
[58]
Figure 8
shows that corrosion in deionized water at a constant
temperature immediately after instantaneous surface
dissolution begins with a fully controlled ion exchange
phase.
As corrosion progresses the impact of hydrolysis
becomes significant with comparable contributions from
both ion exchange and hydrolytic reactions. Finally,
glass corrosion in demonized water is fully controlled by
hydrolysis. The characteristic time that indicates the
duration of pure ion exchange phase is given by Eq. [19].
Table IX gives characteristic times s(T) for several
glasses corroding in non-Si-saturated near-neutral water
solutions. Note that an increase of contacting water pH
would decrease characteristic transition times.
VIII. EFFECTS OF SELF-IRRADIATION
Although radiation is not believed to have a signif-
icant effect on glass corrosion rates,
[65]
it can influence
glass stability through the formation of corrosive
radiolytic products in the contacting water solution,
alteration of glass structure, and radiation-enhanced
diffusion. For the expected times of water–glass contact
(‡10
3–4
years), radiation dose rates are likely to be low
so that no intensive radiolysis is expected. Significant
alteration of the glass structure is also not expected as
the glass is originally amorphous and no more disorder
arises from radiation damage than originally is present
in the glass structure. The formation of gas bubbles
observed in glasses under irradiation as well as redistri-
bution of alkalis is an effect that results from radiation-
induced diffusion rather than from alteration of glass
structure. Hence, the most important radiation-induced
effects in nuclear waste glasses for times of water–glass
contact are those that result from radiation-enhanced
diffusion. Between the two basic corrosion mechanisms,
ion exchange is controlled directly by the diffusion of
species in the glass, whereas hydrolysis can be affected
only indirectly by radiation-enhanced diffusion. The rate
of ion exchange for an irradiated glass is given by
Eq. [11], however, with parameters corresponding to an
irradiated glass. Because the diffusion coefficients of
species are higher compared with nonirradiated glasses,
the rates of ion exchange are higher. The higher the
absorbed dose and the lower the temperature, the higher
the increase in the ion exchange rate.
[66]
This remains
true only until the ion exchange is significant in the
corrosion of glasses, e.g., in silica-saturated conditions
(when dissolution rh =0) and at relatively low temper-
atures and pH <9. Because of enhanced diffusion
coefficients, the selectivity of leaching should be higher
compared with unirradiated glasses. In several experi-
ments, the effects of irradiation were observed clearly
and characterized. Static leach tests conducted with
PNL 76-78 glass immersed in deaerated and demonized
water demonstrated the highest pH increase and release
rates for Si, B, and Na at the lowest test temperature
323 K (50 C) and lowest differences at the highest test
temperature 363 K (90 C) for c-radiation tests at a dose
rate 1.75 10
4
Gy/hour compared with nonirradiated
glass.
[67]
Moreover, cation releases were most incongru-
ent at 323 K (50 C) and were almost congruent at
363 K (90 C).
[67,68]
Highly incongruent glass dissolu-
tion was observed in c-irradiated in situ tests of waste
glasses in Belgian Boom clay; moreover, the glass
corrosion mechanism becomes a more diffusion-con-
trolled process in the presence of the radiation field.
[69]
The hydrolytic mechanism of corrosion is hardly
affected by self-irradiation. Hence, practically no change
in corrosion behavior is expected when the dominant
mechanism of corrosion is hydrolysis. This conclusion is
confirmed by many experiments; for example, unaf-
fected corrosion behavior has recently been found for a
Pu-bearing borosilicate glass over a pH interval of 9 to
12 at 353 K to 361 K (80 Cto88C).
[70]
Leach tests of
French SON68 glass in silica-saturated solutions showed
that ion-exchange rates are increased after irradiation
whereas hydrolysis remained unchanged.
[71]
Similar
effects were observed in irradiated phosphate glasses.
[16]
Summarizing the data available leads to the conclu-
sion
[72]
that the irradiation had a detectable and even
significant impact in cases when corrosion occurred via
diffusion-controlled ion exchange. These conditions are
characterized by relatively low temperatures (£323 K
Fig. 8—Corrosion mechanisms for British Magnox-waste glass in
deionized water.
Table IX. Transition Times for the Intermediate Stage
of Glass Dissolution
Glass T, K (C) s(T)
British Magnox-waste 363 (90) >28 days
USA SRL131A, SRL202A 298 (25) >240 days
Russian Bs-10 284 (11) 3.2 years
Russian K-26 277.5 (4.5) 16.4 years
Roman IF
(Archaeological)
287 to 288
(14 to 15)
~1800 years
METALLURGICAL AND MATERIALS TRANSACTIONS A VOLUME 42A, APRIL 2011—849

[50 C]), low and medium pH (£8), relatively high
absorbed doses (‡610
6
Gy), and for diluted solutions
over short corrosion times (t£s(T)). The most impor-
tant consequences of irradiation in these cases were
enhanced leaching rates and incongruency in glass
corrosion. In contrast, when hydrolysis controls the
glass corrosion, practically no differences were found in
the corrosion behavior of nonirradiated and irradiated
glasses. These conditions prevail at high temperatures
(>323 K [50 C]) and high pH of contacting water (>9),
as well as long corrosion times in dilute solutions
(t‡16s(T)) when, because of cationic depletion of
near surface glass layers, ion exchange reactions are
diminished.
IX. CONCLUSIONS
Glass is a solid-state material that behaves like a
solid crystalline material but has a topologically disor-
dered internal structure. Although, compared with
crystalline materials of the same composition, glasses
are metastable materials, their relaxation to crystalline
structures is kinetically impeded so that practically no
crystallization occurs within times that for oxide glasses
are longer than the universe lifetime. The physical and
chemical durability of glasses combined with their high
tolerance to compositional changes makes glasses
irreplaceable when highly toxic wastes such as long-
lived and highly radioactive wastes need reliable immo-
bilization for safe long-term storage, transportation and
consequent disposal. Immobilization of radioactive
wastes in glassy materials using vitrification has been
used successfully for many years, although novel glassy
wasteforms are still being developed, and studies of
their properties are performed. Nuclear waste vitrifica-
tion is attractive because of its flexibility, the large
number of elements that can be incorporated in the
glass, its high corrosion durability, and the reduced
volume of the resulting wasteform. Vitrification is a
mature technology and has been used for HLW
immobilization for more than 40 years in France,
Germany and Belgium, Russia, UK, India, Japan,
and the United States. Vitrification involves melting of
waste materials with glass-forming additives so that the
final vitreous product incorporates the waste contam-
inants in its macrostructure and microstructure.
Hazardous waste constituents are immobilized either
by direct incorporation into the glass structure or by
encapsulation when the final glassy material can be in
form of a GCM. Both borosilicate and phosphate
glasses are used currently to immobilize nuclear wastes;
moreover, in addition to relatively homogeneous
glasses, novel GCMs are used to immobilize problem-
atic waste streams. The spectrum of wastes that are
currently vitrified increases from HLW to LILW such
as legacy wastes in Hanford, WA and nuclear power
plant operational wastes in Russia and Korea. Glassy
wasteforms in the form of relatively homogeneous
glasses or as GCM incorporating crystalline disperse
phases are currently the most reliable hosts used for
nuclear waste immobilization.
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