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1010 Charged Particle and Photon Interactions with Matter

concentrations were quickly achieved on the surface (Nielsen etal., 2008a), and this assumption was

employed in the further model calculations of radiolytic oxidative dissolution (Nielsen etal., 2008b).

In the repository, it is predicted that anoxic corrosion of metal canisters by water will generate

hydrogen and the concentration of dissolved hydrogen would be very high under connement pres-

sures up to 5MPa. It has been shown that dissolved hydrogen signicantly reduces the dissolution

rates of spent nuclear fuels (Sunder etal., 1990; Spahiu etal., 2000, 2004a,b; Röllin etal., 2001;

Ollila etal., 2003; Carbol etal., 2005, 2009; Loida etal., 2005, Muzeau etal., 2009). The corrosion

potential of UO

in dissolved H

under gamma radiation decreased to a level lower than that of the

unirradiated sample and reduction of the oxidized surface by hydrogen atom (King and Shoesmith,

2004; Shoesmith, 2008).

While the effect of dissolved hydrogen on dissolution rates has been experimentally established,

its mechanism is still a matter of controversy. Since dissolved hydrogen showed almost no effect on

the 5MeV helium ion beam radiolysis of aqueous hydrogen peroxide solutions (Pastina and LaVerne,

2001), the effect probably involves surface reactions. The effect has been observed in the case of

alpha-doped UO

indicating a direct effect of hydrogen (Spahiu etal., 2004a,b; Carbol etal., 2009).

However, dissolved hydrogen showed no effect on the decomposition of H

on neat UO

surfaces

without ionizing radiation. Noble metals (Pd, Mo, Ru, Tc, Rh) as ssion products in spent nuclear fuels

are observed as nanosized metallic particles called ε-particles. Reduction of U(VI) in UO

was facili-

tated by ε-particles, which act as catalysts for the oxidation of dissolved hydrogen. This was observed

as lower corrosion potentials of SIMFUEL in the presence of ε-particles (Broczkowski etal., 2005).

The catalytic effect of noble metal particles was studied with UO

pellets doped with different amount

of Pd particles (Trummer etal., 2008). The oxidative dissolution of UO

by H

was reduced by

increased concentrations of Pd particles under anoxic conditions. The oxidative dissolution of UO

dissolved oxygen was also catalyzed by Pd particles (Nilsson and Jonsson, 2008). Radiolytic dissolu-

tion under gamma radiation was almost inhibited in 3wt% Pd doped UO

in an N

atmosphere. The

dissolution rate equation was expressed as the difference between the oxidation and reduction rates of

, and the former was divided into direct and noble metal–catalyzed reactions. The reaction rate

constants of oxidation and reduction were derived by comparison with the calculated and measured

dissolution rates. They were close to diffusion-limited values (Trummer etal., 2009).

The effects of solutes in groundwater on the radiolytic dissolution of UO

have also been studied.

Reduction by dissolved Fe

(Johnson and Shoesmith, 1988) acts as a radical recombiner as follows:

Fe OH Fe OH

2 3+ + −

+ → +

•

(34.90)

Fe H Fe H

3 2+ + +

+ → +

•

(34.91)

The same effect has recently been found to explain the effect of bromide ion which also suppressed

the effect of dissolved hydrogen by recombining hydroxyl and hydrogen radicals in the same way

(Metz etal., 2008). In aerated condition, 2-propanol increased the corrosion rate. The alcohol scav-

enges OH

•

to form isopropyl radical, which then reduces dissolved oxygen to generate hydrogen

peroxide. In contrast, 2mol dm

–3

−

decreased the rate, which was attributed to lower G

)

and

lower

dissolved oxygen concentration at higher ionic strength (Roth and Jonsson, 2009).

34.4.3.2

dissolution

of hlw g

lass

In contrast to the case of spent nuclear fuel, the oxidation of the waste form is not critical with

regard to dissolution of HLW glass. The effects of ionizing radiation on HLW glass were at rst

studied mainly under gamma radiation (McVay and Pederson, 1981; Pederson and McVay, 1983;

Burns etal., 1982; Barkatt etal., 1983). Increased leach rates were attributed to increased acidity,

possibly due to the formation of nitric acid in the gas phase of the vessel. A comprehensive review

of radiation effects on HLW glasses are given by Weber (1988) and Weber et al. (1997). Changes

Radiation Chemistry inNuclear Engineering 1011

in chemical bonding, migration of elements, gas generation, volume and specic surface area, and

physical properties have been studied. Their effects on the leach rates were reviewed. It was con-

cluded that the radiation effect on the leach rate would be less than a factor of 10, assuming uncer-

tainties

in the measured leach rates.

Effects

of alpha radiation have also been studied and increased leach rates were observed in

alpha-doped glasses (Weber etal., 1985). However, no enhancement of leach rates by alpha radia-

tion was observed in ion-beam irradiation experiments (Abdelouas etal., 2004) or alpha-doped

R7T7

glasses (Peuget etal., 2007).

Leach

tests of HLW glasses doped with actinides were performed in situ, embedded in the back-

ll materials under Co-60 gamma radiation (European Commission, 2007). Certain differences

were observed in the diffusion proles of redox-sensitive nuclides in the buffer material result-

ing from gamma radiation. The leachate became slightly acidic under gamma-irradiation and the

mechanism

is not clearly understood.

34.4.4 redox front Migration induced by groundwater radiolySiS

In a diffusion–reaction system, when oxidants/reductants migrate into the region in which reduc-

tants/oxidants are xed, redox reactions occur almost instantaneously. A sharp front may be formed

which separates the region in which redox reactive species are consumed and xed. Formation and

migration of such “redox fronts” in natural formations have been studied as natural analogues of the

geological disposal systems that might experience changes in the redox condition during operation

(e.g., Yoshida etal., 2008).

Redox fronts in the geological disposal system might be formed in the earliest phases of enclo-

sure as the excavated walls of the repository are exposed to air and oxygen. Intrusion of oxygen

containing groundwater to the repository might also form a redox front. Groundwater radiolysis

was assumed to be another possible mechanism of redox front formation in the repository as follows

(Neretnieks, 1983): After the waste package loses its integrity and porewater comes into contact

with the waste form, alpha-radiolysis of water would occur. While radiolytically generated H

relatively inert and would dissipate without chemical reactions, oxidizing species such as H

and O

would gradually diffuse out of the waste package. They are then advectively transported

by groundwater ow into the bedrock through fractures, and eventually diffuse into rock matrices

through fracture surfaces. H

and O

are reduced by Fe(II) contained in iron bearing minerals

such as biotite in granite, and a redox front is formed in the rock matrix. The extent of the movement

of the redox front in a repository was estimated to be about 60 m after 10

years, by assuming a H

formation rate on the spent nuclear fuel surface by alpha-radiolysis, geometries of the fractures and

groundwater ow rate, and the molar percentage of ferrous ion in the rock matrices. However, the

generation of H

might have been overestimated by using the G-value of pure water alpha-radi-

olysis, as radiolytic generation of H

may be signicantly suppressed with Fe

in groundwater

(Christensen and Bjergbakke, 1982a,b). Decomposition of oxidants in backll materials was also

neglected, while they also contain iron-bearing minerals (e.g., pyrite) and smectites, which act as

reductants for radiolytic oxidants. In an additional study, the migration distances of a redox front in

the

backll materials surrounding a canister were estimated (Andersson etal., 1982).

The

experimental verication of formation and migration of the radiolytic products of water in engi-

neered barrier materials were performed. A beta-emitter (

147

Pm) disk was contacted with a compacted

bentonite disk saturated with synthetic groundwater and hydrogen generated by radiolysis of porewa-

ter in bentonite was measured after it diffused out of the opposite face of the bentonite disk (Eriksen

and Jacobsson, 1983). The same setup was applied with an alpha-emitter (

241

Am) disk (Eriksen etal.,

1987, Eriksen and Ndalamba, 1988). The G-values of hydrogen peroxide generation and Fe

reduction

in bentonite were estimated to be 0.069 and 0.40 μmol J

–1

, respectively. These results are supposed

to have provided a basis for performance assessments of geological disposal systems to consider the

redox-front formation in the repository due to groundwater radiolysis (e.g., SKI, 1996).

1012 Charged Particle and Photon Interactions with Matter

The model of the reaction and migration of radiolytic products in the repository was further

developed by taking into account various processes, e.g., the effect of dissolved oxygen, Fe

, the

geometry of fuel cladding, precipitation of dissolved U(VI), and alpha-radiolysis of water in the

buffer (Liu and Neretnieks 2001a,b,c, 2002; Liu etal., 2003). One-dimensional diffusion-reaction

model calculations of the processes in the canister were also performed to estimate the dissolu-

tion rate by taking into account the corrosion of the iron canister material (Shoesmith etal.,

2003). The effect of dissolved Fe

from corrosion of the waste package on the formation and

decomposition of oxidative species by alpha-radiolysis was also considered, and the migration

of redox fronts in the buffer material was predicted to be stopped (Johnson and Smith, 2000).

However, redox front migration by groundwater radiolysis in the near eld has not been consid-

ered in recent performance assessments. Based on recent results concerning spent nuclear fuel

dissolution under reducing conditions, it is believed that radiation-induced oxidative dissolution

is strongly suppressed (e.g., SKB, 2006b).

aCknowledgments

Bruce J. Mincher’s contribution to this work was supported by the U.S. Department of Energy,

Ofce

of Nuclear Energy, under DOE Idaho Operations Ofce Contract DE-AC07-05ID14517.

Makoto

Yamaguchi’s contribution was based in part on a project for assessment methodol-

ogy development of chemical effects on geological disposal systems funded by the Ministry of

Economy, Trade and Industry, Japan. He thanks Morimasa Naito for carefully reading a draft of the

manuscript and giving him valuable comments and corrections.

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