8.2.3. Release of radionuclides from waste forms

There are three principal ways in which the radionuclides in radioactive waste can escape from the confines of a repository:

by transport in groundwater – the groundwater pathway;

by transport as a gas – the gas pathway;

through human or natural intrusion.

With few exceptions, for deep disposal the most important of these pathways is the first.

8.2.3.1. Solubility and sorption

Many important radionuclides – the actinides in particular – have low solubilities, so that the rate at which they can be leached from the waste is constrained; such radionuclides are said to be solubility-limited. Other radionuclides may have a high solubility or else be present in sufficiently small quantities that all of the radionuclide inventory could, in principle, exist as a solute; such radionuclides are said to be inventory-limited. For many radionuclides, an important phenomenon is that of sorption on near-field materials. Sorption lowers the concentration of radionuclides in solution so that the rate at which they can be transported in groundwater is reduced. This simple solubility/sorptionlimited conceptual model underlies many safety assessments (see also comments in Chapter 6).

A great deal of R&D has been done to measure and catalogue radionuclide solubilities under near-field conditions (e.g., Chambers et al., 1995; NEA, 2005a) and to measure the sorption of radionuclides on near-field materials (e.g., Wieland and van Loon, 2002; Bradbury and Baeyens, 2003).

8.2.3.2. Waste form dissolution

The solubility/sorption limited model is, in essence, a thermodynamic model that takes no account of the kinetics of dissolution. And yet, waste forms such as vitrified HLW are specifically formulated so that they have low solubility and low rates of dissolution in water. Consequently, waste form dissolution is a subject of considerable interest for R&D. Where it is intended that suitably contained SF is to be directly disposed of, the waste form of interest is SF itself. Irradiation of the fuel leads to the formation of fission products and to structural changes and these have a direct impact on the dissolution behaviour, making it necessary for the experiments to be conducted on real SF. Experimental data (obtained from ‘‘bare’’ fuel fragments) are often interpreted in terms of an ‘‘instant release fraction’’ followed by a slower release. The instant fraction is considered to consist of highly soluble radionuclides close to water-accessible fuel surfaces; the slower release results from dissolution of the uranium dioxide fuel matrix. Because of the importance of solubility and sorption, fuel dissolution can be strongly influenced by the chemistry of the surrounding groundwater (see Poinssot et al., 2004 for details).

Where SF is reprocessed to recover the re-usable uranium and plutonium, the waste form of interest is vitrified HLW. Again, considerable research has been done to examine the dissolution of this waste form. A frequent experimental observation is that the rate of release of radionuclides reduces with time. This has been shown to be caused by the formation of an alteration film on the surface of the glass consisting of a gel of corrosion products. This film creates a diffusion barrier that hinders further attack.

If the chemistry of the water in contact with the glass changes, this temporarily destabilises the gel, but then a new gel forms and the dissolution process continues as before (see Ribet et al., 2004 for an overview).

A third type of radioactive waste requiring deep disposal is long-lived ILW (sometimes termed TRU), a category of waste that is extremely varied (see Hooper et al., 2005 for details). Immobilisation is mostly by encapsulation in a cement grout. Lesser quantities of ILW are encapsulated in organic resins and some older wastes were placed in bitumen, a practice that has now largely ceased (see comments in Chapter 2). The encapsulated wastes can be held in containers made from mild steel or stainless steel that are a few millimetres thick. Many of these wastes are metallic and will generate small quantities of gas as they corrode. For this reason, the container may be designed with a vent to prevent pressure build-up. Absolute physical containment is less likely to be an important feature of post-closure SA than it is for HLW and SF and here containment is achieved through chemical containment, a concept that aims to retard the migration of radionuclides by providing conditions in which radionuclide mobility is reduced through low solubility and strong sorption (Atkinson et al., 1986). Recently, 20 dm3 scale ‘‘equilibrium leach tests’’ using real radioactive wastes have demonstrated the effectiveness of this concept (Angus and Tyson, 2002). These experiments, carried out in a shielded facility (Fig. 8.3) entailed remote sampling and analysis of water and gas chemistry and of dissolved/particulate radionuclides.

8.2.3.3. Colloids

Some circumstances can occur that may perturb the solubility/sorption-limited near-field model and one of these is the formation of colloids. Although the formation, stability and transport of colloidal material is well-known in the paint and pharmaceutical industries, this is much less so in radwaste and they have thus been the subject of much R&D. Two main possibilities emerge: First, the oxides or hydroxides of the actinides may themselves be present in a colloidal state and this possibility is routinely assessed when measuring actinide solubility values (e.g., Baston et al., 2003; Geckeis et al., 2004). Although this could actually reduce actinide releases by filtering the colloidal form in the bentonite backfill, this is usually ignored in SA and any uncertainties are covered by assuming conservative values for actinide solubilities.

The second possibility is that radionuclide ions may sorb onto colloidal material in the near-field. This colloidal material could be derived from the degradation of near-field engineering materials such as steel, cement or bentonite (see, e.g., Pusch, 1999; Swanton and Myatt, 2003). For HLW/SF repository designs which incorporate a bentonite buffer (see Chapter 5 for details), it is, as noted above, considered unnecessary to incorporate colloids into the modelling (e.g., Nagra, 1994, 2002a; SKB, 1999). For some cementitious repository designs (e.g., Nirex, 1997a), it was found that the effect of colloids was small when the calculations included the perturbations caused by organic degradation products (q.v.), but this is not the case for all designs (e.g., those which do not include much organic waste), so a significant R&D effort is currently ongoing in the field of cementitious colloids (see, e.g., Wieland et al., 2004).

8.2.3.4. Organic degradation products

A second important factor that can perturb the solubility/sorption-limitation model is the degradation of organic – especially cellulosic – wastes that may form part of the