The situations dealt with in this chapter are interventions for areas that have been contaminated as a result of human activities and that could cause prolonged radiation exposure. In this context, the term ‘areas’ is used in its broadest sense and can include land, forests, urban environments and indus­trial sites. These areas may have been contaminated as a result of inade­quate practices for radioactive waste management and disposal, radioactive discharges to the environment that did not meet regulatory requirements, nuclear accidents, atomic weapon tests, incidental releases of radionuclides by users of radioactive material or past practices that were not adequately controlled. This chapter also applies to radioactive discharges from facilities that were managed in accordance with less stringent requirements than those that are applied today (IAEA, 2006a). However, this chapter is not specifically intended for the management of huge amounts of uranium/ thorium mill tailings or naturally occurring radioactive material (NORM), which have their own specific circumstances and management options.

Some examples of contamination that might be encountered are given below. The list is not exhaustive but is intended to show the wide range of problems that might be found. Finally, although this chapter is intended for radioactively contaminated sites, the need for environmental remediation also includes non-radioactive, toxic contaminants that may be associated.

RAW management requires a systemic approach: all the stages of RAW management are the components of one overall system, and all the techniques and procedures adopted in each separate stage are connected and have a common goal of radiation safety.

A comprehensive analysis of the status and existing problems in the field of RAW management in Ukraine was carried out in 2006-2007 within the framework of the TACIS project — U4.03/04: Development of the National Strategy and Concept for State Programme for Radioactive Waste

Management in Ukraine, including a Strategy for National Company Ener — goatom Radioactive Waste Management. The results of the project imple­mentation were published in Shestopalov et al. (2008) and the following conclusions regarding radioactive waste management in Ukraine were made (as of 2008).

Ukraine has accumulated significant amounts of RAW. The volume of waste and the rate of its accumulation will continue to grow in the future due to the extension of the operating period and decommissioning of the existing units and commissioning of new NPP units. The contribution of NPP to the current accumulation of RAW in Ukraine is about 95%. In general, wastes have not been sorted or reprocessed taking into account the need for further conditioning and disposal. Separation of waste into short­lived and long-lived is not carried out at NPPs.

Unprocessed RAW from the non-nuclear sector continues to accumulate without being buried. Safety of the already buried waste has not been con­firmed. The issues of storage and disposal of vitrified HLW as a result of reprocessing in the Russian Federation of Ukrainian NPP SF and long-lived waste of Chernobyl origin have not been addressed. The existing system of RAW management is not focused on the final disposal of all types and categories of RAW. The organization responsible for developing and imple­menting the technical policy in the field of RAW disposal has not been identified.

Stable funding of the design, construction and operation of infrastructure facilities for RAW management has not been ensured, and a special state fund for RAW management has not been established. The existing classifi­cation of RAW in Ukraine ensures the safety of RAW management at the stages of their collection and storage, but is economically inefficient in terms of achieving its ultimate objective the safe disposal of RAW. This is particu­larly true regarding the problems of disposal of the large amounts of waste of Chernobyl origin, which contain significant amounts of long-lived radionuclides.

The amount of funding for RAW management provides the minimum acceptable level of safety. State investment in RAW management infrastructure upgrades is virtually zero and a reassessment of the safety of storage facilities has not been made. Production of containers for the storage of RAW has only commenced in recent years. However, trans­portation of waste has not yet been provided by licensed transport containers.

Most of the problems of RAW management in Ukraine arose due to the lack of a single state policy that would allow a systematic solution to the problems of RAW management (including disposal) and the problems of stable funding.

Radioactive waste is material that contains, or is contaminated with radio­nuclides at concentrations or activities greater than the clearance levels set by the regulators, and for which no use is foreseen. The hazard associated with radioactive wastes depends on the concentration and nature of the radionuclides with those emitting higher energy radiation or being more toxic to life, being the most hazardous.

Radiotoxicity is the harmful effect of chemical substances as a result of their containing radioactive elements. The effect of ionising radiation emitted by the elements leads to changes in the metabolism and structure of living organisms. It is a measure of how harmful a radionuclide is to health. The type and energy of rays, absorption in the organism, residence time in the body, etc., all influence the degree of radiotoxicity of a radionuclide.

Alpha particles (He atoms) are very strongly ionising, so if they come into contact with atoms in a living tissue they can cause mutations, unusual chemical reactions in the cell and possibly cancer. Although the most ionis­ing, it is the least dangerous form of radiation as long as it is not ingested or inhaled, because it is stopped by, for example, a sheet of paper or skin so that it cannot penetrate into your body. Alpha radiation is most com­monly used in smoke detectors generated by americium.

Beta radiation is made up of an electron with high energy and speed. Beta radiation is more hazardous because it can also cause ionisation of living cells. Although it is less ionising than alpha radiation, it has the capa­bility to pass through living cells and can be stopped by an aluminium sheet. If beta radiation hits a molecule of DNA it may cause spontaneous muta­tion and cancer. It is used industrially in thickness measurement such as in paper mills and aluminium foil production.

Gamma rays are high frequency, very short wavelength electromagnetic waves with no mass and no charge. They are emitted by a decaying nucleus so that it can release energy allowing it to become more stabilised as an atom. Gamma rays have the highest penetrating power, only being stopped by a few centimetres of lead or a few metres of concrete. They are the least ionising of the radiations but this does not mean that they are not danger­ous. Gamma rays are likely to be emitted alongside alpha and beta radiation but some isotopes only emit gamma radiation. Gamma rays are useful because they can kill living cells and so be used to sterilise by, for example, destroying harmful bacteria. Gamma rays are also used in radiotherapy to kill off cancerous cells. They are also used to sterilise medical equipment, which is particularly useful in tools that would be melted by heat sterilisa­tion or compromised by bleaches and other disinfectants.

Radioactive waste is accompanied by significant levels of radiation, hence it requires not only immobilisation to prevent radionuclides spreading around the biosphere, but also shielding and, in some cases, remote handling. A waste with activity concentrations equal to, or less than, clear­ance levels is considered non-radioactive. Radioactive wastes are either controlled or uncontrolled.

Controlled wastes are largely a product of the nuclear fuel cycle (NFC) used to generate electricity for civil use (Fig. 1.1). Wastes are generated during ore mining and processing to access the uranium metal or oxide, its enrichment and synthesis into fuel (the front end of the NFC), the operation and running of the reactor (operations wastes) and from fuel removal, treat­ment and disposal (the back end of the fuel cycle). Front end waste is contaminated basically with naturally occurring radionuclides, whereas operational waste also contains fission and activated products (typically low level waste (LLW) and to a lesser extent intermediate level waste (ILW); these are defined in Section 1.3). Front end wastes include contaminated mining wastes and uranium hexafluoride tails from enrichment. Opera­tional wastes include spent filters and ion exchange resins, evaporator

concentrates and absorber rods. Back end wastes include sludges from storage ponds, typically cemented ILW and vitrified HLW from reprocess­ing or spent fuel if direct disposal is planned.

During the early part of the nuclear era, consideration was not given to disposal of radioactive waste. As a result some NFC wastes (now termed legacy or historic wastes) are ill-characterised and stored under conditions which are far from ideal. They comprise a vast range of materials, e. g. Pu — contaminated materials (PCM) such as paper, wood and plastics, fuel clad­ding, damaged and corroded fuel elements, old tools and equipment and assorted test samples often mixed together. Sometimes these have been stored under water and have degraded over time to form complex sludges and supernatant liquids.

Controlled non-NFC wastes include those from various applications of radionuclides in research, medicine and industry including spent sealed radioactive sources (SRS) of isotopes used in medical applications.

Uncontrolled wastes arise when unexpected events occur or where the level of care was not taken that would be expected today. At Hanford, one of two sites where the US stores its military (defense) wastes, poorly char­acterised highly active sludges were stored in massive single shell steel tanks

that eventually leaked (Fig. 1.2a). At Sellafield in the UK, some materials were stored in inappropriate open ponds (Fig. 1.2b) where ingress of atmos­pheric (salty) rain and organic matter (bird droppings, etc.) has added to the complexity of the problem. Uncontrolled wastes also arise from acci­dents such as at Chernobyl, Ukraine (Fig. 1.2c) and Fukushima, Japan. The financial cost of cleaning up such sites and others where accidental releases of radioactivity have occurred, such as in Fukushima, is enormous. Nonethe­less, the nuclear industry is now developing smart clean-up programmes and concepts (Fig. 1.3) and the knowledge gained from these mistakes has helped us be more proactive in dealing with uncontrolled waste.

Nonetheless, armed with sufficient resources, the results of decades of inten­sive research and international support progress can be made (Fig. 1.3).

Intermediate level waste is the waste that, because of its higher radioactivity concentration and/or higher concentration of long-lived radionuclides, does not fit into the previous LLW category. ILW may contain long-lived radio­nuclides, in particular, long-lived fission products and alpha emitting radio­nuclides that will not decay to a level of activity concentration acceptable for near-surface disposal during the time for which institutional controls can be relied upon. However, ILW needs no provision, or only limited provi­sion, for heat dissipation during its handling, storage and disposal. The activity concentration of bulk radionuclides, as well as minor long-lived radionuclides used to distinguish between LLW and ILW are not universally agreed upon. Moreover, the upper activity concentration limits for ILW are not universally agreed upon. These limits and concentrations are site — specific and they shall be established in each individual case by the regulator, based on a safety analysis of the disposal option being considered. Some guidelines about the limiting values for long-lived radionuclide activity concentrations can be found in Table 2.1.

ILW is typically generated at NPP as a result of treatment (concentra­tion) of primary waste. Another significant source is reprocessing of spent fuel. ILW requires a greater degree of containment and isolation than LLW and disposal in subsurface repositories at depths of the order of tens of metres to a few hundred metres.

3.1.2 Joint Convention on the Safety of Spent Fuel Management and on the Safety of Radioactive Waste Management and other conventions

The safety of RAW is a universal concern, as illustrated by the fact that an international treaty was established in 1997 committed to achieving and maintaining a high level of safety worldwide in the management of RAW and spent nuclear fuel. This is the first and only legally binding international convention on these materials and is based on the IAEA Fundamental

Safety Principles. In 2011 the convention was supported by 58 contracting parties [49].

Implementation of any industrial operation to recycle one or more compo­nents from irradiated fuel will result in the generation of a number of waste streams that must be properly managed to mitigate environmental conse­quences. The primary wastes from aqueous reprocessing of spent fuel include cladding hulls and hardware, undissolved solids, gaseous fission products, and HLW raffinate. A host of job control, maintenance, and opera­tional secondary wastes (typically low or intermediate level) are also gener­ated. In this section, the nature of these waste streams and the methods in which they can be immobilized for disposal are summarized.

The most important requirement for a waste form is its chemical durability, expressed as a dissolution rate. It should be noted that for some radionu­clides, solubility limits the dissolution rate while others are completely soluble, e. g. "Tc, 129I, or 135Cs. These soluble radionuclides are released at the maximum forward rate of dissolution. For the production of durable nuclear waste forms, it is desirable for the waste forms to be highly insoluble in the long term to minimize release to the environment, i. e. to have the slowest forward dissolution rate possible. Since no ‘durability test’ can be carried out on these geologic timescales, dual approaches are taken:

1. Durability test parameters such as surface area (SA), time (t), tempera­ture (T), or a combination such as (SA)^(t) are used to ‘accelerate’ dissolution as long as the acceleration parameter(s) used does not change the dissolution mechanism. To ensure that the mechanism is not ‘altered’ by the acceleration modes of the experiments, natural analogs are usually tested simultaneously.

2. Models are used to predict waste form dissolution from parameters that can be measured such as the activation energy of dissolution, forward rate of dissolution, and from an understanding of the dissolution mecha­nisms. Predictive and/or transport models for waste form performance on extended time scales (1,000-1,000,000 years) has led to various ther­modynamic and kinetic models (see [160], [187]).

Thus, there are no ‘waste form-specific’ durability tests, but a suite of tests that must be performed to understand the leaching mechanism(s) of a waste form and to derive the parameters necessary for the particular predictive or transport model(s) being applied.

In order to determine if a particular waste form is acceptable, it must be demonstrated that the waste form performance in the disposal system is adequate. Such evaluations in the US are known as total system perform­ance assessments (TSPA) for HLW and performance assessments (PA) for shallow land disposal of immobilized LAW known as ILAW in the inte­grated disposal facility (IDF). The TSPA or PA includes all of the testing and performance modeling that has been gathered on the waste form and the TSPA is intended to provide a technical basis that a waste form is acceptable for deep geological disposal.

For HLW in many countries the geological disposal sites have not been determined while wasteform producers have already made many canisters of vitrified waste (see Table 6.1). Due to the mismatch in timing between the need to stabilize HLW and when a geological repository will be chosen and ready to receive the wasteforms, the US devised a strategy to addresses vitrified waste acceptance based on production control. Production control is intended to determine how the production of a waste form material affects (or controls) its performance and identify the ranges for processing variables that result in an acceptable waste form. The primary role of most of the waste acceptance product specifications (WAPS) developed in the US for vitrified HLW waste forms verify that the properties of a specific waste form product are consistent with the existing regulations and thus will be acceptable for disposal, either by direct measurement or through process control.

Therefore, waste acceptance testing is, for the most part, focused on comparing a specific waste form product to the range of waste forms that are (1) considered to have acceptable performance based on performance modeling and (2) produced within the production control limits. What will be acceptable with respect to waste form performance and processability will depend on the disposal site and engineered system and cannot be com­pletely quantified at the time the waste form is made. The range of accept­able waste form compositions will depend on the required performance [167].

While the predicted long-term durability of a waste form is a necessity for its ‘qualification for shallow land burial’ or ‘deep geologic disposal’, there is also a need for short-term testing that can be related to acceptable performance by the following linking relationships [168]:

process control ^ composition control ^ dissolution rate control ^ performance control ^ acceptable performance.

This approach allows a waste form producer to ensure that the waste form that they are producing on a tonnage per year basis will be acceptable to long-term performance instead of having to test each and every canister or form produced. For HLW glass (alkali borosilicate glass) in the US, the manner in which this was done is given below in a brief stepwise fashion and explained in more detail in Refs [11, 169-173]:

1. Develop an acceptable waste form durability based on HLW perform­ance modeling (fractional dissolution rates between 10~4 to 10“6 parts per year (i. e., the glass waste form would take 10,000 to 1,000,000 years to totally dissolve [174]).

2. The middle of the range determined by HLW performance modeling was adopted as the waste form specification; if the long-term fractional dissolution rate of a wasteform was <10-5 parts per year for the most soluble and long-lived radionuclides, then borosilicate glass would provide acceptable performance for any repository site or concept.

3. Develop an understanding of the glass durability mechanisms from a combination of the test protocols (ASTM C1220 which was previously known as MCC-1, ASTM C1285 which is known as the Product Consist­ency Test (PCT) [175, 176] , ASTM C1662 which is the SPFT test, and ASTM C1663 which is the Vapor Hydration Test or VHT).

4. Develop a glass standard, the Environmental Assessment (EA) glass [177, 178] that bounded the upper release rate found to be acceptable from the HLW repository modeling from step 1 above.

5. Generate databases for modeling the maximum radioactive release rate(s) by relating the release of "Tc, 129I, and 135Cs to the release of non-radioactive species such as Na, Li, and B which leach at the same rate (congruently); this is part of the ASTM C1285 (PCT) test protocol.

6. Develop a short-term test and process control strategy for ensuring that every glass produced has a dissolution rate less than that of the EA glass at the l95% confidence level based on Na, Li, B which in turn ensures acceptable performance control.

7. Continue to qualify that the radionuclide response of production glasses verify that production glass radionuclide releases are consistent with the releases predicted by Na, Li, and B.

Therefore, a suite of the existing durability tests (those for affinity control, solubility control, and/or diffusion control) must be performed on a waste form to determine the mechanisms, and determine the parameters neces­sary for the mechanistic model(s) being developed, e. g. the transition state theory (TST) models used in the TSPA for HLW geological disposal or the PAs for shallow land burial. Different durability tests are used for a diffu­sion model, for example for cement. However, one cannot apply a glass standard that leaches by an affinity limited mechanism to cement that leaches by diffusion, nor can one apply a borosilicate glass standard to non — borosilicate-type glasses since it is not known whether the radionuclides in non-borosilicate glasses leach by the same degradation mechanism and whether the leaching of Na, Li, and B remain congruent with the leaching of the radionuclides. In these cases, new standards need to be developed and qualified and the leaching mechanisms understood.

For glasses, the advances in the measurement of medium range order (MRO) in glass waste forms has led to the understanding that the molecular structure and composition of a glass, like the molecular structure and com­position of minerals, controls the waste form durability by establishing the distribution of ion exchange sites, hydrolysis sites, and the access of water to those sites. During the early stages of glass dissolution, a ‘gel’ layer resembling a membrane forms through which ions exchange between the glass and the leachant (Fig. 6.7). The hydrated gel layer exhibits acid/base properties which are manifested as the pH dependence of the thickness and nature of the gel layer. Advances in the understanding of the dissolution mechanisms of borosilicate glasses proposed for nuclear waste solidification were extensively studied in the 1980s-1990s [22, 179-186] and such mecha­nisms are still being studied [160, 187-190]. At least four operative mecha­nisms have been shown to control the overall glass durability as shown in Fig. 6.7. These four mechanisms are ion exchange, matrix dissolution, accel­erated matrix dissolution, and surface layer formation (possibly of a protec­tive or passivating nature).

One can bound or model the shorter term durability of a glass using kinetic or thermodynamic models to describe the impacts of ion exchange and matrix dissolution or hydrolysis by examining either time-temperature data (Fig. 6.8) or release vs time or accelerated release, expressed as SA/VHime (Fig. 6.9), but these underlying mechanisms become modified if surface layers form and/or if, over very long periods of time, the gel layer ages in situ into clay or zeolite minerals or the leachate becomes satu­rated with respect to a clay or zeolite phase. If zeolite mineral assemblages (higher pH and Al3+ rich glasses) form, the dissolution rate increases (Fig. 6.9) which is undesirable for long-term performance of glass in the environment.

The current theories of glass dissolution [ 159] suggest that all glasses typically undergo an initial rapid rate of dissolution denoted as the ‘forward rate’ (Figs 6.8 and 6.9). However, as the contact time between the glass and the leachant lengthens, some glasses come to ‘steady state’ equilibrium and corrode at a ‘steady state’ rate, while other glasses undergo a disequilibrium reaction with the leachant solution that causes a sudden change in the solu­tion pH or the silica activity in solution [191] [ The ‘return to the forward rate’ (Fig. 6.9) after achieving ‘steady state’ dissolution is undesirable as it

6.7 ( a) Schematic diagram of glass dissolution mechanisms (ion exchange and matrix dissolution) in aqueous solution, coupled with both hydrated amorphous surface layer formation and crystallization/ precipitation from solution [179, 402]. (b) Schematic diagram of the glass dissolution mechanism known as ‘accelerated matrix dissolution.’ In this mechanism, the excess strong base in the leachate released by the ion exchange mechanisms attacks the glass surface layers, including the gel layer, and makes the glass appear to have little or no surface layer.

can cause a glass to return to the rapid dissolution characteristic of initial dissolution.

The initial rate is often referred to as Stage I dissolution in the US litera­ture, but it encompasses zones where multiple mechanisms are operative including regimes that are interdiffusion controlled, hydrolysis controlled, and a rate drop that is diffusion or affinity controlled [159] . The ‘steady

state’ rate (also known as the residual or final rate) that signals the end of the alteration phase and/or a pseudo-equilibrium between the alteration and re-condensation reactions [159, 192] is known as Stage II dissolution, and the return to a forward rate (or resumption of alteration) is known as Stage III dissolution. Diffusion controlled dissolution of network modifiers and/or radionuclides during Stage I and Stage II normally follow a math­ematical function related to the square root of the test duration as observed in many burial studies [190], while other radionuclides are solubility limited, entrapped in the gel layer, or complexed in secondary alteration phases that form from the leachate solution.

A reaction zone is formed as the leached layer solution interface progresses into the glass (Fig. 6.7 a). The front of the reaction zone repre­sents the region where the glass surface sites interact with the ions in solu­tion [193]. The top of the gel reaction zone represents the leached layer-glass interface where a counter-ion exchange occurs [193] . The glass dissolution rate is modified by the formation of the hydrated amorphous gel layers and/ or secondary precipitates, e. g., metal hydroxo and/or metal silicate com­plexes that have reached saturation in the leachate and can precipitate on the surface of the gel layer [22, 179, 181, 182, 194, 195] . These ‘back reac­tions’ have been attributed to formation of silanol bonds as surface

adsorption sites which were modified by changes in solubility of the species in solution and surface (zeta potential) considerations [22, 196].

The gel layer may, under certain conditions, act as a selective membrane [194, 197] or as a protective/passivating layer [22, 159, 180-182, 184-186, 192, 198]. The slowing of glass dissolution to a steady state rate by solution saturation (affinity) of glass matrix elements or reaction through a surface layer has been referred to as Stage II dissolution including residual rate dissolution, steady state dissolution, or the final dissolution rate. Recent mechanistic modeling of glass durability, including the slowing of the dis­solution rate due to affinity and/or surface layer effects, was first modeled by Grambow and Muller [199] and is referred to as the GM2001 model. The GM2001 model combines the effect of glass hydration by water diffu­sion with ion exchange and affinity-controlled glass network corrosion (Figs 6.8 and 6.9). The slowing of dissolution due to the effect of a growing surface gel layer is represented by a mass transfer resistance for silica by this layer. At the interface between the glass and the gel layer, a different ‘gel layer’ is assumed to be hydrated glass that allows diffusion of H2O in and boron and alkali atoms out of the glass (similar to Fig. 6.7). A 2003 modification of the GM2001 model, known as the GM2003 model [159], treats silica dis­solution and silica diffusion through the gel separately from water diffusion, and boundary conditions are specified at the gel/diffusion layer and the gel/ solution interfaces. Recently, the GRAAL (glass reactivity with allowance for the alteration layer) model [187, 189] has been proposed, which is dependent on the composition and the passivating nature of the gel layer, called the passivating reactive interphase (PRI). The leached layer has been found experimentally to be zoned (5-7 zones) and the GRAAL model assigns various mechanisms to different zones within the PRI.

The resumption of alteration (Stage III) causes the long-term dissolution rate to reaccelerate to a rate that is similar to the initial forward dissolution rate for some glasses. This unexpected and poorly understood return to the forward dissolution rate has been shown to be related to the formation of the Al3+-rich zeolite, analcime, and/or other calcium silicate phases. Moreo­ver, the presence of Al3+ and Fe3+ in the HLW glass, in the leached layer, and in the leachant has been shown to influence whether a glass maintains Stage II dissolution or reverts to the forward rate of dissolution, e. g., Stage III dissolution. Van Iseghem and Grambow 3 191] demonstrated that an Al3+-rich zeolite (analcime) formed on certain glasses during dissolution but not on others. Van Iseghem and Grambow also demonstrated that a change in solution pH accompanied the return to the apparent forward rate when analcime formed. Likewise, Inagaki et al. [200] demonstrated that solution pH and solution concentrations of Na and K were also involved in the formation of undesirable analcime versus Na-bedellite (a smectite clay). Other zeolites and smectite clays that are rich in Fe3+ compared to Al3+ do not appear to accelerate glass corrosion [191, 201, 202].

Since many long-term durability models are still being refined and an international study group [203] is actively working on a refined understand­ing of the PRI, a variety of leaching tests are being used to facilitate an integrated understanding of these stages of durability.

Over the past decade, a number of remediation techniques have been developed worldwide to deal with the environmental clean-up of radioac­tively contaminated sites. These techniques vary in terms of sophistication and costs and must be selected on a case-by-case basis. However, the devel­opment of a successful remediation programme does not only rely on the availability of technology and expertise. Good management plans are also needed as well as appropriate communication with the various stakehold­ers, as pointed out in previous sections.

One important factor determining the selection of remediation technology(ies) is the area radiation levels. It is clear that excavations under low contamination/radiation conditions can be performed ‘hands on’, whereas at Fukushima many of the contaminated streams cannot be handled this way due to excess radioactivity. In general in-situ treatments (see below) entail less exposure to the workers and would be preferable in case of high contamination/radiation.

ER may face specific challenges not only because of the lack of resources but also because of the lack of appropriate technology, or the lack of experi­ence in using new or imported technologies. These aspects altogether can end up constituting important barriers for project implementation. However, experience has shown that with appropriate planning and assistance, reme­dial actions are more likely to be implemented. As such the interaction of more experienced countries in ER with less experienced ones facilitated by international organizations and other donors may lead to better conditions for full implementation of projects.

This section presents particulars on ER technologies (control and treat­ment). The technologies addressed can be categorized as follows:

• in-situ treatment,

• removal of contamination; and

• ex-situ treatment.

Details on these technologies can be found in IAEA (1999b) and Hamby (2012).

1. In 2002, a project was initiated at the Kurchatov Institute with the aim of remediating contaminated objects and sites [34]. The principal tasks of this project were the removal of the accumulated RAW, the decom­missioning of old repositories and the remediation of the contaminated territory. The volumes involved were evaluated at 4,000 m3 of SRAW,

20.0 m3 of contaminated soils, and RAW activity of up to 700 Ci. In 2006, the decontamination of 10 old repositories was completed. More than 3,400 m3 of SRAW with activity >500 Ci was removed: of this, over

3. m3 was removed for long-term storage at MosNPO ‘Radon’, and 300 tons of metal low-level waste were sent for further melting at the ‘Ekomet-S’ facility.

2. A project to examine and decontaminate the buildings of Moscow’s ‘Zavod Polimetallov’ plant was carried out between 1999 and 2003. There were 32 buildings and more than 9,000 m2 of land. Decontamina­tion of 17 buildings revealed an RAW volume of over 400 m3, with radionuclides composed of 232Th and 226Ra.

3. The radiochemical laboratory of the Vernadsky Institute of Geochem­istry and Analytical Chemistry (Russian Academy of Sciences) was in use from 1966 until the end of the 1980s. The contaminated objects at this site were seven hot cells with an operating area of 240 m’ ’ as well as 15 auxiliary premises, and a ventilation system. After a partial decon­tamination and dismantling of the equipment, it was decided to preserve the premises for 50 years.

4. The JSC ‘Koltsugtsvetmet’ (Vladimir Region) contained a workshop for the production of luminescent substances based on soluble 2 26Ra bro­mides. This facility was built and put into operation in the mid-1950s. The two-storey building with a total area of 1,200 m’ was decontami­nated; the building and ventilation system were then dismantled and the surrounding territory remediated.