3.3.1 Precipitation Processes

The earliest separations were developed in support of the WW2 Manhattan Project and exploited the different solubilities of Pu(iv) and U(vi). The primary centre for these activities was the Hanford site in Washington State, USA.7 Most fission products are soluble in strongly acid conditions, whereas the fluorides and phosphates of the tri — and tetravalent actinides are not.

Clint A. Sharrad, Laurence M. Harwood and Francis R. Livens — Total for fuel

— Fission and activation products

Thus, these separations were based on dissolution of fuel in nitric acid, with control of the uranium oxidation state to vi and of the plutonium oxidation state to III or IV. On addition of Bi, followed by phosphoric acid, BiPO4 was precipitated, carrying plutonium and separating it from uranium and many fission products. The crude precipitate was then dissolved in strongly oxidising conditions, for example with BiO32 or MnO4 in the presence of Cr2O72 holding oxidant, which switched the plutonium to oxidation state VI. Further purification was achieved by lanthanum fluoride precipitation, which elimi­nated lanthanide fission products and isolated plutonium.

These processes were carried out on an industrial scale from the end of 1944 until the early 1950s and were used to manufacture tonne quantities of pluto­nium from hundreds of tonnes of uranium (the irradiated fuel contained about 250 mg Pu tonne 1U). In the early 1950s, precipitation separation was replaced by solvent extraction, leading to the development of the Purex process and a substantial increase in throughput. By the early 1960s, Hanford was separating between 1.5 and 2 tonnes of plutonium from 7000 tonnes of irradiated uranium each year, 80% of this from the Purex process.

Sediment washing is a relatively simple, typically ex situ technique involving the cleaning of contaminated soils with various reagents. Depending on the nature of the contaminant, a number of additives can be employed in the washing process including acid washing (e. g. H2SO4 and HNO3) and chela­ting agents [e. g. EDTA, diethylenetriaminepentaacetic acid (DTPA) and ethylenediamine-N, N’-disuccinic acid (EDDS)] to assist with the solubilisation and desorption of the metal from the sediment. This technique is useful for weaker bound metals, those associated with the exchangeable, carbonate and reducible oxide fractions of the soil, but is inefficient at removing metals in the residual fraction.93 Acid washing can be applied through a variety of abiotic and biological ex situ techniques.70,94 Chelating agents can be used in soil washing to remove contaminants from sediments through the formation of stable metal chelate complexes which can then be removed in solution.95 EDTA has been studied extensively for use as a chelating agent for use in soil

washing96,97 and can enhance metal mobilisation via two mechanisms: fast thermodynamically favourable complexation between EDTA and certain cationic metals, as well as slow driven EDTA-dissolution. The former involves the breakdown of some weak soil-metal bonds while the latter can partially disrupt the soil structure thus mobilising metals bound to oxides and organics.98

Radionuclides may enter the atmosphere as gas, aerosol or particulate matter. The transport of suspended radionuclides is dependent on particle size; larger particles will settle and deposit faster than smaller particles. Following the release into the atmosphere, the dispersion of radionuclides is mainly controlled by meteorological conditions (i. e. winds, turbulence, advection and wet and dry precipitation), radioactive decay and diffusion.

For atmospheric nuclear weapons tests, transport of the radioactive debris will depend on the height and yield of the explosion, the nature of the debris, the location of the test site and prevailing meteorological conditions. Refractory radionuclides, such as plutonium, 95Zr and 144Ce, are released mainly in parti­culate form,5,48 and so will tend to be deposited more rapidly, and be less widely dispersed, than more volatile radionuclides, such as 137Cs and 131I.49 During testing, radioactive debris will be injected into the atmosphere at different heights, and this will depend primarily on the height of the test and the explosive yield; low yield tests will tend to release debris into the troposphere, with the quantity of radioactive material released into the stratosphere increasing with yield.50 For tests conducted near the surface, it is estimated that around 50% of the debris is deposited locally or regionally, with the remainder more widely dispersed.5,49 Debris released into the troposphere (the lowest level of the atmosphere) can be transported up to several thousand kilometres from the test site over 1-2 weeks, as a result of the turbulent air movements that occur there.49,50 Removal of particulate debris from the troposphere is mainly caused by precipitation but dry deposition of radionuclides can also occur.50 Radioactive debris released into the stratosphere remains in the atmosphere for much longer periods of time (> 1 year) than material released at lower altitudes, and so will be dispersed over a much greater area, with precipitation the main mechanism for deposition.6,10 As a result, global radioactive contamination arising from deposition of material from the stratosphere will consist of longer-lived radionuclides, compared to local and regional contamination.6,49 Simon et al. (2004)6 investigated the geo­graphical distribution in the USA of radionuclide fallout arising from tests at the Nevada Test site (NTS) and global fallout. The distribution of radioactive debris from the low yield tests at the NTS depended on the wind patterns and local rainfall events at the time of the test, but in general the highest levels of deposition were in the region immediately east of the site. With global fallout, higher levels were deposited in the eastern and mid-western regions than the south-western states, reflecting the relative levels of precipitation in these regions.

More localised atmospheric transport of radionuclides occurs with the use of DU weapons and uranium mining and milling. When a DU munitions hits its target, an estimated 10%-35% (maximum of 70%) of the DU mass is converted into aerosol, with most of the dust particles < 5 pm.26 The transport of DU particles will depend on particle properties (i. e., size and density) and on prevailing meteorological conditions.51 Surveys of the post-conflict zone in Kosovo and Bosnia and Herzegovina reported DU contamination up to 200 m away from the point of impact.52,53 Lloyd et al. (2009)54,55 investigated the dispersion of aerosols formed during the combustion of waste metal at a ura­nium and DU processing factory in Colonie (NY, USA). The distribution of the DU aerosol was controlled by prevailing winds, with DU contamination found up to 600 m from the factory. It has been estimated that at least 3.4 tonnes of uranium was deposited within 1 km of the factory.56 Resuspension of DU dust has also been found to occur by wind or human disturbance.

From uranium mining and milling, radon gas will be released into the atmosphere, but in arid climates, windborne dispersion of fine radioactive particulate wastes can also be a problem.16,57 Lottermoser and Ashley (2006)58 investigated the physical dispersion of radioactive waste from a rehabilitated uranium mine in South Australia. Under the semiarid conditions at this site, there had been significant wind dispersion of radioactive particulates from the site. Around the main tailings storage facility, tailings material up to 10 cm thick was spread up to 80 m from the source in the northeast and southeast sides, reflecting the prevailing wind directions at the site. Around this source an area of 1km2 had uranium concentrations >100 ppm, with another 2 km2 con­taminated with 10-100 ppm of uranium. Radon, generated in the subsurface or in waste materials, is mainly released into the atmosphere by diffusion, but advection caused by wind and changes in barometric pressure can also play a role and mining activities will enhance rates of release into the atmosphere.24,59

Radioactive materials released into the atmosphere from the accident at the Fukushima Nuclear plant were detected globally but at very low levels. Mon­itoring undertaken by the Preparatory Commission for the Comprehensive Nuclear-Test-Ban Treaty Organisation (http://www. ctbto. org/press-centre/ highlights/2011/fukushima-related-measurements-by-the-ctbto/fukushima-related- measurements-by-the-ctbto-page-1/) detected traces in eastern Russia on 14th March 2011, three days after the earthquake and tsunami that damaged the reactors. Radiation was detected on the west coast of USA by 16th March and all across the northern hemisphere 15 days after the accident. The equator acts as a dividing line between the northern and southern air masses, and so the dispersal of radioactive materials was initially limited to the northern hemisphere; however, by 13th April, radiation from Fukushima had spread to the southern hemisphere.

So how is all of this extensive advice and guidance supposed to be implemented and actually applied to different radiation exposure situations? It goes without saying that the detailed interpretation can be complex indeed, but by and large it all follows much the same pattern, which can be summarized as follows:

(i) A characterisation of the possible situations where radiation exposure may occur (planned, emergency and existing exposure situations).

(ii) A classification of the types of exposure (those that are certain to occur and potential exposures, as well as occupational exposure, medical exposure of patients and public exposure).

(iii) An identification of the exposed individuals (workers, patients and members of the public).

(iv) A categorisation of the types of assessment, namely source-related and individual-related.

(v) A precise formulation of the principles of protection: justification, optimisation of protection, and the application of dose limits.

(vi) A description of the levels of individual doses that require protective action or assessment (dose limits, dose constraints and reference levels).

(vii) A delineation of the conditions for the safety of radiation sources, including their security and the requirements for emergency pre­paredness and response.

Radiological protection is undertaken within the UK under various pieces of legislation, including a number of regulations, the developments of which have been recorded in some detail by O’Riordan.19

By the time of the oil crisis in 1973, seventeen countries had 167 power reactors with a capacity of nearly 61 000 MWe.47 The nuclear industry thought the four­fold increase in oil would make nuclear more economic than coal. For a time this was indeed the case. However, the worsening inflationary environment led inexorably to a steady raising of interest rates making capital-intensive projects like nuclear stations very costly. Moreover, the deep global recession that began to bite as the 1970s drew to a close led to a steep fall in electricity demand. This negatively impacted on some countries’ nuclear ambitions, with some countries like France willing and able to bear the increased costs more than others. Although there were substantial orders placed in the US for new nuclear sta­tions during 1973 and 1974, many were never completed.48 Indeed, no new stations were ordered in America after 1978, with many plants cancelled when 90% complete.49 This was a rapid reversal in fortunes from the golden age, announced only in 1973, when the US AEC predicted that by the year 2000 there would be 1000 operational nuclear reactors in the US alone.

Similarly in the UK the Central Electricity Generating Board (CEGB), partly in response to the oil crisis, drew up plans during 1973 to build 32 PWRs (only one — Sizewell B — was ever built) and during 1974 The Department of Energy ordered six Steam Generating Heavy Water Reactors, a programme which was mothballed in 1978 as recession deepened with £145 million having already been sunk into the project.50

During the late 1970s protests against the construction of nuclear plants steadily began attracting increasing numbers of people particularly in continental Europe — from the 60 000 who demonstrated against the construction of a fast breeder reactor in Malville, France, in July 1977 where one protester was killed, to the large scale protests across West Germany in which for the first time the police began deploying water cannons and tear gas to disperse protesters. These gathering protests were labelled by one writer at the time as the ‘‘anti nuke explosion’’.51

In the late 1970s, several countries also decided to abandon the more ambitious parts of their nuclear programmes, particularly fast breeder and reprocessing technology. This development was led by the United States when President Carter issued a decree in 1977 halting reprocessing, formally moti­vated by nuclear proliferation concerns but also responding to the enormous technical difficulties and financial costs that had arisen with regard to devel­oping fast breeder reactors and reprocessing plants.52

From the late 1970s to 2002 the nuclear power industry suffered relative decline, particularly in Europe and North America. The number of reactors coming online from the mid 1980s little more than matched retirements, though capacity increased by nearly one third and output increased 60% due to capacity plus improved load factors in existing reactors. The share of nuclear in the world electricity market from the mid 1980s was fairly constant at 16-17%. During the period 1986-2005 there were only 71 new nuclear plants constructed, with only a small number of these being in the ‘‘old’’ nuclear power countries; in the preceding 20 year period there had been 436 new nuclear plants. As a result, but also due to an increase in secondary supplies, the uranium price dropped.

Moreover, in the USA, in what had been one of the largest markets for nuclear power stations, nuclear reactors became increasingly uncompetitive in comparison to coal-fired power stations as well as to the emerging combined cycle gas turbine plants.53 Gradually ‘‘the optimism of the early 1970s turned to pessimism about the future of nuclear power’’.54 Although at first the US Clean Air Act increased the attractiveness of nuclear power in comparison to coal through the 1970s, this was more often than not offset by the effects of capital cost escalation and increasing time delays that plagued the construction of new nuclear plants. From the mid 1970s to the mid 1980s 100 nuclear plants were cancelled in the US alone.55

There were two accidents, however, which sounded the death knell for the nuclear industry, reinforcing the negative public and utility perception toward nuclear power that had emerged during the 1970s. Firstly, the partial core meltdown in 1979 at the reactor at Three Mile Island, Pennsylvania, which, although did not lead to any loss of life, caused over 100 000 people to flee their homes and cost a great deal of public money ($2.5 billion). This hastened moratoria on building new nuclear stations in a number of European countries (Italy, Belgium and Sweden) as well as contributing to killing off the industry in the USA.

Secondly, the fire and core melt down at one of the RBMLK reactors at Chernobyl during 1986 which sent radioactivity into the atmosphere contaminating dozens of countries, causing 50 immediate casualties and

thousands of cancer victims in the years afterwards, and displacing 100000s of people across the former Soviet Union from the Ukraine and Belarus. The accident led to an intensification of anti nuclear sentiment across Europe, in Germany in particular,56 as well as contributing to Finland (a relatively pro nuclear country) to shelve plans for a new nuclear reactor. Opinion polls throughout Europe captured the impact Chernobyl had on public attitudes toward nuclear power. Even in Finland, a country with relatively higher levels of public support for nuclear power, the amount of people who wanted to phase out nuclear power after Chernobyl increased from 21.3% in 1983 (ref. 58) to 34.5% in 1986 (ref. 59). In the UK a de facto moratorium on new nuclear builds was declared in 1989, pending a five year review.59,60 Chernobyl marked the end to the ailing nuclear industry’s hopes of recovery after the decline of the late 1970s. Nuclear power became increasingly framed in terms of its risks rather than its feel this is a good description benefits. The ‘‘new’’ framing around the risks of nuclear energy did not erase, but came to compete against and co-evolve with the original framing around the ‘‘promise of civil nuclear energy’’.61

The economic case against nuclear power became increasingly central to the debate during the 1980s. Security of supply faded to the background in a number of countries, as the influence of OPEC declined, oil prices fell, and the North Sea oil and gas fields provided cheap domestic energy. Moreover, the privatization of electricity networks also exposed the fragi­lity of nuclear economics. The hidden costs of nuclear were starkly exposed when nuclear power stations were exempt from the privatization of the UK electricity industry in 1989, given the refusal ofthe private sector to take on the risk of ageing nuclear stations with potentially massive liabilities.™11 In the UK, as a result of the 1995 white paper The Prospects for Nuclear Power in the United Kingdom, seven AGR stations and the one PWR were floated on the stock market as British Energy for the sum of just £2.1 billion in 1996 (Sizewell B alone had cost £2.8 billion). The private sector had got eight stations for the price of one,65 reflecting the desire of government to get rid of the nuclear plants as quickly as possible. xxvin It became clear that private utilities found it more economical to build gas-fired plants than either nuclear or coal-fired plants. It also became transparent that, bereft of a subsidy, the utilities in liberalised energy markets did not view nuclear stations as the most cost effective investment.[24] [25] [26]

There is relatively little information available on social and psychological consequences of the Windscale accident. This likely reflects a lower level of awareness, both in the public and in scientists and decision makers, of these issues at the time. A great deal of secrecy surrounded the Windscale plant and accusations have been made of a ‘‘cover-up’’ of the accident consequences by the operators and authorities.13 The fact that Wolff5 noted that the con­taminated milk from the accident ‘‘…could have been used for manufacturing purposes or the feeding of livestock but, because of public apprehension, it was decided not to salvage the milk’’ suggests significant public concern over the contamination, and official awareness of that concern. An article in Scientific American a few months after the accident stated that ‘‘The accident produced something approaching panic among the local population’’.14

KATHERINE MORRIS*, GARETH T. W. LAW AND NICK D. BRYAN

ABSTRACT

In the UK, there is a nuclear waste legacy associated with over 50 years of nuclear power generation that is currently stored at the Earth’s surface. This is a global phenomenon in which many nations are now facing up to the radioactive waste legacy of several decades of nuclear power gen­eration. As society considers new nuclear power as a low carbon, secure source of energy, it is apparent that geological disposal of higher activity radioactive wastes is now the favoured route for management of this highly radioactive legacy material. Timely implementation of geological disposal is therefore a current challenge facing the UK and other nuclear nations if we are to demonstrate safe management of these materials for future generations. In this chapter, we review the type and characteristics of the higher activity wastes that the UK needs to dispose of; examine the concept of a geological disposal facility in the context of UK and inter­national experience; and discuss the proposed implementation pathway for UK higher activity waste geodisposal in the context of our large and complex nuclear legacy. Finally, we discuss the environmental chemistry research challenges that we see as vital to the safe management and dis­posal of these legacy radioactive wastes.

1 Introduction

Radioactive waste management is now a pressing issue for the UK: there is an extensive legacy of higher activity wastes (HAW), some of which have been treated for storage and geological disposal; there is also a need to demonstrate * Corresponding author

Issues in Environmental Science and Technology, 32 Nuclear Power and the Environment Edited by R. E. Hester and R. M. Harrison © Royal Society of Chemistry 2011

Published by the Royal Society of Chemistry, www. rsc. org that wastes generated from any new nuclear power reactors which are proposed as a secure, low-carbon energy source, can be managed in the future. In England and Wales, geological disposal has been chosen as the long-term management pathway for HAWs. The implementation programme for a national geological disposal facility (GDF) was launched in 2008 (ref. 1). Indeed, the scientific and societal challenges facing the UK in legacy manage­ment are being echoed at a global scale. In this chapter, we discuss the types and quantities of radioactive wastes that are likely to be disposed of to a GDF and the likely planning and design concept for this type of facility. We also highlight selected environmental chemistry research challenges in geological disposal as they relate to the reduction in uncertainties for the GDF safety case.

The mobility of most radionuclides from soil to other organisms is predominantly via plant root uptake, which will be largely determined by physicochemical factors influencing the distribution of radionuclides between the solid and solution phases of soil. The uptake of most elements by plant roots occurs mainly from the soil solution.

The important interactions of any chemical species in solution, which can influence its mobility in soils and eventual root uptake, include: charge
interactions; complexation and precipitation reactions with other chemical species (e. g. organic and inorganic ligands); oxidation-reduction (redox) transformations; and specific interactions with soil components including soil biota. Soil factors influencing the mobility of some of these radionuclides are outlined in the following paragraph.

The extent of sorption in soil is described by the solid-liquid distribution coefficient (Kd).

Simple Kd-based models assume that the radionuclide on the solid phase is in equilibrium with that in solution. However, Kd can change with time as the sorption process ‘‘ages’’.

The Kd for a radionuclide may vary within various orders of magnitude depending on the combination of radionuclide and soil type.21 The use of a cofactor approach can decrease the variability of the ranges of Kd values associated with a soil type. For example, Kd is affected by the radiocaesium interception potential (RIP), K and NH4+ status for radiocaesium, the cation exchange capacity (CEC), Ca and Mg concentrations for radiostrontium, and the pH for heavy radionuclides.21

Caesium is strongly sorbed in soil by ion exchange, some of which is irrever­sible, or fixed, with fixation being influenced by clay mineralogy. A number of models relating the availability of radiocaesium to soil properties have been proposed, including increasing soil-plant transfer with increasing soil organic matter;22,23 decreasing soil-plant transfer with increasing soil solution potas­sium;24 a semi-mechanistic approach using soil clay and organic matter contents, exchangeable K status, pH and NH4+ concentration;25 and, more recently, the use of RIP of soils and exchangeable potassium concentration to predict caesium uptake.21 Characterisation of the soils in the areas of Japan which are receiving radionuclide deposition from Fukushima should enable reasonable predictions of the long term availability of radiocaesium to foodstuffs.

Recent comparisons of data21 showed that the Kd of strontium for sand, loam, clay and organic soil groups were similar, although the value for the sand group was significantly lower. For radiostrontium, the key soil characteristics determining sorption were CEC and calcium and magnesium concentrations.

The transuranic radionuclides, americium and plutonium, have relatively low mobility due to their strong tendency to sorb onto soil particles.26 Americium generally exists in the iii and/or iv valence state, whereas plutonium often exists in the iv state, but can be found in any of four oxidation states (iii, iv, v or vi) depending on the redox conditions of the soil system.

After extraction from ore, a uranium concentrate (“yellowcake”) is manu­factured and shipped to facilities where it is enriched (the proportion of 235U is increased from the natural 0.72 atom%) if required and fabricated into fuel. Enrichment is needed for modern reactor fuels which are made from UO2, and generally requires conversion of the uranium into UF6, a relatively volatile compound which is attractive for enrichment because fluorine is monoisotopic, followed by multiple stages of membrane diffusion or centrifugation. The enrichment process, as used for fuel production, creates two uranium streams, one enriched, typically to 3-5 atom% 235U (referred to as low enriched ura­nium), and one depleted to around 0.2 atom% 235U. Typically, therefore, pro­duction of 1 kg of low enriched uranium creates 5 or 6 kg of depleted uranium, for which there is little current use. In total, about 1.2 M tonnes of depleted uranium exist worldwide. A significant proportion of the global depleted ura­nium inventory is still in the form of UF6, a reactive, corrosive material which is not suitable for long term storage or disposal.

For use in current reactors, enriched uranium is “deconverted” from UF6 into UO3, reduced to UO2, a durable ceramic, and formed into pellets. The pellets are loaded into metal tubes, generally of stainless steel or zircaloy (a range of zirconium-based alloys, often containing tin or niobium), depending on reactor type, and are then suitable for loading into a reactor. Early reactors, such as the first Hanford production reactors in the USA, and the UK Magnox reactors, were designed to use fuel of natural isotopic composition, which obviously avoids the difficulty and cost of enrichment. However, such reactors cannot tolerate the dilution of fissile isotopes which occurs in UO2, and have to use uranium metal, whose properties limit reactor operating temperatures and efficiency.

A large and complicated nuclear legacy has been left behind by the break up of the Soviet Union with numerous nuclear facilities located in Russia and other former Soviet states. Although civilian activities have contributed to this legacy, the majority of contamination issues in the former Soviet Union were created by military nuclear facilities used for the production of nuclear weap­ons. This problem was exacerbated by a previously relaxed attitude towards environmental issues with regards to nuclear waste disposal. Three nuclear facilities, Chelyabinsk-65, Tomsk-7, and Krasnoyarsk-26 operated in secret in the Ural mountains during the Cold War and were not subject to strict environmental practices.16 Of these sites, Chelyabinsk-65 (Mayak) is the most publicised regarding its former activities, revealing a long history of accidental release and discharges, contributing to significant environmental contamina­tion which will be discussed in greater detail below.

The worst nuclear power plant accident in history, the only level 7 event on the International Nuclear Event Scale to have occurred, happened on the 26th of April 1986 at Chernobyl, when a test was carried out to determine the ability of a turbine generator to provide power in the event of a station blackout. Serious violations of safety procedures and operating rules resulted in a steam explosion, cutting cooling channels on both sides of the reactor core resulting in a further explosion.17 The release of 137Cs from the explosion is estimated to have been around 85 PBq with an estimated 1760 PBq of 131I, 10 PBq of 90Sr and 3 PBq of plutonium isotopes also released.18,19