"PROCESS AND ABSORBER MATERIAL FOR REMOVING RADIOACTIVE ISOTOPES FROM WATER"

_{********************}

FIELD OF THE INVENTION

The present invention refers to a process and an absorber material for decontaminating water containing ions of radioactive isotopes; the process and the material of the invention are particularly effective in removing cesium ions from water.

STATE OF THE ART

It is known that nuclear plants produce high volumes of water contaminated by radioactive isotopes; the production of water containing these isotopes occurs both as a result of the regular plant operation and as a consequence of catastrophic events.

A first cause of production of large amounts of contaminated water in the regular operating cycle of nuclear plants is the water reactor cooling operation. The best example is offered by the electro-nuclear power plants, which constitute the vast majority of currently active nuclear power plants, and typically use this type of cooling. The cooling waters of nuclear reactors become radioactive during their operational cycle for various reasons, ranging from contact with different products (radioactive due to fission or activation), to direct activation due to exposure to radiation. These cooling waters are generally contaminated by a significant amount of fission and/or activation products, of which cesium isotopes (uranium and Plutonium fission products) frequently constitute the main fraction.

A second source of contaminated waters regularly produced by nuclear plants are the storage pools, i.e. the water tanks used for cooling and shielding used radionuclei ("spent fuel"). These pools are an example of temporary, convenient and safe storage located in the nuclear power plant complex itself. The common radioactive contamination of water in the storage pools still includes soluble fission and/or activation products, among which again cesium, and in smaller amounts strontium, predominate.

A third cause of origin of waters contaminated by radioactive isotopes is the use of water in large temporary repositories of nuclear waste waiting for appropriate treatment and conditioning for their disposal; in these repositories, large volumes of water are present as highly radioactive supernatant. Typically, these waters are heavily contaminated by cesium and strontium.

Finally, contaminated waters may be produced due to situations not resulting from planned activities, following exceptional and often catastrophic events. In these cases, again, the content of cesium and strontium in the contaminated waters is dominant.

Water contaminated by radioactive isotopes is a fundamental issue for the management of nuclear plants, due to the huge volumes produced. The water radioactivity issue directly involves the safety of the personnel managing the plant. The consequent radioactivity abatement in plant waters, including the storage pools, significantly affects the operating costs of the plant. The costs for disposal of the slag generated by the radioactive isotopes removed from the water in the decontamination work must be added to the costs of eliminating radioactivity from the waters. The presence of radioactive contaminants also interferes with the sophisticated processes of water purification from other contaminants, not necessarily radioactive, whose concentrations must be kept at minimum levels (typically ppb) because they are potentially harmful to the stability and efficiency of the plant.

The decontamination of water contaminated by water-soluble radioactive isotopes, in particular cesium and strontium, is therefore a constant issue in the nuclear industry. Moreover, in the major accidents occurred in electro-nuclear power plants in the last decades, the decontamination of large volumes of water contaminated by radioactive elements represented one of the high priority issues in emergency interventions for the control and rehabilitation of the plants themselves.

The methods used for the removal of radioactive isotopes, in particular for decontamination of cooling water contaminated with cesium, are typically based on solid absorbers, that are necessarily limited to those not water-soluble. Zeolites and ion exchange resins constitute the two large families of materials used for this purpose.

Zeolitic structures (typically aluminum silicates) generally have the ability to immobilize cations through ion exchange; for the decontamination of radioactive waters, synthetic zeolites of various types have also been developed.

In particular, clinoptilolite, a type of natural zeolite, has specific ion exchange properties for cesium cations. Furthermore, being zeolites porous, they in principle allow to host molecules or ions in their cavities, which can also be thermally activated so as to develop, within the same cavities, acidic functions of Bronsted type, Lewis type or both types, useful for fixing molecules, typically organic ones; see the article "Adsorption Behavior of Cesium and Strontium on Synthetic Zeolite P", H. Mimura e K. Akiba, Journal of Nuclear Science and Technology, (1993) 30:5, 436-443.

Zeolites, especially the naturally occurring ones, have the advantage of being an abundant material available at low cost (typically fractions of euro/kg). On the other hand, zeolites have a low absorption efficiency, in the order of 0.05 - 0.1 equivalents of species to be absorbed per kg of zeolite (eq/kg); furthermore, the definitive storage of radionuclei requires their immobilization in glass of high chemical resistance, and the vitrification of zeolites is relatively expensive, because zeolites cannot be vitrified as such and must be incorporated into borosilicate glass; see, for example, the report by the Pacific Northwest Laboratory (Richland, USA) "Summary of radioactive operations for zeolite vitrification demonstration program", G. H. Bryan et al., (1984).

Ion-exchange resins are the most widely used material for decontaminating cooling water in nuclear plants: see the article "Radioactive waste management at nuclear power plants" V. M. Efremenkov, IAEA BULLETIN, 4/1989, pp. 37-42; and the article "Ion Exchange Resins For Use In Nuclear Power Plants", J.J. Wolff (article revised and updated in January 2012). The predominant use of ion-exchange resins is justified by their high efficiency (from 2 to 10 eq/kg), high reliability, relatively low cost (in the order of 1 euro/kg), and the possibility of repetitive use through the regeneration of the same. Given these advantages, resins have however the issue of being virtually not vitrifiable; the calcination of the organic component of the resin "charged" with radioactive isotopes involves high secondary contamination due to fumes, with high abatement costs. Conversely, if the resins are regenerated, a liquid radioactive waste (in HNO3) is produced, which does not completely eliminate secondary contamination: typically, after about 10-12 recycles, a non-disposable residue is left, which obliges to dispose of the resin.

In addition to zeolites and resins, other materials have been proposed in the literature for the absorption of radioactive ions, in particular cesium, from contaminated waters. The main materials evaluated are:

- Ammonium MolybdoPhosphate (AMP) - it is a reagent usable only in the pH range from acid to neutral and very active as a cesium absorber, with a material capacity, used alone, up to 0.68 eq/kg. Its use as such for column applications, however, is not convenient, because it is microcrystalline. AMP is then usually used incorporated in a resin, thus obtaining an absorber material known as ALIX (Advanced Lyophilic Ion-exchanger), as reported in the article "Recovery of cesium from high level liquid nuclear waste by an advanced polymer composite", L. Varshney et al., Bare Newsletter, No. 327 (2012), pp. 26-30 (available at: http://www.barc.gov.in/publications/nl/2012/201207081 1.pdf); in this case, the capacity is 0.34 eq/kg. It has been successfully used, incorporated into polyacetonitrile (AMP)-(PAN), for oceanographic research. It has been proposed, incorporated in polysulfone, for the decontamination of high level radioactive nuclear waste (see the article "Extraction of cesium in seawater off Japan using AMP-PAN resin and quantification via gamma spectroscopy and inductively coupled mass spectrometry", S. M. Pike et al., J Radioanal Nucl Chem (2013) 296:369). This material has the following issues: the AMP compound is not vitrifiable as such but only after appropriate conditioning; Cs absorption by AMP is not reversible; and, finally, the AMP reagent cost is relatively high, and the application thereof in ion- exchange resins is expensive (also because it is not reversible), and in any case it would suffer from the same drawbacks of the resins (not vitrifiable material);

- Potassium ferrocyanide - it has been known for over half a century in the nuclear industry for its properties as (reversible) cesium absorber from liquid solutions (see, for example, US Patent 3,453,214, 1969); it is probably the most commonly cited reagent in the decontamination activity of plants cooling water. Operatively, potassium can be advantageously replaced by transition metals (Cu, Co, Zn). In the decontamination of cooling water, it is not convenient to use the adsorber as is (in powder form), but in the form incorporated in polymers (ion- exchange resin), which can operate in suitable columns. For this purpose, the substitute metal may play a critical role in fixing specific organic chelates; the capacities are 1 .22 eq/kg, in the case of cobalt ferrocyanide, and 0.55 eq/kg with zinc ferrocyanide. The drawback of this material is that its use is dependent on the organic resin incorporating it, and the issues are those typical of ion-exchange resins (it is not vitrifiable);

- Silicotitanates - these materials have good activity as reversible cesium absorbers, and are available in crystalline and non-crystalline form. In the crystalline form, the material has been proposed as a "tank reagent" for the abatement of cesium in high concentration (HLLW) for the pool park of the US Department of Energy (DOE). The same technique would not be feasible for the decontamination of cooling water with a low concentration of radioactive isotopes. A considerable research and development activity has been instead dedicated to its incorporation (also by inorganic materials) for decontamination in column applications. In this form, it has been proposed for the decontamination of supernatants containing high cesium concentration in tanks with a strongly alkaline pH. The advantages of this material are a high absorption efficiency and the possible vitrification (in the version with incorporation by inorganic material); the issues it poses are that, as such, it is not directly vitrifiable; in the application in ion-exchange resins it has the typical drawback of resins, i.e. it is not vitrifiable; while its incorporation into glass is relatively expensive;

- finally, layered metal sulfides - in relatively recent times, new inorganic materials, based on mixtures of potassium, manganese, tin, and sulfur in various ratios, have been tested. In these materials, potential ion-exchange properties could be favored both by their layered morphology (which allows wide internal spaces), and by the mobility of K ⁺ ion within the same structure. In this regard, a remarkable ion exchange activity, selective for the Sr ²⁺ ion, has been reported, see the article "Highly Selective Removal of Cesium and Strontium Utilizing a New Class of Inorganic Ion Specific Media - 9267", M. S. Denton et al., presented at the WM Conference in March 2009 in Phoenix. The inorganic composition allows the possible conditioning for disposal (vitrification by incorporation in vitrifiable material), but the material cannot be directly vitrified.

The need to have a method for the absorption of radioactive ions (in particular, cesium) from contaminated water, which has a combination of advantageous characteristics for the intended application, is therefore still felt in the sector.

The object of the present invention is to provide a process that allows high efficiency absorption of radioactive isotopes in ion form from water, and direct vitrification, without having to resort to incorporation of the material resulting from said absorption into third materials.

SUMMARY OF THE INVENTION

This object is achieved with the present invention which, in a first aspect thereof, relates to a process for the absorption of radioactive ions from water, comprising the steps of:

a) preparing a mixture of one or more iron(lll) phosphates, iron(ll) phosphates or iron(lll)/iron(ll) mixed phosphates with phosphoric acid, in an amount such that the P/Fe atomic ratio is of between 1 .2 and 3.5, and subsequently treating the mixture of said one or more iron phosphates and phosphoric acid at a temperature of between 45 and 300 °C for a time of between 1 and

48 hours, thus obtaining an absorber material;

b) contacting the absorber material obtained in step a) with water containing ions of radioactive isotopes, causing the absorption of said ions by said absorber material; and

c) after absorption, physically separating said absorber material from the water.

In a second aspect thereof, the invention relates to the material obtain as a product in step a) of the process described above.

BRIEF DESCRIPTION OF THE FIGURES

The invention will be described in detail with reference to the Figures, in which:

- Figure 1 shows in graphical form the behavior of the percentage of cesium absorbed as a function of the increasing ratio between the initial cesium amount present in the solution and the amount of absorber material used;

- Figure 2 shows, in graphical form, the behavior of cesium absorption capacity of a material of the invention, in terms of cesium absorbed mass (normalized with respect to absorber material mass), as a function of the ratio between the initial cesium amount present in the solution and the amount of absorber material used. DETAILED DESCRIPTION OF THE INVENTION

The inventors observed that an absorber material obtained according to the method described in step a) is insoluble in water, and intimately contacted with water containing typical radioactive contaminants present in nuclear plants' cooling systems, for example cesium ions, interacts therewith forming mixed phosphates, for example iron-cesium mixed phosphates, which are also highly insoluble in water, thus allowing the effective removal of these ions from the aqueous phase.

In the following description and claims, reference is made to water, but the invention may be applied in the same way also to water-based liquid phases containing small amounts of other liquids, such as organic solvents, particularly if completely miscible with water.

Iron phosphates that may be used as starting materials in the invention are all forms of iron(lll) phosphates, iron(ll) phosphates, or iron(lll)/iron(ll) mixed phosphates, as well as mixtures thereof.

It is known that phosphates can occur in various aggregation forms, hydrated or not; for example, iron(lll) phosphate, FePO4, exhibits four polymorphs and two polymorphs of the dihydrate form, FePO4 2h Q. Iron(il) phosphate, Fe3(PO4)2, exists in anhydrous form or in the octahydrate form, Fe3(P04)2-8H20, also present in nature as vivianite mineral; the mixed form, {Fe ² VxFe ^3* x)(PO4)2(OH)x-(8-x)H20, corresponding to metavivianite mineral, wherein Fe(II) e Fe(HI) are co-present, is also known. All these compounds are essentially insoluble in water.

In step a) of the process of the invention, a mixture of one or more iron(lll) phosphates, iron(ll) phosphates, or iron(lll)/iron(ll) mixed phosphates with phosphoric acid, in an amount such that the P/Fe atomic ratio is of between 1.2 and 3.5, and preferably of between 1 .5 and 2.0, is prepared.

This mixture may be produced starting from an iron phosphate having a P/Fe ratio lower than 1.2, or from a mixture of iron phosphates with an average composition lower than said value. In this case, the phosphate or phosphate mixture is added with phosphoric acid in the amount needed to bring the P/Fe ratio in the range of between 1.2 and 3.5.

Alternatively, it is possible to produce a native mixture having already the P/Fe ratio in the desired range. This condition can be obtained, for example, by preparing a solution of Fe(lll) and/or Fe(ll), for example by dissolving iron chlorides or nitrates in water, and adding phosphoric acid in excess to the amount of iron ions to be precipitated, and such that the P/Fe atomic ratio in the solution has a value in the aforementioned range. A solution of this type can also be produced by etching of iron or steel with phosphoric acid, as described in the patent application WO 2013/1681 18 A2 in the name of the Applicant.

The mixture prepared according to any of the two alternative methods reported above, containing an excess of phosphoric acid, is then subjected to the heat treatment of step a).

Step a) is carried out within the time and temperature ranges reported above; within these ranges, the duration of the phosphate and phosphoric acid mixture heat treatment decreases with increasing temperature.

In evaluating the amount of phosphoric acid to be used to reach an atomic

P/Fe ratio in the aforementioned range, the phosphorus already present in the starting phosphate (or phosphate mixture) must be taken into account.

In the case where a mixture of several iron(ll) and/or iron(lll) phosphates is used as starting phosphate, an analysis of the percentage content by weight of iron and phosphorus in the phosphates mixture is carried out before the acid treatment, to determine the number of moles of atomic iron and phosphorus in the sample; from this data it is then possible to determine the amount of phosphoric acid needed to have the P/Fe atomic ratio required in the absorber material of the invention.

Conversely, when the iron phosphate is a pure salt, the calculation of the amount of phosphoric acid required is made simply by obtaining the moles of atomic iron and phosphorus by dividing the weight of the sample by its molecular weight (thus obtaining the number of moles of phosphate), and possibly multiplying the value obtained by the number of iron and phosphorus atoms in the brute formula of the phosphate.

For example, in the preferred use as starting material of iron phosphate dihydrate, FePO4 ^■ 2H2O, the initial atomic P/Fe ratio is 1 ; to bring the ratio into the useful range of the invention, between 0.2 and 2.5 (preferably between 0.5 and 1 .0) moles of orthophosphoric acid, H3PO4, per mole of initial phosphate are added.

The inventors observed that also absorber materials obtained by treatment of one or more iron phosphates with phosphoric acid, and subsequent heat treatment, are essentially insoluble in water.

In step b) of the process of the invention, the absorber material produced in the previous operation is brought into contact with water containing radioactive isotopes.

The absorber materials of the invention proved to be effective in the absorption and fixing of radioactive isotopes ions present in water, particularly in the case of cesium; the result of a treatment of an aqueous phase containing cesium ions with an absorber material of the invention is the removal of the cesium ions from the solution, with consequent decontamination of the same water. The phenomenon may be observed with cesium salts aqueous solutions, e.g. cesium nitrate in water, typically in the pH range of 2.5 to 9. The more finely ground is the absorber material of the invention the more efficient is the removal, and (in the initial absorption phase) it is a direct function of the time by which the same is kept in intimate contact with the water containing the cesium ions.

Since the absorption rate is a function of the absorber material exposed surface, this is preferably used in granule or powder form, having a size of between 0.01 and 5 mm; with particle sizes bigger than this range, the absorption times are undesirably elongated, while smaller particle sizes make subsequent operations, for example the filtration to separate the absorber material from the decontaminated water, less easy. For a concentration of radioactive ions that typically, in these applications, is of between about 1 10 ^-3 and 1 10 ^-9 mol/L, the absorption times range between 50 minutes and 50 hours per liter of treated solution.

The ability of the absorber materials of the invention to remove cesium ions from aqueous solutions and concentrate them in the solid state appears high. In terms of Cs milligrams removed per gram of absorber material used, in experiments carried out by the inventors, cesium concentration values in iron phosphate were obtained, after absorption, equal to or higher than 10% by weight of solid absorber material used, i.e. in the order of 100 g of cesium per kg of absorber material, i.e. about 0.75 eq/kg. In the context of a coordinated process for decontamination of radioactive water and disposal of the waste produced, these cesium concentration data in the solid absorber material may be directly translated into concentrations in the final disposal glass, as shown in the examples given. The cesium absorption capacity of the absorber materials of the invention is higher than that of zeolites and resins, the materials most commonly used for this purpose; it is instead inferior to that of materials such as AMP and cobalt ferrocyanide, which however have the issues seen above, particularly the impossibility of direct vitrification.

A fundamental property of the absorber materials of the invention is their insolubility in water at the conditions of use, which allows recovery thereof by simple phase separation after absorption, for example by filtration. As an example of this property, the material obtained by treatment of neutral ferric phosphate FeP04 ^■ 2H2O in granules, which is virtually insoluble in water at pH values higher than 0.5 (solubility of about 3x10 ^"8 mol/L at room temperature), and can therefore be fully recovered after any contact thereof with aqueous solutions, may be mentioned.

At the end of the absorption phase (which can be monitored by real-time analysis), in step c) of the process, the absorber material loaded with radioactive ions can be separated from water through simple operations, such as filtration, centrifugation, or similar.

The solid thus recovered may be sent to simple disposal operations. The third very important feature of the absorber materials employed in the present invention is, in fact, their direct vitrification. After absorption of the radioactive ions, in fact, the iron phosphates can be melted at temperatures of between 900 °C and 1200 °C, and preferably of between 950 °C and 1 100 °C, to form dense bodies, without giving rise to ashes, fumes, or other by-products, and without the need for the addition of further compounds or incorporating materials (such as cements, bitumen, or glass used for the disposal of other types of nuclear waste); the material is therefore self- incorporating, and constitutes itself the high chemical resistance glass suitable for the disposal of long-lived radioactive waste. The final volume of the glass produced is extraordinarily reduced and reasonably predictable. For example, 1 cubic meter (1000 liters) of water contaminated by 100 ppm of ¹³⁷ Cs requires about 1000 g of absorbent that, vitrified, will have a volume of 1000 g/(3 g/cm ³ ) = 0.333 liters of glass.

This feature is particularly important for environmental protection, because it allows to avoid, through direct vitrification, any secondary contamination downstream of possible chemical treatment and/or conditioning operations for the disposal of absorbers and/or ionic exchangers that are not directly vitrifiable.

A practical example of this third property is the glass obtained from ferric phosphates (IPG - Iron Phosphate Glass) containing absorbed ions of ¹³⁷ Cs or other radioactive isotopes, that demonstrates excellent chemical resistance.

The waters recovered after separation of the absorber material, no longer containing radioactive ions, may be sent to standard purification operations from non-radioactive contaminants; the waters thus purified may be reused for nuclear plant cooling.

In the present description, evaluations of radiation intensity related to the concentration of radionuclei have not been included. Any concentration limits are indicated by safety regulations and the stability of the glass itself. The concentrations shown can in principle be considered obtainable in ferro-phosphate glass, according to the procedures described. Obviously, limiting the radionuclei concentrations in the glass is not a technical problem.

In the second aspect thereof, the invention relates to the absorber material obtained as a product in step a). This material has a complex composition and, so far, it has not been possible to determine its structure. So far, the inventors have only been able to observe that the P/Fe atomic ratio in the absorber material obtained at the end of step a) is always slightly lower than that of the initial iron phosphates/phosphoric acid mixture; for example, with a starting mixture containing a P/Fe atomic ratio equal to 1 .5, the final absorber material has a P/Fe ratio of 1 .43, while with an initial ratio of 2.0, a P/Fe atomic ratio equal to 1 .83 is obtained in the final absorber material. The reason for this phenomenon could be a partial evaporation of species derived from the acid during the heat treatment in the production of the material.

The invention will be further described by the following experimental part.

EXAMPLE 1

This example relates to the absorption of cesium ions by direct immersion of an absorber material of the invention in an aqueous solution.

An absorber material of the invention was prepared by treating 5.87 g of iron phosphate dihydrate, FePO4-2H2O, reagent grade (Aldrich, catalog number 43601 1 ), with 2.5 g of an 85% by weight orthophosphoric acid aqueous solution (corresponding to 2.13 g of orthophosphoric acid). The mixture was treated at 105 °C for 6 hours; 8.00 g of absorber material were obtained which, analyzed by X-ray fluorescence (XRF), showed a P/Fe atomic ratio equal to 1 .43.

Separately, a cesium aqueous solution was prepared by adding 0.146 g of cesium nitrate, CsN03, reagent grade (Aldrich, catalog number 289337, 99% purity) in a hermetically sealed polyethylene bottle containing 1000 ml_ of demineralized water; the resulting solution contains 100 ppm of Cs in 1000 g of solution. The bottle was subjected to vigorous stirring until a clear solution was obtained.

100 g of the solution thus obtained were transferred into a second polyethylene bottle to which all the previously prepared absorber material was added. The bottle was sealed and subjected to vigorous stirring for a total of 6 times in 15 hours.

The suspension was subjected to filtration with IF5A filter; an initial rate of one drop every 3 s was observed, with a progressive decrease over time to one drop every 12 s at the end of the filtration, which lasted 2.5 hours. The filtrate was clear and, by weighing, was found to have a total mass of 76.29 g.

The net weight of the material left on the filter (wet mud) was equal to 28.79 g, while 2.00 g of solution remained in the bottle (measurement obtained by weighing the bottle at the beginning and at the end of the procedure).

The mass balance resulted to be: 76.29 g + 28.79 g + 2.00 g = 107.08 g. The missing mass with respect to the initial 108.00 g, equal to 0.92 g, is probably due to water evaporation during filtration.

The solution obtained by filtration was analyzed to determine the residual cesium concentration; the measurement was performed by Atomic Emission Spectroscopy, with a flame excitation source; the cesium concentration in the filtered solution resulted to be 3.2 ppm. The percentage of cesium absorbed by the phosphate is therefore equal to:

[(100 ppm - 3.2 ppm)/100 ppm]x100 = 96.7%.

The ratio between the Cs absorbed mass and the absorber material mass is equal to:

0.00967 g/ 8.00 g = 0.00121 , i.e. 1 .21 mg of Cs absorbed per gram of absorber material. EXAMPLE 2

This example relates to the absorption of cesium ions from an aqueous solution by an absorber material of the invention charged in an ion exchange column.

The absorber material used was prepared starting from a mixture of highly water-insoluble iron-phosphate salts, obtained by precipitation with orthophosphoric acid from an aqueous solution containing Fe ²⁺ and Fe ³⁺ ; after a separate analysis, this mixture resulted to be comprised for the most part (more than 80%) by FeP04-2H ₂ 0 and FeH ₃ (P04)2 4H ₂ 0, plus minor amounts of Fe(H ₂ P04)3, Fe ₄ (P04)3(OH)3, FeP04 and other iron phosphates. This mixture was subjected to heat treatment at 105 °C for 6 hours. The obtained absorber material was measured by X-ray fluorescence (XRF) and a P/Fe atomic ratio of 1 .66 was measured.

The column had a useful length of 40 cm, an internal diameter of 32 mm, and was supplied with a ceramic filter at the base, a lower tap, a connection joint at the upper end with a tap for vertical communication, a side valve for checking the internal pressure, and an additional 1500 ml_ upper tank connected to the column through a joint.

9.25 g of the absorber material prepared as described above were transferred into the column; then, 100 g of the aqueous cesium solution at 100 ppm concentration prepared as described in Example 1 were transferred into the column.

3 hours were required for the solution to pass through the column, with an initial drip rate of one drop every 2 s and a slowing down of the rate after about an hour and a half (one drop every 10 s). At the end of the drip, 78.91 g of solution were collected in the container under the column.

An Atomic Emission Spectroscopy was performed on this solution, which revealed a residual concentration of cesium equal to 0.2 ppm.

The percentage of cesium absorbed by the phosphate is therefore equal to:

[(100 ppm - 0.2 ppm)/100 ppm]x100 = 99.8%.

The ratio between the Cs absorbed mass and the absorber material mass is equal to:

0.00998 g/9.25 g = 0.00108, i.e. 1.08 mg mg of Cs absorbed per gram of absorber material. EXAMPLE 3

Three standard cesium solutions were prepared in tridistilled water, in a graduated flask, as follows:

- Solution A: 1 .268 g of cesium chloride (RPE Carlo Erba, purity > 95%) were weighed and dissolved in 1 L of tridistilled water, in a graduated flask, thus obtaining a cesium concentration equal to 1000 mg/L;

- Solution B: 50 mL of Solution A were withdrawn and brought to volume with tridistilled water, in a 500 mL graduated flask, thus obtaining a cesium concentration equal to 100 mg/L;

- Solution C: 50 mL of Solution B were withdrawn and brought to volume with tridistilled water, in a 500 mL graduated flask, thus obtaining a cesium concentration equal to 10 mg/L.

7 samples were prepared using the standard solutions A, B and C, each having a volume of 40 mL, in 50 mL polyethylene bottles with screw cap. A different amount of absorber material, prepared as in Example 2, was transferred into each sample, as specified in Table 1 . An eighth sample was prepared with a substantially higher solution volume (400 mL of solution A) and a minimum amount of absorber material (0.04 g).

All samples were subjected to mechanical stirring for 12 hours. Subsequently, each sample was filtered through a 0.25-micron membrane.

The cesium concentration in the filtered solutions and standard solutions was measured using Atomic Emission Spectroscopy with flame excitation source.

The conditions of the various measurements and the results obtained are shown in Table 1 .

Table 1

The results obtained indicate that, in an equilibrium time of 12 hours and at the experimental conditions, the dominant parameter for the absorption phenomenon is the ratio between the cesium initial mass and the absorber material mass, as better highlighted in the following Table 2.

Table 2

The results in the two Tables above are also shown in graphical form in Figure

1 .

EXAMPLE 4

A series of five repetitions of the Example 1 test were performed increasing the initial cesium amount in solution to a test constant value of 12.1 mg per gram of absorber material, and varying the absorption times. The results of this series of tests are shown in Table 3.

Table 3

Results comment

The examples show that iron phosphates subjected to treatment with phosphoric acid and heat, used in the process of the present invention, are able to absorb almost all of the cesium present in a solution. The absorption is more effective when the solution goes through a bed of absorber material, because in this way all the cesium is forced to come into contact with said material; in case of addition of the phosphates treated according to the invention to a solution, the cesium transport towards the absorber takes place by concentration gradient, which is a slower phenomenon and it is probably responsible for the slightly worse results, in terms of absorption efficiency, obtained in the Example 1 .

In both cases, however, there was a high absorption efficiency, with percentages of cesium removal from water well above 95% (and, specifically, 99.8% in case of use of the absorber material in an ion exchange column); in the case of the column, the efficiency (time required to obtain the effect) is higher due to the aforementioned reason. In these two Examples, the capacity of the absorber material was used to the full, due to the lack of Cs availability that could not saturate said capacity.

In Example 4, the amount (availability) of Cs was increased compared to the first tests; in this case, it was confirmed that the absorber material of the invention has absorption capacity always higher than 95%, and in most cases higher than 99% of Cs present in solution.

Example 3, instead, explores the cesium absorption capacity up to its saturation. A consideration of marked applicative impact is offered by the first four values in the columns "Cs Initial Mass/Absorber Material Mass" and "Percentage of Cs absorbed" in Table 2: from these data, it can be concluded that up to an amount equal to or less than 10 mg of cesium per gram of absorber, at the experimental conditions, substantially total absorption is obtained. The maximum cesium absorption capacity by the material is obtained from the cesium absorption diagram per absorber unit as a function of cesium availability, reported in Figure 2.

From these functional values, it is realistic to foresee highly competitive costs for the decontamination of large volumes of water contaminated with radioactive cesium.