actinide speciation in environment and their...

Actinide Speciation in Environmentand Their Separation Using FunctionalizedNanomaterials and Nanocomposites

N. Priyadarshini, K. Benadict Rakesh, and P. Ilaiyaraja

ContentsIntroduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3Background: Hydrolysis of Metal Ions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5Hydrolysis of Actinides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6Plutonium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

Complexities in Plutonium Aqueous Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7Hydrolysis and Polymerization of Tetravalent Plutonium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

Parameters Influencing the Formation of Plutonium Polymers/Colloids . . . . . . . . . . . . . . . . . . . . 9Probing the Formation and Structure of Pu(IV) Polymer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10Size and Morphology of Pu(IV) Colloid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13Hydrolysis and Solubility of Pu(IV) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Methods Adopted to Remove Polymers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15Extraction of Pu(IV) Polymer/Colloid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 16Depolymerization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17Proposed Methods for Depolymerizing Pu(IV) Polymer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17Uranyl Nitrate Effect . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

Hydrolysis of Th(IV) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 18Probing the Formation and Structure of Th(IV) Hydrolyzed Species . . . . . . . . . . . . . . . . . . . . . . . 18Size and Morphology of Hydrolyzed Th(IV) Polymer/Colloid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19Hydrolysis and Solubility of Th(IV) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19Stability of Th(IV) Hydrolyzed Colloids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20Uranyl Effect on Polymerization of Th(IV) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20Molecular Weight of Th(IV) Hydrolyzed Polymer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

Uranium . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21Hydrolysis and Polymerization of Uranium(IV) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 21

Probing the Formation and Structure of U(IV) Hydroxide/Oxide . . . . . . . . . . . . . . . . . . . . . . . . . . . 21Size and Morphology of U(IV) Colloid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22Hydrolysis and Solubility of U(IV) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22

N. PriyadarshiniDepartment of Chemistry, SSN College of Engineering, Kalavakkam, TN, Indiae-mail: [email protected]

K. Benadict Rakesh · P. Ilaiyaraja (*)Department of Physics, Indian Institute of Technology Madras, Chennai, TN, India

# Springer International Publishing AG 2018C. M. Hussain (ed.), Handbook of Environmental Materials Management,https://doi.org/10.1007/978-3-319-58538-3_143-1

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Molecular Weight of Hydrolyzed U(IV) Polymer . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22Hydrolysis of Uranium(VI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23Nanomaterials for Separation of Actinides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24Carbon Nanostructures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25Carbon Nanotubes (CNTs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25Ligand-Impregnated MWCNTs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28Graphene Oxide (GO) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28Magnetic Nanoparticles for Sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29Nanocomposites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30Magnetic Nanocomposites . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32Mesoporous Materials for Sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 33Mesoporous Membrane for Sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34Reverse Osmosis (RO) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34Nanofiltration (NF) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 35Ultrafiltration (UF) Combined with Sorption/Precipitation/Complexation . . . . . . . . . . . . . . . . . . . . . . 35Nanopolymer/Dendrimer-Assisted Ultrafiltration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 36Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 37Cross-references . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

AbstractActinides are the major and most important environmental contaminants associ-ated with anthropogenic activities such as mining and milling of uranium ores,and generation of nuclear energy resulting in the production of nuclear reactorwastes. These actinides have greater migrating ability in aquifer systems. Butactinides in aqueous environment exhibit an inordinately complex chemistry. Asa result, the chemical interactions of actinides in the environment are difficult tounderstand unless a detailed knowledge on their chemical speciation, oxidationstate, redox reactions, sorption characteristics, temperature and pressure profiles,pH, and redox potential (Eh) is available. The solubility and migration behaviorare also related to these factors. To predict how an actinide might spread throughthe environment and how fast that transport might occur, we need to characterizeall local conditions, including the nature of site-specific minerals, and ligandconcentrations. A quantitative knowledge of the competing geochemical pro-cesses that affect the actinide’s behavior is also mandatory. Once actinides enterthe environment; they pose major risk and hence safe management of radioactivewaste with minimum impact to environment gains major importance in additionto speciation of actinides in environment. The major aim of radioactive wastemanagement is to identify the chemical form of long-lived alpha-emitting radio-nuclide species spread in the environment and to separate them. Functionalnonmaterial has gained wide attention and its increasing advancements in multi-disciplinary research will make it a good candidate for separation of radionu-clides. It is mainly because of their unique structure and exceptional properties. Inthis chapter, we discuss in detail the basic research progress in the speciationstudies and separation of actinides in environment with the emphasis on applica-tion of functionalized nanomaterials and nanocomposites for the separation ofradionuclides from the environment.

2 N. Priyadarshini et al.

KeywordsActinides · Nanomaterials · Speciation · Separation and radionuclides

Introduction

The actinides comprise 15 metallic elements with atomic numbers from 89 to103, i.e., actinium through lawrencium. Among them actinium, thorium, protactin-ium, and uranium exist naturally and the remaining so-called transuranium elementsare produced by man through a variety of nuclear transformation processes such asneutron irradiation, atom bombardment, and radioactive decay. The amounts of thenatural actinides on the earth are in the range of 3 � 1023 g for thorium and3 � 1022 g for uranium (Cornelis et al. 2005). Actinide series are categorized intotwo groups, the light actinides (Th, Pa, U, Np, Pu, Am) and heavy actinides (Cm, Bk,Cf, Es, Fm, Md, No, Lr). Usually light actinides are the main environmentalcontaminants associated with anthropogenic activities such as mining and millingof uranium ores, generation of nuclear energy, and storage of nuclear waste resultingfrom the manufacturing and testing of nuclear weapons (Nagasaki and Nakayama2015). Increased global demand for electricity production resulted in adoptingnuclear technologies which further led to generation of large quantities of radioactivewaste. Moreover majority of the radioactive wastes originate from nuclear reactorsand the suitable way to be dealt with is by reprocessing of spent nuclear fuel. Butreprocessing of spent fuel faces a major challenge due to hydrolysis of actinides thatprevails in low-acid conditions leading to many consequences and the important onebeing the criticality hazard due to accumulation of plutonium metal. For radioactivematerials stored in the ground or in geological formations, the most probabletransport medium is the aqueous phase and several experimental evidences indicatethat colloids and hydrolyzed species of actinides can be the major factor in rapidmigration of radioactivity in groundwater. The low acidity of groundwater favors theformation of above-said species and thus provides a potential transport pathway formigration of actinides and mainly plutonium away from the repository (Cleveland1967). That is the reason why the radioactive waste is a serious environmentalproblem for which there is, as yet, no universally accepted solution. Thus a com-prehensive understanding of the complex chemistry of actinides in aqueous media isrequired. There are two important factors which need deep insights in order to avoidthese consequences. Firstly speciation studies have to be carried out. Secondlyadvance methods have to be identified for the extraction of the actinide speciesprevailing in the environment. Most of the speciation studies are carried out withactinides from thorium to curium and the concentration range for speciation studieswith actinides is mostly below 1 � 10�5 M (Rand et al. 2007). At high concentra-tions, the radioactivity of these elements can cause radiolysis of actinide species inthe environment. Actinides can exist in variety of oxidation states and most

Actinide Speciation in Environment and Their Separation Using Functionalized. . . 3

commonly in trivalent and tetravalent states. The element thorium exists only in thetetravalent oxidation state. In this oxidation state, the solubility generally is very lowand hydrolysis is the most important reaction (Knope and Soderholm 2013b). Theelements uranium, neptunium, and plutonium show a wide variety of oxidationstates. Pentavalent ions are known from the elements protactinium to americium.In the pentavalent oxidation state, oxo-cations are formed. Uranium, neptunium, andplutonium are stable in the +6 oxidation state. A stable form of a heptavalent ion isknown from neptunium, and plutonium can also be oxidized to the octavalentoxidation state (Tananaev et al. 2007). Actinide ions in aqueous medium often arenot in a state of thermodynamic equilibrium and their solubility and migrationbehavior are related to the form in which the nuclides are introduced into the aquaticsystem. The solubility, transport properties, bioavailability, and toxicity of actinidesare dependent on their speciation, composition, oxidation state, molecular levelstructure, and nature of the phase in which the contaminant element or moleculeoccurs. The most important reactions of actinides in environment are hydrolysis andcomplexation with inorganic ligands such as carbonate, sulfate, phosphate, chloride,and organic ligands as well as humic substances (Henry et al. 1992).

For the extraction or removal of radioactive nuclides from the aqueous environ-ment, a variety of technologies have been investigated. To mention a few solventextraction, precipitation, membrane filtration, sorption, electrodeposition, steam/solar evaporation, and phytoremediation are appropriate methods (Weigel et al.1997). Among these, sorption, precipitation, and membrane filtration techniquehave been widely used because they are easily carried out, economical, and can beapplied on a large scale for practical applications. The selection of an appropriatetechnology for radioactive waste processing depends mainly on the characteristics ofthe radioactive waste, the scale of waste production, and the final form of the waste(Nagasaki and Nakayama 2015). Advances in synthetics chemistry such as inventionof nanomaterials, nanostructured materials, and nanocomposite are providingunprecedented opportunities to develop effective separation processes for treatmentof radioactive waste. Nanomaterials with their fascinating physiochemical propertiescan be used to circumvent many limitations of bulk material and thus represent aparadigm shift in our thinking. Nanomaterials differ from bulk materials not only insize, but also in terms of physicochemical properties that can be used to develop newtechnologies and improve the existing ones (Shi et al. 2012).

This chapter is mainly contributed for the detailed discussion of speciation ofactinides in the aqueous systems which are considered as a most prevailing conditionof the environment. It also discusses the methods used to remove the actinide speciesfrom the environment by using nanotechnology.


Background: Hydrolysis of Metal Ions

The chemical behavior of metallic element in aqueous solution is determined by thenature of the ionic or molecular species that it forms. In the absence of stronglycomplexing ligands and depending on the acidity of the solution the simple cation ofthe element often reacts with water itself (hydrolyzes) to form hydroxo complexes.This is due to the fact that in aqueous medium, the metal cations Mz+ are solvated bydipolar water molecules giving rise to aquo ions [M(OH2)n]

z+. Charge transferoccurs through the M-OH2 σ bond. Electron density is transferred from molecularorbital of coordinated water molecule to empty orbitals of metal cation. This resultsin the weakening of O-H bond and finally deprotonation takes place leading towardsthe formation of bonds with oxygen and OH� ligand (Henry et al. 1992). Hydroxideions are always present in water at concentrations varying from>1 to<10�14 M as aresult of small self-dissociation constant of water. The resulting complexes may becations, neutral molecules, or anions. They may be mononuclear or polynuclear.That is, they may contain one metal atom or several (Baes and Mesmer 1976). Theability of hydrated metal ions to hydrolyze depends on the ionic potential (charge tosize ratio) of the metal. Small and highly charged metal ions are highly prone tohydrolysis. The hydrolysis constant increases in the order M4+ >M6+ >M5+ >M3+

mainly due to decrease in ionic radius and increase in Lewis acidity of the metal ion(Altmaier et al. 2013). The initial step in the hydrolysis of a cation is usually theformation of the mononuclear species MOH(z-1)+ and in general it is represented as

Mzþ þ H2O , MOH z�1ð Þþ þ Hþ (1)

Complete reaction is represented by including water of solvation:

M OH2ð Þzþn , M OHð Þ OH2ð Þ z�1ð Þn�1 þ Hþ (2)

In the above reaction a proton is lost by a solvating water molecule and on furtherhydrolysis,

M OHð Þ OH2ð Þ z�1ð Þn�1 ! M OHð Þ2 H2Oð Þz�2

n�2 þ Hþ (3)

M OHð Þ OH2ð Þ z�2ð Þn�2 ! M OHð Þ2 H2Oð Þz�3

n�3 þ Hþ (4)

and so on. But in reality complete hydrolysis does not happen; instead otherchemical reactions or aggregation of two or more hydrolyzed species takes placeand in general it is represented as follows.


(H2O)d-2MOH

OH2

(y-1)+

+ M(H2O)d-2

H2O

HO

(y-1)+

(H2O)d-2M M(H2O)d-2

O

O

H

H

(y-1)2+

+ 2H2O

ð5Þ

When this reaction further continues, then there is a possibility of formation ofhigh-molecular-weight polynuclear chains. The determination of the identity andstability of dissolved hydrolysis products has proven to be a difficult and challengingtask primarily for two reasons:

I. The hydroxide complexes formed are often polynuclear, that is, they containmore than one metal ion. It will be readily perceived that this can result in theformation of a greater variety of species during the hydrolysis of cations. Morehydrolysis products present simultaneously in appreciable amounts.

II. The range of pH over which the formation of soluble hydrolysis products can bestudied is often limited by the precipitation of the hydroxide or the oxide of themetal cation (Baes and Mesmer 1981).

Hydrolysis of Actinides

Hydrolysis reactions are the primary complexation reactions of actinide elements inaqueous solution. It is mainly correlated with the electrostatic interaction energybetween the actinide ion and the OH� ligand (Choppin 1983; Grenthe andPuigdomenech 1997; Neck et al. 2001). The general order is Pu4+ > Np4+ >U4+ > Pa4+ > Th4+. The effective charge of the actinide ions decreases in theorder of An4+ > AnO2

2+ > An3+ > AnO2+. Due to high electric charge, tetravalentactinides can hydrolyze even under very acidic conditions (pH < 3 depending onactinide) (Walther et al. 2007). In the near-neutral pH, An(OH)(aq) species dominatesin the absence of other complexing ligands. Actinide ions in the trivalent andtetravalent states in acidic solutions are in the form of the simple hydrated ionsAn3+ and An4+, whereas An5+ and An6+ exist as trans-dioxo cations. This is due tolarge positive charges which make them readily strip oxygen atoms from water


molecules. They are extremely stable and have a symmetric and nearly linearstructure [O = An = O]n+ where n = 1,2. This structure decreases the effectivecharge on the central actinide ion making it less prone for hydrolysis compared totetravalent actinides (Choppin 1999).

Plutonium

Among the actinides, plutonium is the most chemically diverse and fascinatingelement in the periodic table. Plutonium is generated by transmutation of uraniumin nuclear reactors in large scale. Plutonium has 18 isotopes. The most importantamong these are 238Pu (t1/2 = 86 years) and 239Pu (t1/2 = 24,110 years) (Clark et al.2006). The ground-state electronic configuration of atomic plutonium is [Ru]5f67s2.Plutonium ions in solution exist in +3, +4, +5, and +6 oxidation states as Pu3+, Pu4+,PuO2

+, and PuO22+ (Cleveland 1970). The chemistry of plutonium is in great

contrast to the light elements of periodic table and it is very difficult to control.Some of the complexities of plutonium chemistry in aqueous medium are listed asfollows.

Complexities in Plutonium Aqueous Chemistry

Three reaction pathways have to be considered while investigating plutonium inaqueous systems: hydrolysis, polymerization and colloid formation, and redoxreactions of different oxidation states of plutonium. It is difficult to study oneparticular reaction without the interference of the other. Because all the reactiontakes place simultaneously, that is, mainly due to the same redox potentials of thecouples, Pu (III)/Pu (IV) (E0 = �1.031 V) and Pu (IV)/Pu (V) (E0 = �0.936 V), anequilibrium of two or more oxidation state may occur in solution. Differences inredox potentials for different oxidation states of plutonium are given in Fig. 1.

In other words, plutonium acts both as oxidizing and reducing agent at the sametime. This is called disproportionation reaction. The disproportionation reactionmainly happens at elevated temperature. The equations governing the redox reac-tions for plutonium ions under acidic conditions are given as Eqs. 6, 7 and 8:

2Pu4þ þ 2H2O , Pu3þ þ PuOþ2 þ 4Hþ (6)

Pu4þ þ PuOþ2 , Pu3þ þ PuO2þ

2 (7)

2PuOþ2 þ 4Hþ , Pu4þ þ PuO2þ

2 þ 2H2O (8)

• Conversely, under some conditions, two plutonium ions of different oxidationstates can react by means of a reproportionation reaction. The two ions aresimultaneously oxidized and reduced to form two ions of the same oxidationstate.


• Another important complexity is that all plutonium isotopes are radioactive. Onemilligram of plutonium emits about 106 alpha particles per second, and theradioactive decay is constantly adding energy to the plutonium solution. Thisleads to radiolytic decomposition of water-producing redox reagents such asshort-lived radicals �H, �OH, and �O. These radicals recombine to form H2, O2,and H2O2. Thus radiolysis tends to reduce Pu(VI) and Pu(V) to Pu(IV) and Pu(III)states (Clark 2000).

Hydrolysis and Polymerization of Tetravalent Plutonium

The tendency towards hydrolysis, polymerization, and further to colloid formation isstrongest for tetravalent plutonium as compared to other oxidation states due to itshigh effective charge. The effective charges for different oxidation state of pluto-nium are Pu4+(~ + 4)> PuO2

2+ (~ + 3.3)> Pu3+ (~ + 3)> PuO2+ (~ + 2.2). Pu(IV) is

stable only at very high acidities and considerable amounts of Pu(III), Pu(V), and Pu(VI) are formed at pH > 0 as shown in Fig. 2 (Walther 2008).

Only at pH< 0.5, Pu4+(aq) is the dominant species whereas with increasing pH thehydrolysis reaction leads to the quantitative formation of mononuclear complexes Pu(OH)n

4�n where n = 1–4. Rapid polymer formation takes place even in 1 � 10�3

and 1� 10�4 M Pu(IV) solutions and close to solubility (Kumar and Koganti 1997).The polymeric plutonium refers to a hydrolytic form which is characterized by itsbright emerald green color. The only means by which polymer formation can beavoided is to detect colloids that are formed initially in the solution. This is becausethe freshly formed polymer can be depolymerized where as aged ones are difficultdue to conversion of amorphous to crystalline form which involves the modificationof hydroxyl bridges to oxo bridges as shown in Eq. 9 (Seaborg and Loveland 1990).

Pu(VI)

PuO22+

Pu(V)PuO2+

Pu(IV)

Pu4+

Pu(III)Pu3+ Pu(0)

0.99 V -1.25 V

0.94 V 1.04 V 1.01 V -2.00 V

Fig. 1 The redox potential differences for the plutonium aquo ions in 1 M perchloric acid, as wellas the potential difference between the plutonium aquo ions and pure Pu (Cleveland 1970)


xPu(OH)n4-n - yH 2O

Pu

O

O

H

H

Pu

O

O

H

H

Pu

m+

- yH 2O

Pu

O

O

Pu

O

O

Pu

m+

Hydroxy bridged

oxo bridged

- yH 2OPuO2

ð9Þ

Parameters Influencing the Formation of Plutonium Polymers/Colloids

The formation and properties of polymers and colloidal hydroxides are directlydependent on pH of the solution. At alkaline pH, some hydroxides can form anioniccomplexes and possibly polymers. With increase in pH the degree of polymerizationand formation of colloidal aggregates also increases (Schuelein 1975). Ionic strength

Fig. 2 Plutonium oxidation-state distribution dependingon –log[H+] [21] (Walther2008)


of the electrolyte affects the solubility. Increase in ionic strength increases thecoagulation. Concentration of metal ion has direct effect on polymerization andcolloid formation. With decrease in Pu(IV) concentration, the pH of colloid forma-tion increases (Bitea et al. 2003b). The rate of polymerization is a function oftemperature. Polymerization can occur even at room temperature in solution lessacidic than 0.3 M H+ and in 1.26 M H+ at boiling temperatures. The amount ofpolymer formed at 75 �C is ~3.5 times higher than the amount at 25 �C when theinitial acidity is 0.075 M (Costanzo et al. 1973).

The free energy of plutonium polymer is estimated as about �341.9 kcalmol�1

whose solubility product of Pu(OH)4 corresponding to this number is 2.5 � 10�56

(Silver 1983). This shows that the polymerization phenomenon is a very rapid andirreversible process (Toth et al. 1981). Under conditions of low acidity and elevatedtemperature, the polymeric species require drastic treatment for conversion to theionic state. The presence of low concentrations of Pu in natural waters and indisposal streams is of much interest to those concerned with reactor and radiationsafety. These solutions are usually nearly neutral and any plutonium is probably inthe polymeric state.

Formation of plutonium colloids can be detected by spectroscopic studies as eachoxidation state of plutonium and polynuclear plutonium has its own characteristiccolor (Fig. 3a). The absorption spectrum of polymeric Pu(IV) state is different fromthose of other plutonium species (Fig. 3b). At pH 0–2, there is a steady decrease inthe characteristic absorption band at 470 nm characteristic of ionic Pu(IV). They alsoshow relatively broad, low-intensity bands and rather intense with increasing absorp-tion below 460 nm (Rabideau and Kline 1960). Polymerization of Pu(IV) leads toalterations in redox equilibria and additional easily identifiable spectral featurearound 620 nm appears for the presence of small Pu(IV) colloids. Absorption spectraof aged polymers do not show any difference with that of fresh polymers except thepronounced 620 nm peak (Cleveland 1967).

Probing the Formation and Structure of Pu(IV) Polymer

Though first step in the formation of polynuclear species is the condensation ofmononuclear species, there are no reports on the solution structure of Pu(IV) mononuclear hydroxides. As an evidence for the presence of dimer, an X-raydiffraction study has reported a solid-state structure of Pu2(OH)2(SO4)3.4H2O withdimeric [Pu2(OH)2]

6+ units linked by bridging sulfate ligands (Wester 1982).Electrospray ionization mass spectrometry (ESI-MS) studies on dimers, trimers,and tetramers revealed the presence of mixed oxidation states of plutonium, i.e.,Pu(III) and Pu(V) (Walther et al. 2009). This gives an insight into the coexistence ofmonomeric and pentavalent plutonium species with colloids and also their interfer-ence in redox reactions (Newton et al. 1985; Rai and Swanson 1981; Neck et al.2007). A hexanuclear [Pu6(OH)4O4]

12+ core with glycine ligands containing bothoxo and hydroxo bridges was also reported (Knope and Soderholm 2013a). Thesame linkage was also found in case of lanthanides (Wang et al. 2000). The


formation of the polynuclear species is assumed to involve hydroxide or oxygenbridging of plutonium ions but some reviews consider such species to be thermo-dynamically unstable (Lemire et al. 2001; Guillaumont et al. 2003). As an alternativethey favor colloid formation from mononuclear hydroxide complexes. Thus Pu(IV) colloids are more stable in contrast to other tetravalent species such as Th(IV) and U(IV). The most widely accepted mechanism of formation of plutoniumcolloid involves the condensation of [Pu(OH)n]

(4�n)+ through an olation reaction toyield hydroxo-bridged species. On further condensation, the hydroxo-bridged olig-omers produce mixed plutonium oxide hydroxides (Knope and Soderholm 2013b).X-ray absorption fine structure (XAFS) and laser-induced breakdown detection(LIBD) measurements proposed that the formation of Pu(IV) colloids follows bystacking eightfold coordinated Pu units. It involves the formation of a trinuclearspecies through hydrolysis and condensation of a monomeric Pu(OH)2(H2O)6

2+ unitwith a binuclear species to form single-edge sharing, leading to the formation ofneutrally charged species as product or a double-edge sharing species with commoncorner. The dimers and trimers agglomerate and eventually form nm-sized colloids(Fig. 4). The predominant species contain highly asymmetric oxygen coordinationwhich indicates the presence of different Pu-O bond lengths from different coordi-nated oxygen atoms (–O, –OH, –OH2). PuxOy(OH)4x�2y(H2O)z (0 � y � 2x) hasbeen proposed as colloid structure formed by stacking of mononuclear or polynu-clear building blocks on top of each other.

During this process, the cubic subunits are preserved and form a distorted fluoritelike plutonium lattice (Rothe et al. 2004). Single crystals of 38 plutonium nano-clusters [Li14(H2O)20[Pu38O56Cl54(H2O)8] and Li2[Pu38O56Cl42(H2O)]�15H2O]

100

0

10

400

abso

rptiv

ity (

M–1

cm–1

)

500 600 700wavelength (nm)

800

20

30

40

50

60

a

b

Pu(IV) Pu(III)

colloids

Pu(V)

Pu(VI)

000

Fig. 3 (a) Characteristic colors of each oxidation state of Pu and Pu colloids (Clark 2000). (b)Absorption spectra of different oxidation state of Pu and Pu colloids (Rabideau and Kline 1960)


were isolated for structural characterization using single-crystal X-ray diffraction.Both structures consist of the [Pu38O56]

40+ core, wherein 38 8-coordinate Pu(IV) cations are bridged exclusively via oxo linkages. The data clearly shows thatintracluster packing and structural topology of clusters of composition[Pu38O56Cl54(H2O)8]

40+ are identical with PuO2 and the surface is occupied bychloride ion and water molecules (Soderholm et al. 2008; Wilson et al. 2011).However the solid-state structure of the crystal does not contain Pu-OH moietieswhich contradict the structure proposed by Rothe et al. (2004). Small-angle neutronscattering (SANS) experiments were carried out on colloidal Pu(IV) suspensions inaqueous phase and polymer extracted into C6D6 solutions by alkyl esters of

Fig. 4 Schematics showing mechanism of hydrolysis and polymerization of Pu4+(aq) leading toformation of nm-size colloids Rothe et al. (2004)


phosphoric acid. The results revealed the presence of long, thin rodlike polymer inboth the phases. However the diameter of ellipsoids of extractant-Pu(IV) polymercomplexes in C6D6 is larger than that of aqueous polymer. This increase in diameteris due to the attachment of extractant molecules around the rod-shaped Pu polymer.The length of aqueous Pu(IV) polymer is larger than the polymer extracted in C6D6

(Thiyagarajan et al. 1990). A comparative study was also made between freshlyprepared and 5-year-old Pu(IV) colloids using Pu L3-edge extended X-ray absorp-tion fine structure (EXAFS). The aged colloidal particles contain only 3–4 Pu atomswith a structure very similar to solid Pu(IV) oxide but with shorter Pu-O and Pu-Pudistances which was caused due to partial oxidation of Pu(IV) to Pu(V)/Pu(VI) (Ekberg et al. 2013).

Size and Morphology of Pu(IV) Colloid

Polymeric plutonium was observed in two forms: amorphous and crystalline primaryparticles and secondary particles which appeared to be aggregates of the primaryparticles. X-ray diffraction patterns of precipitated plutonium show a faint PuO2

structure, suggesting that the hydroxide is in reality hydrated PuO2 rather than Pu(IV) hydroxide. The precipitate becomes more crystalline when aged at elevatedtemperatures (Haire et al. 1971). Similarly the precipitates aged in a basic or neutralmedium display a crystalline pattern that corresponds to the cubic PuO2 structure andwas confirmed by electron diffraction pattern. Colloids formed in highly acidicmedia differ from those formed at higher pH (Bell et al. 1973). Thus colloidmorphology varies strongly depending on the method of preparation, conditions offormation, and age of the solutions. For instance, fast neutralization of Pu(IV) solutions with a base or water usually forms less ordered structures (Reedet al. 2006). Scanning electron microscopy (SEM) and transmission electron micro-scopic (TEM) images show amorphous morphology for these polymers (Neu et al.1997). The IR spectrum of Pu(IV) polymer showed major bands at 360 and3400 cm�1. They are assigned to Pu-O vibration similar to crystalline PuO2 and toOH stretching vibration of a water group which is either hydrated or occluded in thepolymer structure. The comparison of the IR spectrum of PuO2 and the Pu(IV) polymer precipitated and dried in air showed close resemblance of the polymerbands in the 250–600 cm�1 region suggesting a similar lattice structure in the twocompounds. Unlike PuO2, water and possibly hydroxyl groups are present in thepolymer (Toth and Friedman 1978). The size of Pu(IV) colloids ranges from a fewnanometers to almost micrometers, depending on the conditions of generation(Lloyd and Haire 1978; Triay et al. 1991). Electron micrograph shows that theprimary particles are extremely small of the order of 5–20 Å whereas the amorphousprimary particles are of the order of 10 Å (Lloyd and Haire 1973). TEM analysis ofdried solutions shows evidence of small PuO2-like clusters with a diameter of about2 nm. LIBD experiments gave mean particle size of 5–12 nm. Particle size mayincrease to the point of precipitation. In acidic solutions (pH 0.4–1) the weightedmean colloid size increases from 12 to 25 nm with increasing degree of


oversaturation with respect to amorphous Pu(IV) hydroxide. They are said to becomposed of small crystalline particle covered by an amorphous layer (Walther andDenecke 2013). The small-angle X-ray diffraction (SAXS) and XRD affirmed thepresence of <40 Å PuO2 clusters. Dynamic light scattering (DLS) and field flowfractionation (FFF) indicate that the hydrodynamic radii of colloids vary on the orderof 20–200 Å. Also the colloid density is slightly lower than the density of PuO2

(Rundberg et al. 1987). As a matter of fact, the size of plutonium colloids shouldincrease with ageing time. But in contrary, a recent report suggested that there wasreduction in the size of 5-year-aged colloid due to partial oxidation of Pu(IV) resulting in shorter Pu-O and Pu-Pu bond distance. By means of TEM it wasobserved that Pu(IV) forms small spherical PuO2 colloid of 2 nm. A recent TEManalysis on intrinsic plutonium nanocolloids generated in the absence of goethite orquartz mineral shows 2–5 nm diameter (Fig. 5a, b) (Powell et al. 2011).

Hydrolysis and Solubility of Pu(IV)

The hydrolysis reactions of Pu4+ ions are generally as in Eq. 10:

xPu4þ þ nH2O , Pux OHð Þ4x�nn þ nHþ (10)

The hydrolysis constants Kxn0 in a given medium and Kxn

o (at infinite dilution) aregiven by Eq. 11:

K0xn ¼

Pux OHð Þ 4x�nð Þn

h iHþ½ �n

Pu4þ� �x ¼ Ko

xn γPu4þð Þx awð Þn

γPux OHð Þ4x�nn

� �γHð Þn

(11)

Fig. 5 TEM image of (a) intrinsic Pu nanocolloid cluster. (b) TEM image of intrinsic Punanocolloid and tubular crystal of goethite (Powell et al. 2011)


And the corresponding formation constants βxy0 and βxyo:

β0xn ¼Pux OHð Þ4x�n

n

h i

Pu4þ� �x

OH�½ �n ¼βoxn γPu4þð Þx γOH�ð Þn

γPux OHð Þ4x�nn

� �γHð Þn

(12)

γ is the activity coefficient and aw is the activity of water. The value of log β011 wasfound as 14.6 from experiments at trace levels of plutonium where the formation ofpolymeric species is less likely (Metivier and Guillaumont 1972). Solubility oftetravalent actinide is an important factor to be considered for mobility of plutoniumcolloids in aquifer systems. Since Pu(IV) hydroxide phases or colloids span a widerange of structural features and surface properties, such as amorphous/crystallinities,size distributions (Fanghanel and Neck 2002), and stabilities it therefore exhibits awide range of solubilities (Neck and Kim 2001; Runde 2000) and thus the reportedsolubility product shows considerable discrepancies (Kim and Kanellekopulos 1989;Rai and Ryan 1982; Strickert et al. 1984). In general the solubility products of theamorphous An(IV) hydroxide decrease in the series, Th(IV) > U(IV) > Np(IV) > Pu(IV), due to decrease in An-O distance in the lattice (Neck and Kim2001; Guillaumont et al. 2003). The same trend is observed for crystalline An(IV) dioxides. The solubility products Ksp

0 and Kspo (at infinite dilution) of amor-

phous Pu(IV) oxide or hydroxide, Pu(OH)4(am), are given by Eq. 13:

Pu OHð Þ4 amð Þ , Pu4þ þ 4OH- (13)

Ksp0 ¼ Pu4þ

� �OH�½ �4 and Ksp

o ¼ Ksp0 γPuð Þ γOHð Þ4

log Kspo for crystalline PuO2 is�64.0 and for amorphous hydrated hydroxide�58.3

[30]. Thus crystalline PuO2 is far less soluble than its amorphous hydroxide. Theslope of the solubility curve obtained from LIBD studies (Fig. 6) provides furtherevidence that the dihydroxo-complex (Pu(OH)2

2+) is the prominent species in therange of 1.0< pHc< 1.6. The slope= �2 for the solubility curve infers that colloidformation proceeds from Pu(OH)2

2+ via consumption of 2OH� per Pu(IV) ionsuggesting Eq. 14 (Walther et al. 2007):

Pu OHð Þ22þ þ 2OH� $ Pu OHð Þ4 am=collð Þ (14)

Methods Adopted to Remove Polymers

Two ways can be adopted to avoid the polymers formed in the system. They can beeither extracted or depolymerized.


Extraction of Pu(IV) Polymer/Colloid

Attempts were made to extract plutonium polymers or colloids using neutral extra-ctant TOPO (trioctylphosphine oxide), CMPO (octylphenyl-N,N-diisobutylcarbamoylmethylphosphine oxide), and the bifunctional extractantsDHDECMP (dihexyl-N,N-diethylcarbamoylmethyl phosphonate). But in most ofthe cases there is crud formation. They also co-extract other oxidation-state metals(Muscatello et al. 1983). Back extraction of the polymer either was done in a silicasol or required large concentrations of salts to reverse the extraction (Chaiko 1992).A novel approach was made by dissolving the [Pu38O56Cl54(H2O)8]

14� nanoclusterin aqueous solution of LiCl producing green solution of colloidal Pu which wasconfirmed by the optical spectra. Addition of 6 M HCl changes green-coloredsolution to red with change in the spectrum. The rapid change involves ion- orligand-exchange reaction between water and Cl�. But the core moiety Pu38O56

remains unperturbed. Trichloroacetic acid (TCA) in n-octanol was chosen as extra-ctant due to its immiscibility with water. The well-defined surface of Pu colloids canbe altered and manipulated which can involve in ion- and ligand-exchange reactionsresulting in selective extraction of only plutonium colloid leaving other plutoniumspecies in aqueous phase.

Fig. 6 Solubility of amorphous Pu(OH)4(am) comparing available literature data Walther et al.(2007)


Depolymerization

Depolymerization of plutonium polymer is a slow process. Freshly formed polymercan be depolymerized whereas aged or polymers heated to higher temperatures aredifficult to depolymerize and it needs drastic conditions. In other words it leads tovery low depolymerization rates due to conversion of amorphous to crystallineprimary particles which involves more extensive cross-linking (Hunter and Ross1991; Brunstad 1959). Some of the reported methods applied for depolymerizing thepolymer if formed accidentally are listed below.

Proposed Methods for Depolymerizing Pu(IV) Polymer

• Pu(IV) polymers can be depolymerized in the presence of reductants and oxidantssuch as hydroquinone, H2O2, hydroxylamine, U(IV) and KMnO4, H2O2 + Fe(NO3)3, Na2S2O8 + catalysts, and Co(III) in 0.533 M HNO3. This is because Pu(III) and Pu(VI) formed due to reduction and oxidation are not prone to polymer-ization (Ermolaev et al. 2001; Oak et al. 1983).

• Passing direct current through an aqueous nitric acid solution which contains Pu(IV) polymer converts the polymer to Pu3+ and PuO2

2+ ions in turn are convertedto Pu4+ ions in the solution, followed by the step in which Pu4+ ions are extractedfrom the solution. This can be done by contacting the acid solution with a solutioncontaining about 30% tri-n-butyl phosphate and 70% hydrocarbon diluent (suchas dodecane). It can also be done with ion-exchange resin (Tallent et al. 1982).

• To avoid the complications of colloid formation, plutonium has often beeninjected as the citrate complex, prepared by adding Pu(NO3)4 in concentratednitric acid to a 0.3–2.0% solution of trisodium citrate followed by adjustment withbase to pH 4–7. The rate of depolymerization is directly proportional to citrateconcentration, time, and acidity (Lindenbaum and Westfall 1965).

• Depolymerization is accelerated by introducing fluoride, sulfate, and other ionsforming strong complexes with Pu4+ (Weigel et al. 1997).

• The process of depolymerization is accelerated on exposure to UV radiation (Belland Friedman 1976; Friedman et al. 1977).

Uranyl Nitrate Effect

• Solutions of plutonium are commonly used in the presence of large amounts ofU. Therefore the effect of uranyl ion on the polymer chemistry is of considerableinterest. Much uncertainty exists about the effect of UO2(NO3)2 on the formationrate of Pu(IV) polymer because this solute provides nitrate ions which couldeither complex with Pu(IV) and stabilize it with respect to hydrolysis andpolymerization or, through complexation with Pu(IV), shift the disproportion-ation equilibrium to the left, thereby decreasing the real acid concentration of thesolution and causing a corresponding increase in the rate of hydrolysis and


polymerization. UO22+ functions as a chain terminator through the formation of

hydroxyl bridges to terminal uranyl groups (Quiles and Burneau 1998). Also, Pu(IV) polymer formed in the presence of uranium is shorter than those which areformed in the absence of uranium. This shows that presence of uranium in apolymerizing Pu(IV) solution reduces the rate of polymer formation about 30%without appreciably entering the polymer.

• Costanzo and Biggers (1963) prepared polymer in 5 N HNO3 and aged for severalmonths. It is observed that for aged polymers the depolymerization halftime is320 h and it is only 20 h for freshly prepared polymer.

• Depolymerization occurs easily at 90 �C in 6 M HNO3. Also, Pu(IV) polymer canbe dissolved using Na2S2O8. Decomposition of S2O8 yields H2SO4 favoring Pu(IV) depolymerization (Savage and Kyffin 1986).

Hydrolysis of Th(IV)

Th(IV) exhibits a strong tendency towards hydrolysis and subsequent polymeriza-tion over a wide range of pH and concentrations. Thorium has an ionic radius of1.09 Å in nine coordinations making it the largest stable tetravalent metal ion and inspite of its high valence it is least susceptible to hydrolyze when compared to U(IV) or Pu(IV) (Torapava et al. 2009). Many studies have reported Th(IV) hydrolysisand the species formed in solution depending on pH and concentration. Th4+(aq) ioncan be found as a predominant species only at pH < 3. It has been observed thatmononuclear ionic complexes dominate the species distribution at very low concen-trations as well as at high acidity (Khazaei et al. 2011). At near-neutral pH, and as thethorium concentration approaches the solubility limit of the amorphous hydroxideTh(OH)4, polymers become the dominant species in solution, starting with dimerswhich grow and form pentamers. Pentamers have been found to be the mostabundant and stable complex rather than tetramers and hexamers (Walther et al.2012). The polymeric species are found to contain one or more thorium atoms linkedby hydroxy bridges (Bacon and Brown 1969). With increasing pH and thoriumconcentration it forms thorium nanoparticles or the so-called colloids.

Probing the Formation and Structure of Th(IV) Hydrolyzed Species

Th(IV) hydrolysis and polymerization reactions follow the same pattern as Pu(IV).Th(IV) colloid formation is preceded by polynuclear hydroxide complexes such asTh2(OH)2

6+, Th2(OH)35+, Th4(OH)8

8+, Th4(OH)128+, Th6(OH)14

10+, andTh6(OH)15

9+. But in contrast to Pu(IV) polynuclear hydroxide complexes, Th(IV) polynuclear hydroxide complexes are thermodynamically unstable (Waltheret al. 2008). Structural studies on aqueous solution of the hydrolysis products of Th(IV) have identified three different types of hydrolysis species: a μ2O-hydroxodimer, [Th2(OH)2(H2O)12]

6+; a μ2O-hydroxo tetramer, [Th4(OH)8(H2O)16]8+; and

a μ3O-oxo hexamer, [Th6O8(H2O)n]8+ (Rand et al. 2007). Single crystals of oxo- and


hydroxo-bridged haxameric [Th6(μ3-O)4(μ3-OH)4]12+ (Takao et al. 2009; Knopeet al. 2011) and octameric [Th8O4(OH)8]

16+ (Knope et al. 2012) have also beenisolated by addition of complexing ligands. SAXS investigation revealed the pres-ence of microcrystalline particles with the thorium oxide structure in highly hydro-lyzed Th(IV) solutions (Magini et al. 1976). In another study, when a thoriumchloride solution was left to evaporate at room temperature, it produced hydroxo-bridged Th dimers. Thorium structure of Th(IV) pentamers was investigated byXAFS and by quantum chemical calculations. It was considered that the mostfavorable structure contains two Th(IV) dimers linked by a central Th(IV) cationthrough hydroxide bridges. These polymers are then subjected to continuous hydro-lysis; that is, the number of hydroxide ligands increases with increase in pH. Theultimate end product of Th hydrolysis is an oxo-bridged ThO2 with fluorite structure(Sellers et al. 1954).

Size and Morphology of Hydrolyzed Th(IV) Polymer/Colloid

Freshly formed small particles have a mean diameter in the range of 16–23 nm(Rothe et al. 2002; Bundschuh et al. 2000). Three different oxides/oxo-hydroxidesolid phases have been reported for thorium, that is, crystalline ThO2 whose particlesize is >50 nm, microcrystalline ThO2.xH2O of 15–30 nm, and amorphous ThO2.xH2O of 2–5 nm. Upon heating a highly hydrolyzed thorium solution the particlesize was estimated to reach as large as 20–50 Å in diameter. But if that same solutionwas left to age without evaporation it produced small ThO2 particles with sizes ofabout 20 Å. The small 20 Å ThO2 particles are X-ray amorphous (Neck et al. 2002).DLS measurement shows that well-defined colloids were formed at two different pHdomains. The colloids formed at pH~0.8 were smaller, when compared to thoseformed at pH~2 (Priyadarshini et al. 2016).

Hydrolysis and Solubility of Th(IV)

M-O bond length of Th(IV) is largest compared to Pu(IV) due to actinide contractionand hence anionic ligand forms much stronger for Pu(IV) than for Th(IV). Accord-ingly the first hydrolysis constant of Th(IV) is smaller than Pu(IV). Some of thereported values of formation constants for the first mononuclear hydroxide complexM(OH)3+ are logβ1,1� = 11.8 � 0.2 and logβ1,1� = 11.5 � 0.5. In order to measurethe solubilities, two regions with respect to pH and Th(IV) concentration need to beconsidered, which correspond to different solid phases, microcrystalline and X-rayamorphous ThO2 (Ekberg et al. 2000). Rai et al. (2000) determined the solubilityproduct of crystalline ThO2 in acidic solution (pH< 2.5) as log Ksp

o=�54.2� 1.3.A combined coulometric and pH titration in the pH range 1.5–2.5 with LIBD gave asolubility product of log Ksp

o = �52.8 � 0.3. The ambiguity in the solubility data ismainly due to the presence of polynuclear species or intrinsic colloids prevailing inthe system. With increase in pH and close to the solubility limit, where Th(OH)4 is


the predominant aqueous species, the solubility product of crystalline ThO2 issimilar to that of amorphous ThO2 (log Ksp�(Th(OH)4 (am;hyd)) = �47.8).Hence it can be concluded that in near-neutral solutions the bulk crystalline ThO2

is covered with an amorphous and more soluble surface layer. Solubility studiesconducted on Th(IV) intrinsic colloids in concentrated NaCl (0.5 and 5.0 M) andMgCl2 (0.25, 2.5, and 4.5 M) solutions within the pH range 8.8–10.8 revealed thatthe hydrophilic oxo-hydroxide intrinsic colloids formed through chemical poly-nucleation reactions are stable equilibrium species and contribute to the total solu-bility of Th(IV).

Stability of Th(IV) Hydrolyzed Colloids

Laser-induced breakdown (LIBD) spectroscopy combined with ultrafiltration wasused to investigate the generation of Th(IV) colloids in the concentration range from10�5 to 10�2 M at pH 3–5 in 0.5 M NaCl. Surprisingly the colloids generated bycoulometric titration are found to be small in size and remain stable up to 400 days ofinvestigation, without a tendency towards agglomeration or precipitation. There wasno change in size and concentration of colloids during the course of investigationwhich shows their extreme stability. These suspensions on dilution lead to anequilibrium between colloids and ionic species (Bitea et al. 2003a).

Uranyl Effect on Polymerization of Th(IV)

Previously it was shown that the presence of uranyl ion retards the rate of Pu(IV) hydrous polymer formation. Raman spectroscopic studies were done on Th(IV) in order to characterize the uranyl-thorium(IV) interaction. By monitoring thechanges in the frequency of the uranyl symmetric stretching vibration, that is, theshift in 869 cm�1, Raman band corresponding to unassociated UO2

2+ to 851 cm�1

for UO22+ attached to Th(IV) polymer. In the same solutions when aged, there was

change in bridging structure (i.e., hydroxyl to oxygen) which was confirmed by theappearance of band at 665 cm�1 and by the change in color of dispersed polymerfrom a uranyl pale yellow to a urinate golden orange (Toth et al. 1984).

Molecular Weight of Th(IV) Hydrolyzed Polymer

Static light scattering measurements showed that the Th(IV) colloids formed at pH~0.8 have a molecular weight of ~3449 Da and it shows that ~15 Th atoms arepresent in the polymeric network at the initial stage. The second virial coefficient isnegative. It implies that there was aggregation in the system leading to precipitation.The colloids formed at acidic pH tend to agglomerate causing precipitation in thesystem. But, the colloids formed at pH > 2.0 have a molecular weight of ~6223 Dawith positive second virial coefficient. This polymeric colloid formed due to


polynucleation contains around 20 atoms of Th when initially formed. The solute-solvent interaction in the system is higher leading to greater stability of the system.The colloids upon ageing for 7 days did not show any considerable change in theirmolecular weight but when it was aged for 55 days the molecular weight increased to43,573 Da (Priyadarshini et al. 2016).

Uranium

In nuclear reactors, uranium is handled in large quantities when compared to otheractinide elements. The behavior of uranium in radioactive waste repositories and themigration of uranium in the environment depend on its oxidation state and itsspeciation is determined by pH, presence of complexing anions, and total saltbackground. Uranium exists in different oxidation states (from +3 to +6). Theknown redox states of uranium are U(II), U(III), U(IV), U(V), and U(VI) and theircorresponding reduction potentials are given in Fig. 7.

Uranium(III) can exist only under anaerobic and strongly reducing conditions.U(IV) is also considered as toxic waste under strongly reducing conditions that isoften accepted to be present in deep geological repositories. U(V) is unstable anddisproportionates. U(VI) is the most common species in the environment. Undernatural aquatic systems, uranium exists as U(IV) and U(VI).

Hydrolysis and Polymerization of Uranium(IV)

The known redox states of uranium are U(II), U(III), U(IV), U(V), and U(VI). U(IV) aqueous chemistry is of particular interest under reducing conditions whichprevails in both inside of nuclear waste repositories and under deep uranium mines.Due to hydrolysis, U(IV) occurs in the form of [U(OH)x]

(4�x)+. The hydrolysisreactions of U(IV) resulting in the formation of polynuclear hydrolysis complexesare given in Eq. 15:

x U H2Oð Þn� �4þ þ yH2O , Ux OHð Þy H2Oð Þn�y

h i 4x�yð Þþþ yH3O

þ (15)

U(IV) is readily oxidized to U(VI) when reducing conditions are not maintainedcarefully. Due to hydrolysis U(IV) occurs in the form of [U(OH)x]

(4�x)+. There arelimited literature data on hydrolysis studies on U(IV).

Probing the Formation and Structure of U(IV) Hydroxide/Oxide

In contrast to UO22+ for which several oxides and hydroxides are known, only few

U(IV) oxide/hydroxide structures are known. Uranium(IV) dioxide adopts a fluoritestructure. The compound prepared by hydrothermal hydrolysis of a U(IV) sulfate


solution at 100–150 �C showed a structure isostructural with that described forthorium and consists of infinite chains of hydroxo-bridged (U(OH)2)n

2n+ units withU�U distances of 3.90 Å.

Size and Morphology of U(IV) Colloid

Attempts were made to produce colloids of U(IV) by electrochemical reduction of U(VI). EXAFS investigations were performed and compared with solid UO2 andamorphous UO2.nH2O. They showed correspondence with the latter (Opel et al.2007). Around 20 nm colloidal particle was detected in 2.5 mM U(IV) solutions atpH ~3.6 by DLS experiment. It decreased to ~5 nm when the concentration of U(IV) was increased to 19 mM (Priyadarshini et al. 2014).

Hydrolysis and Solubility of U(IV)

The solubility product of crystalline UO2(cr) formed at pH ~1 and amorphoushydrous oxide UO2.xH2O(am) formed at pH ~3 was determined by combinedcoulometric titration and LIBID experiments as log Ksp

o = �59.6 � 1 and logKsp

o = �54.1 � 1 (Manfredi et al. 2006). Compounds containing U(IV) areinsoluble in mildly acidic and alkaline medium. U(OH)2SO4 was prepared at100–150 �C from the hydrothermal hydrolysis of a U(IV) sulfate solution. Thestructure is isostructural with that described for thorium and consists of infinitechains of hydroxo-bridged (U(OH)2)n

2n+ units (Qiu and Burns 2013).

Molecular Weight of Hydrolyzed U(IV) Polymer

Light scattering measurements performed on freshly formed polymers of U(IV)showed a weight average molecular weight (Mw) of 1820 Da and it increased to13,000 Da when aged for 3 days. Around 40–50 atoms of U are considered to bepresent in the aged polymer. Positive value of second virial coefficient shows that thesolute-solvent interaction is high leading to stable suspension of aged polymers(Priyadarshini et al. 2014).

UO22+ UO2

+U4+ U3+

0.28 V -1.35 V

0.19 V 0.37 V -0.61 V -4.7 V

U

-0.1 VU2+

-1.60 V

Fig. 7 The redox potential differences for the uranium ions and pure U


Hydrolysis of Uranium(VI)

Under aerobic conditions, hexavalent uranium is more relevant in aqueous medium.U(VI) is thermodynamically most stable form with relatively high solubility (Zankeret al. 2007; Zhao and Steward 1997). Being a hard Lewis acid, it shows stronginteractions with hard donor ligands resulting in a tendency to undergo hydrolysis,formation of soluble complexes with carbonates and phosphates, and the mostimportant is its ability to form colloids which will increase the migration behaviorin natural systems (Eliet and Bidoglio 1998). Hence it becomes mandatory to studythe speciation and related reaction products of uranium in aqueous systems, underwell-defined conditions such as pH, ionic strength, concentration, and presence ofother ligands. The hydrolysis equilibrium reaction of uranyl ion in general can bewritten as given in Eq. 16:

mUO2þ2 þ nH2O , UO2ð Þm OHð Þn

� � 2m�nð Þþ þ nHþ (16)

The solution chemistry of U(VI) is dominated by the linear dioxo cation UO22+

(Kirishima et al. 2004). In aerated aqueous solutions at pH � 2.5, the uranyl ion isvery stable. The hydrolysis of uranyl ion starts at pH values greater than 3 (Clarket al. 1999) and also if the concentration of uranyl exceeds 10�3 M (Steppert et al.2012). UO2OH

+ is the only mononuclear hydrolysis product present at appreciableconcentration up to pH ~7 in 10�5 M or more concentrated U(VI) solutions (Brookeret al. 1979). The well-established dinuclear species (UO2)2(OH)2

2+ is present overthe pH range of 3–6 and accounts for at most 40% of the total U(VI) in 10�2 Msolutions. (UO2)3(OH)5

+, (UO2)4(OH)7+, and (UO2)3(OH)7

� are the dominant poly-nuclear species from pH 5 to 9 in solutions of 10�2 to 10�5 M U(VI), accounting forup to 100% of the total uranium in solution. Potentiometric measurement (Palmerand Nguyen-Trung 1995), time-resolved fluorescence spectroscopy (Eliet et al.2000), Raman spectroscopy (Quiles and Burneau 2000), attenuated total reflectioninfrared (Muller et al. 2008), and electrospray mass spectrometry (Clark et al. 1999)studies reveal that in acidic solutions of pH levels (2 � pH � 5) UO2

2+,(UO2)2(OH)2

2+, and (UO2)3(OH)5+ are the predominant species. Complexes

(UO2)3(OH)7�, (UO2)3(OH)8

2�, (UO2)3(OH)104�, (UO2)3(OH)11

5�, and(UO2)3(OH)4

2� have also been suggested by various methods (Kirishima et al.2004). There is consensus that several polynuclear species coexist in alkalinesolution. A number of structures that contain chains or extended structures ofhydroxo- and oxo-bridged UO2

2+ have been reported.The hydrolyzed U(VI) species aggregate and reach colloidal dimensions and exist

as suspended sols in the aqueous medium. Particle size is an important property todefine a colloidal system. It influences the behavior of colloids during diffusion,aggregation, etc. DLS studies were performed on different concentrations of uranylsolution diluted to neutral pH. A peak around 32–36 nm was observed on all theconcentration range of U(VI) in DLS (Priyadarshini et al. 2014). Hence it has beenconfirmed that the colloids have size of ~32–36 nm. With increasing concentration


of uranyl ions the pH for colloid formation decreases considerably. The colloidsformed were not stable for a long period of time. There is no change in averageparticle size of the colloids as the concentration of U(VI) increases. Well-definedcolloids are formed when the concentration reaches the precipitation point. Themolecular weight of U(VI) colloid was determined as ~763 Da which was muchless when compared to U(IV) and Pu(IV) (Priyadarshini et al. 2014). The calculatedmolecular weight shows that there may be 2–3 units of uranyl groups bridged byhydroxyl groups. Hence it is only oligomer that has been formed. Figure 8 representsthe relative abundance of U(VI) species as a function of pH in the aqueous solutionat atmospheric normal environmental.

Nanomaterials for Separation of Actinides

Safe management of radioactive waste with minimum impact to environment isemerging as one of the major challenges in separation science. Accumulation oflarge volume of the radioactive wastes is highly risky. The risk is directly propor-tional to volume of the waste. Therefore, these waste streams need to be treated toreduce their activity to a level at which they are permitted as per national regulations.This leads to more demanding efforts in removal of radionuclides from aqueoussolution for environmental concern. It is one of the most urgent technologicalproblems that mankind face worldwide. Decay is the only natural way of reducingradioactivity. Since radionuclides have decay rates ranging from seconds to thou-sands of years, proper segregation of wastes depending on their half-lives andconditioning is an important factor. The major aim is to separate long-lived alpha-emitting actinide radionuclides. Hence, safe and efficient management of long-livedactinide is of utmost importance. Under this situation, versatile materials andtechnologies that can remove the radionuclides and remediate the environment areof severe demand. Although a wide range of separation techniques are employed for

100UO2

2+ UO2(CO3)(OH)3–

UO2(CO3)22–

β-UO2(OH)2(s)

(UO2)3(OH)5+

UO2(OH)+

UO2(CO3)34–

80

60

40

Rel

ativ

e d

istr

ibu

tio

n (

%)

20

03 4 5 6 7 8 9

pH10

Fig. 8. Speciation diagram ofU(VI) ions as a function of pHin aqueous solution atatmospheric condition (Scierzand Zanker 2009)


removal of radionuclide, amalgamation of nanotechnology into conventional tech-niques has revolutionized the separation processes. Nanomaterials offer high surfacearea, high sorption capacities, selectivity, and reusability (Dash and Chakravarty2017). In recent years, nanomaterials such as carbon nanostructures, meso porousmaterials, and nanopolymers/dendrimers have been extensively investigated foractinide separation. The potential application of aforementioned nanomaterials inremoval of long-lived actinides from aqueous media is briefly discussed in thischapter.

Carbon Nanostructures

Carbon nanostructures, including carbon nanotubes (CNTs), nanodiamonds, fuller-enes, graphene, and other carbon materials, have been the most effective materialsfor environmental remediation, and have attracted great attention recently. The largespecific surface area, the outstanding thermal and chemical stabilities, and the recentdevelopment in large-scale synthesis make carbon nanostructures attractive as inter-esting possible solid-phase extraction (SPE) materials (sorbents) for treatment ofradioactive waste. Various types of bulk sorbents such as organic (Sureshkumar et al.2010), inorganic (Li et al. 2012), biosorbent (Ghasemi et al. 2011), composites (Gaoet al. 2010), and other carbon-based material (Starvin and Rao 2004) have beendeveloped for the recovery of radionuclides from aqueous systems. However, thelow sorption capacities or efficiencies of these materials have obviously restrictedtheir applications. Search for alternatives to conventional bulk adsorbents hasintrigued intense interest on use of nanomaterial-based adsorbents due to theirunique physicochemical properties as compared to their bulk equivalents. Nano-materials circumvent many limitations of bulk material-based adsorbents. Carbonnanostructure characteristics such as large surface area, high specificity, fast adsorp-tion kinetics, and ability to interact with different chemical species make themexcellent candidates for radionuclide separation.

Carbon Nanotubes (CNTs)

Carbon nanotube (CNT) is a well-known carbon nanostructure. CNTs can bedescribed as a graphite sheet rolled up into a nanoscale tube; they are typicallyseveral nanometers (nm) in diameter and both ends are normally capped byfullerene-like structures. There are two main types of CNTs: (i) single-walledCNTs and (ii) multi-walled CNTs named based on the number of graphene sheet.CNTs exhibit unique properties such as strong tensile strength, large elastic module,high heat conductivity and electrical conductibility, and large surface area (Belloniet al. 2009). Due to its extraordinary physicochemical properties, CNTs have beenwidely applied in various scientific areas such as electronic, power engineering,medical, and catalysis. With a continuously decreasing cost of production andrequired properties (large surface area, hollow structure, strong affinity for ionic


and organic species, chemical and radiation stability), its application has beenstretched to sorbent material for extraction of radionuclides. For example, pristineMWCNTs were investigated as a sorbent for europium (Tan et al. 2008), americium(III) (Wang et al. 2005), thorium (Chen et al. 2007a), uranium, and plutoniumradionuclides due to the strong complexation of sorbates on the MWCNT surfacein the aqueous and NaClO4 medium. Functional groups like –COOH, –C=O, and–OH are incorporated on the surface of MWCNTs by treating with strong oxidizingagents like conc.HNO3 or mixture of conc.HNO3 and conc.H2SO4 at elevatedtemperature (Chen et al. 2007b; Darmstadt et al. 2003). The oxidized MWCNTsoffer a more hydrophilicity and surface area. The oxidized MWCNTs show fastkinetics and a high sorption activity towards uranium, thorium, and plutoniumradionuclides in weakly acidic to weakly basic solutions (pH 3–8) (Myasoedovaet al. 2009). The pH of aqueous solution plays an important role; it influences boththe metal ion speciation and total surface charge on sorbents. The isoelectric point ofpristine and oxidized MWCNTs is ~7.1 and ~5, respectively (Wang et al. 2005). Theisoelectric point of MWCNTs (ξ-potential) depends on their preparation, pre-treatment methods, and different functionalities on MWCNTs. At pH less thanisoelectric point (ξ-potential), the surface of MWCNTs is positively charged due toprotonation reaction and negatively charged due to deprotonation at pH greater thanisoelectric point. The electrostatic repulsion between the positively charged surfaceof the sorbent and positively charged species of uranium/thorium/plutonium resultedin a very low sorption capacity at lower pH. In the pH range of 5–8, the transition inthe surface polarity of MWCNTs favored the sorption of hydrolyzed species ofuranium (UO2(OH)

+, (UO2)2(OH)22+, (UO2)3(OH)5

+, UO2(OH)2)/thorium (Th(OH)3+, Th(OH)2

2+, and Th(OH)4) and maximized the uptake capacity. But at higherpH values (pH > 8), the electrostatic repulsion between the negatively chargedMWCNTs surface and the negatively charged uranium species (UO2(CO3)2

2� andUO2(CO3)3

4�) resulted in a rapid decrease in sorption. In case of thorium at pH> 4,deposition of precipitated thorium complexes onto the MWCNTs resulted indecrease of sorption. Plutonium is known to take part in disproportionation, hydro-lysis, and polymerization reactions and form concurrently polymeric forms indifferent oxidation states (Walther 2008). The oxidized CNTs are capable of extra-cting both the polymeric Pu4+ and ionic forms (Pu3+, Pu4+, PuO2

+, PuO22+) from

weakly acidic to weakly basic solutions (Perevalov and Molochnikova 2009). Pusorption on oxidized MWCNTs follows the Pu4+ > PuO2

2+ > PuO2+ trend due to

their charges. Both the pristine- and acid-pretreated CNTs were found to be usefulfor actinide sorption; however their removal efficiency and selectivity towardsparticular radionuclides remain quite limited. In recent years, much attention hasbeen diverted towards making the MWCNTs highly selective and less hydrophobic.Due to higher hydrophobicity, the dispersion of MWCNTs in an aqueous medium isless, resulting in aggregation of MWCNTs due to van der Waals interaction and thushindering in the sorption process to a greater extent. To enhance the dispersibility inaqueous solution, sorption performance, and selectivity towards particular radionu-clides, CNTs have been functionalized without much alteration of their physical andchemical properties. Modification of CNTs is accomplished through covalent or


noncovalent functionalization. MWCNTs were easily dispersed in solution by plasma-induced grafting of carboxymethyl cellulose on MWCNT (MWCNT-g-CMC).MWCNT-g-CMC has much higher adsorption capacity for U(VI) compared to theraw MWCNT or oxidized MWCNT (Wang et al. 2009). Similarly, MWCNTs graftedwith chitosan (MWCNT-g-CS) showed high sorption performance for uranium (Shaoet al. 2010).

MWCNTs functionalized with organic ligands are shown to be more selectivethan pristine and oxidized MWCNTs. Specifically, the CNTs are functionalized withamide- and phosphorus-based organic ligands used in the liquid extraction. Thesorption behavior of functionalized MWCNTs depends on the functionality and themechanism adopted for the sorption. Sorption performances of diglycolic acid(DGA)-functionalized MWCNTs (DGA-MWCNTs) (Deb et al. 2012, 2013) havebeen evaluated for U(VI) and Th(IV) in aqueous solution. The sorption behavior ofU(VI) and Th(IV) on DGA-MWCNTs is highly influenced by pH of the solution.Effective sorption of U(VI) and Th(IV) on DGA-MWCNTs was found to be atpH 5–7 and 4, respectively. DGA-MWCNTs have been investigated for U(VI), Am(III), and Pu(IV) from nitric acidic medium (Sengupta et al. 2017a; Deb et al. 2012).The sorption of U(VI), Am(III), and Pu(IV) increases gradually with nitric acidconcentration and reaches maximum at 3–5 M HNO3. Further increase in acidconcentration does not alter much the sorption. At acidic condition (pH < 2),DGA group in MWCNTs remains unionized and acts as neutral chelating groupwhich can bind effectively with neutral metal nitrate species. The increase information of M(NO3)n (Pu(NO3)4, Am(NO3)3, UO2(NO3)2) species with nitricacid concentration effectively coordinates with DGA group; thereby increase insorption was observed:

Mnþ þ nNO3� $ M NO3ð Þn M ¼ Am3þ=Pu4þ=UO2

2þ and n ¼ charge� �

(17)

M NO3ð Þn þ DGA-MWCNTs $ M NO3ð Þn:DGA-MWCNTs (18)

Sorption of U(VI), Pu(VI), Am(III), and Pu(IV) on dihexylamide (DHA)-functionalized MWCNTs (DHA-MWCNTs) also found to be increasing with nitricacid concentration (Gupta et al. 2017b). However, DHA-MWCNTs have shown lowsorption performance for UO2

2+ and Th4+ compared to DGA-MWCNTs. Sorption ofPu4+ or Am3+ on poly(amidoamine) (PAMAM) dendrimer of generation 1 and2 functionalized MWCNTs (MWCNT-PAMAMG1 and MWCNT-PAMAMG2)was studied (Kumar et al. 2017). PAMAM-functionalized MWCNTs (MWCNT-PAMAMG1 and MWCNT-PAMAMG2) have shown much higher sorption capacityfor Pu4+ and Am3+ compared to DGA-MWCNT or DHA-MWCNT sorbents.PAMAM dendrimer has more number of amide and amine functional groups andthey increase with dendrimer generation MWCNT-PAMAMG(n = 1,2. . .). Theseamide and amine functional group of PAMAM forms strong complex with f-blockelement leads to have higher sorption capacity than other functionalized MWCNTs.

The amidoamine-functionalized MWCNTs (MWCNT-AA) were also employedfor the sorption of Pu4+,NpO2

+, PuO22+, and NpO2

2+ in an acidic aqueous medium


(Sengupta et al. 2017b). In aqueous medium neptunium exists as NpO2+ without

disproportion with lower chemical potential making it very difficult to separate.However, the sorption performance of MWCNT-AA for Np was found ten timeshigher than ligand-impregnated MWCNTs making it an efficient solid-phase sorbentfor the separation of Np from nuclear waste solution. The sorption of neptuniumspecies (NpO2

+ and NpO22+) increases with increase in aqueous nitric acid concen-

tration and remains almost constant above 3M HNO3 concentration. However, anexponential increase in sorption of Pu4+ and PuO2

2+ on MWCNT-AAwas observedwith aqueous nitric acid concentration (Sengupta et al. 2017b).

Ligand-Impregnated MWCNTs

Ligands which showed high distribution coefficient along with selectivity for aparticular oxidation state of actinides in liquid–liquid extraction have been impreg-nated on MWCNTs. Various phosphorus-based ligands like diphenyl(dibutylcarbamoylmethyl) phosphine oxide (CMPO), tri-n-octylphosphine oxide(TOPO), tri-n-butylphosphate (TBP) [Murali and Mathur 2001; Yaftian et al. 2003],nitrogen-based ligands like N,N‘-dimethyl-N,N‘-dioctylhexylethoxymalonamide(DMDOHEMA) (Serrano-Purroy et al. 2005), and task-specific ionic liquids liketetra alkyl ammonium hydrogen phthalate (Gupta et al. 2017a) and phosphoniumionic liquid Cyphos IL-101 (Quinn et al. 2013) were shown distribution ratio andselectivity towards actinides in the presence of various other fission products. Hence,these ligands were impregnated on the MWCNTs by stirring a suspension ofMWCNTs and a ligand directly in nitric acid solution. Impregnation is mostly acombination of pore filling and surface adsorption. The extractant gradually fills thepore space starting with the smallest pores and moving up to pores of about 10 nm andthen surface adsorption becomes the dominant force. Interaction between the extra-ctant and support is usually quite weak, consisting of only the attractive forces betweenalkyl chains and/or aromatic rings of the ligand and those of the support.

The sorption of Np(V), U(VI), Pu(IV), Am(III), and Eu(III) onto CMPO-, TOPO-, and Cyphos IL-101-impregnated MWCNTs (Taunit) had shown some promisingresults in nitric acid solution (Zakharchenko et al. 2012). Taunit-TBP and Taunit-DMDOHEMA sorbents are also useful for the separation of plutonium from acidicnuclear waste solution (Zakharchenko et al. 2012).

Graphene Oxide (GO)

Graphene oxide (GO) is one of the most important graphene derivatives. It hasunique structure and exceptional physicochemical properties. The remarkable prop-erties of graphene include high Young’s modulus, fracture strength, thermal con-ductivity, mobility of charge carriers, specific surface area (theoretical calculatedvalue of 2630 m2 g�1), high chemical stability, large pore volume structure, andfascinating transport phenomena (Chen et al. 2012). GO has many oxygen-


containing functional groups in the form of epoxy, hydroxyl, and carboxyl groups onits basal plane and edges. These oxygen-containing groups can bind metal ions andorganic pollutants through coordination, electrostatic interaction, hydrogen bonding,etc., which ensures its potential application in environmental remediation(Nupearachchia et al. 2017).

GO is hydrophilic in nature due to existence of many oxygen-containing func-tional groups. In addition, the delocalized π electron systems of graphene layer asLewis base can form electron donor–acceptor complexes with radionuclides asLewis acid through Lewis acid–base interaction. The U(VI), Am(III), Sr(II), andPu(IV) radionuclides were removed using GO sorbent (Romanchuk et al. 2013). GOwas shown as an efficient sorbent for Th(IV) due to large surface area and sufficientexposure of active sites of abundant oxygen-containing functional groups with Th(IV) (Bai et al. 2014). Removal of U(VI) ions from an aqueous system usingfew-layered GO nanosheets was reported. However, the aggregation and accumula-tion of few-layered GO nanosheets resulted in limited sorption of U(VI) which wasobserved in the aqueous systems (Zhao et al. 2012). GO-based materials are easier toprocess and easily functionalized. Therefore, polymers including polyaniline, poly-acrylamide, cyclodextrins, chitosan, and amidoxime were grafted onto GOs tointroduce various functional groups and to enhance their dispersibility in solutionsand thereby improve the removal ability of some radionuclides (Wang et al. 2016).Such approaches, which added functionality to groups already present on the GOsurfaces, make GO a more versatile sorbent (Dreyer et al. 2010). Cyclodextrin-modified GO nanosheets (CD/GO) were synthesized using an in situ polymerizationmethod (Song et al. 2014). The mutual effects on the simultaneous removal of U(VI) and humic acid from an aqueous system by CD/GO were investigated. Thesorption of U(VI) and humic acid (HA) sorption on CD/GO highly depend on pHand ionic strength. The presence of HA enhanced U(VI) sorption at low pH andreduced U(VI) sorption at high pH, while the presence of U(VI) enhanced HAsorption.

Magnetic Nanoparticles for Sorption

Magnetite nanoparticles (MNPs) have received much attention recently to removeradioactive metal ions due to the easy separation of sorbent from solution with theaid of magnetic field. The bisphosphonate-modified magnetite Fe3O4 nanoparticleswere used to remove U(VI) from blood at pH 7.0 (Wang et al. 2006). However, theFe3O4 nanoparticles are unstable at acidic condition, which limits their applicationfor nuclide sorption from acid nuclear wastewater. Aggregation and decompositionof Fe3O4 nanoparticles under extreme pH conditions hinder the separation process ofactinides. This situation can be circumvented by designing core-shell type of nano-materials (Kaur et al. 2013a). The magnetic core material is protected by a shellmade up of nonmagnetic materials such as polymer, silica, carbon, and gold. Thissort of core-shell dimension reduces aggregation as well as improves extractionefficiency. For instance, silica-coated magnetic nanoparticles were developed for the


https://www.ncbi.nlm.nih.gov/pubmed/?term=Wang%20L%5BAuthor%5D&cauthor=true&cauthor_uid=17031939

separation of actinides from acidic nuclear waste, in which the Fe2O3 nanoparticleswere coated with the silica, followed by covalent attachment of the actinide-specificchelators. The silica-coated MNPs are stable even in 1 M HCl solution, and showenhanced actinide separation efficiency compared to the uncoated counterparts (Hanet al. 2010).

The affinity of the magnetic nanoparticle towards a particular metal ion isenhanced by functionalizing it with -thiol, -amine, and -carboxyl functional groups.These functional groups provide active sites for the capture of metal ions which getcomplexed by chelate formation, ion-exchange process, or electrostatic interactions.Therefore, surface-functionalized Fe3O4 magnetic nanoparticles are preferred overpristine Fe3O4 nanoparticles for the separation of metal ions (Singh et al. 2013). Theselectivity of the chelators towards particular metal ion can be tailored by selectingligands with oxygen, nitrogen, and sulfur donor atoms. This is done based on hard-soft nature of donor atoms and metal ions. Generally hard and soft metal ions formstrong complexes with hard and soft ligands, respectively (Kaur et al. 2013a). Thisconcept is exploited for the selective extraction of metal ion using ligands.Diethylene triamine penta acetic acid (DTPA) chelator covalently tethered tosilica-coated magnetic nanoparticles was found to be efficient for the recovery ofAm(III), Np(V), Pu(IV), and U(VI) under acidic conditions (Kaur et al. 2013b). N,N,N0,N0-tetraoctyl diglycolamide (TODGA)- and bis(2-ethylhexy)phosphoric acid(HDEHP)-coated superparamagnetic Fe3O4 nanoparticles (NP) were investigatedfor the extraction of Am(III) and Pu(IV) from the solution of 3–4 M HNO3 (Ojhaet al. 2017). TODGA-coated magnetic nanoparticles extracted both Am(III) and Pu(IV) whereas HDEHP-coated magnetic nanoparticles selectively extracted Pu(IV).This sort of magnetic nanoparticle-extractant framework can be effectively utilizedfor the pre-concentration of metal ions followed by their easy separation.

Nanocomposites

Nanocomposite is a multiphase solid-state material with at least one of the phases innano-dimension which might be zero-dimensional (quantum dots) orone-dimensional (nano rods) or two-dimensional (nanosheets) (Okpala 2013).Also, the multiphase materials with their phases separated by a distance of nanoscaleare termed as nanocomposites. The idea of incorporating nanoscale materials is topromote the synergism among the constituents that emerges into novel propertiescapable of meeting desired specifications. Multifunctional capability of nano-composites arises due to the benefit of nanoscale properties, huge interphase zone,and chemical functionalization. Different types of metal-, carbon-, ceramic-, andpolymer-based nanocomposites possess a variety of properties (Camargo et al.2009). The properties of nanocomposites can be modified by choosing the rightparameters. These parameters include nanoscale dimension, loading, degree ofdispersion, shape, orientation, matrix material, and interaction among the constitu-ents. These parameters are meticulously selected to prepare materials with desiredflexibility, ductility, strength, conductivity, compatibility, magnetism, and surface


activity (Camargo et al. 2009). Nanocomposites are used in catalysts, sensors, opto-electronic devices, magnetic devices, food packaging, fuel cells, dentistry, construc-tion, coating, medicine, and automobiles (Okpala 2014). Nanocomposites also foundpotential applications in the separation of heavy metal ions. The prospective ofnanocomposites over the separation of actinides has emerged recent times.

CNT nanocomposite was prepared by using hydroxylated fullerene (C60(OH)n),carboxylated fullerene (C60(C(COOH)2)n), and oxidized-MWCNTs (Wang et al.2013). The presence of hydroxylated fullerene (C60(OH)n) and carboxylated fuller-ene (C60(C(COOH)2)n) in MWCNTs increased the sorption of Th4+ onto oxidizedMWNCTs at pH < 4. However, the sorption of Th4+ lowered at pH > 4 with theincreasing concentration of C60(OH)n or C60(C(COOH)2)n (Wang et al. 2013). Thisstriking behavior was due to the good sorption affinity of C60(OH)n or C60(C(COOH)2)n onto oxidized MWCNTs by strong π–π electron donor–acceptor inter-actions between the flat surfaces of both aromatic C60(OH)n/C60(C(COOH)2)n andoxidized MWCNTs. Thus, the increasing concentration of fullerene created a com-petitive pathway for the sorption of Th4+. Polyvinyl alcohol MWCNT(PVA-MWCNT) composite was evaluated for the removal of U(VI) (Abdeen andAkl 2015). The maximum sorption of U(VI) on PVA-MWCNTs was observed atpH 3 and the adsorbed uranium could be desorbed by using 0.1 M EDTA.

Various GO-based composites were synthesized and used for environmentalremediation. GO-supported sepiolite (GO@sepiolite) composites were synthesizedapplied for the removal of U(VI) (Cheng et al. 2013a). GO-activated carbon felt(GO–ACF) composites were prepared by an electrophoretic deposition and subse-quent thermal annealing(Chen et al. 2013b). GO–ACF showed high sorption capac-ity for U(VI) due to presence of carboxyl functional groups in GO–ACF.

The sorption of U(VI) on GO-supported polyaniline (PANI@GO) nano-composites was investigated. The PANI has a strong affinity for radionuclides andheavy metal ions due to the large number of amine and imine functional groups;these functional groups can form very strong complexes with radionuclides on thenanocomposite surfaces (Sun et al. 2013). The chemical binding of radionuclideswith the nitrogen-containing functional groups is much stronger than that of radio-nuclides with the oxygen-containing functional groups. The sorption of U(VI) ionson GO-polypyrrole (GO/PPy) composites was much higher than that of U(VI) oneither GO or PPy in an aqueous system (Hu et al. 2014). The highly selectivesorption capacity towards U(VI) was observed due to the strong coordinationbetween the U(VI) ions and nitrogen donor atoms on the GO/PPy composites.GO/PPy composites could be regenerated and reused. Hierarchical three-dimensional composite (layered double hydroxide/graphene) was obtained via insitu growth of layered double-hydroxide (LDH) nanosheet arrays onto graphenesheets and sorption of U(VI) was performed from aqueous media (Tan et al. 2015c).Poly(amidoxime)-reduced GO (PAO/RGO) composites synthesized by in situ poly-merization showed excellent sorption capability for U(VI) (Chen et al. 2014). Thesorption capacity of PAO/RGO composites for U(VI) showed highest sorptioncapacity (872 mg g � 1) compared to other GO-based composites.


Magnetic Nanocomposites

Even though nanomaterials possess extraordinary features, it is practically compli-cated to remove the suspended nanoparticles from the solution. This problem can betackled by introducing magnetic nanocomposites that can be easily removed by theapplication of magnetic field. Design of chelator-tethered magnetic nanomaterialsrevolutionized the process of separation of metal ions. They work on the simpleprinciple of magnetism but they are versatile with high efficiency in the separation.The main advantage of using magnetic materials is the easy separation of the sorbentfrom the solution after the extraction of metal ions (Teja and Koh 2009).

The MWCNT nanocomposites were prepared by incorporation of Fe2O3 orCoFe2O4 magnetic particle. These magnetic Fe2O3-MWCNT (Liu et al. 2014) andCoFe2O4-MWCNT (Tan et al. 2015a) composite were evaluated for the sorption ofTh(IV) and U(VI) radionuclides, respectively. The removal of U(VI) using nano-scale zero-valent iron (nZVI) and reduced graphene oxide-supported nanoscale zero-valent iron (nZVI/RGO) was studied (Sun et al. 2014). The U(VI) species such asUO2

2+, UO2OH+, and (UO2)3(OH)5

+) were efficiently removed in the weak aqueousacid medium. GOs have high dispersion properties in aqueous solution due to theirhydrophilic nature; thereby it is difficult to separate GOs from aqueous solution byusing the traditional separation techniques after the GOs are applied as adsorbents inthe sorption process. The magnetic Fe3O4/GO composite was prepared from naturalflake graphite and employed as adsorbent to remove U(VI) ions (Zong et al. 2013).After U(VI) sorption, magnetic Fe3O4/GO composite was easily removed fromaqueous solution due to magnetic separation. Amidoximated magnetite/ GO(AOMGO) composites were also synthesized and applied to U(VI) sorption, andhad a more sorption capacity compared to Fe3O4/GO composite (Zhao et al. 2013).MnO2–Fe3O4–RGO was successful for the removal of U(VI) (Tan et al. 2015d). Inaddition, U(VI)-loaded MnO2–Fe3O4–RGO was efficiently regenerated and reused.The CoFe2O4-reduced graphene oxide nanocomposite was used for the extraction ofuranium (Tan et al. 2015b). The superparamagnetic GO-Fe3O4 nanocomposite wassynthesized and evaluated for the uptake of Am3+, UO2

2+, Th4+, and Pu4+ in mildlyacidic solutions (Gadly et al. 2017). The uptake of tetravalent metal ion reduced withincrease in hydrolysis due to the reduction in the effective nuclear charge. Thesecomposites are recyclable and therefore can be reused. Magnetic ferberite-GOnanocomposite was utilized for the extraction of U(VI) from aqueous solution(El-Maghrabi et al. 2017). This composite possesses high recycling capacity andthe morphology remained unaltered after several stages of reuse.

Fe3O4-encapsulated ZIF-8 nanocomposite was used for the separation of U(VI)from Ln(III) under various experimental conditions (Min et al. 2017). The fastkinetics and high U(VI) uptake capacity were observed for this kind of metal organicframework composite. Attapulgite-Fe2O3 nanocomposites prepared by chemicalroute were used for the separation of U(VI) from its aqueous solution (Chen et al.2013). At higher pH, sorption of U(VI) occurs through inner sphere surface com-plexation and at low pH sorption of U(VI) follows outer-sphere surfacecomplexation.


Mesoporous Materials for Sorption

Besides nanomaterials, mesoporous materials such as activated carbon (Carboniet al. 2013), silica (Vivero-Escoto et al. 2013), and polymeric membrane with poresize in the range of nano dimensions (size between 2 and 100 nm) also gainedprominence in the separation of actinides. Ordered mesoporous carbon and silicamaterials (MCM-Mobil Composition of Matter-41,48 and SBA-15-Santa BarbaraAmorphous-15) are novel families of the fascinating porous solids, which have theadvantages of large surface area, well-defined pore size, excellent mechanicalresistance, non-swelling, exceptional chemical stability and radiation tolerance, aswell as extraordinarily wide possibilities of functionalization (Darmstadt et al. 2003;Lee et al. 2011). These advantages make the ordered mesoporous carbon and silicacompounds attractive for nuclear waste disposal.

The U(VI) was entrapped using MCM-41 and MCM-48 molecular sieves basedon direct template ion exchange (Vidya et al. 2001, 2004). It was found that theentrapment of U(VI) was facilitated by the large pore size and the high surfactantcontent in the as-synthesized host materials. However, these sorbents normally showpoor selectivity and slow sorption kinetics for U(VI). To promote the sorptionselectivity, and achieve higher sorption capacity and faster sorption kinetics,functionalized mesoporous materials were synthesized. 4-Acetophenone oxime-functionalized ordered mesoporous carbon CMK-5 (Oxime-CMK-5) was developedfor U(VI) sorption (Li et al. 2011). The U(VI) sorption by Oxime-CMK-5 was foundto be rapid and pH dependent (Fryxell et al. 2005; Lin et al. 2005). It was found thatthe composite can be reused without considerable loss in sorption capacity. The5-nitro-2-furaldehyde-modified mesoporous silica (MCM-41) was used for extrac-tion of U(VI) (Yousefi et al. 2009). The sorbent exhibits good stability, reusability,high sorption capacity, and fast rate of equilibrium for sorption/desorption of U(VI).Recently, the phosphonate-functionalized MCM-41(NP10) (Yuan et al. 2011) andamino-functionalized SBA-15 (APSS) (Liu et al. 2012) were synthesized byco-condensation and grafting method, respectively. These synthesized materialswere used as sorbents for the removal of U(VI) from aqueous solution. These newsorbents offered large sorption capacity and possess ultrafast sorption kinetics.Furthermore, these silica materials show a desirable selectivity for U(VI) ions overa wide range of competing metal ions.

Ordered mesoporous carbon (OMC) was synthesized and investigated for thesorption and desorption of Pu(VI) and Eu(III) (Parsons-Moss et al. 2014). Extractionof uranium in aqueous media by nanocomposite of two-dimensional layered doublehydroxide with silica nanoparticles revealed that sorption of U(VI) is through inner-sphere surface complexation and electrostatic interaction (Yang et al. 2017a). Lay-ered double-hydroxide carbon dot composite was used for the extraction of U(VI) and Am(III) from aqueous solutions (Yao et al. 2018). The dispersion of carbondots into layered double hydroxides prevents agglomeration of carbon dots andthereby increased surface area and extraction efficiency of actinides.


Self-assembled monolayer on mesoporous support (SAMMS) was prepared byassembling monolayers of molecules in the mesoporous support and functionalizingthe free end of the monolayers (Fryxell et al. 1999). The large surface area of thesefunctionalized mesoporous materials provides more active sites for the adsorption ofmetal ions. Moreover, the high porosity enhances the diffusion of metal ions andsorption kinetics. This kind of ceramic support is stable under extreme acidic,corrosive, and oxidizing environment. Glycinyl-urea-, carboxylate-, andphosphonate-functionalized SAMMS were investigated for the extraction of U(VI), Pu(IV), and Am(III) from aqueous solutions containing several foreign metalions (Fryxell et al. 2005). Isomers of hydroxypyridinone-functionalized SAMMSwere investigated to understand the extraction behavior of Pu(IV), Np(V), and U(VI) (Lin et al. 2005). The multiple ligand–metal interactions and intramolecularhydrogen bonding enhanced the extraction of actinides.

The surface ion-imprinted mesoporous silica was investigated for the separationof U(VI) from highly acidic solutions (Yang et al. 2017b). The sorbent was synthe-sized using U(VI) (uranyl) as the template and diethylphosphatoethyltriethoxysilaneas functional group. The surface ion-imprinted mesoporous silica have nano-sizedthree-dimensional cavity that is specific for U(VI). Using this sorbent the fastkinetics and higher selectivity for U(VI) were observed when compared to non-ion-imprinted similar mesoporous materials. The material is also resistant to radia-tion such that the performance did not deteriorate after five consecutive usages of theion-imprinted polymer.

Mesoporous Membrane for Sorption

Removal of radionuclides from aqueous waste by membrane separation usingnanofiltration (NF)/reverse osmosis (RO)/ultrafiltration (UF) membrane has becomeone of the emerging technologies with a rapid growth. It has drawn the attention ofresearchers in the field of separation technology with its better performance com-pared to conventional separation processes. In these filtration methods, the feedmixture is pressurized to pass through nano/micro-sized pores of the membrane andthen separated into a retentate (part of the feed that does not pass through themembrane) and a permeate (part of the feed that passes through the membrane).During this process, the actinides are concentrated in the retentate by sieving andadsorption.

Reverse Osmosis (RO)

Reverse osmosis (RO) can effectively remove nearly all inorganic contaminantsfrom water. RO process was applied for the treatment of effluents containinguranium. In this process, uranium content in permeate is reduced to less than 1 mgL�1, where the removal of uranium is 99.5% (Hsiue et al. 1989; Prabhakar et al.1992). RO process was conducted to remove uranium, technetium, tritium,


strontium, and cesium from liquid effluents and groundwater (Garrett 1990; GamalKhedr 2013). It was demonstrated that all of the above contaminants could beremoved to less than the regulated limits, except tritium. Recently, RO was used inthe framework of the Fukushima-Daiichi accident to treat contaminated seawater(Fournel et al. 2012).

Nanofiltration (NF)

Nanofiltration membranes have a nominal pore size of approximately 0.001 micronsand a molecular weight cutoff (MWCO) of 200–1000 daltons (Da) (Yong Du et al.2016). NF membranes are very commendable for retention of multivalent cationsand organic solutes. NF membranes are generally manufactured from celluloseacetate or polyamide materials. The polyamide-based NF membrane was used fortreatment of radioactive effluents containing specific activity levels in 10 kBq/L. Itwas observed that the level of radioactive contaminants could be reduced from10 kBq /L to few Bq/L. Studies on selective removal of uranium from aqueoussolution using nanofiltration membranes showed relatively high selectivity, despitethe high concentration of other divalent and monovalent cations (Reguillon et al.2008; Raff and Wilken 1999).

Ultrafiltration (UF) Combined with Sorption/Precipitation/Complexation

Ultrafiltration has a pore size of approximately 0.002–0.1 microns, an MWCO ofapproximately 10,000–100,000 daltons. UF membranes are constructed from a widevariety of materials, including cellulose acetate, polyvinylidene fluoride, polyacry-lonitrile, polypropylene, polysulfone, polyethersulfone, or other polymers. Each ofthese materials has different properties with respect to the surface charge, degree ofhydrophobicity, strength, flexibility, pH, and oxidant tolerance. UF membranes havelarge pore size than NF membrane and they are not effective in retaining radioactivemetal ions. Hence, in order to retain the radioactive ions, size enhancement ofradioactive species by means of sorption/precipitation/complexation is done priorto UF. UF experiment was conducted with 239,240Pu and pseudo-colloids of Fe inwater having dissolved organic carbon (DOC) in the range of 10–60 mg/L. Thepresence of DOC increases the colloidal size of Fe and acts as a matrix for highsorption of plutonium (Singhal et al. 2009). Studies on ferric hydroxide precipitationemployed for the treatment of low-level waste (LLW) containing 241Am showed>99.9% removal of 241Am (Gao et al. 2004). The recovery of 241Am was alsoaccomplished using sodium dodecyl sulfate (SDS) micellar solution at pH > 3(Kedari et al. 2009).


Nanopolymer/Dendrimer-Assisted Ultrafiltration

Polymer-assisted ultrafiltration (PAUF) is a technology under development forselective concentration and recovery of valuable metal ions from wastewater.PAUF is a combination of ion exchange or chelating property of a functionalizedwater-soluble polymer and sieving power of an ultrafiltration membrane. As themolecular weight of metal polymer complex is larger than the molecular weightcutoff of the membrane, they are concentrated in the retentate and permeate is freefrom metal ions. The polymers could be regenerated by changing pH of the solution.

With the recent significant advantages in nanoscience and nanotechnology, var-ious nanostructured polymers/dendrimers were devised for a wide range of advancedapplications. Examples include the use of polymer nanoparticles as drug deliverydevices, polymer nanofibers as conducting wires, and polymer thin film as optoelec-tronic devices. Like most other nanomaterials, nanostructured polymers also possessinteresting mechanical, electronic, optical, and even magnetic properties that aredifferent from those of the bulk materials, depending on their size, shape, andcomposition.

Dendrimers are a new class of polymeric nanomaterials (after linear, cross-linked,and branched polymers) having unique properties such as high degree of branchingunit, high density of surface functional group, and narrow molecular weight distri-bution. They are spheroid or globular nanostructures that are precisely engineered tocarry molecule encapsulated in their interior void space or attached to the surface.Size, shape, and reactivity are determined by generation (shell) and chemicalcomposition of the core, interior branching, and surface functionalities. Dendrimersare constructed through a set of repeating chemical synthesis procedures that buildup from the molecular level to the nanoscale region under conditions that are easilyperformed in a standard organic chemistry laboratory. The dendrimer diameterincreases linearly whereas the number of surface groups increases geometrically.Dendrimers are very uniform with extremely low polydispersities, and are com-monly created with dimensions incrementally grown in approximately nanometersteps from 1 to over 10 nm. The control over size, shape, and surface functionalitymakes dendrimers one of the “smartest” or customizable nanotechnologies commer-cially available (Tomalia et al. 1985). Figure 9 provides a schematic of a generalizeddendrimer structure.

Among the uses of dendrimer the important one is that they can be used toselectively bind to, or react with, a particular element, ion, or molecule of choice.They are very large, yet soluble, macroligands, and well-defined sizes and shapescan be made, with hundreds or even thousands of complexing sites and reactivechain ends. They can also be covalently linked to each other or to other macromol-ecules to form supramolecular assemblies of various size, shape, and topologies.Dendritic macromolecules can also be functionalized with surface group that makethem soluble in selected solvents or bind to selected surface (Diallo et al. 2009).Currently, dendrimers are under investigation as metal-sequestering agents for wasteremediation technologies (Diallo 2006). The highly branched molecules with numer-ous functional groups can be formulated to provide the required properties of water


solubility, selectivity, high loading capacity, and so on. The use of poly(amido)amine(PAMAM) and poly(propylene)imine (PPI) dendrimers and their derivatives incombination with ultrafiltration technique have shown potential applications inremoval of metal ions such as Cu(II), Ni(II), Co(II), Pd(II), Pt(II), Zn(II), Fe(III),Ag(I), Au(I), and U(VI) from dilute aqueous solution (Rether and Schuster 2003;Diallo et al. 2008; Arkas et al. 2003).

Removal of U(VI) and Th(IV) ions from aqueous solution by ultrafiltration(UF) and dendrimer-assisted ultrafiltration (DAUF) was investigated (Ilaiyarajaet al. 2014). Regenerated cellulose acetate ultrafiltration membrane was used duringultrafiltration and PAMAM dendrimer chelating agent used as a complexing agent inDAUF. In UF process, removal of U(VI) and Th(IV) metal ions is observed to bebased on adsorption/mass deposition on membrane. PAMAM dendrimer contains alarge number of nitrogen and oxygen atoms as amino and amide groups on branchesand terminal surface showing strong chelating action towards actinides. In DAUF,the water-soluble PAMAM dendrimer effectively forms complex with hydrolyzed U(VI) and Th(IV) species in aqueous solution and gets concentrated in the retentate. Inboth UF and PAMAM-DAUF, the U(VI) and Th(IV) were more than 95% in theweakly acid medium (pH 5–6). However, in UF, the membrane fouling (plugging ofmembrane pores) was observed on repeated use. PAMAM-DAUF was shown to beeffective in the removal of U(VI) and Th(IV) at pH > 4.

Conclusion

Separation of radionuclides from aqueous streams is currently one of the mostsignificant and challenging problems. The main challenge involved is the separationof various long-lived actinide species that exist simultaneously in aqueous streamsand their effective separation in the presence of a large excess of competing ionic

Fig. 9 Schematic of a generalized 2D and 3D view of dendrimer structure (Kukowska-Latalloet al. 1996)


species. Therefore, materials to be used for separation of actinides species arerequired to be more efficient and specific enough. This chapter revealed the factthat there has been a large increase in utilization of nanomaterials over the last fewdecades. All above-mentioned studies have dealt with separation of radionuclidesusing nanomaterials. These studies emphasize that nanomaterials show severalinteresting aspects for application in nuclear waste disposal and environmentalremediation. Due to nanostructural and surface features, there is a tendency forvarious interactions (hydrophobic, dipole, π–π interactions, formation of hydrogenand other bonds), good sorption capacity, kinetic properties, and thermal andchemical stabilities. Moreover, their modified or functionalized forms are promisingcandidates for specific actinide separation. The magnetic nanoparticles are incorpo-rated into sorbent to facilitate the easy removal of suspended sorbent from aqueoussolution after the extraction of metal ions. This approach promises industrial levelscaling up of separation process. However, some nanomaterials are not quite stableunder the conditions of ionizing radiation and cannot perform for relatively longerperiod of time in complex chemical environments. Furthermore, the issues onhazardous nature of nanomaterials itself to environment also need to be evaluatedand clarified. The dendrimer-assisted ultrafiltration possesses potential application indecontamination of radioactive wastes. However, the synthesis of dendrimerinvolves a number of steps that are time consuming and expensive. Hence, attentionshall be provided for reducing the number of steps involved in dendrimer synthesis.In these regards, full realization of a true potential of nanomaterials for radionuclideseparation requires further studies concentrating on developing highly efficient,selective, radiation-resistant, renewable, economic, and environment-friendly nano-materials. Therefore this chapter provides a deeper insight into speciation andseparation of actinide species in aqueous stream.

Cross-references

▶Advanced Treatment Technologies▶Design of Conceptual Repository of Nuclear Waste in Granitic Rock for Manage-

ment of Radioactive Waste in India▶Environmental Treatment Technologies-Adsorption▶Management of Radioactive Wastes▶Nanomembranes for Environment▶ Surface Functionalization of Magnetic Nanoparticles for Water and Wastewater

Remediation

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